Citation
Biodegradation of selected phenolic compounds in a simulated sandy surficial Florida aquifer

Material Information

Title:
Biodegradation of selected phenolic compounds in a simulated sandy surficial Florida aquifer
Creator:
Lin, Chen Hsin, 1953- ( Dissertant )
Miller, W. Lamar ( Thesis advisor )
Bolch, Emmet W. ( Reviewer )
Chadik, Paul A. ( Reviewer )
Delfino, Joseph J. ( Reviewer )
Spangler, Daniel P. ( Reviewer )
Place of Publication:
Gainesville, Fla.
Publisher:
University of Florida
Publication Date:
Copyright Date:
1988
Language:
English
Physical Description:
vii, 182 leaves : ill. ; 28 cm.

Subjects

Subjects / Keywords:
Adsorption ( jstor )
Azides ( jstor )
Bacteria ( jstor )
Biodegradation ( jstor )
Carbon ( jstor )
Groundwater ( jstor )
Phenols ( jstor )
Sodium ( jstor )
Soils ( jstor )
Sorption ( jstor )
Aquifers ( lcsh )
Dissertations, Academic -- Environmental Engineering Sciences -- UF
Environmental Engineering Sciences thesis Ph. D
Pentachlorophenol -- Biodegradation ( lcsh )
Phenols -- Biodegradaton ( lcsh )
Genre:
bibliography ( marcgt )
theses ( marcgt )
non-fiction ( marcgt )

Notes

Abstract:
Phenolic compounds are commonly found contaminants in groundwater systems. In this research the sorption and biodegradation of phenol, 2 , 4-d ichlorophenol (2,4-DCP) and pentachlorophenol (PCP) were investigated. The soil materials used were characterized as fine grained sands with negligible organic carbon contents. Freundlich sorption coefficients of 0.0158 for phenol and 0.0547 for 2,4-DCP were found. Pentachlorophenol was more strongly adsorbed with an adsorption coefficient of 1.12. In multi-compound systems competitive sorption was evident, and adsorption capacities were reduced by a margin ranging from 70% for phenol to 30% for both DCP and PCP. All three compounds exhibited nonlinear sorption behavior with a range of exponent values from 0.56 to 0.7. Desorption coefficients showed little difference from adsorption for phenol and 2,4-DCP, but were significantly different for PCP, indicating hysteresis of PCP sorptions. The retardation factors were 1.03 for phenol, 1.16 for 2,4- DCP and 2.26 for PCP. In batch biodegradation studies using indigenous soil bacteria phenol degraded quickly (t.,_ = 12 hours) and was completely destroyed within three days. 2,4-DCP was also completely degraded but had taken 23 days (t,/_ = 7 days) . PCP was resistant to biodegradation with an average halflife of 120 days. In multi-compound systems, phenol degradation rates dropped off to 0.4 day (t , = 1.7 days) but PCP degradation rates increased to 0.008 day (t . = 86 days) . Biodegradation rates in column studies were obviously greater than in batch experiments, with the rate increase for PCP degradation being especially noticeable (t , = 12 days) , because of larger bacterial populations and the dynamic flow conditions made the substrates more available to the bacteria. When controlled under an aerobic environment by the addition of hydrogen peroxide, all three phenolic compounds degraded fastest. under anoxic conditions both the microbial population buildup and the rate of phenolic compound degradation were slower but not by a wide margin. Bacterial growth in the columns did not reduce the hydraulic conductivity of the system, indicating the feasibility of applying in-situ biodegradation techniques to groundwater contamination problems.
Thesis:
Thesis (Ph. D.)--University of Florida, 1988.
Bibliography:
Includes bibliographical references.
Additional Physical Form:
Also available on World Wide Web
General Note:
Typescript.
General Note:
Vita.
Statement of Responsibility:
by Chen Hsin Lin.

Record Information

Source Institution:
University of Florida
Holding Location:
University of Florida
Rights Management:
Copyright [name of dissertation author]. Permission granted to the University of Florida to digitize, archive and distribute this item for non-profit research and educational purposes. Any reuse of this item in excess of fair use or other copyright exemptions requires permission of the copyright holder.
Resource Identifier:
024932052 ( AlephBibNum )
20117325 ( OCLC )
AFM5810 ( NOTIS )

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Full Text












BIODEGRADATION OF SELECTED PHENOLIC COMPOUNDS
IN A SIMULATED SANDY SURFICIAL FLORIDA AQUIFER





BY

CHEN HSIN LIN


A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA
IN PARTIAL FULFILLMENT OF THE REQUIREMENTS
FOR THE DEGREE OF DOCTOR OF PHILOSOPHY


UNIVERSITY OF FLORIDA


1988


7_ 1. 77
E J-


.. p C'
-_, L& cU-'nA L.J
















ACKNOWLEDGEMENTS


I would like to express my thanks to Dr. W. Lamar

Miller, the chairman of my supervisory committee, for his

support during three years of my study. Also, my sincere

thanks go to the rest of the committee members, Dr. W.

Emmett Bolch, Dr. Paul A. Chadik, Dr. Joseph J. Delfino, and

Dr. Daniel P. Spangler for their generous assistance and

thoughtful criticism. Special thanks go to Dr. Ben L.

Koopman for the use of his equipment, and to Mr. Bill Davis

for his assistance with high performance liquid

chromatography.

This work could not have been completed without the

love of my wife, Lily, and the support of my family.













TABLE OF CONTENTS


ACKNOWLEDGEMENTS...... .......... ....................... ii

LIST OF TABLES......................................... v

LIST OF FIGURES........................................ vii

ABSTRACT............................................... x

CHAPTERS

I INTRODUCTION..................................... 1

II OBJECTIVES....................................... 6

III LITERATURE REVIEW ................................ 7

3.1 Environment Significance of the Phenolic
Compounds ............................... 7
3.2 Sorption of the Phenolic Compounds.......... 9
3.3 Degradation of the Phenolic Compounds....... 16
3.3.1 Photolysis............................ 16
3.3.2 Oxidation ........................... 18
3.3.3 Hydrolysis. ........................... 19
3.3.4 Volatilization........................ 19
3.3.5 Biodegradation........... ............ 20
3.3.6 PCP Degradation Mechanisms........... 31
3.4 Summary..................................... 33

IV MATERIALS AND METHODS....... ................ ... 34

4.1 Materials.................................... 34
4.1.1 Soil.................................. 34
4.1.2 Chemicals ............................ 34
4.1.3 Contaminated Water................... 35
4.1.4 Microorganisms ..................... 35
4.2 Analytical Methods... ....................... 35
4.2.1 Chemical Concentration Determinations 35
4.2.2 Soil Characterization................ 37
4.2.3 Sludge Characterization.............. 40
4.2.4 Biological Activity Measurement...... 41
4.3 Experimental Design.......................... 42
4.3.1 Batch Sorption Studies................ 42
4.3.2 Column Sorption Studies............... 44
4.3.3 Batch Biodegradation Studies.......... 48
4.3.4 Column Biodegradation Studies........ 53










V RESULTS AND DISCUSSION ........................... 56

5.1 Soil Characterization. ....................... 56
5.2 Batch Sorption................................ 57
5.2.1 Single Compound Batch Adsorption...... 57
5.2.2 Mixed Compound Batch Adsorption....... 61
5.2.3 Batch Desorption...................... 64
5.3 Column Sorption............................... 72
5.4 Batch Biodegradation ........................ 75
5.4.1 Nutrient Requirement. ................. 76
5.4.2 Single Compound Biodegradation....... 76
5.4.3 Multiple Compounds Biodegradation.... 91
5.5 Column Biodegradation........................ 110
5.5.1 Column Biodegradation I.............. 110
5.5.2 Column Biodegradation II............. 115
5.5.3 Column Biodegradation III............. 119
5.6 Hydraulic Conductivity....................... 124

VI SUMMARY AND CONCLUSIONS........ .................. 125

6.1 Summary... .................................. 125
6.1.1 Sorption.............................. 125
6.1.2 Batch Biodegradation.................. 126
6.1.3 Column Biodegradation ................ 128
6.1.4 Hydraulic Conductivity Tests......... 128
6.2 Conclusions... .............................. 129

APPENDICES

A BATCH SORPTION DATA............................... 132

B COLUMN BREAKTHROUGH DATA......................... 146

C BATCH BIODEGRADATION DATA.......................... 148

D COLUMN BIODEGRADATION DATA....................... 166

E PROCEDURES TO CALCULATE K ...................... 170
oc
REFERENCES...... ........ ............................... 171

BIOGRAPHICAL SKETCH. ................................... 182












LIST OF TABLES


Table Page

3-1 Physical properties of the phenolic compounds.... 9

4-1 Experimental scheme for adsorption study......... 43

4-2 Experimental scheme for phenol biodegradation
and nutrient requirement studies............. 49

4-3 Experimental scheme for 2,4-DCP biodegradation
study......................................... 50

4-4 Experimental scheme for PCP biodegradation
and enzyme induction studies................. 51

4-5 Experimental scheme for mixture biodegradation
and co-degradation studies................... 52

4-6 Experimental scheme for PCP co-degradation in the
presence of phenol........................... 53

4-7 Experimental scheme for column study II
(co-degradation of PCP and phenol) ........... 55

5-1 Selected physical properties of the soil.......... 57

5-2 Adsorption regression parameters of phenolic
compounds in single-compound system on
plain soil.................................... 58

5-3 Adsorption regression parameters of phenolic
compounds in single-compound system on
soil with sludge.. ............................ 58

5-4 Calculated adsorption parameters of phenolic
compounds in single-compound system based on
organic carbon............................... 60

5-5 Adsorption regression parameters of phenolic
compounds in multi-compound system on
plain soil.................................... 61

5-6 Adsorption regression parameters of phenolic
compounds in multi-compound system on
soil with sludge.. ............................ 61

5-7 Calculated adsorption parameters of phenolic
compounds in multi-compound system based on
organic carbon................................ 62









5-8 Single-compound system to multi-compound system
ratios of Freundlich sorption coefficients
for phenolic compounds... .......... ............ 63

5-9 Desorption regression parameters of phenolic
compounds in single-compound systems ....... 64

5-10 Desorption regression parameters of phenolic
compounds in multi-compound systems ......... 71

5-11 Retardation factors of mixed phenolic compounds
calculated by various methods................ 75

5-12 Apparent biodegradation rate constants for phenol 77

5-13 Apparent biodegradation rate constants for
2,4-DCP...................................... 78

5-14 Apparent biodegradation rate constants for PCP... 82

5-15 Conservative estimations of the biodegradation
rate constants for PCP...................... 89

5-16 Apparent biodegradation rate constants for phenol
in multi-compound systems... ....... .......... 92

5-17 Apparent biodegradation rate constants for PCP
in multi-compound systems..... ................ 100

5-18 Conservative estimations of the biodegradation
rate constants for PCP in multi-compound
systems ........................................ 107

5-19 Apparent biodegradation rate constants for PCP
co-metabolized with phenol.................... 108

5-20 Conservative estimations of the biodegradation
rate constants for PCP co-metabolized with
phenol ....................................... 108

5-21 Column biodegradation study I results............. 110

5-22 Column biodegradation study II results............ 116

5-23 Column biodegradation study III results........... 119

5-24 DHA data for column degradation study III......... 120

5-25 Column hydraulic conductivity test results....... 124


vi










LIST OF FIGURES


Figure Page

3-1 Structural formulas for some common PCP
degradation products........................ 32

4-1 Experimental setup for hydraulic conductivity test 39

4-2 Experimental setup for conservative tracer test.. 46

4-3 Experimental setup for column degradation studies 47

5-1 Phenol sorption isotherms on plain soil........... 65

5-2 Phenol sorption isotherms on soil with sludge.... 66

5-3 2,4-DCP sorption isotherms on plain soil.......... 67

5-4 2,4-DCP sorption isotherms on soil with sludge... 68

5-5 PCP sorption isotherms on plain soil.............. 69

5-6 PCP sorption isotherms on soil with sludge........ 70

5-7 Column breakthrough curves for phenol, 2,4-DCP
and PCP ......................................... 74

5-8 Phenol degradation curves in single-compound
systems (initial concentration 5 ppm) ........ 79

5-9 Phenol degradation curves in single-compound
systems (initial concentration 1 ppm) ........ 80

5-10 2,4-DCP degradation curves in single-compound
systems (initial concentration 5 ppm) ........ 83

5-11 2,4-DCP degradation curves in single-compound
systems (initial concentration 1 ppm) ........ 84

5-12 PCP degradation curves in single-compound
systems (initial concentration 5 ppm) ........ 85

5-13 PCP degradation curves in single-compound
systems (initial concentration 1 ppm)........ 86

5-14 PCP degradation curves using bacteria which are
acclimated to phenol and 2,4-DCP
(initial concentration 5 ppm) ................ 87








5-15 PCP degradation curves using bacteria which are
acclimated to phenol and 2,4-DCP
(initial concentration 1 ppm) ................ 88

5-16 Phenol degradation curves in multi-compound
systems (initial concentration 5 ppm) ........ 93

5-17 Phenol degradation curves in multi-compound
systems (initial concentration 1 ppm) ........ 94

5-18 Phenol co-degradation curves in multi-compound
systems (initial concentration 5 ppm) ........ 95

5-19 Phenol co-degradation curves in multi-compound
systems (initial concentration 1 ppm) ........ 96

5-20 Phenol degradation curves in multi-compound
systems using acclimated bacteria (initial
concentration 5 ppm) .. .................... 97

5-21 Phenol degradation curves in multi-compound
systems using acclimated bacteria (initial
concentration 1 ppm)................. ......... 98

5-22 PCP degradation curves in multi-compound
systems (initial concentration 5 ppm) ........ 101

5-23 PCP degradation curves in multi-compound
systems (initial concentration 1 ppm) ........ 102

5-24 PCP co-degradation curves in multi-compound
systems (initial concentration 5 ppm) ........ 103

5-25 PCP co-degradation curves in multi-compound
systems (initial concentration 1 ppm) ........ 104

5-26 PCP degradation curves in multi-compound
systems using acclimated bacteria (initial
concentration 5 ppm) ......................... 105

5-27 PCP degradation curves in multi-compound
systems using acclimated bacteria (initial
concentration 1 ppm) ......................... 106

5-28 PCP degradation curves in multi-compound
systems with different phenol to PCP
concentration ratios.......................... 109

5-29 Phenol degradation curves in column
biodegradation study I ........... ... ......... 112

5-30 2,4-DCP degradation curves in column
biodegradation study I .................. ..... 113


viii








5-31 PCP degradation curves in column biodegradation
study I ...................................... 114

5-32 2,4-DCP degradation curves in column
biodegradation study II........................ 117

5-33 PCP degradation curves in column biodegradation
study II..................................... 118

5-34 Phenol degradation curves in column
biodegradation study III....................... 121

5-35 2,4-DCP degradation curves in column
biodegradation study III..................... 122

5-36 PCP degradation curves in column biodegradation
study III...................................... 123













Abstract of Dissertation Presented to the Graduate School ol
the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy




BIODEGRADATION OF SELECTED PHENOLIC COMPOUNDS
IN A SIMULATED SANDY SURFICIAL FLORIDA AQUIFER


By

CHEN HSIN LIN


December 1988

Chairman: Wesley Lamar Miller
Major Department: Environmental Engineering Sciences

Phenolic compounds are commonly found contaminants in

groundwater systems. In this research the sorption and

biodegradation of phenol, 2,4-dichlorophenol (2,4-DCP) and

pentachlorophenol (PCP) were investigated. The soil

materials used were characterized as fine grained sands witl

negligible organic carbon contents.

Freundlich sorption coefficients of 0.0158 for phenol

and 0.0547 for 2,4-DCP were found. Pentachlorophenol was

more strongly adsorbed with an adsorption coefficient of

1.12. In multi-compound systems competitive sorption was

evident, and adsorption capacities were reduced by a margin

ranging from 70% for phenol to 30% for both DCP and PCP.

All three compounds exhibited nonlinear sorption behavior

with a range of exponent values from 0.56 to 0.7.








Desorption coefficients showed little difference from

adsorption for phenol and 2,4-DCP, but were significantly

different for PCP, indicating hysteresis of PCP sorptions.

The retardation factors were 1.03 for phenol, 1.16 for 2,4-

DCP and 2.26 for PCP.

In batch biodegradation studies using indigenous soil

bacteria phenol degraded quickly (tl/2 = 12 hours) and was

completely destroyed within three days. 2,4-DCP was also

completely degraded but had taken 23 days (t1/2 = 7 days).

PCP was resistant to biodegradation with an average half-

life of 120 days. In multi-compound systems, phenol

degradation rates dropped off to 0.4 day- (t /2 1.7 days)

but PCP degradation rates increased to 0.008 day-1 (tl/2= 86

days).

Biodegradation rates in column studies were obviously

greater than in batch experiments, with the rate increase

for PCP degradation being especially noticeable (tl/2= 12

days), because of larger bacterial populations and the

dynamic flow conditions made the substrates more available

to the bacteria.

When controlled under an aerobic environment by the

addition of hydrogen peroxide, all three phenolic compounds

degraded fastest, under anoxic conditions both the

microbial population buildup and the rate of phenolic

compound degradation were slower but not by a wide margin.








Bacterial growth in the columns did not reduce the

hydraulic conductivity of the system, indicating the

feasibility of applying in-situ biodegradation techniques to

groundwater contamination problems.
















CHAPTER I
INTRODUCTION

Groundwater contamination by trace organic compounds is

a widespread problem but only recently has the public become

aware of the seriousness of this problem. The problem is

serious largely because groundwater does not have the self-

cleaning mechanisms commonly seen in surface water, and

because with an increasing dependency, about half of the

population in the United States are now depending on

groundwater for drinking. In some areas such as Florida the

dependence on groundwater is more than 90 percent of the

population, and the demand of groundwater supply is expected

to increase 25 percent per decade (DeHan, 1981).

Groundwater contamination results from various types of

sources, such as disposal of hazardous wastes into unlined

landfills, accidental spills of chemicals and leakage of

underground storage tanks. Industry-related sources include

chemical leaks from storage areas, accidental spills, and

vapor condensate from solvent-recovery systems.

Nonindustrial sources include road runoff, municipal

landfills, junk yards, septic tanks, and domestic waste

water. Numerous organic chemicals have been detected in

groundwaters as contaminants nationwide. Phenol and

substituted phenols, which are some of the most frequently








found organic chemicals, are accountable for many of the

groundwater contamination cases (Plumb, 1985; Pye and

Patrick, 1983). This is especially true in the southeastern

United States because of the high concentration of wood-

preserving industry located in this region and the wide use

of these chemicals in this industry.

To properly assess a groundwater contamination problem,

it is necessary to understand the transport and fate of the

contaminants in the subsurface environment. Once the

contaminants enter the system, their transport and fate are

determined by the chemical, physical, and biological

properties of both the chemical compounds and the aquifer

materials. Dilution advectionn and dispersion), sorption

(adsorption and desorption), and degradation (biotic and

abiotic), are three major forces governing the fate of the

contaminants (Mackay et al., 1985; Newsom, 1985). The one-

dimensional equation proposed by Bear can be used to

describe these phenomena (Bedient et al., 1985; Skopp et

al., 1981):


3C aC a C p iS
SD ---- --- --- ---- (1-1)
t L ax a x n at

where C = aqueous phase concentration of compound (M/L3)

t = time (T)

D = longitudinal dispersion coefficient (L2/T) = a*v

a = dispersivity

x = distance in flow direction (L)

v = average seepage velocity (L/T)







p = density of bulk dry soil

n = porosity

S = adsorbed phase concentration of compound (M/M)

The terms on the right hand side of Equation (1-1) are

referred to as dispersive transport, convective transport,

and adsorption, respectively. In a linear adsorption the

distribution coefficient Kd= S/C, since aS/ at = Kd*(aC/at),

Equation (1-1) can be written as

aC 2C ac
R ---- = D -- --- (1-2)
St L 3x2 ax


P Kd
R = 1 + ------- (1-3)
n

where R is the retardation factor. If degradation (decay)

is incorporated then the equation becomes

aC a2C aC
R -- = D V* -- K C (1-4)
a t L ax ax D


where KBD is the degradation rate (1/T). Many numerical

solutions have been presented by various researchers

(Amoozegar-Fard et al., 1983; Fuller and Warrick, 1985; Van

Genuchten, 1981). As an example, if given the boundary

conditions C=C0 at x=0 and C=0 at x=infinity, and with an

initial condition C=0 at t<0, Equation (1-4) can be solved

by a numerical solution proposed by Sauty (1980):

C = C0/2 { exp[(v-u)x/2D] erfc[(Rx-ut)/(4DRt)1/2

+ exp[(v+u)x/2D] erfc[(Rx+ut)/(4DRt)1/2] (1-5)

where u = (v + 4 kD R D)/2. With this solution and the
BD








required parameters, the fate of contaminants in groundwater

can be predicted.

Sorption is a measure of partition between the aqueous

phase and solid phase in the aquifer. It is known to be

important to the fate and transport of organic compounds in

groundwater systems. The degradation term of Equation (1-4)

may be contributed by photolysis, hydrolysis, abiotic

oxidation and biotic degradation. Among these pathways

biodegradation is the most important degradation process in

groundwater systems for phenolic compounds. Indigenous

microorganisms can utilize organic compounds as carbon

sources to generate energy for their maintenance requirement

and increase cell mass, provided that adequate nutrients and

electron acceptors are available. When the concentration of

a contaminant becomes very low (which is not unusual in

groundwatLr), the microorganisms may not be able to derive

enough energy to support the maintenance requirement. If

this condition occurs, the population will decline, and

consequently the organic compounds may persist at trace

concentrations (Alexander, 1981, 1985).

A system that was developed to simulate a Florida sandy

aquifer in a natural environment was used to evaluate the

adsorption and desorption coefficients and biological

degradation rates of phenol, 2,4-dichlorophenol, and

pentachlorophenol.

With the data from this research, the behavior and fate

of these important phenolic compounds in a shallow Florida





5

sandy aquifer can be better predicted, and thus lead to the

development of treatment methods for remediating aquifers

contaminated by such phenolic compounds.















CHAPTER II
OBJECTIVES


The objectives of this study were as follows:

(1) To determine estimates of the sorption parameters

for phenol, 2,4-dichlorophenol and pentachlorophenol when

present alone and in mixtures in a sandy Florida soil.

(2) To evaluate the rates of biodegradation of these

phenolic compounds when present alone, and when present as

mixtures under simulated field conditions in a sandy Florida

soil.

(3) To determine the effects of co-degradation, enzyme

induction and sludge amendment on in situ biological

treatment of groundwaters that are contaminated with

phenolic compounds under simulated field conditions.

(4) To evaluate the change of hydraulic conductivity of

soils before and after in situ biological treatment under

simulated field conditions.
















CHAPTER III
LITERATURE REVIEW



This chapter presents a review of the pertinent

literature about the characteristics and the environmental

significance of selected phenolic compounds and their

sorption and degradation processes known to occur.


3.1 Environmental Significance of the Phenolic Compounds


Phenol, 2,4-dichlorophenol (2,4-DCP), and pentachloro-

phenol (PCP) are chosen as the contaminants in this study

because phenolic compounds are commonly found contaminants

in groundwater. This is especially true in Florida and the

southeastern United States because of the high density of

wood-preserving industries in this area of the country.

Among the phenolic compounds, phenol, 2,4-dichlorophenol and

pentachlorophenol have the most commercial importance

(Goldfarb et al., 1981).

Phenol was first isolated from coal tar in 1834, (Moore

and Ramamoorthy, 1984), but today almost all phenols are

manufactured by the cumene hydroperoxide process (Kirk and

Othmer, 1985). It has been used in many commercial products

including resins, nylons, plasticizers, antioxidants, oil

additives, polyurethanes, drugs, pesticides, explosives,








dyes, and gasoline additives. In 1981 alone, more than 1.15

million metric tons of phenol were produced in the United

States (U.S. International Trade Commission, 1982). All 17

possible chlorinated phenols are commercially available.

Monochlorophenols are used mainly in the production of

higher chlorinated phenols. 2,4-DCP is used primarily in the

manufacture of the widely used agricultural pesticide 2,4-

dichlorophenoxy acetic acid (2,4-D). When 2,4-D breaks

down, 2,4-DCP will be present as one of the products (USEPA,

1986). Pentachlorophenol has been extensively used as a

wood preservative because of its fungicidal properties.

Phenol is fairly soluble in both water and nonpolar

solvents as shown in Table 3-1. Alkaline salts of phenol

are also readily soluble in water. Generally the volatility

and the aqueous solubility decreases with the increasing

number of chlorine atoms on the benzene ring. Electron

withdrawal by the ring chlorines causes pentachlorophenol to

be acidic and a relatively weak nucleophile, while making

its salts fairly stable. Physical properties of selected

phenolic compounds are listed in Table 3-1. The organo-

leptic properties of the chlorophenols are manifested by

imparting odor to water and tainting fish flesh (Lee and

Morris, 1962). As a group, the chlorophenols are highly

toxic. Although insufficient information exists on the

carcinogenicity of most chlorophenols, 2,4,6-trichlorophenol

has been shown to be an animal carcinogen, and para-chloro-

phenol is a suspected carcinogen based on mutagenicity





9

screening tests (Moore and Ramamoorthy, 1984). Accordingly,

phenol, 2,4-DCP and PCP are listed as priority pollutants by

the U.S. Environmental Protection Agency (USEPA, 1979).


Table 3-1. Physical properties of the phenolic compounds
(Verschueren, 1977; USEPA, 1979)


Parameters Phenol 2,4-DCP PCP

M.W. 94.1 163.0 266.4

pKa 10.02 7.85 4.74

Melting Pt.(C) 41 45 190

Boiling Pt.(C) 182 210 310

Vapor Density 3.2 5.62

Vapor Pressure (Torr) 0.529 0.12 0.00011

Solubility (mg/1) 93000 4600 14

Sp. gravity 1.07 1.38 1.98

Log Kw 1.46 2.75 5.01
ow
At 200 C
** Aqueous solubility at 200 C


3.2 Sorption of the Phenolic Compounds


Sorption is the process of the mass transfer of a

chemical between the solid phase and a liquid phase, such as

between soil and water mixture, which may be described as

S = Kd C (3-1)

where S is the concentration in the solid phase, C is the

concentration in the aqueous phase, and Kd is the

distribution coefficient. For the purpose of this research,

the term sorptionn" refers to the processes of adsorption








and desorption in general. Sorption is an important factor

in the determination of the fate of hydrophobic compounds,

("hydrophobic compounds" is defined as compounds with Kow

value greater than 5.0) (Doucette and Andren, 1987), in a

water/soil system. Adsorption tends to retard the migration

rate of contaminants in subsurface environment. It may

provide precious time to respond to accidental spills before

the contamination spreads. On the other hand, soils that

slowly desorb contaminants will become constant sources of

groundwater contamination (Delfino, 1977; Delfino and Dube,

1976), greatly prolonging the time required for an effective

cleanup, and increasing the cost of remedial actions.

Various reports indicate that the equilibrium

relationship between soil and solution phase solute

concentrations was found to be described best by the

nonlinear Freundlich isotherm model (Artiola-Fortuny and

Fuller, 1982; Boyd, 1982; Lagas, 1988; Laquer and Manahan,

1987; Means et al., 1980; Miller and Weber, 1986), which is

expressed as

X/m = KF Cb (3-2)

where X is the mass of solute adsorbed to soil surface, m is

the mass of soil, C is the solute concentration at

equilibrium in the aqueous phase, and KF and b are

constants. Freundlich sorption coefficient (K ) is a

measure of the degree of strength of adsorption, while b is

an indication of whether adsorption capacity remains

constant, i.e. when b=l, sorption is linear within the range








of solution concentrations used in a particular study. Note

that in this case the equilibrium Freundlich partition

coefficient, KF, is the same as the distribution

coefficient, Kd, and X/m equals S in Equation (3-1).

Neither Freundlich partition coefficients nor distribution

coefficients are universally transferable because they

depend heavily on both the properties of chemical compounds

and the characteristics of the soil matrix. Enormous

efforts have been devoted to making these relationships more

useful and easier to apply to soils with different

characteristics. Karickhoff et al. (1979) demonstrated that

for a dilute solution (i.e. concentration of the contaminant

less than half of its solubility in water), partition

coefficients based solely on organic carbon in the soil

matrix, Koc, correlate closely to KF/foc as

Kc = K / fc (3-3)

where f is the fraction of organic carbon in the soil
oc
matrix. This relationship ignores any influence of the soil

itself but does facilitate the use of partition coefficients

or distribution coefficients from the literature as long as

the fraction of organic carbon in the soils are documented.

Chiou et al. (1979) and Karickhoff et al. (1979) reported

that Koc could be related to water solubility. They also

reported a relationship of the octanol/water partition

coefficient, Kow (ml/g), as

log K = log K 0.21 (3-4)
log Koc 5 0.67 ow
log K = 5 0.67 log WS (3-5)
oc








WS is the solubility of a chemical in water in umol/l.

These relationships are convenient to use since values for

solubility and the octanol/water partition coefficient are

either well established by various workers or easy to

measure in a laboratory. However, these relationships all

have their limitations. Banerjee et al. (1980) suggested

that for compounds with high melting points, Equation (3-5)

may be invalid, and proposed another correlation between K
ow
and WS which incorporated a melting point correction term as

log K = 6.5 0.89 ( log WS ) 0.015 ( MP ) (3-6)
ow
where MP is the melting point in degrees centigrade. Rao

and Jessup (1983) cautioned that Equations (3-3), (3-5) and

(3-6) may not apply to soils containing less than 0.1

percent of organic carbon.

Both pH and ionic strength have significant influence

on sorption of phenolic compounds. Schellenberg et al.

(1984) showed that sorption of the unionized phenols and

their conjugate bases (phenolates) can occur. They

suggested that in natural waters of low ionic strength (i.e.

ionic strength < 10-3 M) and of pH values not greater than

the pKa values of the phenolic compounds by one unit,

phenolate sorption can be neglected. Based on this theory,

the conjugate bases of phenol and 2,4-DCP need not be

considered in the natural environment.

Phenol has a relatively small Kow value (log K =1.46),

which suggests only a slight tendency to become adsorbed

onto the organic detritus. As a comparison, PCP has a low








water solubility (14 mg/1 at 200C) and a higher K value of
ow
5.01, where imply a strong tendency for PCP adsorption onto

organic matter.

Laboratory experiments have shown a phenol desorption

of almost 100% from a thin layer of montmorillonite clay

exposed to 40% humidity for one week (Moore and Ramamoorthy,

1984). But Isaacson and Frink (1984) reported that phenol,

2-chlorophenol and 2,4-DCP were extensively sorbed onto

sediments, desorption was slower than adsorption, and in

some cases up to 90% of the sorbate was irreversibly held.

This contradiction may have been caused by differences in

the reaction pH, ionic strength, and the percent organic

carbon content of the sorbents in the two experiments.

Hydrophobicity (defined as the lack of the capacity of a

compound to dissolve in water) as indicated by Kow is not

the only factor controlling the sorption of phenolic

compounds (Boyd, 1982; Isaacson and Frink, 1984), hydrogen

bonding may also play an important role. Boyd (1982)

suggested that the phenolic hydroxyl group formed hydrogen

bonds by acting as a proton acceptor. Westall et al. (1985)

found that the more highly substituted chlorophenols are

subject to larger influence by ionic strength. Because the

sorption of molecular pentachlorophenol is much greater than

of ionized pentachlorophenolate, he concluded that pH and

ionic strength play more important roles in PCP sorption to

soil than in the less substituted compounds. Kaiser and

Valdmanis (1982) reported a wide range of Ko values for PCP
0w








from 4.84 at pH 1.2 to 1.30 at pH 10.5 and pH 11.5. This

higher partition coefficient at lower pH suggests a greater

affinity for the organic part of the soil as the pH

decreases.

A number of areas of research in the region of sorption

chemistry remain controversial, such as reversibility,

extent of reversibility, rate of attaining equilibrium, and

the effect of competitive solutes in sorption equilibria.

In many cases sorption is considered to be reversible

(Angley, 1987). However, Miller and Webber (1984) reported

that many researchers disagree about the reversibility of

sorption with various chemicals. Laquer and Manahan (1987)

reported that the sorption of phenol onto a siltstone showed

differences in adsorption and desorption isotherms, an

effect termed hysteresis. Rogers et al. (1980) found that

once sorbed, benzene tends to resist desorption. Van

Genuchten et al. (1977) suggested that values of the

equilibrium desorptive constant should be different from

that of adsorption while Equation (3-2) holds for both

cases. Nathwani and Phillips (1977) drew the same

conclusions on some hydrocarbons in crude oil, and found

that the percentage of hydrocarbon component desorbed varied

inversely with the amount of organic matter in the soil

matrix. Researchers have also shown diverse results about

the rate of attaining equilibrium. Miller and Webber (1986)

observed equilibrium occurring after several days for

nitrobenzene and lindane. Rao and Davidson (1980) found








that sorption reactions for many organic compounds were 60

to 80% completed within one minute. Ogram et al. (1985)

stated that greater than 98% of the 2,4-D sorbed at

equilibrium was sorbed within the first five minutes, and

Means et al. (1980) reported that equilibrium for some

polynuclear aromatic hydrocarbons was achieved in 20 hours

or less. According to this evidence, it is reasonable to

expect the sorption of phenolic compounds onto sandy soils

to reach equilibrium within 24 hours. In cases when more

than one compounds are present in a mixture, the effect of

competitive sorption should be considered. Theoretically,

if these compounds have similar Kow values, it is likely

that they will compete with each other for sorption sites

unless the concentration of these compounds is low and the

sorption surfaces relatively high. This phenomenon is

called competitive sorption (Kinniburgh, 1986). When the

compounds have very diverse partition coefficient values, an

increase in water solubility has been observed for the more

hydrophobic compound as a result of cosolvent effects. This

results in a decrease of adsorption or an increase of

desorption of the affected compounds. This effect may be

directly related to a compound's availability for

biodegradation (Thomas et al., 1986).

Whether sorption will enhance or decrease microbial

degradation rates in groundwater depends upon whether the

sorbed phenolic compounds are available to the microbes.

When contaminants are irreversibly sorbed to soil organic






16

matter, they are isolated from the degrading organisms and

are protected from intracellular degradation. On the other

hand, bacteria may also be sorbed. If bacteria and

contaminants are sorbed on adjacent sites on the soil

surface, the uptake of the contaminants by the sorbed

bacteria is facilitated (Ogram et al. 1985).

Isotherm models can be used to predict the sorption and

desorption behavior of the contaminants, and thus help to

design groundwater/soil reclamation programs. Although

equilibrium may not be reached in reality, the prediction

may serve as a guide to the direction of mass transfer.


3.3 Degradation of Phenolic Compounds

Many processes can contribute to the degradation of

phenolic compounds in the environment. Among these are

photolysis, chemical oxidation, hydrolysis, volatiliza-tion

and biodegradation. Each process needs special conditions

in order to proceed and has its own role in the degradation

of these compounds from the subsurface environment.


3.3.1 Photolysis


Phenol has long been known to form reddish high

molecular weight material when exposed to sunlight and air.

It can undergo photolysis either in the phenolate anion form

(maximum absorbance at 270 nm) or in the undissociated

molecule (maximum absorbance at 310 nm). Experimental

irradiation of phenol at 254 nm in the presence of oxygen

yields a phenoxy radical intermediate that subsequently give








substituted biphenyls, hydroquinone (m-dihydroxy benzene),

and catechol (o-dihydroxy benzene) (Moore and Ramamoorthy,

1984). Photolysis of phenol to hydroquinone occurs under

both natural sunlight and commercial sun lamps (USEPA,

1979).

Assuming a first order reaction, the rate of

disappearance of an organic compound by direct photolysis

from surface water is

-dC/dt = K [C] = k 10 (e-qz) [C] (3-7)

where C is the concentration of the compound, K is the

apparent first-order photolysis rate constant, k is a

constant of proportionality which includes the quantum yield

of the reaction, I0 is the solar radiation intensity at

photochemically active wave lengths incident on a water

surface, q is the extinction coefficient of the water (which

is a function of dissolved and particulate absorbers), and z

is the depth (Pignatello et al., 1983). Equation (3-7) can

be converted into a mathematically calculable form:

In ( C / Co ) = k 1 (e-qz) [C] (3-8)

One EPA report (1979) stated that 2,4-DCP and PCP do

undergo photolysis but its significance and environmental

importance is uncertain. However, Hwang et al. (1986)

indicated that in summer time K values for 2,4-DCP and PCP

were 1.0 and 0.37 h-, respectively at a depth of 3cm while
-l
compared to 0.016 h- for phenol. A similar result for PCP

photolysis was reported by Pignatello et al. (1983). Crosby

(1981) concluded that in either water or organic solvents,








PCP can be photolytically reduced to isomeric tri- and

tetrachlorophenols, and, in dilute aqueous solutions exposed

to sunlight, PCP or its salts undergo the replacement of

ring chlorines by hydroxyl groups to form corresponding

chlorohydroquinones, which are subsequently oxidized to

chlorobenzoquinones and then dechlorinated and/or ring

cleaved. Pentachlorophenol is a moderately acidic compound

and thus will exist primarily as an anion in natural waters.

This is environmentally significant because the anion

absorbs well beyond 310 nm (sunlight spectrum) leading to

more effective photolytic reactions. Wong and Crosby (1978)

reported that the rate of photolysis of pentachlorophenolate

anion was much faster than that of the undissociated

compound.

For the phenolic compounds in groundwater no photolysis

occurs naturally. However, this process can be useful in a

remedial action when spraying and recirculating is involved,

and needs to be considered as an option when performing a

feasibility study in a groundwater reclamation project.


3.3.2 Oxidation


Phenol has been oxidized by passing molecular oxygen

into an aqueous solution at 250C and pH 9.5-13. This

suggests a possibility of nonphotolytic oxidation in highly

aerated waters. Little information is available pertaining

to the oxidation of chlorinated phenols but usually highly








chlorinated organic compounds are resistant to oxidation

under natural environmental conditions (USEPA, 1979).


3.3.3 Hydrolysis


The rate of hydrolysis of a chemical compound can be

calculated by

-dC/dt = kA [H+] [C] + kB [OH ] [C] + kN [C] (3-9)

where kA and kB =second-order acid and base hydrolysis

constants, respectively; and kN= first-order hydrolysis rate

constant for pH independent reactions (Moore and

Ramamoorthy, 1984).

The covalent bond of a substituent attached to an

aromatic ring is usually resistant to hydrolysis because of

the high negative charge density of the aromatic nucleus.

Therefore, hydrolysis of phenolic compounds in a natural

groundwater environment will not be a significant process

(USEPA, 1979; Moore and Ramamoorthy, 1984).


3.3.4 Volatilization


The rate of volatilization for general organic

compounds is described by Smith et al. (1980) as
-1
-dC/dt = k [C) = C/L [1/k1 + R T/H k ]1 (3-10)
-1
where k = volatilization rate constant (hr )

L = depth of aqueous layer

kl= transfer coefficient in the liquid phase (cm/hr)

H = Henry's law constant (torr/M)

k = transfer coefficient in the gas phase (cm/hr)
g









R = ideal gas constant

T = absolute temperature.

The low vapor pressure and the high aqueous solubility

of phenol indicates that there is little tendency for

volatilization from water. Chlorinated phenols are less

soluble in water, but the higher acidity increases the

proportion of the ionized form (which is much less volatile

than its unionized counterpart) in the natural environments

and causes them to be highly solvated. Thus, volatilization

will not have a significant contribution for loss of most

chlorophenols in aquatic environments.


3.3.5 Biodegradation


As early as 1946 Claude E. ZoBell (ZoBell, 1946)

reported that more than 100 species representing 30

microbial genera had been shown to have the ability to

utilize organic compounds as carbon and energy sources, and

that such microorganisms are widely distributed in nature

(Atlas, 1981). Bartha and Atlas (1977) listed 22 genera of

bacteria, one algal genus and 14 genera of fungi that had

been demonstrated to contain members which utilize petroleum

hydrocarbons. All of these microorganisms were isolated

from an aquatic environment. In soil samples Jones and

Eddington (1968) found that 11 genera of fungi and six

genera of bacteria were the dominant microbial genera

responsible for hydrocarbon oxidation.








Ghiorse and Balkwill (1983) found 5x106 microbes per

gram of dry subsurface material by direct count using

epifluorescence microscopy. This result is very similar to

what Wilson et al. (1983) have found, 3xl06 to 9x106

microbes per gram of dry material, in soils taken from

various depths below the surface of the ground. They and

others further showed that those microorganisms can degrade

several hydrocarbons (Stetzenbach et al., 1985; Roberts et

al., 1980; Yaniga, 1982).

Bouwer and McCarty (1985) reported that 91% of

chlorobenzene can be biodegraded from a concentration of 11

ug/l by a biofilm grown with 1 mg/l of acetate after a 20

day acclimation period. In their studies ethylbenzene was

also cometabolized with acetate as a secondary substrate.

Tabak et al. (1981) have done a series of biological

degradation studies with organic priority pollutant

compounds under aerobic conditions. They followed a static-

culture flask-screening procedure with settled domestic

wastewater as microbial inoculum, and found that, at a

concentration of 5 mg/l, phenol, 2,4-DCP and 2,4,6-TCP can

be biodegraded 60 to 100% with rapid acclimation while PCP

showed only 19% reduction after seven days of incubation.

When the concentration was increased to 10 mg/l, a slight

decline in the rates of degradation was observed. Brown et

al. (1986) also found that 600 mg/l of ionized PCP can be

continuously biodegraded without affecting steady-state

growth in a fixed-film bioreactor containing a pCP-adapted








Flavobacterium. On the contrary, Klecka and Maier (1985)

reported that PCP degradation was inhibited at much lower

concentrations (800-1600 ug/1). Watanabe (1973a, 1973b)

examined PCP degradation in soil perfused with 40 ppm of PCP

and observed, after an eight day lag period during which

essentially no degradation occurred, chloride ion liberation

was initiated, and was complete within three weeks.

Subsequent additions of PCP were degraded more rapidly with

no lag period. Most of these degradation studies were

conducted under aerobic conditions. Boyd and Shelton

(1984), Smith and Novak (1987), and Ehrlich et al. (1982)

demonstrated that chlorophenols can also be degraded

anaerobically. However, rates of anaerobic degradation for

most organic contaminants are significantly slower than

those under aerobic conditions (Delfino and Miles, 1985),

and the anaerobic reductive dechlorination of PCP seemed to

stop at 3,5-dichlorophenol (Mikesell and Boyd, 1985).

Increased chlorination of the phenolic compounds

increased stability to oxidation and enzymatic degradation

(Cserjesi, 1967), therefore, highly chlorinated phenols tend

to be more resistant to degradation.

Many factors can influence the rate of biodegradation,

such as temperature, genus of the microorganisms, nutrients,

electron acceptor, pH, soil matrix, chemical concentration

of the compounds, and enzymes.


Temperature. Although biodegradation can occur over a

wide range of temperatures, temperature greatly influences








the rate of biodegradation. Within the ambient temperature

range, rates of biodegradation are faster at higher

temperatures than at lower temperatures. ZoBell (1969)

found that hydrocarbon degradation was over an order of

magnitude faster at 250 C than at 5 C. Larger

microorganism populations as well as higher assimilation

rates at higher ambient temperatures both contribute to this

increase. Vela and Ralston (1978) found that at higher

temperatures more phenol was metabolized per cell than was

required to support growth. A modified Arrhenius

mathematical model is available to estimate the effects of

temperature on biodegradation rate constants:
(T2-T1)
K2 = K1 (T2-TI) (3-11)

where K1 and K2 are the rate constants at temperature T1 and

T2 respectively, and g is a coefficient. Typical values for

g are from 1.01 to 1.04 in wastewater treatment systems

(Benefield and Randall, 1980).


Genus of microorganisms. Many microorganisms in the

natural environment are capable of degrading organic

compounds. Although the microorganisms may prefer some

particular compounds, they can rapidly adapt in order to

utilize available substrates (Hollibaugh, 1979; Haller,

1978; Hutchins et al., 1984).

Spain et al. (1984) found the microorganisms in a pond

were successfully acclimated to degrade p-nitrophenol after

a 6-day lag period. Healy and Young (1979) indicated that

microbial populations acclimated to a particular compound







can be simultaneously acclimated to other compounds, and

that a microbial population can metabolize several compounds

at the same time. However, Shimp and Pfaender (1985a,

1985b) reported the organic substances to which the

microorganisms have already been exposed can significantly

influence the ability of microorganisms to degrade other

organic compounds. They observed that exposure to amino

acids, carbohydrates or fatty acids enhances the ability of

microorganisms to degrade certain phenolic compounds while

exposure to humic materials had a negative effect.

More than 25 species of microorganisms were reported

capable of degrading PCP (Engelhardt et al., 1986). They

were isolated from soils, municipal wastewater sludges,

surface waters and groundwaters. Among these microorganisms

Arthrobacter, Trichoderma virgatum, Flavobacterium sp., and

Pseudomonas were most reported (Brown et al., 1986; Crosby,

1981; Edgehill and Finn, 1983; Kaufman, 1978; Mikesell and

Boyd, 1985; Stanlake and Finn, 1982; Steiert et al., 1987;

Suzuki, 1975; and Suzuki, 1977).


Nutrients. Microorganisms need nitrogen, phosphorus

and some trace minerals for incorporation into biomass, so

the availability of these nutrients is critical to

biodegradation. A general formula for microorganism

composition was proposed as C60H87023 12P (Benefield and

Randall, 1980). This formula reveals a C:N:P ratio of

23:5.3:1 in microorganism cells. However, optimal C:N and

C:P ratios for marine oil-degrading microorganisms were








found to be 10:1 and 100:1 respectively (Atlas, 1981).

Since carbon is utilized for both energy (non-growth) and

synthesis requirements (growth) while nitrogen and

phosphorus are used essentially for synthesis of new cells,

the optimal N:P ratio is somewhat less variable than the C:N

and C:P ratios, and 10:1 seems to be a reasonable value to

choose when supplying nutrients to microorganisms.

The optimal C:N or C:P ratio will need to be determined

experimentally for each specific case, because they are

largely dependant on the carbon-energy conversion efficiency

of the tested microorganisms. Dibble and Bartha (1979)

indicated that a C:N:P ratio of 800:13:1 was found to be

optimum and cost-effective for oil sludge biodegradation in

a "landfarming" process, but this ratio is far removed from

the theoretical values. They also reported that addition of

micronutrients and organic supplements (such as yeast

extract) were not beneficial to biodegradation.

The form of phosphorus or nitrogen is not critical for

the growth of microorganisms. However, it has been

recommended that an ammonia-nitrogen source is preferable to

a nitrate-nitrogen source because ammonia-nitrogen is more

easily assimilated by microorganisms (USEPA, 1985).

Kaufman (1978), on the other hand, stated that yeast

extracts accelerated PCP degradation, whereas glucose at 100

ppm suppressed degradation, and the substitution of ammonium

sulfate for sodium nitrate as a nitrogen source also

suppressed degradation. Because Kaufman studied the








degradation of different chemical compounds, the responsible

microorganisms could have been totally different from the

experiment described in the EPA's report.


Electron acceptor. Oxygen is required as an electron

acceptor in the energy metabolism of the aerobic

heterotrophic organisms. A portion of the organic material

removed is oxidized to provide energy for the maintenance

function (non-growth) and another portion for the synthesis

function (growth). Any oxidation must be coupled with

reduction. Oxygen satisfies this requirement in aerobic

biodegradation.

In traditional wastewater treatment, a minimum of 2

mg/l dissolved oxygen (D.O.) concentration is required for

aeration equipment to ensure a sufficient oxygen supply.

Because the microorganism concentrations in groundwater are

far less than those in an aeration tank, a lower residual

D.O. requirement should be expected. Borden et al. (1984)

have found that 0.25 mg/l seemed to be a threshold D.O.

concentration in groundwater for napthalene degradation.

In many cases the rate and extent of biodegradation of

many organic materials in a subsurface environment appear to

be limited by the availability of oxygen. Yaniga and Smith

(1985) reported that instead of traditional aeration, dilute

hydrogen peroxide is a good alternative for elevating

dissolved oxygen concentration in groundwater. Hydrogen

peroxide decomposes to oxygen and water. In the subsurface,

hydrogen peroxide decomposition is catalyzed by chemical and








biological factors. The decomposition can occur so rapidly

that oxygen bubbles out near the point of injection and

oxygen is not made available to the distant portions of the

needed zones. Research has shown that a high concentration

(10 mg/l) of phosphates can stabilize hydrogen peroxide for

prolonged periods of time in the presence of ferric

chloride, an aggressive catalyst for the decomposition.

However, such a high concentration of phosphate may cause

precipitation problems and render the soil impermeable.

Another problem is that hydrogen peroxide is cytotoxic, but

research has demonstrated that it can be added to some

cultures at up to 1000 mg/l concentration without toxic

effects (USEPA, 1985). Yaniga and Smith (1985) reported a

successful aquifer restoration project using 100 mg/l of

hydrogen peroxide as a dissolved oxygen source.


pH. Dibble and Bartha (1979) found that a pH of 7.5 to

7.8 was best for oil sludge degradation. This coincides

with the optimal pH for most microorganism growth. The pH

also has a great influence on the ionization of phenolic

compounds, since at higher pH conditions phenolates are

predominant, causing a decrease in adsorption and/or an

increase in desorption. This phenomenon is more significant

to higher chlorinated phenolic compounds. The effects of

ionization on biodegradation of phenolic compounds are not

clearly understood.

pH also influences toxicity of phenolic compounds.

Unionized PCP is apparently more toxic to both fish and








microorganisms than its ionized salts (Stanlake and Finn,

1982). Typical groundwater has pH values of 6.0 to 8.0

(Davis and Dewiest, 1966), however, 5.5 is a more common pH

value for surfacial aquifer waters in Florida. Laboratory

biodegradation studies should be performed in the same pH

range as that of groundwater before field application of

this treatment technique.


Chemical concentrations. Concentration of the compound

may be a significant factor which affects its susceptibility

to microbial attack. Organic compounds may persist in some

environments as a result of low prevailing concentration or

low solubility in water (Thomas et al., 1986). For example,

evidence exists that solubility limits the rate of bacterial

growth using a series of polycyclic aromatic compounds, and

some normally biodegradable substrates may not be

metabolized when the compounds are present at concentrations

lower than that required for maintenance of the

microorganisms (i.e. the threshold concentration) (Boethling

and Alexander, 1979; Bouwer, 1984).

Rittmann and McCarty (1980a, 1980b) have reported that

the threshold concentration, Cmin can be evaluated by the

relationship:

Cin = K* kD / (Y* k k ) (3-12)

where Ks is the Monod Half-maximum-rate concentration, kBD

is the first-order decay constant, k is the maximum specific

rate of substrate utilization by the microorganisms, and Y

is the cell growth yield.








Many organic contaminants in groundwater are present at

concentrations below Cmin and would apparently go

unutilized. However, simultaneous utilization of several

different substrates is possible. Sometimes microorganisms

can metabolize these trace compounds in the presence of

other substrates, called primary substrates which support

the long-term biofilm growth. This process is termed

secondary utilization, or cometabolism. It is a mechanism

which allows microorganisms to degrade compounds that could

otherwise not provide enough energy to sustain the microbial

culture (McCarty et al., 1981).

Studies have shown that the extent of biodegradation of

polychlorinated biphenyls was enhanced by adding sodium

acetate as a primary carbon source. The effect was

especially significant on higher-chlorinated isomers (Clark

et al., 1979). Marinucci and Bartha (1979) also found a

slight stimulation of 1,2,4-trichlorobenzene mineralization

was caused by the addition of primary substrates. Schmidt

and Alexander (1985) observed that the presence of acetate

has a negative effect on phenol degradation, and the delay

was lengthened by increasing acetate concentrations because

acetate is easier to degrade than phenol. Bouwer and

McCarty (1985) suggested that secondary substrate (i.e.

target contaminants) removal rates increase with time but

not with the increase of primary substrate concentrations

beyond a limiting concentration, and the overall residual








concentration of the target contaminants can be largely

reduced by cometabolism.

Laboratory biodegradation experiments should use

concentration ranges similar to those actually found in the

field or the results may not correctly reflect what will

take place in the field (Alexander, 1985; Wang et al.,

1984).


Soil matrix. Soil matrix affects the biodegradation of

phenolic compounds mainly as a function of the organic

matter in the soil which contributes to the adsorption. It

is unclear whether the compounds that have been sorbed to

the soil particles are subject to biodegradation. The

compounds may be biodegraded while sorbed to the soil, or,

as the aqueous concentration decreases as a result of

biodegradation, some of the sorbed compounds may be desorbed

to restore equilibrium, and then be available for

biodegradation (Smith and Novak, 1987). Ogram et al. (1985)

indicated that sorbed 2,4-D was completely protected from

biodegradation by both sorbed and suspended bacteria. No

equivalent data were found on phenolic compounds. However,

Crosby (1981) stated in his review paper that PCP degraded

faster in soils with high rather than low organic content.


Enzymes. The biodegradation of phenolic compounds was

shown to be highly responsive to enzyme induction, yet, it

is a topic little studied. Not many enzymes that are

responsible for degradation of phenolic compounds have been








isolated (USEPA, 1986). The enzymes necessary for PCP

degradation appeared to be inducible. Steiert et al.

(1987) demonstrated that a suspension of cells grown in the

presence of 2,4,6-trichlorophenol or 2,3,5,6-tetrachloro-

phenol did not show a lag period for degradation of 2,4,6-

trichlorophenol, 2,3,5,6-tetrachlorophenol or PCP,

indicating that one enzyme system can be induced for the

biodegradation of multiple compounds. Chu and Kirsch (1973)

and Karns et al. (1983) also reported similar observations.


3.3.6 Pentachlorophenol Degradation Mechanisms


Pentachlorophenol is very resistant to biodegradation

and may produce less chlorinated phenols as the degradation

products, therefore, its degradation pathway deserves more

study. Three significant mechanisms appear to account for

the biological degradation of PCP in soils: (1) reductive

dechlorination; (2) oxidative dechlorination; (3)

methylation. Conceivably an aggregate of microorganisms

should be more efficient in mineralizing phenolic compounds

(to CO2) than any of the pure cultures. The structural

formulas of those involved compounds are presented in Figure

3-1.


Reductive dechlorination. Bacteria such as

Flavobacterium sp. can utilize PCP as a sole source of

carbon and energy. Thus reductive dechlorination under

anaerobic conditions forms less chlorinated phenolic
















Qi




Pentachlorophenol




c t l


I
Tetrachlorocatechol


C;

Pentachioroanisole

/--'








Tetrach!orobenzoquinone


Tetrachlorohydroquinone


Figure 3-1.


Structural formulas for some common PCP
degradation products.


i


I








compounds. But this process seemed to stop at isomeric

trichlorophenols (Weiss et al., 1982).


Oxidative dechlorination. Watanabe (1973a) and Suzuki

(1977) found Pseudomonas sp. is capable of oxidizing PCP to

CO2 along with chlorohydroquinones and chlorocatechols as

intermediate metabolites. Because chlorohydroquinones and

chlorocatechols are less toxic to fungi than PCP, this

mechanism can be considered a detoxifying process

(Engelhardt et al., 1986).


Methylation. Fungus Trichoderma virgatum and bacterium

Arthrobacter sp. are the most commonly seen microorganisms

that methylate PCP and other chlorinated phenolic compounds

to form the corresponding chloroanisoles (Cserjesi and

Johnson, 1972). Chloroanisoles are also reported to have

less toxic effects on microorganisms than the corresponding

chlorinated phenolic compounds (Weiss et al., 1982).


3.4 Summary


This section reviewed the processes that are most

likely to occur in the natural environment to degrade

phenols. It also supports the feasibility of enhanced

biodegradation as a treatment method for chlorinated

phenolic compounds in groundwater systems.
















CHAPTER IV
MATERIALS AND METHODS


This chapter discusses the materials, analytical

methods and experimental design employed in this research.


4.1 Materials


4.1.1 Soil


The sandy soil used in these studies came from an

unused cell at the north pit of the Southwest Landfill,

Archer, Florida. This sandy soil is representative of soil

conditions in surficial aquifers which supply drinking water

in large areas of Florida. Advective fluxes tend to be

greater through granular horizons of this type than through

other soil formations, thereby facilitating contaminant

transport over wide geographic areas.


4.1.2 Chemicals


Three chemicals involved in this study: phenol (Fisher

Scientific Supplies, 92.9%), 2,4-dichlorophenol (Eastman

Kodak Co.) and pentachlorophenol (Aldrich Chemical Co.,

Inc., 99%) were purchased and used without further

purification.








4.1.3 Contaminated Water


Contaminated water was prepared by dissolving chemical

compounds into distilled deionized water. This approach was

chosen because it is easier to control concentrations and

would tend to have a consistent characteristic throughout

the course of the study.


4.1.4 Microorganisms


The microorganism seeds were taken from the return

sludge in the aeration tank of the University of Florida's

wastewater treatment plant. Unless otherwise specified, the

sludge was used without further treatment. The supernantant

of sludge, if used, was siphoned out after the sludge was

blended by a blender and settled.


4.2 Analytical Methods


4.2.1 Chemical Concentration Determinations


Organics analyses. Phenol, 2,4-dichlorophenol and

pentachlorophenol concentrations were analyzed by a high-

performance liquid chromatography (HPLC), Perkin-Elmer Model

LC-100, with a ZORBAX C-8 column (4.6 mm I.D. x 15 cm), a

LC-75 Spectrophotometric Detector and a Fisher Series 5000

Recorder. For phenol and 2,4-dichlorophenol determination,

a 58/42 (v/v) mixture of methanol/water mobile phase was

used and the wavelength of the UV detector was set at 197

nm. A 72/28 (v/v) mixture of methanol/water mobile phase








and 220 nm wavelength were selected for the analysis of

pentachlorophenol and the mixture of all three phenols.

Both mobile phase solutions were adjusted to pH 2 with

phosphoric acid (approximately 0.15% by volume), filtered,

and degased. Mobile phase flow rates were 2 to 3 ml per

minute.

Analytes were identified by comparing the retention

times of the standards and the retention times of the

samples while concentrations were calculated by comparing

the peak heights of the standards and the peak heights of

the samples. At 2 ml/min flow rate with 72% methanol in the

mobile phase, retention times for phenol, 2,4-DCP and PCP

were 1.3, 1.9 and 5.1 minutes, respectively. At 2 ml/min

flow rate when 58/42 methanol/water mixture was used as the

mobile phase, retention time was 1.6 minutes for phenol and

3.9 minutes for 2,4-DCP. The detection limit for phenol was

0.01 mg/1, for 2,4-DCP was 0.02 mg/l and PCP was 0.03 mg/i.


Optical absorbance measurement. Optical absorbance of

INTF in the biological activity assay (Section 4.2.4) was

measured by a Perkin-Elmer Model 552 Spectrophotometer with

the wavelength set at 465 nm.


Dissolved oxygen measurement. A Yellow Spring

Instrument, (YSI) Model 54 Oxygen Meter was used for the

measurement of dissolved oxygen concentration. A strip

chart recorder was connected when continuous monitoring was

required.








Specific conductivity measurement. A Yellow Spring

Instrument, (YSI) Model 33 S-C-T Meter was used for the

measurement of specific conductivity. A strip chart

recorder was connected when continuous monitoring was

required.


Nutrient analyses. Nitrogen and phosphorus

concentrations were measured by Technicon AutoAnalyzer II.


Weight measurement. All weighing were made on a

Mettler Model AE 160 balance unless the weight exceeded its

capacity of 160 grams. In that case a Mettler Model PR 1200

balance was used.


4.2.2 Soil Characterization


Organic carbon determination. Organic carbon

determinations were performed in the Soil Science

Department, University of Florida, following the Walkley-

Black procedure described in Section 29.3 of Methods of Soil

Analysis, Part 2 (Nelson and Sommers, 1982).


Soil water content determination. The direct method

with oven drying as described in Section 21-2.2 of Methods

of Soil Analysis, Part 1 (Gardner, 1986) was used to

determine the percent soil water content, which equals

([wt. of wet soil]/[wt. of dry soil])-l.


Soil porosity determination. A sample of 200 ml of

well mixed soil were dried in an oven at 105 C for 24 hours








to evaporate any moisture. The soil was placed in an 1-

liter graduated cylinder and the height of the soil was

marked. Then water was added by such that the water level

just coincided with the original soil level. Mixing was

provided to eliminate air pockets.

Porosity = [amount of water added] / 200 (4-1)


Hydraulic conductivity determination. Hydraulic

conductivity of the soil columns was determined by a

constant head permeameter shown in Figure 4-1. Soil

retention screens made of a few layers of glass fiber

supported with a stainless steel mesh were placed both on

top and at the bottom of the soil column. A piece of 3/4

inch tygon tubing was used to feed water from the constant

head reservoir. The frictional head losses from the tubing

and the fixtures were negligible compared to that caused by

the soil column. The hydraulic conductivity (K) was

calculated according to the formula:

K = Q L / dh A (4-2)

where K is the hydraulic conductivity (cm/sec), Q is the

measured flow rate (ml/sec), L is the length of the soil

column (cm), dh is the total head loss through the

permeameter (cm) (which is the difference in elevation

between the inflow and outflow water levels), and A is the

cross sectional area of the soil column (cm2) (McWhorter and

Sunada, 1977).













'rom Fi uOcet


O1P I


Over fiow


K= QL / AhA


Figure 4-1.


Experimental setup for hydraulic conductivity
test.


A h




t


Column
ArecJa=A








Hydraulic conductivity values were evaluated before and

after a set of column biodegradation experiments. This test

was designed to determine the effect of bacterial population

increases on hydraulic conductivity.


4.2.3 Sludge Characterization


Total volatile solids measurement. The following

procedure provided the total volatile solids (TVS)

measurements. Dry sludge weight and ash weight from 100 ml

of wet sludge were measured after drying in an oven at 103 C

overnight and combusting in a muffle furnace at 5500C,

respectively (Sawyer and McCarty, 1967).

TVS = ([wt. of dry solid] [wt. of ash]) x 10 (4-3)


Organic carbon determination. In the sorption isotherm

studies the organic carbon content of the sludge was

determined in order to calculate K values. In sludges
oc
from municipal wastewater plants the volatile solids are

mainly biomass. Accordingly, it was assumed that organic

matter was the same as volatile solids, and corresponds to

the biomass formula C60H 87023N12P (Benefield and Randall,

1980). Therefore, organic carbon, OC, is

OC (mg/l) = TVS x (720/1374) = 0.52 x TVS (4-4)

This factor, 0.52, agrees with the values 0.40-0.53 as

suggested in Methods of Soil Analysis (Nelson and Sommers,

1982).








4.2.4 Biological Activity Measurement


Biological activity was assessed by INT-Dehydrogenase

assay. INT (2-[p-iodophenyl]-3-[p-nitrophenyl]-5-

phenyltetrazolium chloride) is reduced by the electron

transport system of active microorganisms via dehydrogenase

activity (DHA) to form water insoluble, red INT-formazan

(INTF) crystals (Koopman and Bitton, 1987).

The procedure was modified as follows. After adjusting

the sample pH to 7.6, 1 ml of 0.2% INT solution was added to

a 5 ml sample and incubated in the dark until a pink color

developed. The incubation time was recorded, the sample

was filtered through a 0.45 um pore size membrane, and the

filter extracted with 5 ml of DMSO (dimethylsulfoxide). The

extract was centrifuged. The absorbance of INTF, which is

proportional to DHA, was measured by a spectrophotometer at

465 nm wavelength. The controls were prepared the same

procedure as the samples except 1 ml of formaldehyde was

added in order to kill the microorganisms in the controls.

DHA is expressed in equivalent oxygen uptake units (mg

02/l/day).

DHA = [ 905* Ve* (Ds Dc)] / (t* Vs) (4-5)

where Ve is the volume of DMSO, Ds is the optical absorbance

of the sample, Dc is the optical absorbance of control, t is

the incubation time (minutes), and Vs is the volume of

sample filtered.








4.3 Experimental Design


4.3.1 Batch Sorption Studies.


Unless otherwise specified, all experiments in the

sorption studies were performed using 40 ml glass vials with

screw caps and teflon lined septa as the reactors.


Batch adsorption. The objectives of this study were:

(1) to determine the adsorption partition coefficients of

phenol, 2,4-dichlorophenol and pentachlorophenol in an

aquifer with very low organic matter, (2) to show the

effects of mixing of phenol, 2,4-DCP and PCP on sorption,

and (3) to determine the effects of adding sodium azide

(NaN3). The chemicals were tested both individually and as

a mixture. In each vial, 40 grams of sandy soil and 20 ml

of solution were mixed and tumbled continuously by a rotator

at room temperature (approximately 23 C) for 24 hours to

ensure complete mixing. Sodium azide (NaN3) was added in

two different concentrations to selected vials to eliminate

biological degradation. Four chemical concentrations were

used in this study: 10, 7.5, 5 and 1 mg/l. The experimental

matrix consisted of three treatments, four concentration

levels, and four sample categories as listed in Table 4-1.

Replicates were prepared for six of the randomly chosen

treatments for the purpose of quality assurance. The

average number of those treatments was used for the

calculations.








Table 4-1. Experimental scheme for adsorption study.

---------------------------------- ------------ -----------
[Treatment] [Conc.] [Sample]

1 mg/l phenol
2ml sludge + 6mg/l NaN3 5 mg/l DCP
2ml sludge + 2mg/l NaN3 7.5 mg/1 PCP
no sludge + 2mg/l NaN3 (By) 10 mg/l (By) mixture



Batch desorption. The desorption study was performed

following adsorption study. Aqueous portions of the samples

were drained (only about 10 ml could be drained) and the

sample vials refilled with 10 ml of distilled water. The

samples were tumbled continuously by a rotator at room

temperature (approximately 230C) for 40 hours before

analysis of phenolic compound concentrations.


Extraction recovery. The purpose of the extraction

recovery study was to account for all chemical masses and to

determine the efficiency of methanol extraction. This

measurement was necessary in order to determine whether

concentration decreases were caused by biodegradation or by

adsorption later in the degradation experiments. The

extraction recovery was performed following the desorption

study. Aqueous portions of the samples were drained and

refilled with 10 ml of methanol. Samples were tumbled at

room temperature for 3 hours before analysis.


Calculation methods. The amount of each compound

adsorbed to the soil (X/m) were calculated by dividing the

difference between the initial mass in the system and the








mass in the aqueous phase with the amount of soil in the

vial. The amount of each compound desorbed from the soil

(-X/m) were calculated by dividing the difference between

"the mass remained in system after draining the free water,

which was calculated from the results of adsorption

experiment" and "the mass in the aqueous phase" with the

amount of soil in the vial. Sorption coefficients were

calculated by fitting data to the Freundlich model, i.e.,

plotting log(X/m) (log(-X/m) in the cases of desorption)

versus log(C). The slope is the constant b in Equation (3-2)

and ordinate the intercept is log(KFA) or log(KFD), where

KFA (in ml/g) is the adsorption equilibrium partition

coefficient and KFD is the desorption equilibrium partition

coefficient. Recoveries were determined by the

"(calculated) mass of analyte extracted from soil" to

"(calculated) mass of un-desorbed analyte on soil" ratio.

The dilution effect caused by the solutions trapped in the

soil matrix was accounted for in the calculations of

recoveries.


4.3.2 Column Sorption Studies.


Column sorption experiments were performed with three

12 inches long x 3 inches diameter glass columns. A

quantity of 1500 g (about 9 inches high) of sandy soil were

loaded in each column.


Conservative tracer. The experimental setup is shown

in Figure 4-2. Soil in the columns was saturated with





45

distilled water for 24 hours before the study began. A 1 N

ammonium chloride solution was pumped at 6.5 ml/min from a

reservoir, the effluent was monitored by a conductivity

meter and recorded by a strip chart recorder. However,

because ammonium chloride is more dense than water, ion

concentrations higher than influent built up around the

probe rather quickly and caused incorrect readings. To

overcome this problem, the column was saturated with the

ammonium chloride solution then desorbed with distilled

water.


Solute retardation. Columns were saturated with

distilled water for three days, then drained of free water.

Approximately 250 ml of water remained in the soil matrix

after draining. 250 ml of solution consisted of all three

phenolic compounds, each at a concentration of 6 mg/l each

was added to the top of each column which contained 1500 g

soil. The systems were recirculated by a peristaltic pump

(coupled with three heads) with a flow rate of 2.7 ml/min

(Figure 4-3). Samples, 250 ul, were taken at the column

inlets every 15 to 30 minutes and analyzed immediately.


Retardation factor calculation. Retardation factors

were calculated based on the assumption that the retardation

factor (R) equals the number of pore volumes passed through

the column when effluent concentration of each solute




































Pump


-:Cmm6onrd
i::fCthtride*!


Figure 4-2.


Experimental setup for conservative tracer
test.


Column


Probe














Mixture solution '
circulation line
\









Pum r




Iji1


--i r




v1iii iiii

Sampling .
point, o1 |i^ft l

S41
\ T

A ----


Figure 4-3. Experimental setup for column degradation
studies.


. I

i:::-[:::;:
-.-.:.-.-...-
iai!
t:::::::::::::1
"i r

u


- I---- -- - -


I








reached 50% of the influent concentration (Nkedi-Kizza et

al., 1987).


4.3.3 Batch Biodegradation Studies


Batch biodegradation were performed in 40 ml VOC vials

as described in Section 4.3.1. A 10 g soil to 30 ml

solution ratio was used throughout these experiments. The

samples were kept in the dark to avoid photolysis except for

the phenol and DCP degradation studies. Periodically the

caps were opened and 250 ul samples were taken for analyses.

At least duplicate injections were performed for each

analysis.


Nutrient requirement. This test was intended to

determine the effect of nutrients on biodegradation. Phenol

was chosen as the carbon source. Preliminary analysis by

autoanalyzer indicated that 2.36 mg/l of total phosphate and

0.05 mg/l of orthophosphorus can be extracted into 30 ml of

distilled water from 10 g of this soil, which exceeded the

theoretical phosphorus requirement for complete

mineralization of the phenolic compounds. The same analysis

only detected trace amounts of nitrate nitrogen. Therefore

only the nitrogen level was manipulated. The experimental

scheme is listed in Table 4-2.


Phenol biodegradation. Biodegradation of phenol in

soil by indigenous bacteria and by bacteria from municipal

wastewater sludge that was added to the soil was studied.








Sodium azide (NaN3) at a concentration of 2 mg/1 was added

to the control samples to preclude biodegradation. The

experimental scheme, along with the nutrient requirement

study, is listed in Table 4-2.

Table 4-2. Experimental scheme for phenol biodegradation
and nutrient requirement.

-------------------------------------------------------.
No. Conc. N added Added Sludge NaN
mg/1 mg/l C:N ml mg/i

101 5 1.16 10:3 1
102 5 1.16 10:3 1
103 5 0.39 10:1 1
104 5 0 0 1
105 5 0.39 10:1
106 5 1.16 10:3
107 5 0 0 2
108 1 0.23 10:3 1
109 1 0.08 10:1 1
110 1 0 0 2
111 1 0.23 10:3
112 1 0.08 10:1



2,4-DCP biodegradation. Biodegradation of 2,4-DCP in

soil by indigenous bacteria and by bacteria from municipal

wastewater sludge that was added to the soil was examined.

Sodium azide (NaN3) at a concentration of 2 mg/l was added

to #201 and #202, the control samples. The nutrient

requirement experiment indicated no significant differences

in phenol biodegradation at various nitrogen concentrations.

Therefore a 10:1 carbon to nitrogen ratio of ammonia

nitrogen (as nitrogen) was added to all samples. The

experimental scheme is listed in Table 4-3. The sample #203

was a replicate of #204 until t=348 hours when 0.5 ml of a








solution that contained phenol degrading bacteria taken from

the phenol degradation samples was added.


Table 4-3. Experimental scheme for 2,4-DCP biodegradation.


No. Cone. N added Added Sludge NaN
mg/l mg/l C:N ml mg/i

201 5 0.22 10:1 2
202* 5 0.22 10:1 2
203** 5 0.22 10:1 1
204 5 0.22 10:1 1
205 1 0.05 10:1
206*** 1 0.05 10:1
207*** 1 0.05 10:1
208 1 0.05 10:1 1

Replicate of #201
** 0.5 ml phenol degrading bacteria added at t=348 hr
*** Replicate of #205


Pentachlorophenol biodegradation. Biodegradation of

PCP in soil by indigenous bacteria and by bacteria from

municipal wastewater sludge that was added to the soil was

examined. Sodium azide (NaN3) at a concentration of 2 mg/i

was added to the control samples. A 10:1 carbon to nitrogen

ratio of ammonia nitrogen (as nitrogen) was added to all

samples. The experimental scheme is listed in Table 4-4.


PCP degrading enzyme induction. This experiment was to

determine whether the enzyme required for PCP degradation

can be induced by exposing the microorganisms to phenol and

2,4-dichlorophenol. The experimental scheme, along with the

PCP biodegradation study, is listed in Table 4-4. #305 was

a replicate of #306. This experiment was run for t=1440






51

hours, then 0.5 ml of sludge supernatant were added to #304,

#305 and #309.


Table 4-4. Experimental scheme for PCP biodegradation and
enzyme induction studies.

--------------------------------------------------------
No. Cone. N added Added Amendment NaN
mg/l mg/l C:N (1 ml) mg/2
--------------------------------------------------------
301 5 0.14 10:1 2
302 5 0.14 10:1
303 5 0.14 10:1 sludge
304* 5 0.14 10:1 phenol bact.**
305* 5 0.14 10:1 DCP bact.***
306 5 0.14 10:1 DCP bact.***
307 1 0.03 10:1 2
308 1 0.03 10:1
309* 1 0.03 10:1 sludge
310 1 0.03 10:1 phenol bact.**
311 1 0.03 10:1 DCP bact.***
--------------------------------------------------------
0.5 ml sludge supernatant added at t=1440 hr
** The solution that contains phenol degrading bacteria
*** The solution that contains 2,4-DCP degrading bacteria


Mixture biodegradation. Experimental results of

biodegradation of phenol, DCP and PCP when present in a

multi-compound mixture each at an 1 and 5 mg/l initial

concentration was performed and compared with the results of

their degradation when present as single compounds. The

effects of enzyme induction were also examined in this

experiment. The experimental scheme is listed in Table 4-5.


Co-degradation. The co-degradation study was based on

the theory that easily degraded compounds were provided as

primary substrates to build up microorganism populations and

the target phenolic compounds would be co-metabolized along

with the primary substrates. Sodium acetate and glucose








were added to the sludge amended samples and compared to

those without primary substrates. The experimental scheme,

along with the multi-compound biodegradation assay, is

listed in Table 4-5.


Table 4-5. Experimental scheme for biodegradation and co-
degradation studies of phenol, 2,4-DCP and PCP
in multi-compound systems.


No. Conc. N added Added Amendment NaN
mg/l mg/l C:N (1 ml) mg/i

401 5 0.74 10:1 2
402 5 0.74 10:1
403 5 0.74 10:1 sludge
404 5 0.74 10:1 phegol
405* 5 0.74 10:1 DCP
406* 5 0.74 10:1 PCP
407** 5 0.74 10:1 slg/acetate
408 5 0.74 10:1 slg/glucosee
409 1 0.15 10:1 2
410 1 0.15 10:1
411 1 0.15 10:1 sludge
412 1 0.15 10:1 phegol
413 1 0.15 10:1 DCP
414 1 0.15 10:1 PCP
415 1 0.15 10:1 slg/acetate
416 1 0.15 10:1 slg/glucoseg

* Added with 0.5 ml of sludge supernatant at t=529 hours
** Discontinued at t=652 hours
a Solution that contains phenol degrading bacteria
b Solution that contains 2,4-DCP degrading bacteria
c Solution that contains PCP degrading bacteria
d 1 ml sludge plus 5 mg/l sodium acetate
e 1 ml sludge plus 5 mg/l glucose
f 1 ml sludge plus 1 mg/l sodium acetate
g 1 ml sludge plus 1 mg/l glucose


PCP co-degradation in the presence of phenol. Previous

tests revealed that pentachlorophenol is very resistant to

biodegradation, and the presence of phenol seemed to

increase PCP's degradation. The PCP co-degradation in the






53

presence of phenol was performed to further investigate this

phenomenon. The experimental scheme is listed in Table 4-6.


Table 4-6. Experimental scheme for PCP co-degradation in
the presence of phenol.


No. phenol DCP PCP Sludge N added
mg/l mg/l mg/l ml mg/l

601 5 1 1 2 0.46
602 5 1 1 0 0.46
603 5 1 0 2 0.43
604 5 0 1 2 0.43
605 1 1 1 2 0.15



4.3.4 Column Biodegradation Studies


Column biodegradation were performed using the

experimental setups described in Section 4.3.2. Columns

were saturated with distilled water before a shock load of

phenolic compounds was introduced. The recirculation rate

was set at 2.7 ml/min. However, because of continuous

compressions and relaxations, the tubing inside the pump

heads became less flexible and caused the flow rate to be

inconsistent. When left unattended, the recirculation could

be and sometimes was completely stopped in five days.


Column biodegradation I. This study proceeded from the

solute retardation determination described in Section 4.3.2.

Column #1, #2 and #3 were amended with 3 ml, 6 ml and 9 ml

municipal wastewater sludge, respectively, and also with 0.7

mg/l of ammonium nitrogen (as nitrogen). The initial

concentration for phenol, 2,4-DCP and PCP was 3 mg/l after








mixing and dilution of the added solution with the distilled

water in the columns. Sampling and analytical procedures

were the same as described in Section 4.3.2. The

experimental scheme is illustrated in Figure 4-3.


Column biodegradation II. This experiment was designed

to examine the results of PCP co-degradation in the presence

of phenol and was similar to the experiment in Section 4.3.3

except for using recirculation in columns. Columns were

saturated with 500 ml distilled water for three days before

draining all free water. Distilled water was then refilled

so that each column had 350 ml water (including 250 ml

trapped water and 100 ml free water) in it. A 150 ml of

mixed solution with different concentrations of phenol and

PCP were added to the top of each column which contained

1500 g of soil. To enhance the effects of co-degradation,

phenol concentrations with ten fold difference were used.

The same flow rate setting as in the previous experiment

(2.7 ml/min) was maintained. The experimental scheme is

listed in Table 4-7.






55

Table 4-7. Experimental scheme for column II (codegradation
of PCP and phenol).


Condition No. phenol DCP PCP N added
mg/l mg/l mg/l mg/l

Before # 1 6.7 6.7 6.7 1.00
dilution # 2 67 6.7 6.7 5.63
# 3 67 6.7 6.7 5.63

After # 1 2.0 2.0 2.0 0.33
dilution # 2 20 2.0 2.0 1.88
# 3 20 2.0 2.0 1.88



Column biodegradation III. The purpose of this

experiment was to investigate biodegradation of phenolic

compounds in different environments. Column #1 was under an

anoxic environment created by purging the solution in this

column with compressed nitrogen gas. Dissolved oxygen

concentrations were kept below 0.1 mg/l. Column #2 and

column #3 were under an aerobic environment. Column #2 was

aerated with compressed air and maintained at 5.0 mg/l or

higher dissolved oxygen concentrations. Column #3 had 0.2

ml of 30% hydrogen peroxide added (equivalent to 240 mg/l of

hydrogen peroxide in the free solution above the soil level)

whenever the dissolved oxygen concentration fell below 3.0

mg/l. All dissolved oxygen concentrations were measured at

the point just above the soil levels and about 2 inches

below the solution levels. No microorganisms other than

indigenous bacteria were introduced as in the previous

experiments. An equal concentration of 4.5 mg/l for each

compound (after mixing and dilution) was applied.
















CHAPTER V
RESULTS AND DISCUSSION


This chapter presents and reviews the results of all

experiments performed in this research, followed by a

discussion of each topic. Major categories are soil

characterization, batch sorption, column sorption, batch

biodegradation, column biodegradation, and hydraulic

conductivity determination.


5.1 Soil Characterization


The aquifer materials used in this research were dry,

clean, uniformly sized, yellowish brown in appearance, and

predominantly fine grained sands. The selected physical

properties of the soil are presented in Table 5-1.

Soil analysis indicated that there was very little

(non-detectable), if any, organic carbon content in the soil

matrix. For practical purposes it was assumed that the

organic carbon content in the soil matrix was zero. The

bulk density and porosity were not measured under

undisturbed, in-situ conditions.

All the soils were obtained at the same time and stored

in a capped bucket in the laboratory for later use. They

were visually inspected and foreign objects such as grass

roots and wood chips were removed before use.








Table 5-1. Selected physical properties of the soil.


Parameters Values

Particle density 2.52 g/ml
Water content 6.5% by volume
Bulk density 1.45 g/ml
Organic carbon Negligible
Porosity 0.45
Sieve analysis
Passed #30 100%
Retained on #40 0.45%
Retained on #140 96.68%
Passed #140 2.87%



5.2 Batch Sorption


The sorption isotherm data were fitted to the

Freundlich model using the method of least squares

regression analysis. These data are listed in Appendix A.


5.2.1 Single Compound Batch Adsorption.


The Freundlich sorption parameters of the phenolic

compounds on aquifer material are presented in Table 5-2,

and the parameters on aquifer material with sludge addition

are presented in Table 5-3. Notice that PCP concentrations

consisted both the ionized form (pentachlorophenolate) and

the unionized form (pentachlorophenol), thus, the results

for PCP sorption as well as degradation experiments

represent a combination of these two PCP forms.








Table 5-2. Adsorption regression parameters of phenolic
compounds in single-compound system on plain
soil.


Compounds pH log KFA ST.DEV.+ b* ST.DEV. R2

Phenol 5.2 -1.800 0.046 0.642 0.060 0.983
2,4-DCP 5.2 -1.262 0.029 0.696 0.037 0.994
PCP 5.0 0.049 0.054 0.558 0.048 0.985

+ Log standard deviation of log K values b
*Exponent in the Freundlich model: (X/m)=K C


Table 5-3. Adsorption regression parameters of phenolic
compounds in single-compound system on soil
with sludge.

+ 2
Compounds pH log KFA ST.DEV. b ST.DEV. R2

2ppm NaN3
phenol 5.2 -1.097 0.061 0.647 0.077 0.973
2,4-DCP 5.2 -0.849 0.067 0.776 0.083 0.977
PCP 5.0 0.466 0.036 0.724 0.035 0.995
6ppm NaN3
Phenol 5.2 -1.346 0.069 0.672 0.087 0.967
2,4-DCP 5.2 -0.813 0.085 0.787 0.106 0.965
PCP 5.0 0.471 0.064 0.757 0.065 0.985
Average
Phenol -1.346 0.672
2,4-DCP -0.831 0.782
PCP 0.469 0.741

+ Log standard deviation of log KFA values
# Not used, only for reference
$ Not an average number
*Exponent in the Freundlich model: (X/m)=KF C


The sludge used in the sorption studies had an average

total volatile solids (i.e., organic matter) of 4350 mg/l.

Based on the organic carbon/organic matter ratio of 0.524,

the organic carbon content was determined to be 2280 mg/l.

In the applicable batches, 2 ml of sludge were added and

assumed to be fully integrated into the soil matrix, which

had no organic carbon initially. The organic carbon content








then became 4.56 mg as sludge was added to each system, or

0.0114% (4.56 mg organic carbon in 40 g of soil) in the soil

matrix.

In the batch isotherm study of aquifer material with

added sludge, the addition of sodium azide at different

concentration levels, 2 ppm and 6 ppm, did not cause

Freundlich sorption coefficients to significantly change

based on a paired difference t-test at alpha= 0.05

significance level (McClave and Dietrich, 1985), which

indicated that the presence of sodium azide did not

interfere with the sorption behaviors of the phenolic

compounds. Thus the average of those two sets of parameters

was taken and used to calculate the sorption parameters on

organic carbon, with the exception of the phenol data.

Phenol adsorption on aquifer material with sludge and with 2

mg/l sodium azide showed an unusually high KFA value, and

later in the consequent desorption study all the phenol

concentrations were biodegraded to trace amounts, which

indicates that the sodium azide at 2 mg/l was not effective

enough to inhibit all the microorganisms (also, it indicates

that phenol degrades rather quickly). Therefore the data

obtained from this experiments, although presented, were not

used for the calculation of the sorption coefficients.

The Freundlich sorption coefficients for the isotherms

with sludge are quite different from those without sludge

addition. Because both sets of isotherms were performed

under the same conditions (other than the addition of








organic matter), these differences were solely contributed

by the added organic matter. Table 5-4 presents the

calculated sorption parameters that resulted from adding the

organic carbon in the wastewater sludge to the soil along

with some values available in the literature. The

calculation procedure is listed in Appendix E.


Table 5-4. Calculated adsorption parameters of phenolic
compounds in single-compound system based on
organic carbon.


Measured Literature log Koc values

Compounds log K c b (1) (2) (3) (4) (5)

Phenol 2.586 0.682 1.21 3.46
2,4-DCP 2.910 0.826 2.10 3.60 2.54
PCP 4.204 0.824 4.80 3.51 4.84
------------------------------------------------------------
* K values are in ml per gram organic carbon.
* exponent in the Freundlich model: (X/m)=KF C
(1) Boyd (1982).
(2) Isaacson and Frink (1984).
(3) Calculated from K by U.S. EPA (1979).
(4) Calculated from Kow by Kaiser and Valdmanis (1982).
(5) Calculated from Kow values listed by Lagas (1988).
ow


From the results shown in Table 5-2 it is clear that,

although not in great amount, adsorption on soils with

virtually no organic carbon content was still occurring,

demonstrating that Equation (3-3) is not valid in this case.

This result agreed with Rao and Jessup's (1983) suggestion

that Equation (3-3) may not apply to soils containing

organic carbon content less than 0.1 percent. The

calculated Freundlich sorption coefficients listed in Table

5-4 are actually K values for phenol, 2,4-DCP and PCP, and
octhese values will be used throughout this research.
these values will be used throughout this research.








5.2.2 Mixed Compound Batch Adsorption.


The Freundlich sorption coefficients for the mixture of

the three phenolic compounds are shown in Table 5-5, Table

5-6 and Table 5-7, presented in the same order as in Section

5.2.1.


Table 5-5. Adsorption regression parameters of phenolic
compounds in multi-compound system on plain
soil.

------------------------------------~;-----------------_---
Compounds pH log KFA ST.DEV.+ b ST.DEV. R

Phenol 4.79 -2.286 0.119 0.868 0.154 0.941
2,4-DCP 4.79 -1.350 0.057 0.791 0.073 0.983
PCP 4.79 -0.131 0.077 0.682 0.082 0.972

+ Log standard deviation of log K values b
Exponent in the Freundlich model: (X/m)=KF C


Table 5-6. Adsorption regression parameters of phenolic
compounds in multi-compound system on soil with
sludge.


Compounds pH log KFA ST.DEV.+ b ST.DEV. R2

2ppm NaN3
Phenol 4.82 -1.610 0.041 0.840 0.052 0.992
2,4-DCP 4.82 -1.036 0.039 0.923 0.050 0.994
PCP 4.82 0.311 0.041 0.714 0.041 0.993
6ppm NaN3
Phenol 4.88 -1.714 0.080 1.002 0.104 0.971
2,4-DCP 4.88 -0.959 0.057 0.898 0.073 0.987
PCP 4.88 0.320 0.050 0.686 0.048 0.990
Average
Phenol -1.662 0.921
2,4-DCP -0.998 0.911
PCP 0.316 0.700

+ Log standard deviation of log KFA values








Table 5-7. Calculated adsorption parameters of phenolic
compounds in multi-compound system based on
organic carbon.

--*------ --------- -- --- ;- ----- -- ------- ---
Compounds log Ko b

Phenol 2.161 0.936
2,4-DCP 2.690 0.988
PCP 4.064 0.710

K values are in ml per gram organic carbon b
bois the exponent in the Freundlich model: (X/m)=KF C


The differences of sorption behavior between single-

compound and multi-compound systems are significant by t-

test analysis at alpha= 0.05. The Freundlich sorption

coefficients for mixed compounds are 1.5 to 3 times less

than those of single compounds. This phenomenon could have

been the result of either a co-solvent effect or a

competitive sorption effect. Although it is difficult to

identify the appropriate mechanism, the co-solvent effect

will have a greater influence on the co-solute (compounds

with lower solubility in common solvent) than on the co-

solvent (compounds with higher solubility in common solvent)

(Staples and Geiselmann, 1988). The common solvent is water

in this case. On the contrary, competitive sorption effect

should have a greater influence on sorption coefficients of

less hydrophobic compounds since they are less likely to win

the competition with more hydrophobic compounds for the

limited sorption sites. This observation provides a vehicle

to help identify the appropriate mechanism. Comparing the

corresponding Freundlich sorption coefficients from single-

compound and multi-compounds sorption studies, a list of






63

ratios can be calculated and is presented in Table 5-8. The

study shows phenol suffered the greatest loss of adsorption

capacity when mixed with other phenolic compounds, and

indicates that the loss of adsorption capacity was

predominantly caused by competitive adsorption effects.


Table 5-8. Ratios of Freundlich sorption coefficients for
phenolic compounds in single and multiple
compound systems.


Ratio of KFA: Soil Soil+sludge organic carbon

Phenol 3.05 2.88 2.66
2,4-DCP 1.22 1.47 1.66
PCP 1.51 1.42 1.38



In general, Freundlich sorption coefficients for

phenolic compounds increase as the level of chlorination

increases. The less-than-unity values of the Freundlich

exponent, b, indicate the adsorption was not linear, and

higher concentrations resulted in a proportionately less

amount of adsorption. These values of the Freundlich

exponent are in good agreement with the ones reported by

Boyd (1982) (b=0.79 for phenol and b=0.67 for 2,4-DCP),

Lagas (1988) (b=0.86 for PCP), and Laquer and Manahan (1987)

(b=0.65 for phenol).

The calculated K values (Table 5-4 and Table 5-7) are
oc
closer to those predicted by Equation (3-5), which are 2.99

for phenol, 4.03 for 2,4-DCP and 5.86 for PCP. A least

squares linear regression reveals a good correlation between

the log of water solubility (in umole/l) and log of Ko
oc









values for phenol, 2,4-DCP and PCP as

log K = 3.547 0.421 log WS (R2=0.983) (5-1)
oc


5.2.3 Batch Desorption.


The batch desorption data were fitted to the Freundlich

model and the results are presented in Tables 5-9 and 5-10.

Table 5-9 lists the Freundlich desorption parameters of

phenolic compounds in single-compound systems and Table 5-10

are in multi-compound systems. Like the results of

adsorption studies, there are significant differences (t-

test at alpha= 0.05) between single compound and multiple

compound systems.


Table 5-9. Desorption regression parameters of phenolic
compounds in single-compound systems.


Compounds pH log KFD ST.DEV.+ b ST.DEV. R2

[Aquifer material]
Phenol 5.03 -1.739 0.070 0.510 0.089 0.943
2,4-DCP 5.03 -1.129 0.034 0.681 0.044 0.992
PCP 5.01 -0.979 0.052 0.339 0.063 0.936

[Aquifer material + sludge]
2ppm NaN3
Phenol n/a n/a n/a n/a n/a n/a
2,4-DCP 5.03 -0.880 0.065 0.733 0.089 0.971
PCP 5.03 -0.656 0.059 0.696 0.089 0.969
6ppm NaN3
Phenol 5.03 -1.177 0.058 0.681 0.086 0.976
2,4-DCP 5.03 -0.861 0.062 0.898 0.073 0.969
PCP 5.03 -0.654 0.088 0.746 0.132 0.941
Average
Phenol -1.177 0.681
2,4-DCP -0.871 0.707
PCP -0.655 0.721

+ Log standard deviation of log K values
Because of the occurrence of bio degradation






























'" ^ J


,-Y
-11



Sngl-Adsp-
+ Snl-Desp ------
o Mix-Adsp -
S Mix-Desp ----


-1.5 -1.2 -0.8 -0.4


I I i I 1.
0 0.4 0.8 1.2


Log [Ce(ug/mi)]


Figure 5-1. Phenol sorption isotherms on plain soil.















































1 I I I 1 1
-1.2 -0.2 -0.


o' -




E~ -n?-

--

Nr

J


o-


Figure 5-2. Phenol sorption isotherms on soil with sludge.


Sngl-Adsp-
SSngl-Desp -----
| Mix-Adsp -
S Mix-Desp

i I ; I i I I I I
4 O 0.4 0 12 !.

Log [Ce(ug/lm)]


i


i





















-4







Lo I
Figure 5-3. 2,4-DCP sorption isotherms on plain soil.

3 --I -
I



o i
SSngl-Adsp----
+ Sng-Desp------
So Mix-Adsp - -
SnqMix-Desp .
-2. -i-- ---r-- ---------- --,----


-1.2 -O.E -0.4+ 0 0.4 O. 1.2

Log [Ce(ugl/nil)]



Figure 5-3. 2,4-DCP sorption isotherms on plain soil.



















SnI Adsp-
I
Id --

-i-I









I .Mix-Des
I .-
J ii--
F <. '















-. -0.- L /o- 0 142
Log Ce(ug/ )

-1. -- -- .g ---.- 0 0. 0. !.2 .-
Log [Ce(ugl n-l)]


Figure 5-4. 2,4-DCP sorption isotherms on soil with
sludge.














1



0.5



0




rJ











-o.5
--2.




0 -_

-J


-2.5


-1.2 -0.5 -0.4 0

Log [Ce(ug/mI)]


Os


1.2


Figure 5-5. PCP sorption isotherms on plain soil.


--
I-.





--, --"









I --"- _-_.--- --




I ,p
_s.-.11-- I-











SMix-Adsp -
Mix-Desp -
,-i


-1.6





70













-4 T,--- "--
0









SSng- p --
SM-s - -





J II



Se Mix-Des






-1.6 -1.2 -0.- -0.+ 0 C.4 0.S 1.z 1.6
Log solution concentration
-1 .5 i ,. ,




Lo| sot Mix-Adsc t ao n





Log solution concentration


Figure 5-6. PCP sorption isotherms on soil with sludge.








Table 5-10. Desorption regression parameters of phenolic
compounds in multi-compound systems.
...........................................................

Compounds pH log KFD ST.DEV.+ b ST.DEV. R2

[Aquifer material]
Phenol 5.01 -1.952 0.068 0.535 0.089 0.948
2,4-DCP 5.01 -1.380 0.052 0.744 0.068 0.983
PCP 5.01 -1.234 0.036 0.183 0.045 0.891
...........................................................
[Aquifer material + sludge]
2ppm NaN3
Phenol 5.01 -1.417 0.031 0.825 0.040 0.995
2,4-DCP 5.01 -1.057 0.009 0.837 0.012 0.998
PCP 5.01 -0.920 0.032 0.357 0.048 0.965
6ppm NaN3
Phenol 5.01 -1.481 0.079 1.009 0.102 0.980
2,4-DCP 5.01 -1.033 0.014 0.647 0.020 0.998
PCP 5.01 -0.887 0.042 0.374 0.062 0.948
Average
Phenol -1.449 0.917
2,4-DCP -1.045 0.742
PCP -0.904 0.366

+ Log standard deviation of log KFA values


Freundlich adsorption and desorption isotherms are

presented in Figure 5-1 through Figure 5-6. Statistical

analyses (t-test at alpha= 0.05) indicated that the

Freundlich coefficients (KFA and K D) of phenolic compounds

in general are significantly different between adsorption

and desorption. However, further analyses on each compound

showed that this difference was mainly contributed by PCP

(no difference for 2,4-DCP, different for phenol at alpha=

0.05 but no difference at alpha= 0.01). This suggested that

significant amounts of PCP were irreversibly held onto the

sorbents, but not phenol and 2,4-DCP. A mass balance

calculation confirmed this observation. Approximately 10%,

30% and 90% of adsorbed phenol, 2,4-DCP and PCP,









respectively, were irreversibly held onto the soil matrix

after desorbing with distilled water for 40 hours. These

are percentages of initially adsorbed masses that were

irreversibly adsorbed (not desorbed). Isaacson and Frink

(1984) reported similar irreversibilities among other

substituted phenolic compounds. The small Freundlich

exponent values of PCP desorption indicating that the

desorption intensities were low.


5.3 Column Sorption


Column sorption experiments were performed on column #1

and column #2. Breakthrough curves for phenol, 2,4-DCP and

PCP from column #1 are shown in Figure 5-7. These curves

were plotted using the data presented in Appendix B. The

breakthrough curve of the conservative tracer, ammonium

chloride, was reconstructed from the data obtained in the

desorption experiment as described in Chapter 4.

Because these column experiments were designed to

simulate a treatment of groundwater contaminated by a point

source such as a spill, only a limited amount of analytes

were spiked onto each column. This method differs from the

traditional way of performing breakthrough curve

experiments, and make the calculation of retardation factors

very difficult. The first task was to determine the initial

concentration, C This value should range from 3.0 mg/l if
o
a complete mixing mode was assumed, to 6.0 mg/l if a plug

flow mode was assumed. However, the highest concentration






73

ever detected was 3.75 mg/l of phenol from column #1 and 3.3

mg/l from column #2. Based on the Freundlich sorption

coefficients obtained from batch sorption studies, phenol

has a very low tendency to be adsorbed on this particular

type of sandy soil. Therefore it was assumed that the C

value of each compound was 3.75 mg/l for column #1 and 3.3

mg/l for column #2. Figure 5-7 was plotted based on the

normalized C/Co values, and the X-coordinates corresponded

to the intersections of the C/C =0.5 line with each

breakthrough curve being measured as the retardation

factors.

For the purpose of comparison, Equation (1-3) and

Equation (5-2) (Nkedi-Kizza et al., 1987) and the batch

sorption data were used to estimate the retardation factors.

Notice that Equation (1-3) assumes linear adsorption, i.e.,

the Freundlich exponent was assumed to be unity.


p KFA Cb-1
R = 1 + (5-2)
n


The calculated and measured retardation factors for

phenol, 2,4-DCP and PCP are listed in Table 5-11. They

agree with each other very well except for the PCP

retardation factor calculated from Equation (1-3). This

difference was caused by omitting the Freundlich exponent

since PCP adsorption deviates the most from linear.











1.2

1.1




0.9-

0.3

0.7 /
0
0.6
O/
0.5 -

0.4 -

0.3

0.2 -

0.1

A I I I I
0 0.4 0.8 1.2

Pore Volume
0 chloride + phenol


1.6 2 2.4 2,8


o PCP DCP


Figure 5-7. Column breakthrough curves for phenol, 2,4-
DCP and PCP.










Table 5-11. Retardation factors of mixed phenolic compounds
calculated by various methods.

--------------------------------------------------------
Compounds # pore vol. Eq.(l-3)* Eq.(5-2)*

Phenol 1.03 1.017 1.014
2,4-DCP 1.16 1.144 1.109
PCP 2.26 3.385 2.566

p=1.45 g/ml, n=0.45, C =3.75 mg/l
phenol : K =0.00518,b=0.868
2,4-DCP : K =0.0447, b=0.791
PCP : KA=0.74, b=0.682
FA


5.4 Batch Biodegradation


The data from all batch biodegradation experiments are

shown in Appendix C. All measured concentrations were

normalized as C/C x 100%. These normalized data were used

to evaluate the degradation rates and to plot figures. The

degradation rates were calculated as apparent rates, which

include the effect of adsorption at the beginning and the

effect of desorption later during the course of the

experiment. The apparent degradation rate constants for

phenol and 2,4-dichlorophenol are very close to the real

values since adsorption of these two compounds was fairly

weak. However, PCP has a much stronger adsorption than

phenol and 2,4-DCP do, which could cause the apparent

degradation rate constants to be high. Therefore another

set of results calculated by subtracting adsorption effects

at the beginning, termed conservative degradation rate

constants, are presented for all PCP degradation data. The

conservative data did not account for the loss of later





76

desorbed PCP due to biodegradation, therefore these two sets

of results define an upper limit and a lower limit for the

degradation rate constants and half-lives for each sample.


5.4.1 Nutrient Requirement.


The purpose of this test was to determine how much, if

any, nutrient is needed for the biodegradation of phenolic

compounds in this particular type of soil. Background

analyses indicated that there was enough soluble phosphorus

in the soil but the nitrogen concentration was at near the

limit of detection (0.01 mg/) Therefore the effects of

nitrogen content (at three levels) were tested. Biological

degradation was confirmed by comparing samples with

controls, and was the main contributor to the decrease of

phenol concentrations. The result showed no differences

among these treatments as presented in Figures 5-8 and 5-9,

indicating that at least part of the phenol was assimilated

for energy but not for growth.


5.4.2 Single Compound Biodegradation.


Batch biodegradation experiments for phenolic compounds

were performed. Treatments with indigenous soil bacteria

and amended with municipal wastewater sludge were included.

The apparent degradation rate constants were calculated

based on first order reaction kinetics.

Phenol degraded rather quickly, with average half-lives

ranging from 9 hours for the C =5ppm group to 15 hours for
0








the C =lppm group. Notice that in a first order reaction

the half-life values are independent of initial chemical

concentrations. However, the results showed different half-

life values for the C =5ppm and C =lppm samples under

otherwise same treatments. This deviation is because the

first order reaction kinetics does not address all the

factors (such as toxic effects and substrate availability)

that are influential to biological degradation reactions.

Table 5-12 lists the apparent biodegradation rate constants

for phenol. High correlation coefficient values for log of

concentration versus time indicate that the first order

reaction kinetics describes phenol degradation quite well.


Table 5-12. Apparent biodegradation rate constants for
phenol.


Sample C K STD ERR t ,
(ppm) (lay) of KBD (2Y) R2

#101 S 5 1.83 0.20 0.38 0.94
#102 S 5 2.03 0.24 0.34 0.93
#103 S 5 1.80 0.19 0.39 0.95
#104 S 5 1.78 0.22 0.39 0.93
#105 5 1.63 0.26 0.43 0.86
#106 5 1.75 0.35 0.40 0.84
#108 S 1 1.20 0.17 0.58 0.90
#109 S 1 1.16 0.09 0.60 0.97
#111 1 1.15 0.17 0.60 0.90
#112 1 1.12 0.17 0.62 0.90

S: Amended with sludge.
Linear correlation coefficient for log C vs. time.


The major difference between degradation rates lies

between the two groups with different initial

concentrations, and no differences were found among the

various nitrogen levels. The higher initial concentration








group seems to degrade faster than the lower starting

concentration group, and this effect was even greater than

the effect of amending with sludge although Figures 5-8 and

5-9 clearly indicated a lag period for the group with

indigenous bacteria (30 hours for the C =5 ppm group and 20

hours for the C =1 ppm group). All samples were degraded to

below or near 0.01 ppm, the limit of detection.

The results of the 2,4-dichlorophenol biodegradation

are presented in Table 5-13 as well as in Figures 5-10 and

5-11. 2,4-Dichlorophenol, as well as phenol, can be

biodegraded to a concentration close to or below the

detection limit, however, with slower rates.


Table 5-13. Apparent biodegradation rate constants for
2,4-DCP.


Sample C K STD ERR t2
(ppm) (l1Bay) of KBD (ay) R2

#203 S 5 0.23 0.02 3.01 0.91
#204 S 5 0.08 0.01 8.66 0.84
#205 1 0.09 0.01 7.70 0.83
#206 1 0.12 0.01 5.78 0.86
#207 1 0.10 0.01 6.93 0.89
#208 S 1 0.16 0.01 4.33 0.93

No samples with soil bacteria in the C =5ppm group.
#205, 206 and 207 are triplicates.
S: Amended with sludge.
Linear correlation coefficient for log C vs. time.
+ 0.5 ml of phenol degrading bacteria added at t=348 hr.


For the group containing a 1 ppm initial concentration,

samples with indigenous soil bacteria (#205, #206, #207) had

an average half-life of 6.8 days, which was shortened to 4.3

days when the sample was amended with sludge (#208). There










































0 [
001


01 + #103


Figure 5-8.


1 2 3 4


o #104


1ime (days
A #105


x #106


v #107


Phenol degradation curves in single-compound
systems (initial concentration 5 ppm).












'1

0,9


0.8


0.7


0.6


0.5


0.4


0.3


0.2


0.1 -


0-
0


0 #108


+ #109


Time (d#ys)
0 #110


A #111


x #112


Figure 5-9.


Phenol degradation curves in single-compound
systems (initial concentration 1 ppm).


1 2 3 4








were no samples with indigenous soil bacteria in the 5 ppm

initial concentration group, but compare #204 (with sludge

amendment in Co=5 ppm group, t/ =8.66 days) with #208, the

half-life for the higher initial concentration sample

appeared longer than its lower concentration counterpart.

The sample #203 was a duplicate of #204 until t=14.5 days

when 1 ml of solution containing phenol degrading bacteria

was added to #203. The effect was drastic as shown in

Figure 5-11 and was evident in half-life values. This was

attributed to either the increase in microorganism

population or to the introduction of some enzymes which were

induced by exposing the bacteria to phenol. The degradation

rate constant and half-life for #203 are only qualitative

because they were a result of the combination of two

treatments.

Pentachlorophenol also appeared to undergo biological

degradation but with a very different pattern from phenol

and 2,4-dichlorophenol. Figures 5-12 and 5-13 illustrate

the degradation of PCP for different initial concentrations.

Table 5-14 lists the apparent degradation rate constants and

half-lives, and Table 5-15 lists the conservative results,

which were calculated based on the data up to t=60 days.

Because of analytical problems, PCP degradation data showed

day to day variability. In order to depict trends in the

PCP degradation data, variations in concentration versus

time data were dampened by using weighted average

concentrations, C(n), at measurement n, where









2 C(n) + C(n-l) +C(n+l)
C(n) = ----------------------- (5-3)
4

Even with the data processed in this form, the results

of a few sample runs still did not indicate a linear

relationship between log of concentration and time. This

poor correlation is indicated by the low correlation

coefficients (R2) in Tables 5-14 and 5-15. Both raw and

normalized data sets and some curves plotted with raw data

are presented in Appendix C.


Table 5-14. Apparent biodegradation rate constants for PCP.


Sample C K STD ERR t R
(ppm) (19ay) of KBD (ay) R2
-3 -3
#302 I 5 5.07x103 1.5x10-3 137 0.46
#303 S 5 6.28x10-3 2.1x10-3 110 0.40
#306 D 5 7.76x10 3 1.4x10-3 89 0.64
#308 I 1 6.51x103 1.2x10-3 106 0.68
-3 -
#310 P 1 2.50x10 1.4x10 277 0.19
#311 D 1 8.35x10- 1.4x103 83 0.74
.-.--..---.----------------.-----------.-----------------
#304* P 5 1.02x10-l l.lxl0-3 68 0.86
-3 -3
#305* D 5 9.33x10 1.2x10 74 0.82
#309* S 1 9.23x10-3 2.3x103 75 0.56

I: Indigenous soil bacteria.
S: Amended with sludge.
P: Amended with phenol degrading bacteria.
D: Amended with 2,4-dichlorophenol degrading bacteria.
0.5 ml supernatant of sludge added at t=60 days.

































\
\\
V\
\ "


Added 0.5 n

ienoiDegrac

Bacteria


__ __ __ __ft __ _


0 4


I 4z1ni


S I I
8 12


4 f#20Z


5 20 24



CO .4-0


Time (daye)
0>


Figure 5-10. 2,4-DCP degradation curves in single-compound
systems (initial concentration 5 ppm).


I


I / I












1-


0.9


0.8


0.7


E 0.6 -

C

0
0.5

C
S 0.4 -
C
0
0



0.3 -







0 4 8 12 16 20 24 28

Time (days)
0 #205 + #206 o #207 A 208


Figure 5-11.


2,4-DCP degradation curves in single-compound
systems (initial concentration 1 ppm).












































80 100 120


o #303


Figure 5-12.


PCP degradation curves in single-compound
systems (initial concentration 5 ppm).


S20I
0 20D


40


o #301


60

Time (ds)
+4- 302


`o









\a














, -------
ii



O- i ",















I- 1


I '
0 \
| \















0. i---






*


.





\i-
r_ -_I

,/ I.. .i


' .. -
-^

Added 0.5 ml

ge Supernatant


-C,


I I I I
40 6o

Time (adve)
30o7 +4- ,~0


Figure 5-13. PCP degradation curves in single-compound
systems (initial concentration 1 ppm).


I I
120


II0,




S 1CO9























-a -~ -t


Added 0.5 ml

Sludge Supernatant


20 40 S


SI 0I I20
S 100 120


T+re (# +e)
+ ,304 6 #305


Figure 5-14.


PCP degradation curves using bacteria which
are acclimated to phenol and 2,4-DCP (initial
concentration 5 ppm).


a #301


M


-r,
-~f---~
ki\




-t~t-
h
--~--


'f-l

LI




Full Text

PAGE 1

:US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI XS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI IUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI :US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI ^US FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI CUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCUS FOCI

PAGE 2

BIODEGRADATION OF SELECTED PHENOLIC COMPOUNDS IN A SIMULATED SANDY SURFICIAL FLORIDA AQUIFER BY CHEN HSIN LIN A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 1988

PAGE 3

UNIVERSITY OF FLORIDA lllllllll 3 1262 08552 4543

PAGE 4

ACKNOWLEDGEMENTS I would like to express my thanks to Dr. W. Lamar Miller, the chairman of my supervisory committee, for his support during three years of my study. Also, my sincere thanks go to the rest of the committee members, Dr. W. Emmett Bolch, Dr. Paul A. Chadik, Dr. Joseph J. Delfino, and Dr. Daniel P. Spangler for their generous assistance and thoughtful criticism. Special thanks go to Dr. Ben L. Koopman for the use of his equipment, and to Mr. Bill Davis for his assistance with high performance liquid chromatography . This work could not have been completed without the love of my wife, Lily, and the support of my family.

PAGE 5

TABLE OF CONTENTS ACKNOWLEDGEMENTS i i LIST OF TABLES v LIST OF FIGURES vii ABSTRACT x CHAPTERS I INTRODUCTION 1 II OBJECTIVES 6 III LITERATURE REVIEW 7 3.1 Environment Significance of the Phenolic Compounds 7 3.2 Sorption of the Phenolic Compounds 9 3.3 Degradation of the Phenolic Compounds 16 3.3.1 Photolysis 16 3.3.2 Oxidation 18 3.3.3 Hydrolysis 19 3.3.4 Volatilization 19 3.3.5 Biodegradat ion 20 3.3.6 PCP Degradation Mechanisms 31 3.4 S umma ry 33 IV MATERIALS AND METHODS 34 4.1 Materials 34 4.1.1 Soil 34 4.1.2 Chemicals 34 4.1.3 Contaminated Water 35 4.1.4 Microorganisms 35 4.2 Analytical Methods 35 4.2.1 Chemical Concentration Determinations 35 4.2.2 Soil Characterization 37 4.2.3 Sludge Characterization 40 4.2.4 Biological Activity Measurement 41 4.3 Experimental Design 42 4.3.1 Batch Sorption Studies 42 4.3.2 Column Sorption Studies 44 4.3.3 Batch Biodegradat ion Studies 48 4.3.4 Column B iodegrada t ion Studies 53

PAGE 6

V RESULTS AND DISCUSSION 56 5.1 Soil Characterization 56 5.2 Batch Sorption 57 5.2.1 Single Compound Batch Adsorption 57 5.2.2 Mixed Compound Batch Adsorption 61 5.2.3 Batch Desorption 64 5.3 Column Sorption 72 5.4 Batch Biodegradat ion 75 5.4.1 Nutrient Requirement 76 5.4.2 Single Compound Biodegradat ion 76 5.4.3 Multiple Compounds Biodegradat ion ... . 91 5.5 Column Biodegradat ion 110 5.5.1 Column Biodegradat ion I 110 5.5.2 Column Biodegradat ion II 115 5.5.3 Column Biodegradat ion III 119 5.6 Hydraulic Conductivity 124 VI SUMMARY AND CONCLUSIONS 125 6 . 1 Summary 125 6.1.1 Sorption 125 6.1.2 Batch Biodegradat ion 126 6.1.3 Column Biodegradat ion 128 6.1.4 Hydraulic Conductivity Tests 128 6.2 Conclusions 129 APPENDICES A BATCH SORPTION DATA 132 B COLUMN BREAKTHROUGH DATA 146 C BATCH BIODEGRADATION DATA 148 D COLUMN BIODEGRADATION DATA 166 E PROCEDURES TO CALCULATE K 170 oc REFERENCES 171 BIOGRAPHICAL SKETCH 182

PAGE 7

LIST OF TABLES Table Page 3-1 Physical properties of the phenolic compounds.... 9 4-1 Experimental scheme for adsorption study 43 4-2 Experimental scheme for phenol biodegradat ion and nutrient requirement studies 49 4-3 Experimental scheme for 2,4-DCP biodegradat ion study 50 4-4 Experimental scheme for PCP biodegradat ion and enzyme induction studies 51 4-5 Experimental scheme for mixture biodegradat ion and co-degradation studies 52 4-6 Experimental scheme for PCP co-degradation in the presence of phenol 53 4-7 Experimental scheme for column study II (co-degradation of PCP and phenol) 55 5-1 Selected physical properties of the soil 57 5-2 Adsorption regression parameters of phenolic compounds in single-compound system on plain soil 58 5-3 Adsorption regression parameters of phenolic compounds in single-compound system on soil with sludge 58 5-4 Calculated adsorption parameters of phenolic compounds in single-compound system based on organic carbon 60 5-5 Adsorption regression parameters of phenolic compounds in multi-compound system on plain soil 61 5-6 Adsorption regression parameters of phenolic compounds in multi-compound system on soil with sludge 61 5-7 Calculated adsorption parameters of phenolic compounds in multi-compound system based on organic carbon 62

PAGE 8

5-8 5-9 5-10 5-11 5-12 5-13 5-14 5-15 5-16 5-17 5-18 5-19 5-20 5-21 5-22 5-23 5-24 5-25 Single-compound system to multi-compound system ratios of Freundlich sorption coefficients for phenolic compounds 63 Desorption regression parameters of phenolic compounds in single-compound systems 64 Desorption regression parameters of phenolic compounds in multi-compound systems 71 Retardation factors of mixed phenolic compounds calculated by various methods 75 Apparent biodegradat ion rate constants for phenol 77 Apparent biodegradat ion rate constants for 2,4-DCP 78 Apparent biodegradat ion rate constants for PCP... 82 Conservative estimations of the biodegradat ion rate constants for PCP 89 Apparent biodegradat ion rate constants for phenol in multi-compound systems 92 Apparent biodegradat ion rate constants for PCP in multi-compound systems 100 Conservative estimations of the biodegradat ion rate constants for PCP in multi-compound systems 107 Apparent b iodegradat ion rate constants for PCP co-metabolized with phenol 108 Conservative estimations of the biodegradat ion rate constants for PCP co-metabolized with phenol "1-08 Column biodegradat ion study I results 110 Column biodegradat ion study II results 116 Column biodegradat ion study III results 119 DHA data for column degradation study III 120 Column hydraulic conductivity test results 124 vi

PAGE 9

LIST OF FIGURES Figure Page 3-1 Structural formulas for some common PCP degradation products 32 4-1 Experimental setup for hydraulic conductivity test 39 4-2 Experimental setup for conservative tracer test.. 46 4-3 Experimental setup for column degradation studies 47 5-1 Phenol sorption isotherms on plain soil 65 5-2 Phenol sorption isotherms on soil with sludge.... 66 5-3 2,4-DCP sorption isotherms on plain soil 67 5-4 2,4-DCP sorption isotherms on soil with sludge... 68 5-5 PCP sorption isotherms on plain soil 69 5-6 PCP sorption isotherms on soil with sludge 70 5-7 Column breakthrough curves for phenol, 2,4-DCP and PCP 74 5-8 Phenol degradation curves in single-compound systems (initial concentration 5 ppm) 79 5-9 phenol degradation curves in single-compound systems (initial concentration 1 ppm) 80 5-10 2,4-DCP degradation curves in single-compound systems (initial concentration 5 ppm) 83 5-11 2,4-DCP degradation curves in single-compound systems (initial concentration 1 ppm) 84 5-12 PCP degradation curves in single-compound systems (initial concentration 5 ppm) 85 5-13 PCP degradation curves in single-compound systems (initial concentration 1 ppm).. 86 5-14 PCP degradation curves using bacteria which are acclimated to phenol and 2,4-DCP (initial concentration 5 ppm) 87

PAGE 10

5-15 PCP degradation curves using bacteria which are acclimated to phenol and 2,4-DCP (initial concentration 1 ppm) 88 5-16 Phenol degradation curves in multi-compound systems (initial concentration 5 ppm) 93 5-17 Phenol degradation curves in multi-compound systems (initial concentration 1 ppm) 94 5-18 Phenol co-degradation curves in multi-compound systems (initial concentration 5 ppm) 95 5-19 Phenol co-degradation curves in multi-compound systems (initial concentration 1 ppm) 96 5-20 Phenol degradation curves in multi-compound systems using acclimated bacteria (initial concentration 5 ppm) 97 5-21 Phenol degradation curves in multi-compound systems using acclimated bacteria (initial concentration 1 ppm) 98 5-22 PCP degradation curves in multi-compound systems (initial concentration 5 ppm) 101 5-23 PCP degradation curves in multi-compound systems (initial concentration 1 ppm) 102 5-24 PCP co-degradation curves in multi-compound systems (initial concentration 5 ppm) 103 5-25 PCP co-degradation curves in multi-compound systems (initial concentration 1 ppm) 104 5-26 PCP degradation curves in multi-compound systems using acclimated bacteria (initial concentration 5 ppm) 105 5-27 PCP degradation curves in multi-compound systems using acclimated bacteria (initial concentration 1 ppm) 106 5-28 PCP degradation curves in multi-compound systems with different phenol to PCP concentration ratios 109 5-29 Phenol degradation curves in column biodegradat ion study I 112 5-30 2,4-DCP degradation curves in column biodegradat ion study I 113 VI 1 1

PAGE 11

5-31 PCP degradation curves in column biodegradat ion study I 11 4 5-32 2,4-DCP degradation curves in column biodegradation study II 117 5-33 PCP degradation curves in column biodegradation study II 118 5-34 Phenol degradation curves in column biodegradation study III 121 5-35 2,4-DCP degradation curves in column biodegradation study III 122 5-36 PCP degradation curves in column biodegradation study III 123

PAGE 12

Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy BIODEGRADATION OF SELECTED PHENOLIC COMPOUNDS IN A SIMULATED SANDY SURFICIAL FLORIDA AQUIFER By CHEN HSIN LIN December 1988 Chairman: Wesley Lamar Miller Major Department: Environmental Engineering Sciences Phenolic compounds are commonly found contaminants in groundwater systems. In this research the sorption and biodegradat ion of phenol, 2 , 4-d ichlorophenol (2,4-DCP) and pentachlorophenol (PCP) were investigated. The soil materials used were characterized as fine grained sands witr negligible organic carbon contents. Freundlich sorption coefficients of 0.0158 for phenol and 0.0547 for 2,4-DCP were found. Pentachlorophenol was more strongly adsorbed with an adsorption coefficient of 1.12. In multi-compound systems competitive sorption was evident, and adsorption capacities were reduced by a margin ranging from 70% for phenol to 30% for both DCP and PCP. All three compounds exhibited nonlinear sorption behavior with a range of exponent values from 0.56 to 0.7.

PAGE 13

Desorption coefficients showed little difference from adsorption for phenol and 2,4-DCP, but were significantly different for PCP, indicating hysteresis of PCP sorptions. The retardation factors were 1.03 for phenol, 1.16 for 2,4DCP and 2.26 for PCP. In batch biodegradat ion studies using indigenous soil bacteria phenol degraded quickly (t.,_ = 12 hours) and was completely destroyed within three days. 2,4-DCP was also completely degraded but had taken 23 days (t,/_ = 7 days) . PCP was resistant to biodegradat ion with an average halflife of 120 days. In multi-compound systems, phenol degradation rates dropped off to 0.4 day (t , = 1.7 days) but PCP degradation rates increased to 0.008 day (t . = 86 days) . Biodegradat ion rates in column studies were obviously greater than in batch experiments, with the rate increase for PCP degradation being especially noticeable (t , = 12 days) , because of larger bacterial populations and the dynamic flow conditions made the substrates more available to the bacteria. When controlled under an aerobic environment by the addition of hydrogen peroxide, all three phenolic compounds degraded fastest. under anoxic conditions both the microbial population buildup and the rate of phenolic compound degradation were slower but not by a wide margin.

PAGE 14

Bacterial growth in the columns did not reduce the hydraulic conductivity of the system, indicating the feasibility of applying in-situ biodegradat ion techniques to groundwater contamination problems.

PAGE 15

CHAPTER I INTRODUCTION Groundwater contamination by trace organic compounds is a widespread problem but only recently has the public become aware of the seriousness of this problem. The problem is serious largely because groundwater does not have the selfcleaning mechanisms commonly seen in surface water, and because with an increasing dependency, about half of the population in the United States are now depending on groundwater for drinking. In some areas such as Florida the dependence on groundwater is more than 90 percent of the population, and the demand of groundwater supply is expected to increase 25 percent per decade (DeHan, 1981). Groundwater contamination results from various types of sources, such as disposal of hazardous wastes into unl ined landfills, accidental spills of chemicals and leakage of underground storage tanks. Industry-related sources include chemical leaks from storage areas, accidental spills, and vapor condensate from solvent-recovery systems. Non industr i al sources include road runoff, municipal landfills, junk yards, septic tanks, and domestic waste water. Numerous organic chemicals have been detected in groundwaters as contaminants nationwide. Phenol and substituted phenols, which are some of the most frequently

PAGE 16

found organic chemicals, are accountable for many of the groundwater contamination cases (Plumb, 1985; Pye and Patrick, 1983) . This is especially true in the southeastern United States because of the high concentration of woodpreserving industry located in this region and the wide use of these chemicals in this industry. To properly assess a groundwater contamination problem, it is necessary to understand the transport and fate of the contaminants in the subsurface environment. Once the contaminants enter the system, their transport and fate are determined by the chemical, physical, and biological properties of both the chemical compounds and the aquifer materials. Dilution (advection and dispersion), sorption (adsorption and desorption) , and degradation (biotic and abiotic) , are three major forces governing the fate of the contaminants (Mackay et al . , 1985; Newsom, 1985). The onedimensional equation proposed by Bear can be used to describe these phenomena (Bedient et al . , 1985; Skopp et al . , 1981) : 3 C 9 c a c dS °L ~ 2 * (1-1 3 X 3 X 3t where C = aqueous phase concentration of compound (M/L ) t = time (T) 2 D = longitudinal dispersion coefficient (L /T) = a*v a = d isper s i v i ty x = distance in flow direction (L) v = average seepage velocity (L/T)

PAGE 17

3 p = density of bulk dry soil n = poros i ty S = adsorbed phase concentration of compound (M/M) The terms on the right hand side of Equation (1-1) are referred to as dispersive transport, convective transport, and adsorption, respectively. in a linear adsorption the distribution coefficient K,= S/C, since gS/ at = K *(sC/at), Equation (1-1) can be written as 3 C 8 2 C dC R = D * ---rV* (1-2) 3 t L dx l 3x p * K R = 1 + (1-3) n where R is the retardation factor. If degradation (decay) is incorporated then the equation becomes 3 C 3 2 C 3C R --= D L * — 2 v* K BD * C (1-4) 3 t 3x 3x where K Dr ^ is the degradation rate (1/T). Many numerical solutions have been presented by various researchers (Amoozegar-Fard et al., 1983; Fuller and Warrick, 1985; Van Genuchten, 1981). As an example, if given the boundary conditions C=C at x=0 and C=0 at x=infinity, and with an initial condition C=0 at t<0, Equation (1-4) can be solved by a numerical solution proposed by Sauty (1980): C = C /2 { exp[ (v-u) x/2D] er f c [ ( Rxut ) / ( 4DRt ) 1/2 ] + exp[ (v+u) x/2D] erfc[ (Rx+ut) /(4DRt) 7 ] } (1-5) 2 1/2 where u = (v +4 k OT ^ R D) . With this solution and the

PAGE 18

required parameters, the fate of contaminants in groundwater can be predicted. Sorption is a measure of partition between the aqueous phase and solid phase in the aquifer. It is known to be important to the fate and transport of organic compounds in groundwater systems. The degradation term of Equation (1-4) may be contributed by photolysis, hydrolysis, abiotic oxidation and biotic degradation. Among these pathways biodegradat ion is the most important degradation process in groundwater systems for phenolic compounds. Indigenous microorganisms can utilize organic compounds as carbon sources to generate energy for their maintenance requirement and increase cell mass, provided that adequate nutrients and electron acceptors are available. When the concentration of a contaminant becomes very low (which is not unusual in groundwater) , the microorganisms may not be able to derive enough energy to support the maintenance requirement. If this condition occurs, the population will decline, and consequently the organic compounds may persist at trace concentrations (Alexander, 1981, 1985) . A system that was developed to simulate a Florida sandy aquifer in a natural environment was used to evaluate the adsorption and desorption coefficients and biological degradation rates of phenol, 2 , 4-d ichlorophenol , and pentachlorophenol . With the data from this research, the behavior and fate of these important phenolic compounds in a shallow Florida

PAGE 19

sandy aquifer can be better predicted, and thus lead to the development of treatment methods for remediating aquifers contaminated by such phenolic compounds.

PAGE 20

CHAPTER II OBJECTIVES The objectives of this study were as follows: (1) To determine estimates of the sorption parameters for phenol, 2 , 4-dichlorophenol and pentachlorophenol when present alone and in mixtures in a sandy Florida soil. (2) To evaluate the rates of biodegradat ion of these phenolic compounds when present alone, and when present as mixtures under simulated field conditions in a sandy Florida so il . (3) To determine the effects of co-degradation, enzyme induction and sludge amendment on in situ biological treatment of groundwaters that are contaminated with phenolic compounds under simulated field conditions. (4) To evaluate the change of hydraulic conductivity of soils before and after in situ biological treatment under simulated field conditions.

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CHAPTER III LITERATURE REVIEW This chapter presents a review of the pertinent literature about the characteristics and the environmental significance of selected phenolic compounds and their sorption and degradation processes known to occur. 3.1 Environmental Significance of the Phenolic Compounds Phenol, 2 , 4-d ichlorophenol (2,4-DCP), and pentachlorophenol (PCP) are chosen as the contaminants in this study because phenolic compounds are commonly found contaminants in groundwater . This is especially true in Florida and the southeastern United States because of the high density of wood-preserving industries in this area of the country. Among the phenolic compounds, phenol, 2 , 4-dichlorophenol and pentachlorophenol have the most commercial importance (Goldf arb et al . , 1981) . Phenol was first isolated from coal tar in 1834, (Moore and Ramamoorthy, 1984) , but today almost all phenols are manufactured by the cumene hydroperoxide process (Kirk and Othmer, 1985) . It has been used in many commercial products including resins, nylons, plast ici zers , antioxidants, oil additives, polyurethanes , drugs, pesticides, explosives,

PAGE 22

dyes, and gasoline additives. In 1981 alone, more than 1.15 million metric tons of phenol were produced in the United States (U.S. International Trade Commission, 1982). All 17 possible chlorinated phenols are commercially available. Monochlorophenol s are used mainly in the production of higher chlorinated phenols. 2,4-DCP is used primarily in the manufacture of the widely used agricultural pesticide 2,4d ichlorophenoxy acetic acid (2,4-D) . When 2,4-D breaks down, 2,4-DCP will be present as one of the products (USEPA, 1986) . Pentachlorophenol has been extensively used as a wood preservative because of its fungicidal properties. Phenol is fairly soluble in both water and nonpolar solvents as shown in Table 3-1. Alkaline salts of phenol are also readily soluble in water. Generally the volatility and the aqueous solubility decreases with the increasing number of chlorine atoms on the benzene ring. Electron withdrawal by the ring chlorines causes pentachlorophenol to be acidic and a relatively weak nucleophile, while making its salts fairly stable. Physical properties of selected phenolic compounds are listed in Table 3-1. The organoleptic properties of the chlorophenol s are manifested by imparting odor to water and tainting fish flesh (Lee and Morris, 1962). As a group, the chlorophenols are highly toxic. Although insufficient information exists on the carcinogenicity of most chlorophenols, 2 , 4 , 6tr ichlorophenol has been shown to be an animal carcinogen, and para-chlorophenol is a suspected carcinogen based on mutagenicity

PAGE 23

screening tests (Moore and Ramamoorthy, 1984). Accordingly, phenol, 2,4-DCP and PCP are listed as priority pollutants by the U.S. Environmental Protection Agency (USEPA, 1979). Table 3-1. Physical properties of the phenolic compounds (Verschueren, 1977; USEPA, 1979) Parameters

PAGE 24

10 and desorption in general. Sorption is an important factor in the determination of the fate of hydrophobic compounds, ("hydrophobic compounds" is defined as compounds with Kow value greater than 5.0) (Doucette and Andren, 1987) , in a water/soil system. Adsorption tends to retard the migration rate of contaminants in subsurface environment. It may provide precious time to respond to accidental spills before the contamination spreads. On the other hand, soils that slowly desorb contaminants will become constant sources of groundwater contamination (Delfino, 1977; Delfino and Dube, 1976) , greatly prolonging the time required for an effective cleanup, and increasing the cost of remedial actions. Various reports indicate that the equilibrium relationship between soil and solution phase solute concentrations was found to be described best by the nonlinear Freundlich isotherm model ( Art iola-For tuny and Fuller, 1982; Boyd, 1982; Lagas, 1988; Laquer and Manahan, 1987; Means et al . , 1980; Miller and Weber, 1986), which is expressed as (3-2) where X is the mass of solute adsorbed to soil surface, m is the mass of soil, C is the solute concentration at equilibrium in the aqueous phase, and K_ and b are F constants. Freundlich sorption coefficient (K^) is a F measure of the degree of strength of adsorption, while b is an indication of whether adsorption capacity remains constant, i.e. when b=l, sorption is linear within the range X/m = K c r

PAGE 25

11 of solution concentrations used in a particular study. Note that in this case the equilibrium Freundlich partition coefficient, K , is the same as the distribution F coefficient, K , , and X/m equals S in Equation (3-1). Neither Freundlich partition coefficients nor distribution coefficients are universally transferable because they depend heavily on both the properties of chemical compounds and the characteristics of the soil matrix. Enormous efforts have been devoted to making these relationships more useful and easier to apply to soils with different characteristics. Karickhoff et al . (1979) demonstrated that for a dilute solution (i.e. concentration of the contaminant less than half of its solubility in water) , partition coefficients based solely on organic carbon in the soil matrix, K , correlate closely to K„/f as oc' J F oc K K_ / f (3-3) oc F oc where f is the fraction of organic carbon in the soil oc matrix. This relationship ignores any influence of the soil itself but does facilitate the use of partition coefficients or distribution coefficients from the literature as long as the fraction of organic carbon in the soils are documented. Chiou et al . (1979) and Karickhoff et al . (1979) reported that K could be related to water solubility. They also oc J J reported a relationship of the octanol/water partition coefficient, K (ml/g) , as ' ow * ' log K = log K 0.21 (3-4) oc ow log K = 5 0.67 log WS (3-5) 3 oc 3

PAGE 26

12 WS is the solubility of a chemical in water in umol/1. These relationships are convenient to use since values for solubility and the octanol/water partition coefficient are either well established by various workers or easy to measure in a laboratory. However, these relationships all have their limitations. Banerjee et al . (1980) suggested that for compounds with high melting points, Equation (3-5) may be invalid, and proposed another correlation between K 1 ' v v ow and WS which incorporated a melting point correction term as log K = 6.5 0.89 ( log WS ) 0.015 ( MP ) (3-6) ow where MP is the melting point in degrees centigrade. Rao and Jessup (1983) cautioned that Equations (3-3), (3-5) and (3-6) may not apply to soils containing less than 0.1 percent of organic carbon. Both pH and ionic strength have significant influence on sorption of phenolic compounds. Schellenberg et al . (1984) showed that sorption of the unionized phenols and their conjugate bases (phenolates) can occur. They suggested that in natural waters of low ionic strength (i.e. ionic strength < 10~ M) and of pH values not greater than the pK values of the phenolic compounds by one unit, phenolate sorption can be neglected. Based on this theory, the conjugate bases of phenol and 2,4-DCP need not be considered in the natural environment. Phenol has a relatively small K value (log K =1.46) , 1 ow ow which suggests only a slight tendency to become adsorbed onto the organic detritus. As a comparison, PCP has a low

PAGE 27

13 water solubility (14 mg/1 at 20 C) and a higher K value of 2 ow 5.01, where imply a strong tendency for PCP adsorption onto organ ic matter . Laboratory experiments have shown a phenol desorption of almost 100% from a thin layer of montmor i 1 Ion i te clay exposed to 40% humidity for one week (Moore and Ramamoorthy, 1984). But Isaacson and Frink (1984) reported that phenol, 2-chlorophenol and 2,4-DCP were extensively sorbed onto sediments, desorption was slower than adsorption, and in some cases up to 90% of the sorbate was irreversibly held. This contradiction may have been caused by differences in the reaction pH, ionic strength, and the percent organic carbon content of the sorbents in the two experiments. Hydrophobicity (defined as the lack of the capacity of a compound to dissolve in water) as indicated by Kow is not the only factor controlling the sorption of phenolic compounds (Boyd, 1982; Isaacson and Frink, 1984), hydrogen bonding may also play an important role. Boyd (1982) suggested that the phenolic hydroxyl group formed hydrogen bonds by acting as a proton acceptor. Westall et al . (1985) found that the more highly substituted chlorophenol s are subject to larger influence by ionic strength. Because the sorption of molecular pentachlorophenol is much greater than of ionized pentachlorophenolate , he concluded that pH and ionic strength play more important roles in PCP sorption to soil than in the less substituted compounds. Kaiser and Valdmanis (1982) reported a wide range of K values for PCP c 3 ow

PAGE 28

14 from 4.84 at pH 1.2 to 1.30 at pH 10.5 and pH 11.5. This higher partition coefficient at lower pH suggests a greater affinity for the organic part of the soil as the pH decreases . A number of areas of research in the region of sorption chemistry remain controversial, such as reversibility, extent of reversibility, rate of attaining equilibrium, and the effect of competitive solutes in sorption equilibria. In many cases sorption is considered to be reversible (Angley, 1987). However, Miller and Webber (1984) reported that many researchers disagree about the reversibility of sorption with various chemicals. Laquer and Manahan (1987) reported that the sorption of phenol onto a siltstone showed differences in adsorption and desorption isotherms, an effect termed hysteresis. Rogers et al . (1980) found that once sorbed, benzene tends to resist desorption. Van Genuchten et al . (1977) suggested that values of the equilibrium desorptive constant should be different from that of adsorption while Equation (3-2) holds for both cases. Nathwani and Phillips (1977) drew the same conclusions on some hydrocarbons in crude oil, and found that the percentage of hydrocarbon component desorbed varied inversely with the amount of organic matter in the soil matrix. Researchers have also shown diverse results about the rate of attaining equilibrium. Miller and Webber (1986) observed equilibrium occurring after several days for nitrobenzene and lindane. Rao and Davidson (1980) found

PAGE 29

15 that sorption reactions for many organic compounds were 60 to 80% completed within one minute. Ogram et al . (1985) stated that greater than 98% of the 2,4-D sorbed at equilibrium was sorbed within the first five minutes, and Means et al . (1980) reported that equilibrium for some polynuclear aromatic hydrocarbons was achieved in 20 hours or less. According to this evidence, it is reasonable to expect the sorption of phenolic compounds onto sandy soils to reach equilibrium within 24 hours. In cases when more than one compounds are present in a mixture, the effect of competitive sorption should be considered. Theoretically, if these compounds have similar Kow values, it is likely that they will compete with each other for sorption sites unless the concentration of these compounds is low and the sorption surfaces relatively high. This phenomenon is called competitive sorption (Kinniburgh, 1986). When the compounds have very diverse partition coefficient values, an increase in water solubility has been observed for the more hydrophobic compound as a result of cosolvent effects. This results in a decrease of adsorption or an increase of desorption of the affected compounds. This effect may be directly related to a compound's availability for biodegradat ion (Thomas et al . , 1986). Whether sorption will enhance or decrease microbial degradation rates in groundwater depends upon whether the sorbed phenolic compounds are available to the microbes. When contaminants are irreversibly sorbed to soil organic

PAGE 30

16 matter, they are isolated from the degrading organisms and are protected from intracellular degradation. On the other hand, bacteria may also be sorbed. If bacteria and contaminants are sorbed on adjacent sites on the soil surface, the uptake of the contaminants by the sorbed bacteria is facilitated (Ogram et al . 1985) . Isotherm models can be used to predict the sorption and desorption behavior of the contaminants, and thus help to design groundwater/so i 1 reclamation programs. Although equilibrium may not be reached in reality, the prediction may serve as a guide to the direction of mass transfer. 3 . 3 Degradation of Phenolic Compounds Many processes can contribute to the degradation of phenolic compounds in the environment. Among these are photolysis, chemical oxidation, hydrolysis, volat i 1 i zat ion and biodegradat ion . Each process needs special conditions in order to proceed and has its own role in the degradation of these compounds from the subsurface environment. 3.3.1 Photolysis Phenol has long been known to form reddish high molecular weight material when exposed to sunlight and air. It can undergo photolysis either in the phenolate anion form (maximum absorbance at 270 nm) or in the und i ssoc iated molecule (maximum absorbance at 310 nm) . Experimental irradiation of phenol at 254 nm in the presence of oxygen yields a phenoxy radical intermediate that subsequently give

PAGE 31

17 substituted biphenyls, hydroquinone (m-dihydroxy benzene), and catechol (o-dihydroxy benzene) (Moore and Ramamoorthy, 1984) . Photolysis of phenol to hydroquinone occurs under both natural sunlight and commercial sun lamps (USEPA, 1979) . Assuming a first order reaction, the rate of disappearance of an organic compound by direct photolysis from surface water is (3-7) -dC/dt = K [C] = k i_ (e qz ) [C] p where C is the concentration of the compound, K is the apparent first-order photolysis rate constant, k is a constant of proportionality which includes the quantum yield of the reaction, I is the solar radiation intensity at photochemically active wave lengths incident on a water surface, q is the extinction coefficient of the water (which is a function of dissolved and particulate absorbers), and z is the depth (Pignatello et al . , 1983). Equation (3-7) can be converted into a mathematically calculable form: In ( C / Co ) = k I (e" qZ ) [C] (3-8) One EPA report (1979) stated that 2,4-DCP and PCP do undergo photolysis but its significance and environmental importance is uncertain. However, Hwang et al . (1986) indicated that in summer time K values for 2,4-DCP and PCP P were 1.0 and 0.37 h , respectively at a depth of 3cm while compared to 0.016 h~ for phenol. A similar result for PCP photolysis was reported by Pignatello et al . (1983). Crosby (1981) concluded that in either water or organic solvents,

PAGE 32

PCP can be photolyt ical ly reduced to isomeric triand tetrachlorophenols , and, in dilute aqueous solutions exposed to sunlight, PCP or its salts undergo the replacement of ring chlorines by hydroxyl groups to form corresponding chlorohydroquinones , which are subsequently oxidized to chlorobenzoqu i nones and then dechlor inated and/or ring cleaved. Pentachlorophenol is a moderately acidic compound and thus will exist primarily as an anion in natural waters. This is environmentally significant because the anion absorbs well beyond 310 nm (sunlight spectrum) leading to more effective photolytic reactions. Wong and Crosby (1978) reported that the rate of photolysis of pentachlorophenolate anion was much faster than that of the undissociated compound . For the phenolic compounds in groundwater no photolysis occurs naturally. However, this process can be useful in a remedial action when spraying and recirculating is involved, and needs to be considered as an option when performing a feasibility study in a groundwater reclamation project. 3.3.2 Ox idat ion Phenol has been oxidized by passing molecular oxygen into an aqueous solution at 25 C and pH 9.5-13. This suggests a possibility of nonphotol yt ic oxidation in highly aerated waters. Little information is available pertaining to the oxidation of chlorinated phenols but usually highly

PAGE 33

19 chlorinated organic compounds are resistant to oxidation under natural environmental conditions (USEPA, 1979) . 3.3.3 Hydrolys is The rate of hydrolysis of a chemical compound can be calculated by -dC/dt = k_ [H + ] [C] + k_ [OH ] [C] + k M [C] (3-9) A B N where k and k =second-order acid and base hydrolysis A B constants, respectively; and k = first-order hydrolysis rate constant for pH independent reactions (Moore and Ramamoorthy, 1984). The covalent bond of a substituent attached to an aromatic ring is usually resistant to hydrolysis because of the high negative charge density of the aromatic nucleus. Therefore, hydrolysis of phenolic compounds in a natural groundwater environment will not be a significant process (USEPA, 1979; Moore and Ramamoorthy, 1984). 3.3.4 Volat i 1 i zat ion The rate of volatilization for general organic compounds is described by Smith et al . (1980) as -dC/dt = k v [C] = C/L [l/k 1 + R T/H c kl" 1 (3-10) where k = volatilization rate constant (hr ) L = depth of aqueous layer k = transfer coefficient in the liquid phase (cm/hr) H = Henry's law constant (torr/M) c J k = transfer coefficient in the gas phase (cm/hr) g

PAGE 34

R = ideal gas constant T = absolute temperature. The low vapor pressure and the high aqueous solubility of phenol indicates that there is little tendency for volatilization from water. Chlorinated phenols are less soluble in water, but the higher acidity increases the proportion of the ionized form (which is much less volatile than its unionized counterpart) in the natural environments and causes them to be highly solvated. Thus, volatilization will not have a significant contribution for loss of most chlorophenol s in aquatic environments. 3.3.5 Biodegradat ion As early as 1946 Claude E. ZoBell (ZoBell, 1946) reported that more than 100 species representing 30 microbial genera had been shown to have the ability to utilize organic compounds as carbon and energy sources, and that such microorganisms are widely distributed in nature (Atlas, 1981). Bartha and Atlas (1977) listed 22 genera of bacteria, one algal genus and 14 genera of fungi that had been demonstrated to contain members which utilize petroleum hydrocarbons. All of these microorganisms were isolated from an aquatic environment. In soil samples Jones and Eddington (1968) found that 11 genera of fungi and six genera of bacteria were the dominant microbial genera responsible for hydrocarbon oxidation.

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21 Ghiorse and Balkwill (1983) found 5x10 microbes per gram of dry subsurface material by direct count using epi fluorescence microscopy. This result is very similar to what Wilson et al . (1983) have found, 3xl0 6 to 9xl0 6 microbes per gram of dry material, in soils taken from various depths below the surface of the ground. They and others further showed that those microorganisms can degrade several hydrocarbons (Stetzenbach et al., 1985; Roberts et al., 1980; Yaniga, 1982). Bouwer and McCarty (1985) reported that 91% of chlorobenzene can be biodegraded from a concentration of 11 ug/1 by a biofilm grown with 1 mg/1 of acetate after a 20 day acclimation period. In their studies ethylbenzene was also cometabol ized with acetate as a secondary substrate. Tabak et al . (1981) have done a series of biological degradation studies with organic priority pollutant compounds under aerobic conditions. They followed a staticculture flask-screening procedure with settled domestic wastewater as microbial inoculum, and found that, at a concentration of 5 mg/1, phenol, 2,4-DCP and 2,4,6-TCP can be biodegraded 60 to 100% with rapid acclimation while PCP showed only 19% reduction after seven days of incubation. When the concentration was increased to 10 mg/1, a slight decline in the rates of degradation was observed. Brown et al . (1986) also found that 600 mg/1 of ionized PCP can be continuously biodegraded without affecting steady-state growth in a fixed-film bioreactor containing a PCP-adapted

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22 Flavobacter iura. On the contrary, Klecka and Maier (1985) reported that PCP degradation was inhibited at much lower concentrations (800-1600 ug/1). Watanabe (1973a, 1973b) examined PCP degradation in soil perfused with 40 ppm of PCP and observed, after an eight day lag period during which essentially no degradation occurred, chloride ion liberation was initiated, and was complete within three weeks. Subsequent additions of PCP were degraded more rapidly with no lag period. Most of these degradation studies were conducted under aerobic conditions. Boyd and Shelton (1984), Smith and Novak (1987), and Ehrlich et al . (1982) demonstrated that chlorophenols can also be degraded anaerobically . However, rates of anaerobic degradation for most organic contaminants are significantly slower than those under aerobic conditions (Delfino and Miles, 1985), and the anaerobic reductive dechlorination of PCP seemed to stop at 3 ,5-dichlorophenol (Mikesell and Boyd, 1985). Increased chlorination of the phenolic compounds increased stability to oxidation and enzymatic degradation (Cserjesi, 1967), therefore, highly chlorinated phenols tend to be more resistant to degradation. Many factors can influence the rate of biodegradat ion , such as temperature, genus of the microorganisms, nutrients, electron acceptor, pH , soil matrix, chemical concentration of the compounds, and enzymes. Temperature . Although biodegradat ion can occur over a wide range of temperatures, temperature greatly influences

PAGE 37

23 the rate of biodegradat ion . Within the ambient temperature range, rates of biodegradat ion are faster at higher temperatures than at lower temperatures. ZoBell (1969) found that hydrocarbon degradation was over an order of magnitude faster at 25° C than at 5° C. Larger microorganism populations as well as higher assimilation rates at higher ambient temperatures both contribute to this increase. Vela and Ralston (1978) found that at higher temperatures more phenol was metabolized per cell than was required to support growth. A modified Arrhenius mathematical model is available to estimate the effects of temperature on biodegradat ion rate constants K 2 = Kl g (T2 " T1) (3-11) where K and K_ are the rate constants at temperature T and T_ respectively, and g is a coefficient. Typical values for g are from 1.01 to 1.04 in wastewater treatment systems (Benefield and Randall, 1980). Genus of microorganisms . Many microorganisms in the natural environment are capable of degrading organic compounds. Although the microorganisms may prefer some particular compounds, they can rapidly adapt in order to utilize available substrates (Hollibaugh, 1979; Haller, 1978; Hutchins et al . , 1984). Spain et al . (1984) found the microorganisms in a pond were successfully acclimated to degrade p-n i trophenol after a 6-day lag period. Healy and Young (1979) indicated that microbial populations acclimated to a particular compound

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24 can be simultaneously acclimated to other compounds, and that a microbial population can metabolize several compounds at the same time. However, Shimp and Pfaender (1985a, 1985b) reported the organic substances to which the microorganisms have already been exposed can significantly influence the ability of microorganisms to degrade other organic compounds. They observed that exposure to amino acids, carbohydrates or fatty acids enhances the ability of microorganisms to degrade certain phenolic compounds while exposure to humic materials had a negative effect. More than 25 species of microorganisms were reported capable of degrading PCP (Engelhardt et al., 1986). They were isolated from soils, municipal wastewater sludges, surface waters and groundwaters. Among these microorganisms Arthrobacter , Trichoderma virgatum, Flavobacter ium sp . , and Pseudomonas were most reported (Brown et al . , 1986; Crosby, 1981; Edgehill and Finn, 1983; Kaufman, 1978; Mikesell and Boyd, 1985; Stanlake and Finn, 1982; Steiert et al . , 1987; Suzuki, 1975; and Suzuki, 1977). Nutr ients . Microorganisms need nitrogen, phosphorus and some trace minerals for incorporation into biomass, so the availability of these nutrients is critical to biodegradat ion . A general formula for microorganism composition was proposed as C,„H o ^0_ , N. P (Benefield and 60 87 23 12 Randall, 1980). This formula reveals a C:N:P ratio of 23:5.3:1 in microorganism cells. However, optimal C:N and C:P ratios for marine oil-degrading microorganisms were MHB

PAGE 39

25 found to be 10:1 and 100:1 respectively (Atlas, 1981). Since carbon is utilized for both energy (non-growth) and synthesis requirements (growth) while nitrogen and phosphorus are used essentially for synthesis of new cells, the optimal N:P ratio is somewhat less variable than the C:N and C:P ratios, and 10:1 seems to be a reasonable value to choose when supplying nutrients to microorganisms. The optimal C:N or C:P ratio will need to be determined experimentally for each specific case, because they are largely dependant on the carbon-energy conversion efficiency of the tested microorganisms. Dibble and Bartha (1979) indicated that a C:N:P ratio of 800:13:1 was found to be optimum and cost-effective for oil sludge biodegradat ion in a " landf arming" process, but this ratio is far removed from the theoretical values. They also reported that addition of micronutr ients and organic supplements (such as yeast extract) were not beneficial to biodegradat ion . The form of phosphorus or nitrogen is not critical for the growth of microorganisms. However, it has been recommended that an ammonia-nitrogen source is preferable to a nitrate-nitrogen source because ammonia-nitrogen is more easily assimilated by microorganisms (USEPA, 1985) . Kaufman (1978), on the other hand, stated that yeast extracts accelerated PCP degradation, whereas glucose at 100 ppm suppressed degradation, and the substitution of ammonium sulfate for sodium nitrate as a nitrogen source also suppressed degradation. Because Kaufman studied the

PAGE 40

26 degradation of different chemical compounds, the responsible microorganisms could have been totally different from the experiment described in the EPA's report. Electron acceptor . Oxygen is required as an electron acceptor in the energy metabolism of the aerobic heterotrophic organisms. A portion of the organic material removed is oxidized to provide energy for the maintenance function (non-growth) and another portion for the synthesis function (growth). Any oxidation must be coupled with reduction. Oxygen satisfies this requirement in aerobic biodegradation. In traditional wastewater treatment, a minimum of 2 mg/1 dissolved oxygen (D.O.) concentration is required for aeration equipment to ensure a sufficient oxygen supply. Because the microorganism concentrations in groundwater are far less than those in an aeration tank, a lower residual D.O. requirement should be expected. Borden et al . (1984) have found that 0.25 mg/1 seemed to be a threshold D.O. concentration in groundwater for napthalene degradation. In many cases the rate and extent of biodegradation of many organic materials in a subsurface environment appear to be limited by the availability of oxygen. Yaniga and Smith (1985) reported that instead of traditional aeration, dilute hydrogen peroxide is a good alternative for elevating dissolved oxygen concentration in groundwater. Hydrogen peroxide decomposes to oxygen and water. In the subsurface, hydrogen peroxide decomposition is catalyzed by chemical and

PAGE 41

27 biological factors. The decomposition can occur so rapidly that oxygen bubbles out near the point of injection and oxygen is not made available to the distant portions of the needed zones. Research has shown that a high concentration (10 mg/1) of phosphates can stabilize hydrogen peroxide for prolonged periods of time in the presence of ferric chloride, an aggressive catalyst for the decomposition. However, such a high concentration of phosphate may cause precipitation problems and render the soil impermeable. Another problem is that hydrogen peroxide is cytotoxic, but research has demonstrated that it can be added to some cultures at up to 1000 mg/1 concentration without toxic effects (USEPA, 1985). Yaniga and Smith (1985) reported a successful aquifer restoration project using 100 mg/1 of hydrogen peroxide as a dissolved oxygen source. pH . Dibble and Bartha (1979) found that a pH of 7.5 to 7.8 was best for oil sludge degradation. This coincides with the optimal pH for most microorganism growth. The pH also has a great influence on the ionization of phenolic compounds, since at higher pH conditions phenolates are predominant, causing a decrease in adsorption and/or an increase in desorption. This phenomenon is more significant to higher chlorinated phenolic compounds. The effects of ionization on biodegradat ion of phenolic compounds are not clearly understood. pH also influences toxicity of phenolic compounds. Unionized PCP is apparently more toxic to both fish and

PAGE 42

28 microorganisms than its ionized salts (Stanlake and Finn, 1982). Typical groundwater has pH values of 6.0 to 8.0 (Davis and Dewiest, 1966), however, 5.5 is a more common pH value for surfacial aquifer waters in Florida. Laboratory biodegradat ion studies should be performed in the same pH range as that of groundwater before field application of this treatment technique. Chemical concentrations . Concentration of the compound may be a significant factor which affects its susceptibility to microbial attack. Organic compounds may persist in some environments as a result of low prevailing concentration or low solubility in water (Thomas et al., 1986). For example, evidence exists that solubility limits the rate of bacterial growth using a series of polycyclic aromatic compounds, and some normally biodegradable substrates may not be metabolized when the compounds are present at concentrations lower than that required for maintenance of the microorganisms (i.e. the threshold concentration) (Boethling and Alexander, 1979; Bouwer , 1984). Rittmann and McCarty (1980a, 1980b) have reported that the threshold concentration, C . , can be evaluated by the min relat ionship : C . = K * k Dr> / (Y* k k__) (3-12) min s BD BD where K is the Monod Half-maximum-rate concentration, k s dU is the first-order decay constant, k is the maximum specific rate of substrate utilization by the microorganisms, and Y is the cell growth yield.

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29 Many organic contaminants in groundwater are present at concentrations below C . and would apparently qo rain c,r 2 3 unutilized. However, simultaneous utilization of several different substrates is possible. Sometimes microorganisms can metabolize these trace compounds in the presence of other substrates, called primary substrates which support the long-term biofilm growth. This process is termed secondary utilization, or cometabol ism . it is a mechanism which allows microorganisms to degrade compounds that could otherwise not provide enough energy to sustain the microbial culture (McCarty et al . , 1981). Studies have shown that the extent of biodegradat ion of polychlor inated biphenyls was enhanced by adding sodium acetate as a primary carbon source. The effect was especially significant on higher-chlorinated isomers (Clark et al., 1979). Marinucci and Bartha (1979) also found a slight stimulation of 1 , 2 , 4tr ichlorobenzene mineralization was caused by the addition of primary substrates. Schmidt and Alexander (1985) observed that the presence of acetate has a negative effect on phenol degradation, and the delay was lengthened by increasing acetate concentrations because acetate is easier to degrade than phenol. Bouwer and McCarty (1985) suggested that secondary substrate (i.e. target contaminants) removal rates increase with time but not with the increase of primary substrate concentrations beyond a limiting concentration, and the overall residual

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30 concentration of the target contaminants can be largely reduced by cometabol i sm . Laboratory biodegradat ion experiments should use concentration ranges similar to those actually found in the field or the results may not correctly reflect what will take place in the field (Alexander, 1985; Wang et al . , 1984) . Soil matr ix . Soil matrix affects the biodegradat ion of phenolic compounds mainly as a function of the organic matter in the soil which contributes to the adsorption. It is unclear whether the compounds that have been sorbed to the soil particles are subject to biodegradat ion . The compounds may be biodegraded while sorbed to the soil, or, as the aqueous concentration decreases as a result of biodegradat ion , some of the sorbed compounds may be desorbed to restore equilibrium, and then be available for biodegradation (Smith and Novak, 1987). Ogram et al . (1985) indicated that sorbed 2,4-D was completely protected from biodegradation by both sorbed and suspended bacteria. No equivalent data were found on phenolic compounds. However, Crosby (1981) stated in his review paper that PCP degraded faster in soils with high rather than low organic content. Enzymes . The biodegradation of phenolic compounds was shown to be highly responsive to enzyme induction, yet, it is a topic little studied. Not many enzymes that are responsible for degradation of phenolic compounds have been

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31 isolated (USEPA, 1986). The enzymes necessary for PCP degradation appeared to be inducible. Steiert et al. (1987) demonstrated that a suspension of cells grown in the presence of 2 , 4 , 6tr ichlorophenol or 2 , 3 , 5 , 6tet rachlorophenol did not show a lag period for degradation of 2,4,6tr ichlorophenol , 2 , 3 , 5 , 6tetrachlorophenol or PCP, indicating that one enzyme system can be induced for the biodegradation of multiple compounds. Chu and Kirsch (1973) and Karns et al . (1983) also reported similar observations. 3.3.6 Pentachlorophenol Degradation Mechanisms Pentachlorophenol is very resistant to biodegradation and may produce less chlorinated phenols as the degradation products, therefore, its degradation pathway deserves more study. Three significant mechanisms appear to account for the biological degradation of PCP in soils: (1) reductive dechlorination; (2) oxidative dechlorination; (3) methylation. Conceivably an aggregate of microorganisms should be more efficient in mineralizing phenolic compounds (to C0_) than any of the pure cultures. The structural formulas of those involved compounds are presented in Figure 3-1. Reductive dechlorination . Bacteria such as Flavobacter ium sp. can utilize PCP as a sole source of carbon and energy. Thus reductive dechlorination under anaerobic conditions forms less chlorinated phenolic

PAGE 46

32 Pentachlorophenoi Oh I n "etrachiorocatechoi OH CI Tetrachiorohydroquinone Pentach.ioroanlsoie 'v^Tetrachlorobsnzoquinone Figure 3-1. Structural formulas for some common PCP degradation products.

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33 compounds. But this process seemed to stop at isomeric tr ichlorophenols (Weiss et al., 1982). Oxidative dechlorination . Watanabe (1973a) and Suzuki (1977) found Pseudomonas sp. is capable of oxidizing PCP to C0 2 along with chlorohydroquinones and chlorocatechols as intermediate metabolites. Because chlorohydroquinones and chlorocatechols are less toxic to fungi than PCP, this mechanism can be considered a detoxifying process (Engelhardt et al., 1986). Methylat ion . Fungus Trichoderma virgatum and bacterium Arthrobacter sp . are the most commonly seen microorganisms that methylate PCP and other chlorinated phenolic compounds to form the corresponding chloroanisoles (Cserjesi and Johnson, 1972). Chloroanisoles are also reported to have less toxic effects on microorganisms than the corresponding chlorinated phenolic compounds (Weiss et al., 1982) . 3 . 4 Summary This section reviewed the processes that are most likely to occur in the natural environment to degrade phenols. It also supports the feasibility of enhanced biodegradat ion as a treatment method for chlorinated phenolic compounds in groundwater systems.

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CHAPTER IV MATERIALS AND METHODS This chapter discusses the materials, analytical methods and experimental design employed in this research. 4 . 1 Mater ials 4.1.1 Soil The sandy soil used in these studies came from an unused cell at the north pit of the Southwest Landfill, Archer, Florida. This sandy soil is representative of soil conditions in surficial aquifers which supply drinking water in large areas of Florida. Advective fluxes tend to be greater through granular horizons of this type than through other soil formations, thereby facilitating contaminant transport over wide geographic areas. 4.1.2 Chemicals Three chemicals involved in this study: phenol (Fisher Scientific Supplies, 92.9%), 2 , 4-d ichlorophenol (Eastman Kodak Co.) and pen tachlorophenol (Aldrich Chemical Co., Inc., 99%) were purchased and used without further purification . 34

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35 4.1.3 Contaminated Water Contaminated water was prepared by dissolving chemical compounds into distilled deionized water. This approach was chosen because it is easier to control concentrations and would tend to have a consistent characteristic throughout the course of the study. 4.1.4 Microorgani sms The microorganism seeds were taken from the return sludge in the aeration tank of the University of Florida's wastewater treatment plant. Unless otherwise specified, the sludge was used without further treatment. The supernantant of sludge, if used, was siphoned out after the sludge was blended by a blender and settled. 4 . 2 Analytical Methods 4.2.1 Chemical Concentration Determinations Organics analyses . Phenol, 2 , 4-dichlorophenol and pentachlorophenol concentrations were analyzed by a highperformance liquid chromatography (HPLC) , Perkin-Elmer Model LC-100, with a ZORBAX C-8 column (4.6 mm I.D. x 15 cm), a LC-75 Spectrophotomet r i c Detector and a Fisher Series 5000 Recorder. For phenol and 2 , 4-dichlorophenol determination, a 58/42 (v/v) mixture of methanol/water mobile phase was used and the wavelength of the UV detector was set at 197 nm. A 72/28 (v/v) mixture of methanol/water mobile phase

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36 and 220 nm wavelength were selected for the analysis of pentachlorophenol and the mixture of all three phenols. Both mobile phase solutions were adjusted to pH 2 with phosphoric acid (approximately 0.15% by volume) , filtered, and degased. Mobile phase flow rates were 2 to 3 ml per minute . Analytes were identified by comparing the retention times of the standards and the retention times of the samples while concentrations were calculated by comparing the peak heights of the standards and the peak heights of the samples. At 2 ml/min flow rate with 72% methanol in the mobile phase, retention times for phenol, 2,4-DCP and PCP were 1.3, 1.9 and 5.1 minutes, respectively. At 2 ml/min flow rate when 58/42 methanol/water mixture was used as the mobile phase, retention time was 1.6 minutes for phenol and 3.9 minutes for 2,4-DCP. The detection limit for phenol was 0.01 mg/1, for 2,4-DCP was 0.02 mg/1 and PCP was 0.03 mg/1. Optical absorbance measurement . Optical absorbance of INTF in the biological activity assay (Section 4.2.4) was measured by a Perkin-Elmer Model 552 Spectrophotometer with the wavelength set at 465 nm . Dissolved oxygen measurement . A Yellow Spring Instrument, (YSI) Model 54 Oxygen Meter was used for the measurement of dissolved oxygen concentration. A strip chart recorder was connected when continuous monitoring was r equ i r ed .

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37 Specific conductivity measurement . A Yellow Spring Instrument, (YSI) Model 33 S-C-T Meter was used for the measurement of specific conductivity. A strip chart recorder was connected when continuous monitoring was requ ired . Nutrient analyses . Nitrogen and phosphorus concentrations were measured by Technicon AutoAnalyzer II. Weight measurement . All weighings were made on a Mettler Model AE 160 balance unless the weight exceeded its capacity of 160 grams. In that case a Mettler Model PR 1200 balance was used. 4.2.2 Soil Characterization Organic carbon determination . Organic carbon determinations were performed in the Soil Science Department, University of Florida, following the WalkleyBlack procedure described in Section 29.3 of Methods of Soil Analysis, Part 2 (Nelson and Sommers, 1982). Soil water content determination . The direct method with oven drying as described in Section 21-2.2 of Methods of Soil Analysis, Part 1 (Gardner, 1986) was used to determine the percent soil water content, which equals ( [ wt . of wet soil]/[wt. of dry soil])-l. Soil porosity determination . A sample of 200 ml of well mixed soil were dried in an oven at 105 C for 24 hours

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to evaporate any moisture. The soil was placed in an 1liter graduated cylinder and the height of the soil was marked. Then water was added by such that the water level just coincided with the original soil level. Mixing was provided to eliminate air pockets. Porosity = [amount of water added] / 200 (4-1 Hydraulic conduct i vity determination . Hydraulic conductivity of the soil columns was determined by a constant head permeameter shown in Figure 4-1. Soil retention screens made of a few layers of glass fiber supported with a stainless steel mesh were placed both on top and at the bottom of the soil column. A piece of 3/4 inch tygon tubing was used to feed water from the constant head reservoir. The frictional head losses from the tubing and the fixtures were negligible compared to that caused by the soil column. The hydraulic conductivity (K) was calculated according to the formula: K = Q L / dh A (4-2) where K is the hydraulic conductivity (cm/sec), Q is the measured flow rate (ml/sec) , L is the length of the soil column (cm), dh is the total head loss through the permeameter (cm) (which is the difference in elevation between the inflow and outflow water levels), and A is the 2 cross sectional area of the soil column (cm ) (McWhorter and Sunada, 1977) .

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39 4 A Ii\. h f T Constant Head

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40 Hydraulic conductivity values were evaluated before and after a set of column biodegradat ion experiments. This test was designed to determine the effect of bacterial population increases on hydraulic conductivity. 4.2.3 Sludg e Characterization Total volatile solids measurement . The following procedure provided the total volatile solids (TVS) measurements. Dry sludge weight and ash weight from 100 ml of wet sludge were measured after drying in an oven at 103 C overnight and combusting in a muffle furnace at 550 C, respectively (Sawyer and McCarty, 1967) . TVS = ([wt. of dry solid] [wt. of ash]) x 10 (4-3) Organic carbon determination . In the sorption isotherm studies the organic carbon content of the sludge was determined in order to calculate K values. In sludges oc from municipal wastewater plants the volatile solids are mainly biomass. Accordingly, it was assumed that organic matter was the same as volatile solids, and corresponds to the biomass formula C,.H. n 0, o N,.P (Benefield and Randall, b0 8/ 23 12 1980). Therefore, organic carbon, OC, is OC (mg/1) = TVS x (720/1374) = 0.52 x TVS (4-4) This factor, 0.52, agrees with the values 0.40-0.53 as suggested in Methods of Soil Analysis (Nelson and Sommers, 1982) .

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41 4.2.4 Biological Activity Measurement Biological activity was assessed by INT-Dehydrogenase assay. INT (2[ piodophenyl ] -3[ p-n i t rophenyl ] -5phenyl tet razol ium chloride) is reduced by the electron transport system of active microorganisms via dehydrogenase activity (DHA) to form water insoluble, red INT-formazan (INTF) crystals (Koopman and Bitton, 1987). The procedure was modified as follows. After adjusting the sample pH to 7.6, 1 ml of 0.2% INT solution was added to a 5 ml sample and incubated in the dark until a pink color developed. The incubation time was recorded. the sample was filtered through a 0.45 um pore size membrane, and the filter extracted with 5 ml of DMSO (dimethylsulf ox ide) . The extract was centrifuged. The absorbance of INTF, which is proportional to DHA, was measured by a spectrophotometer at 465 nm wavelength. The controls were prepared the same procedure as the samples except 1 ml of formaldehyde was added in order to kill the microorganisms in the controls. DHA is expressed in equivalent oxygen uptake units (mg 2 /l/day) . DHA = [ 905* Ve* (Ds Dc) ] / (t* Vs) (4-5) where Ve is the volume of DMSO, Ds is the optical absorbance of the sample, Dc is the optical absorbance of control, t is the incubation time (minutes) , and Vs is the volume of sample filtered.

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42 4 . 3 Experimental Design 4.3.1 Batch Sorption Studies Unless otherwise specified, all experiments in the sorption studies were performed using 40 ml glass vials with screw caps and teflon lined septa as the reactors. Batch adsorption . The objectives of this study were: (1) to determine the adsorption partition coefficients of phenol, 2 , 4-d ichlorophenol and pentachlorophenol in an aquifer with very low organic matter, (2) to show the effects of mixing of phenol, 2,4-DCP and PCP on sorption, and (3) to determine the effects of adding sodium azide (NaN 3 ) . The chemicals were tested both individually and as a mixture. In each vial, 40 grams of sandy soil and 20 ml of solution were mixed and tumbled continuously by a rotator at room temperature (approximately 23°C) for 24 hours to ensure complete mixing. Sodium azide (NaN ) was added in two different concentrations to selected vials to eliminate biological degradation. Four chemical concentrations were used in this study: 10, 7.5, 5 and 1 mg/1. The experimental matrix consisted of three treatments, four concentration levels, and four sample categories as listed in Table 4-1. Replicates were prepared for six of the randomly chosen treatments for the purpose of quality assurance. The average number of those treatments was used for the calculat ions .

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43 Table 4-1. Experimental scheme for adsorption study, [ Treatment ] 2ml sludge + 6mg/l NaN 2ml sludge + 2mg/l NaN^ no sludge + 2mg/l NaN^ (By) [Cone . ] 1 mg/1 5 mg/1 7.5 mg/1 10 mg/1 [ Sample] phenol DCP PCP (By) mixture Batch desorption . The desorption study was performed following adsorption study. Aqueous portions of the samples were drained (only about 10 ml could be drained) and the sample vials refilled with 10 ml of distilled water. The samples were tumbled continuously by a rotator at room temperature (approximately 23°C) for 40 hours before analysis of phenolic compound concentrations. Extraction recovery . The purpose of the extraction recovery study was to account for all chemical masses and to determine the efficiency of methanol extraction. This measurement was necessary in order to determine whether concentration decreases were caused by biodegradat ion or by adsorption later in the degradation experiments. The extraction recovery was performed following the desorption study. Aqueous portions of the samples were drained and refilled with 10 ml of methanol. Samples were tumbled at room temperature for 3 hours before analysis. Calculation methods . The amount of each compound adsorbed to the soil (X/m) were calculated by dividing the difference between the initial mass in the system and the

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44 mass in the aqueous phase with the amount of soil in the vial. The amount of each compound desorbed from the soil (-X/m) were calculated by dividing the difference between "the mass remained in system after draining the free water, which was calculated from the results of adsorption experiment" and "the mass in the aqueous phase" with the amount of soil in the vial. Sorption coefficients were calculated by fitting data to the Freundlich model, i.e., plotting log(X/m) (log(-X/m) in the cases of desorption) versus log(C). The slope is the constant b in Equation (3-2) and ordinate the intercept is log(K .) or log(K„_), where t A c D K„ (in ml/g) is the adsorption equilibrium partition coefficient and K„_. is the desorption equilibrium partition coefficient. Recoveries were determined by the "(calculated) mass of analyte extracted from soil" to "(calculated) mass of un-desorbed analyte on soil" ratio. The dilution effect caused by the solutions trapped in the soil matrix was accounted for in the calculations of recover ies . 4.3.2 Column Sorption Studies . Column sorption experiments were performed with three 12 inches long x 3 inches diameter glass columns. A quantity of 1500 g (about 9 inches high) of sandy soil were loaded in each column. Conservative tracer . The experimental setup is shown in Figure 4-2. Soil in the columns was saturated with

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45 distilled water for 24 hours before the study began. A 1 N ammonium chloride solution was pumped at 6.5 ml/min from a reservoir, the effluent was monitored by a conductivity meter and recorded by a strip chart recorder. However, because ammonium chloride is more dense than water, ion concentrations higher than influent built up around the probe rather quickly and caused incorrect readings. To overcome this problem, the column was saturated with the ammonium chloride solution then desorbed with distilled water . Solute retardation . Columns were saturated with distilled water for three days, then drained of free water. Approximately 250 ml of water remained in the soil matrix after draining. 250 ml of solution consisted of all three phenolic compounds, each at a concentration of 6 mg/1 each was added to the top of each column which contained 1500 g soil. The systems were recirculated by a peristaltic pump (coupled with three heads) with a flow rate of 2.7 ml/min (Figure 4-3). Samples, 250 ul , were taken at the column inlets every 15 to 30 minutes and analyzed immediately. Retardation factor calculation . Retardation factors were calculated based on the assumption that the retardation factor (R) equals the number of pore volumes passed through the column when effluent concentration of each solute

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Column 46 Conductivit) Meter Probe P u m p I )

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47 Mixture soiution circulation line P u m p UJ im | E » Point of Sampling .-.-.-.-.-.-."TT -^H t>>::-:::-::j t SSPiuj i I Figure 4-3. Experimental setup for column degradation studies .

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48 reached 50% of the influent concentration (Nkedi-Kizza et al. , 1987) . 4.3.3 Batch Biodegradat ion Studies Batch biodegradat ion were performed in 40 ml VOC vials as described in Section 4.3.1. A 10 g soil to 30 ml solution ratio was used throughout these experiments. The samples were kept in the dark to avoid photolysis except for the phenol and DCP degradation studies. Periodically the caps were opened and 250 ul samples were taken for analyses. At least duplicate injections were performed for each analys is . Nutrient requirement . This test was intended to determine the effect of nutrients on biodegradat ion . Phenol was chosen as the carbon source. Preliminary analysis by autoanalyzer indicated that 2.36 mg/1 of total phosphate and 0.05 mg/1 of orthophosphorus can be extracted into 30 ml of distilled water from 10 g of this soil, which exceeded the theoretical phosphorus requirement for complete mineralization of the phenolic compounds. The same analysis only detected trace amounts of nitrate nitrogen. Therefore only the nitrogen level was manipulated. The experimental scheme is listed in Table 4-2. Phenol biodegradat ion . Biodegradat ion of phenol in soil by indigenous bacteria and by bacteria from municipal wastewater sludge that was added to the soil was studied.

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49 Sodium azide (NaN ) at a concentration of 2 mg/1 was added to the control samples to preclude biodegradat ion . The experimental scheme, along with the nutrient requirement study, is listed in Table 4-2. Table 4-2. Experimental scheme for phenol biodegradat ion and nutrient requirement. No. Cone mg/1 N added mg/1 Added C:N Sludge ml NaN mg/I 101

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50 solution that contained phenol degrading bacteria taken from the phenol degradation samples was added. Table 4-3. Experimental scheme for 2,4-DCP biodegradat ion . No.

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51 hours, then 0.5 ml of sludge supernatant were added to #304, #305 and #309. Table 4-4. Experimental scheme for PCP biodegradat ion and enzyme induction studies. No.

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52 were added to the sludge amended samples and compared to those without primary substrates. The experimental scheme, along with the multi-compound biodegradat ion assay, is listed in Table 4-5. Table 4-5. Experimental scheme for biodegradat ion and codegradation studies of phenol, 2,4-DCP and PCP in multi-compound systems. No. Cone. N added Added Amendment NaN.. mg/1 mg/1 C:N (1 ml) mg/I 401

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53 presence of phenol was performed to further investigate this phenomenon. The experimental scheme is listed in Table 4-6. Table 4-6. Experimental scheme for PCP co-degradation in the presence of phenol . No.

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54 mixing and dilution of the added solution with the distilled water in the columns. Sampling and analytical procedures were the same as described in Section 4.3.2. The experimental scheme is illustrated in Figure 4-3. Column biodegradat ion II . This experiment was designed to examine the results of PCP co-degradation in the presence of phenol and was similar to the experiment in Section 4.3.3 except for using recirculation in columns. Columns were saturated with 500 ml distilled water for three days before draining all free water. Distilled water was then refilled so that each column had 350 ml water (including 250 ml trapped water and 100 ml free water) in it. A 150 ml of mixed solution with different concentrations of phenol and PCP were added to the top of each column which contained 1500 g of soil. To enhance the effects of co-degradation, phenol concentrations with ten fold difference were used. The same flow rate setting as in the previous experiment (2.7 ml/min) was maintained. The experimental scheme is listed in Table 4-7.

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55 Table 4-7. Experimental scheme for column II (codegradat ion of PCP and phenol) . Condition No phenol mg/1 DCP mg/1 PCP mg/1 N added mg/1 Before

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CHAPTER V RESULTS AND DISCUSSION This chapter presents and reviews the results of all experiments performed in this research, followed by a discussion of each topic. Major categories are soil characterization, batch sorption, column sorption, batch biodegradation, column biodegradat ion , and hydraulic conductivity determination. 5.1 Soil Characterization The aquifer materials used in this research were dry, clean, uniformly sized, yellowish brown in appearance, and predominantly fine grained sands. The selected physical properties of the soil are presented in Table 5-1. Soil analysis indicated that there was very little (non-detectable), if any, organic carbon content in the soil matrix. For practical purposes it was assumed that the organic carbon content in the soil matrix was zero. The bulk density and porosity were not measured under undisturbed, in-situ conditions. All the soils were obtained at the same time and stored in a capped bucket in the laboratory for later use. They were visually inspected and foreign objects such as grass roots and wood chips were removed before use.

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57 Table 5-1. Selected physical properties of the soil. Parameters Values Particle density 2.52 g/ml Water content 6.5% by volume Bulk density 1.45 g/ml Organic carbon Negligible Porosity 0.45 Sieve analysis Passed #30 100% Retained on #40 0.45% Retained on #140 96.68% Passed #140 2.87% 5 . 2 Batch Sorption The sorption isotherm data were fitted to the Freundlich model using the method of least squares regression analysis. These data are listed in Appendix A. 5.2.1 Single Compound Batch Adsorption . The Freundlich sorption parameters of the phenolic compounds on aquifer material are presented in Table 5-2, and the parameters on aquifer material with sludge addition are presented in Table 5-3. Notice that PCP concentrations consisted both the ionized form (pentachlorophenolate) and the unionized form (pentachlorophenol) , thus, the results for PCP sorption as well as degradation experiments represent a combination of these two PCP forms.

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58 Table 5-2. Adsorption regression parameters of phenolic compounds in single-compound system on plain so il . Compounds pH log K ST.DEV.+ r A ST.DEV. R Phenol

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59 then became 4.56 mg as sludge was added to each system, or 0.0114% (4.56 mg organic carbon in 40 g of soil) in the soil matr ix . In the batch isotherm study of aquifer material with added sludge, the addition of sodium azide at different concentration levels, 2 ppm and 6 ppm, did not cause Freundlich sorption coefficients to significantly change based on a paired difference t-test at alpha= 0.05 significance level (McClave and Dietrich, 1985), which indicated that the presence of sodium azide did not interfere with the sorption behaviors of the phenolic compounds. Thus the average of those two sets of parameters was taken and used to calculate the sorption parameters on organic carbon, with the exception of the phenol data. Phenol adsorption on aquifer material with sludge and with 2 mg/1 sodium azide showed an unusually high K value, and later in the consequent desorption study all the phenol concentrations were biodegraded to trace amounts, which indicates that the sodium azide at 2 mg/1 was not effective enough to inhibit all the microorganisms (also, it indicates that phenol degrades rather quickly). Therefore the data obtained from this experiments, although presented, were not used for the calculation of the sorption coefficients. The Freundlich sorption coefficients for the isotherms with sludge are quite different from those without sludge addition. Because both sets of isotherms were performed under the same conditions (other than the addition of

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60 organic matter), these differences were solely contributed by the added organic matter. Table 5-4 presents the calculated sorption parameters that resulted from adding the organic carbon in the wastewater sludge to the soil along with some values available in the literature. The calculation procedure is listed in Appendix E. Table 5-4. Calculated adsorption parameters of phenolic compounds in single-compound system based on organic carbon. Measured Literature log K values 3 oc Compounds log K 1) (2) (3) (4) (5 Phenol 2.586 0.682 1.21 3.46 2,4-DCP 2.910 0.826 2.10 3.60 2.54 PCP 4.204 0.824 4.80 3.51 4.84 * K values are in ml per gram organic carbon. , * exponent in the Freundlich model: (X/m)=K * C (1) Boyd (1982) . (2) Isaacson and Frink (1984). (3) Calculated from K by U.S. EPA (1979). (4) Calculated from K° W by Kaiser and Valdmanis (1982) (5) Calculated from K° W values listed by Lagas (1988). ow From the results shown in Table 5-2 it is clear that, although not in great amount, adsorption on soils with virtually no organic carbon content was still occurring, demonstrating that Equation (3-3) is not valid in this case. This result agreed with Rao and Jessup's (1983) suggestion that Equation (3-3) may not apply to soils containing organic carbon content less than 0.1 percent. The calculated Freundlich sorption coefficients listed in Table 5-4 are actually K values for phenol, 2,4-DCP and PCP, and J oc these values will be used throughout this research.

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61 5.2.2 Mix ed Compound Batch Adsorption . The Freundlich sorption coefficients for the mixture of the three phenolic compounds are shown in Table 5-5, Table 5-6 and Table 5-7, presented in the same order as in Section 5.2.1. Table 5-5. Adsorption regression parameters of phenolic compounds in multi-compound system on plain so il . _ _ _-Compounds pH log K„» ST.DEV. b ST.DEV. R ___ __ _ _E A _ Phenol 4.79 -2.286 0.119 0.868 0.154 0.941 2,4-DCP 4.79 -1.350 0.057 0.791 0.073 0.983 PCP 4.79 -0.131 0.077 0.682 0.082 0.972 + Log standard deviation of log K values . * Exponent in the Freundlich model: (X/m) =K * C t Table 5-6. Adsorption regression parameters of phenolic compounds in multi-compound system on soil with sludge . Compounds

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62 Table 5-7. Calculated adsorption parameters of phenolic compounds in multi-compound system based on organic carbon. Compounds log K Phenol 2.161 2,4-DCP 2.690 PCP 4.064 0.936 0.988 0.710 * K values are in ml per gram organic carbon . b°is the exponent in the Freundlich model: (X/m)=K„ * C t The differences of sorption behavior between singlecompound and multi-compound systems are significant by ttest analysis at alpha= 0.05. The Freundlich sorption coefficients for mixed compounds are 1.5 to 3 times less than those of single compounds. This phenomenon could have been the result of either a co-solvent effect or a competitive sorption effect. Although it is difficult to identify the appropriate mechanism, the co-solvent effect will have a greater influence on the co-solute (compounds with lower solubility in common solvent) than on the cosolvent (compounds with higher solubility in common solvent) (Staples and Geiselmann, 1988). The common solvent is water in this case. On the contrary, competitive sorption effect should have a greater influence on sorption coefficients of less hydrophobic compounds since they are less likely to win the competition with more hydrophobic compounds for the limited sorption sites. This observation provides a vehicle to help indentify the appropriate mechanism. Comparing the corresponding Freundlich sorption coefficients from singlecompound and multi-compounds sorption studies, a list of

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63 ratios can be calculated and is presented in Table 5-8. The study shows phenol suffered the greatest loss of adsorption capacity when mixed with other phenolic compounds, and indicates that the loss of adsorption capacity was predominantly caused by competitive adsorption effects. Table 5-8. Ratios of Freundlich sorption coefficients for phenolic compounds in single and multiple compound systems. Ratio of K_,. : Soil FA Soil+sludge organic carbon Phenol 2,4-DCP PCP 3 .05 1.22 1.51 2.88 1.47 1.42 2.66 1.66 1.38 In general, Freundlich sorption coefficients for phenolic compounds increase as the level of chlorination increases. The less-than-unity values of the Freundlich exponent, b, indicate the adsorption was not linear, and higher concentrations resulted in a proportionately less amount of adsorption. These values of the Freundlich exponent are in good agreement with the ones reported by Boyd (1982) (b=0.79 for phenol and b=0.67 for 2,4-DCP), Lagas (1988) (b=0.86 for PCP), and Laquer and Manahan (1987) (b = 0.65 for phenol) . The calculated K values (Table 5-4 and Table 5-7) are oc closer to those predicted by Equation (3-5), which are 2.99 for phenol, 4.03 for 2,4-DCP and 5.86 for PCP. A least squares linear regression reveals a good correlation between the log of water solubility (in umole/1) and log of K

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64 values for phenol, 2,4-DCP and PCP as log K = 3.547 0.421 log WS (R =0.983) (5-1) oc 5.2.3 Batch Desorption . The batch desorption data were fitted to the Freundlich model and the results are presented in Tables 5-9 and 5-10. Table 5-9 lists the Freundlich desorption parameters of phenolic compounds in single-compound systems and Table 5-10 are in multi-compound systems. Like the results of adsorption studies, there are significant differences (ttest at alpha= 0.05) between single compound and multiple compound systems. Table 5-9. Desorption regression parameters of phenolic compounds in single-compound systems. Compounds

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65 0.5 ? -o.H £ v -1.5 X l_l J -2.5 I I ! _1 fi — 1. + .^ + '• : < /' i Sngl-Adsp + Snal-Deso Mix-Adsp Mix-Desp -0.2 -0.4 ! I 1 6 Log[Ce(ug/m!)3 Figure 5-1. Phenol sorption isotherms on plain soil.

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66 0.5 -I r £ -fl.5-1 \ /> H\ 3 v 05 J -i -J -1-5 H ^-; 4 ,-i Sngi-Adsp + Sngl-Desp Miv-AHcn IlilA nWVH Mix-Desp Log [Ce(ug/mi)i Figure 5-2. Phenol sorption isotherms on soil with sludge

PAGE 81

67 0.5 -

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Ui -0.5 X 0) J -2.5 -I i i i -1.6 -1. i SngrAdsp + Sngi-Desp ivilA nyjp A Mix-Desp I I i i i i o.r —n 4. n Q.S Log (Ce(ug/ml)] 68 Figure 5-4. 2,4-DCP sorption isotherms on soil with sludge .

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0.5 H r £ S 0) X J -o.h 69 --"1 _A— Sngl-Adsp + Sngl-Desp * Mix-Adsp a Mix-Desb i i r~ -1.6 -1.2 i r -Q.5 -0.4 1.2 Log[Ce(ug/mi)i Figure 5-5. PCP sorption isotherms on plain soil.

PAGE 84

70 o>

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71 Table 5-10. Desorption regression parameters of phenolic compounds in multi-compound systems. Compounds pH log K _ ST.DEV. + b ST.DEV. R 2 c D [Aquifer material] Phenol 5.01 -1.952 0.068 0.535 0.089 0.948 2,4-DCP 5.01 -1.380 0.052 0.744 0.068 0.983 PCP 5.01 -1.234 0.036 0.183 0.045 0.891 [Aquifer material

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72 respectively, were irreversibly held onto the soil matrix after desorbing with distilled water for 40 hours. These are percentages of initially adsorbed masses that were irreversibly adsorbed (not desorbed) . Isaacson and Frink (1984) reported similar irreversibilities among other substituted phenolic compounds. The small Freundlich exponent values of PCP desorption indicating that the desorption intensities were low. 5.3 Column Sorption Column sorption experiments were performed on column #1 and column #2. Breakthrough curves for phenol, 2,4-DCP and PCP from column #1 are shown in Figure 5-7. These curves were plotted using the data presented in Appendix B. The breakthrough curve of the conservative tracer, ammonium chloride, was reconstructed from the data obtained in the desorption experiment as described in Chapter 4. Because these column experiments were designed to simulate a treatment of groundwater contaminated by a point source such as a spill, only a limited amount of analytes were spiked onto each column. This method differs from the traditional way of performing breakthrough curve experiments, and make the calculation of retardation factors very difficult. The first task was to determine the initial concentration, C . This value should range from 3.0 mg/1 if o a complete mixing mode was assumed, to 6.0 mg/1 if a plug flow mode was assumed. However, the highest concentration

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73 ever detected was 3.75 mg/1 of phenol from column #1 and 3.3 mg/1 from column #2. Based on the Freundlich sorption coefficients obtained from batch sorption studies, phenol has a very low tendency to be adsorbed on this particular type of sandy soil. Therefore it was assumed that the C value of each compound was 3.75 mg/1 for column #1 and 3.3 mg/1 for column #2. Figure 5-7 was plotted based on the normalized C/C values, and the X-coord inates corresponded o ' c to the intersections of the C/C =0.5 line with each o breakthrough curve being measured as the retardation factors . For the purpose of comparison, Equation (1-3) and Equation (5-2) (Nkedi-Kizza et al., 1987) and the batch sorption data were used to estimate the retardation factors. Notice that Equation (1-3) assumes linear adsorption, i.e., the Freundlich exponent was assumed to be unity. R = 1 + P K r „ C * FA o b-1 (5-2) The calculated and measured retardation factors for phenol, 2,4-DCP and PCP are listed in Table 5-11. They agree with each other very well except for the PCP retardation factor calculated from Equation (1-3). This difference was caused by omitting the Freundlich exponent since PCP adsorption deviates the most from linear.

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74 D chloride Pore Volume + phenol Figure 5-7. Column breakthrough curves for phenol, 2,4DCP and PCP.

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75 Table 5-11. Retardation factors of mixed phenolic compounds calculated by various methods. Compounds # pore vol. Eg. (1-3)* Eg. (5-2)* Phenol 2,4-DCP PCP 1.03 1.16 2.26 1.017 1.144 3.385 1.014 1.109 2.566 * p=1.45 g/ml , n=0.45, C =3.75 mg/1 phenol : K =0 . 00518 ,°b=0 . 868 2,4-DCP : K^ = 0.0447, b = 0.791 PCP K^ = 0.74, b=0.682 5.4 Batch Biodegradat ion The data from all batch biodegradat ion experiments are shown in Appendix C. All measured concentrations were normalized as C/C x 100%. These normalized data were used o to evaluate the degradation rates and to plot figures. The degradation rates were calculated as apparent rates, which include the effect of adsorption at the beginning and the effect of desorption later during the course of the experiment. The apparent degradation rate constants for phenol and 2 , 4-dichlorophenol are very close to the real values since adsorption of these two compounds was fairly weak. However, PCP has a much stronger adsorption than phenol and 2,4-DCP do, which could cause the apparent degradation rate constants to be high. Therefore another set of results calculated by subtracting adsorption effects at the beginning, termed conservative degradation rate constants, are presented for all PCP degradation data. The conservative data did not account for the loss of later

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76 desorbed PCP due to biodegradat ion , therefore these two sets of results define an upper limit and a lower limit for the degradation rate constants and half-lives for each sample. 5.4.1 Nutrient Requirement . The purpose of this test was to determine how much, if any, nutrient is needed for the biodegradat ion of phenolic compounds in this particular type of soil. Background analyses indicated that there was enough soluble phosphorus in the soil but the nitrogen concentration was at near the limit of detection (0.01 mg/1) . Therefore the effects of nitrogen content (at three levels) were tested. Biological degradation was confirmed by comparing samples with controls, and was the main contributor to the decrease of phenol concentrations. The result showed no differences among these treatments as presented in Figures 5-8 and 5-9, indicating that at least part of the phenol was assimilated for energy but not for growth. 5.4.2 Single Compound Biodegradat ion . Batch biodegradat ion experiments for phenolic compounds were performed. Treatments with indigenous soil bacteria and amended with municipal wastewater sludge were included. The apparent degradation rate constants were calculated based on first order reaction kinetics. Phenol degraded rather quickly, with average half-lives ranging from 9 hours for the C =5ppm group to 15 hours for

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77 the C =lppm group. Notice that in a first order reaction the half-life values are independent of initial chemical concentrations. However, the results showed different halflife values for the C =5ppm and C =lppra samples under otherwise same treatments. This deviation is because the first order reaction kinetics does not address all the factors (such as toxic effects and substrate availability) that are influential to biological degradation reactions. Table 5-12 lists the apparent biodegradat ion rate constants for phenol. High correlation coefficient values for log of concentration versus time indicate that the first order reaction kinetics describes phenol degradation quite well. Table 5-12. Apparent biodegradat ion rate constants for phenol . Sample

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78 group seems to degrade faster than the lower starting concentration group, and this effect was even greater than the effect of amending with sludge although Figures 5-8 and 5-9 clearly indicated a lag period for the group with indigenous bacteria (30 hours for the C =5 ppm group and 20 hours for the C =1 ppm group) . All samples were degraded to below or near 0.01 ppm, the limit of detection. The results of the 2 , 4-dichlorophenol biodegradat ion are presented in Table 5-13 as well as in Figures 5-10 and 5-11. 2 , 4-Dichlorophenol , as well as phenol, can be biodegraded to a concentration close to or below the detection limit, however, with slower rates. Table 5-13. Apparent biodegradat ion rate constants for 2,4-DCP. Sample

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79 01 + 1103 Time (days) 104 ' A #105 #106 V #107 Figure 5-8. Phenol degradation curves in single-compound systems (initial concentration 5 ppm) .

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80 D |1I + #109 Time (days) o #110 #111 #112 Figure 5-9. Phenol degradation curves in single-compound systems (initial concentration 1 ppm) .

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81 were no samples with indigenous soil bacteria in the 5 ppm initial concentration group, but compare #204 (with sludge amendment in Co=5 ppm group, t , =8.66 days) with #208, the half-life for the higher initial concentration sample appeared longer than its lower concentration counterpart. The sample #203 was a duplicate of #204 until t=14.5 days when 1 ml of solution containing phenol degrading bacteria was added to #203. The effect was drastic as shown in Figure 5-11 and was evident in half-life values. This was attributed to either the increase in microorganism population or to the introduction of some enzymes which were induced by exposing the bacteria to phenol. The degradation rate constant and half-life for #203 are only qualitative because they were a result of the combination of two treatments . Pentachlorophenol also appeared to undergo biological degradation but with a very different pattern from phenol and 2 , 4-dichlorophenol . Figures 5-12 and 5-13 illustrate the degradation of PCP for different initial concentrations. Table 5-14 lists the apparent degradation rate constants and half-lives, and Table 5-15 lists the conservative results, which were calculated based on the data up to t=60 days. Because of analytical problems, PCP degradation data showed day to day variability. In order to depict trends in the PCP degradation data, variations in concentration versus time data were dampened by using weighted average concentrations, C(n) , at measurement n, where

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82 2 C(n) + C(n-l) +C(n+1 C(n) = (5-3 Even with the data processed in this form, the results of a few sample runs still did not indicate a linear relationship between log of concentration and time. This poor correlation is indicated by the low correlation 2 coefficients (R ) in Tables 5-14 and 5-15. Both raw and normalized data sets and some curves plotted with raw data are presented in Appendix C. Table 5-14. Apparent biodegradat ion rate constants for PCP. Sample

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83 5 -fh I I 4

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84 Time (days) #205 + #206 o #207 A #208 Figure 5-11. 2,4-DCP degradation curves in single-compound systems (initial concentration 1 ppm) .

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85 D #301 Time (days) + #302 o #303 Figure 5-12. PCP degradation curves in single-compound systems (initial concentration 5 ppm) .

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86 I !r« 0.9 -\\\ \\Sk i I * BI 1 1 \ \ % t\ *-T 1 . ** r +-#0.3 -> "•4 jTV. """4' Added 0.5 ml --+.. Sludge Supernatant tsJ Time (days) + #303 if.^nB Figure 5-13. PCP degradation curves in single-compound systems (initial concentration 1 ppm) .

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87 \\Tfc I * — t* — * W^ "-J Hi Added 0.5 ml i\ Sludge Supernatant \ \ 60 I #301 Ttfiie (doye) + #304 6 #3C Figure 5-14. PCP degradation curves using bacteria which are acclimated to phenol and 2,4-DCP (initial concentration 5 ppm) .

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88 \ o E \s C #307 Time (days) A #310 x #311 Figure 5-15. PCP degradation curves using bacteria which are acclimated to phenol and 2,4-DCP (initial concentration 1 ppm) .

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89 Table 5-15. Conservative biodegradat ion rate constants for PCP. Sample

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90 of time (about 30 days) after PCP solutions had been introduced, and then the degradation appeared to cease. Upon addition of an additional 0.5 ml of sludge supernatant at t=60 days (to #304, #305, and #309) the degradation activities resumed for another period of time and stopped again. There were only negligible amounts of organic carbon in the sludge supernatant, so the degradation should result solely from biological activities. This degradation pattern suggests that the microorganisms can degrade PCP but cannot rely on PCP as a sole source of energy and carbon, i.e., the microorganism populations were first in stationary phase then in endogenous phase. Dehydrogenase activity (DHA) tests on these PCP degradation study samples were performed two months after the study ended and showed DHA only at background levels, indirectly confirming the hypothetical degradation pattern. Chu and Kirsch (1973) found one bacteria culture was capable of utilizing PCP as a single source of carbon and energy for growth. However, Stanlake and Finn (1982) suggested that sporadic or continuous addition of PCP-adapted cells to the treatment system might be necessary because yields were low when bacteria utilize PCP as the sole carbon source. Kaufman (1978) also reported in his review that PCP was removed by respiring cells but did not support growth. The degradation pattern shown in Figures 5-12 and 5-13 agreed with the latter two observa t ions .

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91 Figures 5-14 and 5-15 illustrate PCP degradation by bacteria that were acclimated to phenol (#304 and #310) or 2 ,4-dichlorophenol (#305, #306 and #311). Some increase in degradation rates in these samples were observed but the effects were less pronounced than in those samples that were amended with sludge. The average half-life for PCP was about 80 days with the exclusion of #310 because of the low 2 R value. Neither phenol degrading bacteria nor 2,4-DCP degrading bacteria showed conclusive superiority over the other as an aide to PCP degradation. 5.4.3 Multiple Compound Biodegradat ion . A multi-compound biodegradat ion experiment was conducted in order to examine the effects of co-degradation. As in previous experiments, phenol, 2,4-DCP and PCP were the target compounds, and unless otherwise specified, equal concentrations for all three compounds were used in the solutions. For unknown reasons the 2,4-DCP data were so anomalous that had to be discarded. All degradation data, including those for 2,4-DCP, are presented in Appendix C. The results for phenol degradation in multi-compound conditions are presented in Table 5-16 and Figures 5-16 through 5-21. The degradation rate constants were calculated based on eight to ten data points (n) , depending on how soon the compound had reached the detection limit.

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92 Table 5-16. Apparent biodegradat ion rate constants for phenol in multi-compound conditions. Sample

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93 10 #401 Time (dcys) + #402 o #403 Figure 5-16. Phenol degradation curves in multi-compound systems (initial concentration 5 ppm) .

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94 \ £ vy C ID Time (days) f #410 Figure 5-17. Phenol degradation curves in multi-compound systems (initial concentration 1 ppm) .

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95 \ E C Time (dcys) #401 + #403 o #407 Figure 5-18. Phenol co-degradation curves in multicompound systems (initial concentration 5 ppm) .

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96 10 Time (days) #409 + #411 #415 A #416 Figure 5-19. Phenol co-degradation curves in multicompound systems (initial concentration 1 ppm) .

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97 \ E \s C 10 a #401 Time (days) X #405 v #406 Figure 5-2( Phenol degradation curves in multi-compound systems using acclimated bacteria (initial concentration 5 ppm) .

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98 \ Time (days) D #409 A #412 X #413 V #414 Figure 5-21. Phenol degradation curves in multi-compound systems using acclimated bacteria (initial concentration 1 ppm) .

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99 by sludge amendment. Glucose and sodium acetate seemed to have little influence, and the bacteria that were acclimated to both 2,4-DCP and PCP were not significantly effective in aiding phenol degradation. Table 5-17, Table 5-18 and Figures 5-22 through 5-27 show the results for PCP degradation in multi-compound conditions. For the same reason as described in the PCP single-compound degradation study, a weighted average method was used to calculate the rates and to construct the 2 figures. The correlation coefficient (R ) values for some samples in the lower concentration range were low, presumably because the deviations caused by analytical problems were magnified at the lower initial concentrations relative to at the higher initial concentrations. Less weight was put on these samples when assessing the results. Also, #407 was discontinued at t=27 days when the sample was contaminated with PCP standard solution. Therefore, the K, and t , values of #407 cannot be compared with other samples. Notice the samples #402 and #410 in Table 5-17 that the PCP degradation rates in multi-compound systems were significantly greater than in single-compound systems. The average half-life for plain soil samples was about 85 days, compared to 120 days in #302 and #308 in Table 5-14 when PCP was in single-compound systems. This may be that PCP was co-metabolized by the bacterial populations which were increased by metabolizing phenol as a primary substrate . BD

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Table 5-17. Apparent biodegradat ion rate constants for PCP in multi-compound conditions. Sample

PAGE 115

101 #401 Time (days) + #402 o #403 Figure 5-22. PCP degradation curves in multi-compound systems (initial concentration 5 ppm) .

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102 \ Time (dcys) + #410 Mil Figure 5-23. PCP degradation curves in multi-compound systems (initial concentration 1 ppm) .

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103 +.5 -1 1 v-»-^ * #407 stopped at t=27 days &>--.__ '"--&-. 20 40 Ttme ('days j #401 #4o: 60 #407 BO 100 Figure 5-24. PCP co-degradation curves in multi-compound systems (initial concentration 5 ppm) .

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104 \ 20 D #409 + #411 40 60 Time (days) #415 100 #416 Figure 5-25. PCP co-degradation curves in multi-compound systems (initial concentration 1 ppm) .

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105 --i &P 1 I I \ [\ H%^ ^ ec ^ °^ m ' Siuc ^ e Supernatant 1 V/ s ?\ to #405 and #406 \ \ A v r*s 20 40 #+01 Time (dove) A #+04 X ^+05 V #+06 Figure 5-26. PCP degradation curves in multi-compound systems using acclimated bacteria (initial concentration 5 ppm) .

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106 Time (days) #409 A 0412 x #413 #414 Figure 5-27. PCP degradation curves in multi-compound systems using acclimated bacteria (initial concentration 1 ppm) .

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107 Table 5-18. Conservative biodegradat ion rate constants for PCP in multi-compound conditions. Sample

PAGE 122

108 sodium acetate in this study, information about whether phenol and 2,4-DCP can induce the enzyme required for PCP degradation could have been obtained. 2 , 4-Dichlorophenol degrading bacteria had only a moderate effect on PCP degradation . To confirm phenol's role in PCP degradation, another multi-compound experiment was performed. Samples with different phenol to PCP concentration ratios were examined. Only PCP concentrations were monitored and the results are presented in Table 5-19, Table 5-20 and Figure 5-28. Table 5-19. Apparent biodegradat ion rate constants for PCP co-metabolized with phenol. Sample Cone . Ratio* (FPday) STD ERR ° f K BD l&h #601

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109 \ #601 20 40 Time (days) + #602 #604 A #605 Figure 5-28. PCP degradation curves in multi-compound systems with different phenol to PCP concentration ratios.

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110 The sample with a 5:1:1 ratio had the highest degradation rate, followed by the sample with a 5:0:1 ratio, supporting the theory that phenol and 2 , 4-dichlorophenol served as primary substrates and PCP was reduced by codegradat ion . 5 . 5 Column Biodegradat ion 5.5.1 Column Biodegradat ion I The first set of column biodegradat ion studies was intended to examine the effects of different initial microorganism concentrations on degradation of phenolic compounds. Table 5-21 and Figures 5-29 through 5-31 illustrate the results of this experiment. Table 5-21. Column biodegradat ion I results. Compound

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Ill Both phenol and PCP degraded faster in column experiments than in batch experiments. The increase in degradation rates for PCP were approximately five fold faster than in batch experiments. With different added bacteria concentrations, there were essentially no differences in degradation rates. Two reasons could have been contributing to these rate increases. The dynamic flowing condition brought phenolic compounds as carbon substrates to the bacteria and made these substrates available for metabolism, while in static batch conditions bacteria could only utilize the substrates immediately surrounding them or the limited amount made available by diffusion. It is widely believed that bacteria prefer attachment situations to free floating situations because of the energy requirements involved. In order to validate this hypothesis, a DHA test was conducted on a column with clean soil. Sub-samples were taken from the distilled water in the column before and after an intentional mild stir. The clear sub-samples taken before the stirring had an average DHA value of 3.06x10 mg /1/day while the turbid subsamples taken after stirring had 1.30x10 mg 2 /l/day. The higher DHA value in the turbid sample indicate that the hypothesis may be valid. The average standard error for this test was 31.9%. Another contributing factor was that a much larger quantity of soil was used in the column experiments than in the batch experiments. Larger populations of bacteria were therefore included initially.

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112 Time (cto/s) t #2 P Figure 5-29. Phenol degradation curves in column biodegradat ion study I.

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113 \ Time (days) + #2 Figure 5-30. 2,4-DCP degradation curves in column biodegradat ion study I.

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114 \ 0> D §\ Time (days) f |2 o #3 Figure 5-31. PCP degradation curves in column biodegradat ion study I.

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115 This hypothesis can also be supported by the fact that the amounts of sludge added did not affect the degradation rates because the numbers of added bacteria were small compared to the bacterial populations harbored on the soil matrix in the columns . This column study demonstrated a promising results with respect to in-situ biodegradat ion of phenolic compounds but still left unaddressed the fate of the phenolic compounds adsorbed on the soils. To solve this question, 50 g samples of soil were taken from the top, middle and bottom of each column. These were each extracted with 10 ml of methanol and analyzed. This extraction method had an average of 72% recovery for PCP at low levels of adsorbed mass (obtained from previous mass balance calculations, data presented in Appendix A) , and the result showed only trace concentrations (recognizable peaks in the chromatograms , roughly equivalent to 0.05 mg/1, but not measurable with confidence) of PCP in methanol, providing an indication that adsorbed amounts of all three phenolic compounds were later desorbed and biodegraded . 5.5.2 Column Biodegradat ion II . The second column experiment involved different phenol to DCP to PCP concentration ratios, and was designed to examine the effects of PCP co-degradation in the presence of phenol. in this experiment phenol concentrations were not monitored because of the huge differences in concentrations

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116 involved. The results are presented in Table 5-22 and Figures 5-32 and 5-33. Table 5-22. Column biodegradat ion II results. lompound column (??day) STD ERR ° f K BD Uh R 2,4-DCP 1* 2** 3** 8.63x1 7 .58x10 9.36x10 ,-2 -2 6.72x10 5.07x10 5.42x10 -3 -3 -3 9.1 7.4 0.93 0.94 0.96

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117 D §1 Time (days) + §2 13 Figure 5-32. 2,4-DCP degradation curves in column b iodegradat ion study II.

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118 \ D #1(2:2:2) Time (days) + #2(20:2:2) Figure 5-33. PCP degradation curves in column biodegradat ion study II.

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119 5.5.3 Column biodegradat ion III . The third column biodegradat ion experiment was performed to examine the effects of different environments on degradation of phenolic compounds. Column #1 was held under anoxic condition, column #2 was aerated with compressed air and column #3 had been supplied with hydrogen peroxide and maintained at a minimum dissolved oxygen concentration of 3 mg/1 throughout the time of the experiment. The degradation results are presented in Table 5-23, the DHA assay data are presented in Table 5-24, and Figures 5-34 through 5-35. Table 5-23. Column biodegradat ion ill results. Compound

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120 Table 5-24. DHA data for column degradation III. Time Column 1 Column 2 Column 3 AVG. ST.ER. AVG. ST . ER . AVG . ST . ER day mg/l/day ( % ) mg/l/day ( % ) mg/l/day ( % ) 3.07x10":? 61.6 3.07x10*, 18.2 3.36x10", 7.7 5 8.56xl0_:r 23.9 1.45xl0_, 16.1 1.42xl0_, 2.9 16 2.13xl0_^ 19.6 4.09xl0_, 1.1 2.58xl0_, 31.0 43 3.45x10 l 17.4 1.08x10 14.3 1.42x10 3.1 As in previous experiments, phenol degraded rapidly, the half-life was about 11 hours for all three columns. 2 , 4-Dichlorophenol and PCP both degraded slowest in anoxic conditions (column #1) while the system that amended with hydrogen peroxide (column #3) degraded fastest but not by a wide margin. The DHA data correlated generally well with the degradation data. in column #1 (anoxic conditions) the microbial activities increased consistently but were lower than the other two systems except at the end of the experiment. However, between the two aerobic conditions (columns #2 and #3) the DHA data did not show significant difference. Column #2 demonstrated an unusual weak sorption of PCP, but the reason is unclear. The results of this experiment revealed that the three phenolic compounds degrade in both aerobic and anoxic conditions, but the degradation is less effective in an anoxic condition. The microbial populations in anoxic conditions also build up slower than in aerobic conditions.

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121 \ E #1 -Anoxic + #2-Aerated #3-Hydrgn peroxide Figure 5-34. Phenol degradation curves in column biodegradat ion study III.

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122 \ E C #1 -Anoxic Time (days) + #2-Aerated ^3-Hydrgn peroxide Figure 5-35. 2,4-DCP degradation curves in column biodegradat ion study III.

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123 #1 -Anoxic 10 20

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124 5 . 6 Hydraulic Conductivity The purpose of this experiment was to investigate whether or not the bacterial growth during the aquifer reclamation period changes hydraulic conductivity. Todd (1959) indicated some decrease in infiltration rate resulted from microbial growths clogging the soil pores. Hydraulic conductivity values of all three columns were evaluated before and after the first set of column biodegradat ion studies using the method described in Chapter 4. The results are listed in Table 5-25. The apparent hydraulic conductivity after the biodegradat ion study was corrected with the shrinkage of soil column depth (L). Table 5-25. Column hydraulic conductivity results.

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CHAPTER 6 SUMMARY AND CONCLUSIONS 6 . 1 Summary The adsorption, desorption and biodegradat ion of phenol, 2 , 4-dichlorophenol and penta-chlorophenol under simulated subsurface environmental conditions were investigated under single-compound conditions as well as in multi-compound solutions. These experiments were conducted in both batch and column systems. The soil materials were taken from the surficial sandy aquifer in an unused part of the Southwest Landfill in Archer, Florida. They were characterized as fine grained sands with negligible organic carbon content . 6.1.1 Sorption Batch sorption data were fitted with the Freundlich model to calculate the sorption coefficients. In singlecompound systems phenol had a very low adsorption coefficient (K„.) of 0.0158 on this type of sandy aquifer FA material. K for 2,4-DCP was 0.0547, and PCP was most FA adsorbed with a K„, value of 1.119. In multi-compound FA systems the three phenolic compounds compete for adsorption sites, and adsorption capacities were reduced by a margin 125

PAGE 140

126 ranging from 70% for phenol to 30% for both DCP and PCP . All three compounds exhibited non-linear sorption behavior with an average Freundlich exponent value of 0.7. Adding sludge greatly increased the adsorption capacity of the soil for all three compounds. Differences between desorption coefficients and adsorption coefficients for phenol and 2,4-DCP were small or negligible, but were significant for PCP, indicating hysteresis of PCP sorption on soil. Column sorption studies were performed in multicompound systems, the retardation factors obtained from the pore volumes at C/C =0.5 were 1.03 for phenol, 1.16 for 2,4o DCP and 2.26 for PCP. These values closely agreed with the predicted values for phenol and 2,4-DCP using the batch sorption coefficients but were at the lower end for PCP. The low retardation factors of phenol and 2,4-DCP mean that these two compounds will not be highly retarded in this type of Florida surficial aquifer. 6.1.2 Batch Biodegradat ion Batch biodegradat ion experiments determined that no extra phosphorus was needed, and the amount of nitrogen added was not critical for phenol degradation in the soil systems used. Phenol degradation was fast and was complete (below detection limit of 0.01 ppm) within three days. 2,4DCP was also completely degraded although 23 days were required to reach a level of 0.02 ppm (detection limit for

PAGE 141

127 2,4-DCP). PCP was resistant to biodegradat ion and the rates of assimilation were not high enough to support bacterial growth. The populations shifted into endogenous respiration phase and eventually became extinct. Using indigenous soil bacteria, the apparent degradation rates averaged 1.5 day (t . = 0.5 days) for phenol, 0.1 day" (t ]/2 = 7 da y s) for 2,4-DCP, and 0.006 day -1 (t, /2 = 120 days) for PCP. When tested in multi-compound systems, phenol degradation rates dropped off to 0.4 day -1 (t 1/2 = 1.7 days) but PCP degradation rates increased to 0.008 day" ^ fc i/2 = 86 da y s ^This increase was attributed to the effect of co-degradation where bacteria utilized phenol as the primary carbon and energy source to build up a larger population. Another set of batch degradation experiments using different phenol to PCP initial concentration ratios supported this hypothesis. Municipal wastewater sludge did not increase phenol degradation rates but was a good source of a wide variety of bacteria for aiding indigenous bacteria to degrade the other two phenolic compounds. However, periodic addition of sludge was required for more effective PCP removals. The addition of sodium acetate and glucose as primary substrates did not assist biodegradat ion in the sludge amended samples, presumably because the effects were overshadowed by the effects of sludge amendment. This result does not imply that sodium acetate and glucose are not good substrates for co-degradation. The enzymes required for PCP degradation did not seem to be induced by exposing the bacteria to

PAGE 142

128 either 2,4-DCP or to PCP itself. Phenol had an unclear role in inducing the PCP degrading enzymes. 6.1.3 Column Biodegradat ion Column b iodegradat ion were performed under nonsteadystate conditions in continuously recirculating setups. Biodegradat ion rates were obviously greater than in batch experiments, with the rate increase for PCP degradation being especially noticeable, partly because of larger bacterial populations in the columns, and partly because the dynamic flow conditions made the substrates more available to the bacteria. A large percentage of input PCP was initially adsorbed onto the soil but later desorbed and degraded completely. When controlled under different environments, the column which was kept under an aerobic environment by adding hydrogen peroxide degraded all three phenolic compounds fastest. In the anoxic conditions both the microbial population build-up and the rate of phenolic compounds degradation were the slowest among the three systems but not by a wide margin. 6.1.4 Hydraulic Conductivity Tests Hydraulic conductivities of the soils in each column were tested before and after the biodegradat ion experiments The bacterial growth in the columns did not have an impact on the hydraulic conductivity, which suggested the

PAGE 143

129 feasibility of applying in-situ biodegradat ion techniques to groundwater remedial problems. 6 . 2 Conclus ions The following conclusions were drawn from this research project : 1. The sorption behavior of phenol, 2,4-DCP and PCP was non-linear on soils with negligible organic carbon contents . 2. Phenol will not be retarded once introduced to groundwater systems, while higher chlorinated phenolic compounds exhibited higher retardation factors. 3. PCP desorption showed hysteresis but not phenol and 2,4-DCP desorptions. 4. On soils with negligible organic carbon content the phenolic compounds exhibited competitive sorption. 5. Extra ammonia nitrogen did not interfere with phenol biodegradat ion . 6. All three phenolic compounds were biodegradable, with the half-life of phenol in hours, PCP took a much longer time to degrade. 7. PCP alone could not support enough bacterial growth to outpace the decay. 8. Initial bacterial populations influenced both the extent and rate of degradation of PCP. 9. Periodic addition of fresh bacterial sources will be necessary when PCP is the only carbon and energy source.

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130 10. Municipal wastewater sludge was a good source of bacteria for aiding indigenous bacteria in in-situ biodegradat ion techniques. 11. There were substantial amounts of bacterial populations harbored on the soil used in these experiments. 12. In a saturated environment, more bacteria stayed in the attached state than in the free floating state. 13. Biodegradat ion in column (continuous flowing) systems proceeded faster than in batch systems. 14. PCP disappeared faster when co-degraded in the presence of phenol. 15. A PCP degrading enzyme could not be induced by either 2,4-DCP or PCP itself. 16. phenolic compounds disappeared fastest in an aerobic environment amended with hydrogen peroxide, and slowest in the anoxic conditions. 17. Bacterial activities (DHA) increased substantially during a column biodegradat ion experiment. 18. Hydraulic conductivities of the soil columns were unchanged before and after biodegradat ion experiments.

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APPENDIX A BATCH SORPTION DATA

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132 Phenol (single compound) adsorption data Co (mg/1) Ce ( ug/ml X (ug) X/ra (ug/g) log (X/m) log(Ce) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume=22 ml 9.09 6.82 4.55 0.91 8.78 6.53 4.29 0.84 6.84 6.38 5.72 1.52 0.17 0.16 0.14 0.04 -0.767 -0.797 -0.845 -1.421 0.944 0.815 0.632 -0.0757 [2ml sludge, sodium azide 2ppm] m=40g soil, solution volume=22 ml 9.09 6.82 4.55 0.91 8.56 6.35 4.12 0.79 11.68 10.34 9.46 2.62 0.29 0.26 0.24 0.07 -0.535 -0.588 -0.626 -1.184 0.932 0.803 0.615 -0.103 [no sludge, sodium azide 2ppm] m=40g soil, solution volume=20 ml 10.00

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2 , 4-dichlorophenol (single compound) adsorption data Co (rag/1) Ce (ug/ml) X (ug) X/m (ug/gl log (X/m) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume=22 ml log(Ce) 133 9.09

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134 Pentachlorophenol (single compound) adsorption data Co (mg/i: Ce (ug/ml X (ug: X/m (ug/g) log (X/m) log (Ce) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume=22 ml 9.09 6.82 4.55 0.91 1.80 0.99 0.58 0.09 160.40 4.01 128.26 3.21 87.34 2.18 18.018 0.45 0.603 0.506 0.339 -0.346 0.255 -0.00436 -0.237 -1.0458 [2ml sludge, sodium azide 2ppm] m=40g soil, solution volume=22 ml 9

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Phenol (single compound) desorption data ra=40g soil, solution volume= about 18 ml 135 Co

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136 2 , 4-dichlorophenol (single compound) desorption data m=40g soil, solution volume= about 18 ml Co Ce -x -X/m (mg/1) (ug/ml) (ug) (ug/g) log(-x/m) log(Ce) [2ml sludge, sodium azide 6ppm] 3.73

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137 Pentachlorophenol (single compound) desorption data m=40g soil, solution volume= about 18 ml Co

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138 Phenol (multiple compounds) adsorption data Co Ce X X/m log (X/m) (mg/1) (ug/ml) (ug) (ug/g) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume= 22 ml [2ml sludge, sodium azide 2ppm] m=40g soil, solution volume= 22 ml [no sludge, sodium azide 2ppm] m=40g soil, solution volume= 20 ml log(Ce) 9.

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139 2 , 4-dichlorophenol (multiple compounds) adsorption data Co (mg/1) Ce (ug/ml ) X (ug] X/m (ug/g) log (X/m) log(Ce) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume= 22 ml 9.09

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140 Pentachlorophenol (multiple compounds) adsorption data Co (mg/1) Ce (ug/ml) X (ug) X/m (ug/g) log (X/m) log(Ce) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume= 22 ml 9.09 2.43 146.542 6.82 1.67 113.300 4.55 0.80 82.500 0.91 0.11 17.578 3.664 0.564 2.833 0.452 2.063 0.314 0.439 -0.357 0.386 0.223 •0.097 •0.959 [2ml sludge, sodium azide 2ppm] m=40g soil, solution volume= 22 ml 9.09

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141 Phenol (multiple compounds) desorption data Co Ce -X -X/m log(-X/m) log(Ce) (mg/1) (ug/ml) (ug) (ug/g) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume= about 18 ml 4,

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142 2 ,4-dichlorophenol (multiple compounds) desorption data Co Ce -X -X/m log(-X/m) log(Ce) (mg/1) (ug/ml) (ug) (ug/g) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume= about 18 ml 3,

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143 Pentachlorophenol (multiple compounds) desorption data Co (mg/1) Ce (ug/ml -X (ug) -X/m (ug/g) log(-X/m) log(Ce) [2ml sludge, sodium azide 6ppm] m=40g soil, solution volume= about 18 ml 1.17

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144 Mass Balance Calculations (single-compound systems) in 1 . mass ug (1) [Sorpt ion] in mass solu . drain ug ug (2) (3) {PHENOL} 200 193. 150 143. 100 94. 20 18. {PHENOL} 200 188. 150 139. 100 90. 20 17. {PHENOL} 200 197. 150 147. 100 98. 20 19. {2,4-DCP} 200 173. 150 125. 100 79. 20 15. {2,4-DCP} 200 175. 150 127. 100 81. 20 15. {2,4-DCP} 200 190. 150 141. 100 93. 20 18. {PCP} SLG 200 150 100 20 {PCP} 200 150 100 20 {PCP} 200 150 100 20 39. 21. 12. 1. SLG 38. 23. 13. 1. No 99. 64. 32. 3. SLG,AZ=6 16 115. 66 86. 38 57. 48 10. SLG,AZ=2 32 114. 70 85. 64 55. 38 9. No SLG 40 109. 80 99. 00 55. 36 11. SLG,AZ= 36 102. 40 74. 20 46. 62 9. SLG,AZ= 12 102. 38 74. 62 47. 84 8. No SLG 20 103. 00 77. 40 52 . 00 10. ,AZ = 6 60 23. 78 13. 76 7. 98 1. ,AZ = 2 28 22. 32 13. 20 76 SLG 60 40 40 60 90 20 06 84 70 09 62 68 56 03 62 52 6 44 10 80 30 2 68 11 12 86 66 55 30 40 17 40 14 62 46 47 00 28 62 23 ,98 Rcovery= [ (7) + (5) (4 55. 34 18 1 [Desorpt ion] in mass in solu drain system ug ug ug (4) (5) (6) 84.04 45.59 38.51 63.73 34.58 29.27 42.82 23.24 19.70 9.17 4.94 4.22 3.35 1.76 83.53 3.35 1.76 63.19 3.15 1.65 42.73 2.57 1.43 8.89 89.21 46.73 43.72 58.43 34.85 16.13 43.64 23.63 20.75 8.33 4.65 3.83 84.36 43.51 54.05 64.22 33.12 42.82 43.70 21.85 31.35 9.64 4.74 5.96 86.52 43.04 54.28 66.43 32.87 43.06 46.32 23.04 29.84 10.05 5.36 5.79 94.93 50.69 45.65 70.68 37.94 34.51 47.19 24.35 23.35 9.78 4.94 4.98 31.35 16.75 159.85 15.15 8.06 128.77 10.20 5.41 87.20 3.86 2.00 16.86 29.07 15.22 162.16 16.98 8.98 127.56 11.53 6.20 86.34 3.71 1.75 17.25 50.84 26.36 118.36 34.27 18.51 96.87 18.38 9.21 72.56 4.18 2.09 15.93 / [ (5)+(6)-(4)] x 1! [ Extract ion]

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145 Mass Balance Calculations (multi-compound systems)

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APPENDIX B BREAKTHROUGH CURVE DATA Breakthrough curve data of column #1 and column #2 1 pore volume = 310 ml Flow rate = 2.7 ml/min. Bulk density = 1.45 g/ml Porosity = 0.45 Concentrations are in mg/1 Time

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APPENDIX C BATCH BIODEGRADATION DATA

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148 Phenol batch degradation data Initial concentration 5 mg./l Time (hr)

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149 Normalized phenol batch degradation data initial concentration 5 mg/1, Values in C/Co x 100% T(days) #101 #102 #103 #104 #105 #106 #107

PAGE 164

2,4-DCP batch degradation data #201-#204: Initial concentration 5 mg/1 #205-#208: Initial concentration 1 mg/1 Values in mg/1 T(hr) #201 #202 #203 #204 #205 #206 #207 #208 151 9 21 32 50 73 122 241 363 390 412 437 480 509 555 652 50 36 33 45 29 10 04 73 03 74 76 59 51 53 68 55 16 46 50 46 98 3.65 3.55 56 45 56 29 3.10 3.37 3.26 .07 .00 .00 5.00 4.54 4.47 4.54 4.57 4.51 59 50 34 51 47 18 4.06 60 33 07 09 07 3.72 39 79 51 60 0.97 0.47 0.00 91 77 77 69 58 45 31 .36 40 ,28 ,35 ,19 ,20 ,07 ,04 1.00 0.93 0.84 0.84 0.84 0.66 0.53 0.35 0.37 0.38 0.26 0.23 0.06 0.05 0.04 0.02 1.00 0.95 0.93 0.87 0.82 0.69 0.53 0.40 0.34 0.38 0.21 0.19 0.11 0.07 0.04 .00 1.00 0.91 0.84 0.82 0.78 0.76 0.52 0.43 0.21 0.14 0.04 0.06 0.03 0.03 0.00 .00 Normalized 2,4-DCP batch degradation data #201-#204: initial concentration 5 mg/1 #205-#208: Initial concentration 1 mg/1 Values in C/Co x 100% T(d) #201 #202 #203 #204 #205 #206 #207 #20 8 0.0

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Pentachlorophenol batch degradation data Values in mg/1 151 Time #301 #302 #303 #304 #305 #306 #307 #308 #309 #310 #311

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Normalized pentachlorophenol batch degradation data Initial concentration 5 mg/1 Values in C/Co x 100% 152 T(days) #301 #302 #303 #304 #305 #306 .00

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153 Normalized pentachlorophenol batch degradation data Initial concentration 1 mg/1 Values in C/Co x 100% T (days) #307 #308 #309 #310 #311 .00

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154 Phenol degradation data (in mixture) Initial concentration 5 mg/1 Values in mg/1 Time #401 (day) #402 #403 #404 #405 #406 #407 #408 0. 0. 2. 2. 4. 6. 11.0 15.0 63 43 35 33 5 4 4 4 4 4.08 3.99 4.36 4.38 4.21 4.18 58 23 83 74 66 15 02 0.21 0.16 0.13 00 29 22 29 98 86 15 07 07 04 5.0 4.7 4.2 3.7 3.5 0.0 00 67 98 00 97 93 42 10 10 07 06 5.00 4.56 3.95 4.13 3.98 2.49 1.52 0.08 0.07 0.07 0.06 22 76 43 59 20 05 06 02 00 33 68 08 01 39 0.12 0.08 0.06 0.00 . Phenol degradation data (in mixture) Initial concentration 1 mg/1 Values in mg/1 Time (day) #409 #410 #411 #412 #413 #414 #415 #416 0.0 2 0.3 0.9 2.0 2.9 4.0 6.0 11 00 94 76 ,79 78 ,75 ,79 ,75 ,46 ,37 1.00 0.96 0.78 0.75 0.68 0.66 0.11 0.06 0.04 0.01 1.00 0.85 0.71 0.64 0.09 0.07 0.04 .00 0.02 0.01 93 68 36 05 07 05 00 0. 00 95 77 73 53 16 1.00 0.93 0.70 0.67 0.63 0.64 .62 0.65 0.10 0.07 06 05 04 1 .00 0.91 0.81 0.68 0.05 0.05 0.05 0.04 0.02 0.01 1.00 0.88 0.86 0.68 0.06 0.04 0.02 0.06 0.05 0.02

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155 Normalized phenol degradation data (in mixture) Initial concentration 5 mg/1 Values in

PAGE 170

156 2 ,4-dichlorophenol degradation data (in mixture) initial concentration 5 rag/1 Values in mg/1 Time (day) #401 #402 #403 #404 #405 #406 #407 #408 0.3 .9 2.0 2.9 4.0 6.0 8.0 11.0 15.0 19.0 22.0 27 33 40 46 60 74 90 53.0 5.00 4.43 3.50 3.92 4.13 61 13 79 ,54 3.25 48 66 ,83 ,40 ,13 ,51 3.86 3.80 3.79 4.30 3.86 5.00 4.08 3.37 4.08 4.17 3.54 2.92 58 33 16 81 89 75 25 ,02 ,35 ,51 ,40 .48 .29 .14 5.00 3.86 3.17 4.17 87 22 92 96 14 25 36 93 91 4.65 4.31 51 83 84 90 71 14 71 98 95 73 22 78 09 17 23 10 82 05 95 49 64 3.87 3.76 2.65 2.58 2.10 5.00 3.90 3.10 4.04 3.87 46 97 24 40 17 88 ,54 3.52 2.83 69 44 ,54 ,96 03 10 25 87 10 83 12 09 03 19 67 01 35 95 10 32 19 97 2.17 1.68 97 13 90 32 11 25 33 38 21 99 48 90 59 85 93 71 05 ,53 ,05 00 02 06 95 80 78 3.07 3.16 15 20 07 76 40 30 25 80 68 40 33 23 78

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2 , 4-dichlorophenol degradation data (in mixture) Initial concentration 1 rag/1 Values in mg/1 157 Time (day) #409 #410 #411 #412 #413 #414 #415 #416 0.0 0.0; 0.3 0.9 2.0 2.9 4.0 6.0 8.0 11.0 15.0 19.0 22.0 27 33 40 46 53.0 60.3 74.2 90.1 1.00 0.96 0.86 0.77 0.77 0.73 0.80 0.84 0.74 0.71 0.62 0.65 0.75 0.85 0.94 0.77 0.83 0.74 0.69 0.73 0.78 1.00 0.95 0.87 0.78 0.74 0.65 0.68 0.70 0.67 0.65 0.53 0.68 0.85 1.00 0.97 0.74 0.94 0.80 0.76 0.75 0.74 1.00 0.88 0.78 0.73 0.70 0.63 0.71 0.70 0.67 0.65 0.47 0.63 0.81 1.00 1.08 0.70 0.88 0.59 0.49 0.36 0.32 1.00 0.86 0.78 0.72 0.70 0.60 0.69 0.67 0.66 0.63 0.50 0.60 0.87 0.95 1.16 0.73 0.81 0.54 0.43 0.35 0.32 1.00 0.90 0.78 0.73 0.68 0.62 0.74 0.70 0.67 0.66 0.53 0.62 0.79 0.88 0.93 0.65 0.75 0.50 0.38 0.25 0.21 .93 .78 .73 .68 .63 .76 .74 .69 .70 .55 .66 .81 .05 .05 .79 .93 .65 .61 .51 .54 1.00 0.92 0.89 0.72 0.68 0.62 0.74 0.67 0.69 0.67 0.53 0.59 0.85 0.98 1.02 0.74 0.85 0.56 0.43 0.39 0.25 1.00 0.90 0.91 0.69 0.68 0.62 0.72 0.62 0.64 0.63 0.48 0.57 0.87 0.93 1.02 0.73 0.80 0.54 0.39 0.31 0.20

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Pentachlorophenol in mixture degradation data Initial concentration 5 mg/1 Values in mg/1 158 Time #401 (day) #402 #403 #404 #405 #406 #407 #408 0.02 0.3 0.9 2.0 2.9 4.0 6.0 11 15 19 22 27 33 40 46 53 60 74 90 54 93 23 19 42 11 95 18 13 98 43 33 14 89 08 09 90 79 96 77 5.00 3.91 4.08 3.86 3.60 3.54 3.08 3.42 23 69 13 87 2.61 58 39 70 09 46 20 35 82 69 95 16 64 22 92 36 83 09 00 06 42 35 32 54 35 40 60 14 72 05 61 04 1.96 1.35 84 13 27 73 83 23 51 14 50 60 1.15 0.93 1.55 1.04 0.78 98 29 18 73 43 76 71 47 88 93 90 53 77 70 42 50 68 41 18 75 06 42 23 80 32 71 14 74 94 27 81 36 97 77 57 71 71 43 35 92 24 50 08 70 28 61 10 23 40 97 61 23 0.98 93 06 28 70 55 15 49 98 34 47 79 56 06 58 21 62 54 34 08 .67

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Pentachlorophenol in mixture degradation data Initial concentration 1 mg/1 Values in mg/1 159 Time (day) #409 #410 #411 #412 #413 #414 #415 #416 0.0 1.00 0.02 0.89 0.3 0.75 0.9 0.71 2.0 0.73 2.9 0.67 4.0 0.67 6.0 0.65 8.0 0.71 11.0 0.76 15.0 0.63 19.0 0.71 22.0 0.61 27.2 0.65 33.4 0.72 40.1 0.66 46.1 0.71 53.0 0.63 60.3 0.63 74.2 0.73 90.1 0.76 1.00 0.91 0.68 0.71 0.52 0.55 0.51 0.39 0.44 0.53 0.37 0.49 0.53 0.45 0.57 0.28 0.42 0.43 0.43 0.36 0.27 0. 0. 0. 0. 0. 0. 0. 0. 0. 0. 0. 0. 0. 0. 0. 92 55 46 30 36 32 28 34 42 26 31 28 25 32 19 18 26 22 12 1.00 0.94 0.68 0.48 0.40 0.36 0.34 0.32 0.36 0.45 0.29 0.34 0.30 0.30 0.27 0.31 0.34 0.39 0.36 0.28 0.22 1.00 0.91 0.66 0.56 0.38 0.34 0.43 0.34 0.41 0.48 0.29 0.38 0.42 0.46 0.38 0.27 0.35 0.40 0.37 0.35 0.36 1.00 0.93 0.71 0.52 0.39 0.39 0.47 0.35 0.42 0.59 0.34 0.44 0.42 0.55 0.45 0.38 0.48 0.43 0.46 0.50 0.41 1.00 0.88 0.49 0.42 0.27 0.28 0.28 0.26 0.36 0.44 0.23 0.26 0.28 0.35 0.29 0.28 0.22 0.35 0.24 0.17 0.13 1.00 0.90 0.54 0.43 0.30 0.31 0.31 0.31 0.37 0.47 0.25 0.26 0.28 0.36 0.31 0.31 0.24 0.35 0.26 0.15 0.19

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160 Normalized pentachlorophenol degradation data ( in mixture) Initial concentration 5 mg/1 Values in C/C x 100% o Time (day) #401 #402 #403 #404 #405 #406 #407 #408 0.02 0.3 0.9 2.0 2.9 4.0 6.0 11 15 19 22 27 33 40 46 53 60 74 90 100 90.1 83.2 82.9 85.2 85.7 83.0 81.0 82.2 82.1 82.6 85 86 82 80.7 80.8 78.4 77.2 77.4 63.4 100 84.5 79.7 77 .0 73.0 68.8 65.6 65.7 62.8 53.7 49 52 53 50 50 54 56 51 46 43 36 100 76 .7 63.8 59.6 5 3.3 45.0 37.1 3 6 1 8 7 3 2 7 28.8 28.2 28.7 28.7 24.2 18.6 32 35 40 40 37 31 27 27 100 85 .0 73.2 62 .2 51.6 43 .3 36.6 32.5 35.8 36.9 32.0 32 .8 33.1 29.0 27.1 26.5 28.7 31.2 28.7 22.1 15.6 101 81 73 76 75 66 53 43 47 50 48 51 48 38 33 30 30 31 28 22 15.0 101 82 75 78 75 65 54 48 52 54 51 51 47 40 35 33 33 32 29 25 19 100 80.4 69.5 66.6 61.8 53 .9 44.4 38.0 40.2 39.8 35.0 34.8 32.1 24.6 100 79. 66. 61. 56. 49 . 41. 35. 39.0 40 35 33 29 29 29.1 28 29 30 26 20 13

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161 Normalized pentachlorophenol degradation data (in mixture) Initial concentration 1 mg/1 100% Values in C/C x Time (day) #409 #410 #411 #412 #413 #414 #415 #416 0.0 0.02 0.3 0.9 2.0 2.9 4.0 6.0 8.0 11.0 15.0 19.0 22.0 27.2 33.4 40 46 53.0 60 74 100.0 88.2 77.5 72.5 71.0 68.5 66.5 67.0 70, 71 . 68 66 64 65 68 68 67 90.1 65.0 65.5 71.2 76.0 87 74 65 57 53 48 43 45 46 44 47 50 50 46 39 38 42 41 35 27 100.0 5 84.8 62.0 44.3 35.6 33.7 32.0 30.4 34.5 36.0 31.2 28.8 27.9 27.5 27.1 22 .1 20.3 23.0 20.5 13.5 89 69 51 41 36 34 33.5 37.3 38.8 34.3 31.8 31.0 29.4 29.1 31.1 34.6 37.0 34.8 28.5 22.0 87 .0 69.8 54 .0 41.5 37.4 38.8 38.1 41.0 41.5 35 36 41 42 37 31 34 38 37 35 36 100.0 89.3 71.8 53.5 42.2 40.9 42.0 39.8 44 48 42 40.9 45.7 49, 45, 42 44, 45 46 46 41 100.0 81.2 57.0 40.0 31.0 27 27 29 35 36 100.0 29.0 25.9 29 31 30 26 26.8 29.0 25.0 17.7 13.0 83 60 42 33 30 30.6 32.3 38.0 39.0 30.8 26.4 29.5 32.6 31.9 28.9 28.4 30.0 25.5 18.7 19.0

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162 Data for pentachlorophenol co-degradation with various ratios of phenol and 2 , 4-d ichlorophenol T ime (day)

PAGE 177

163 #301 (Control) f #302 Example of an untreated pentachlorophenol degradation curves in single-compound systems (initial concentration 5 ppm) .

PAGE 178

164 D #410, Phenol Time (days) + Regression Line Example of a linear regression plot of phenol degradation data using first-order reaction kinetics.

PAGE 179

APPENDIX D COLUMN BIODEGRADATION DATA

PAGE 180

Column biodegradat ion I data 166

PAGE 181

Column biodegradat ion II data 167

PAGE 182

Column biodegradat ion III data 168

PAGE 183

APPENDIX E PROCEDURES TO CALCULATE K

PAGE 184

171 I . procedures to calculate K oc 1. Assume n number of solutions with initial concentration in the interested range for the compound and in contact with m grams of soil. 2. Based on Freundlich model, calculate the total mass sorbed onto the soil particles in each solution (XI) n using the Freundlich sorption coefficients obtained from the batch study with sludge. 3 Calculate the total mass sorbed onto the soil particles in each solution (X2) using the Freundlich sorption coefficients obtained from the batch study without sludge . 4. The difference between Xl and X2 is the amount sorbed to the organic carbon (o.c.) in the added sludge. 5. Use (X1-X2) values to calculate the equilibrium concentrations (Ce) n . 6. Based on the values of [ (X1-X2) /o .c . ] n and (Ce) n an isotherm can be constructed and K QC can be calculated. II. A program for equilibrium concentration calculations: 10 INPUT M,V,CO 2 READ KF,B 30 REM M=mass of soil (g), V=volume of solution (ml), CO=initial concentration (mg/1) , B=exponent, KF=Freundlich sorption coefficient, X= total mass of compound sorbed onto soil 40 H=CO*V/25 : X2=0 60 X1=X2 : X2=X1+H 70 IF C0<=X2/V THEN GOTO 140 100 Y1=M*KF*(C0-X1/V)~B-X1 110 Yl=M*KF*(CO-X2/V)~B-X2 120 SG=Y1*Y2 : IF SG>0 THEN GOTO 60 140 X=(Xl+X2)/2 : ER=ABS((X2-Xl)/2) 150 Y=M*KF*(CO-X/V) ~B-X 160 IF Y=0 THEN GOTO 290 170 IF ER< = 0. 00001 OR (ER/X)< = 0. 00001 THEN GOTO 290 180 IF Y1*Y <0 THEN GOTO 200 190 X1=X : GOTO 140 200 X2=X : GOTO 140 290 CE=CO-X/V 300 PRINT KF, B, CE, X : GOTO 20 400 DATA KF(1), B(l), KF(2), B(2), , KF(n), B(n)

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REFERENCES Alexander, M. (1981). Biodegradat ion of Chemicals of Environmental Concern. Science 211: 132-138. Alexander, M. (1985). Biodegradat ion of Organic Chemicals. Environmental Science & Technology 19: 106-111. Amoozegar-Fard , A., W.H. Fuller, and A.W. Warrick. (1983). A Simplified Model for Solute Movement Through Soils. Soil Sci. Soc. Am. J. 47: 1047. Angley, J.T. (1987). An Evaluation of the Attenuation Mechanisms for Dissolved Aromatic Hydrocarbons from Gasoline Sources in a Sandy Surficial Florida Aquifer. Ph.D. Dissertation, University of Florida. Artiola-Fortuny, J., and W.H. Fuller. (1982). Adsorption of Some Monohydroxybenzene Derivatives by Soils. Soil Science 133: 18-26. Atlas, R.M. (1981) . Microbial Degradation of Petroleum Hydrocarbons: An Environmental Perspective. Microbiological Reviews 45: 180-209. Banerjee, S., S.H. Yalkowsky, and S.C. Valvani. (1980). Water Solubility and Octanol/Water Partition Coefficients of Organics. Limitations of the Solubility-Partition Coefficient Correlation. Environmental Science & Technology 14: 1227-1229. Bartha, R., and R.M. Atlas. (1977). The Microbiology of Aquatic Oil Spills. Advanced Applied Microbiology 22: 225-266. Bedient, P.B., R.C. Borden, and D.I. Leib. (1985). Basic Concepts for Ground Water Transport Modeling. In Ground Water Quality . C.H.Ward, W. Giger, and P.L. McCarty (eds) . New York, NY: John Wiley & Sons, Inc., pp. 512531. Benefield, L.D., and C.W. Randall. (1980). Biological Process Design for Wastewater Treatment . Englewood Cliffs, NJ: Prentice-Hall, Inc.

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181 Wilson, J. T., J.F. McNabb, D.L. Balkwill, and W.C. Ghiorse. (1983). Enumeration and Characterization of Bacteria Indigenous to a Shallow Water-Table Aquifer. Ground Water 21: 134-142. Wong, A.S., and D.G. Crosby. (1978). Photolysis of Pentachlorophenol in Water. In Pentachlorophenol ; Chemistry, Pharmacology, and Environmental Toxicology . K.R. Rao (ed) . New York, NY: Plenum Press, pp. 19-25. Yaniga, P. (1982). Alternatives in Decontamination for Hydrocarbon-Contaminated Aquifers. Ground Water Monitoring Review 2 (No. 4): 40-49. Yaniga, P.M., and W. Smith. (1985). Aquifer Restoration: In Situ Treatment and Removal of Organic and inorganic Compounds, in Proceedings of Symposium on Ground Water Contamination and Reclamation . American Water Resources Association, pp. 149-164. ZoBell, C.E. (1946). Action of Microorganisms on Hydrocarbons. Bacter iolog ial Reviews 10: 1-49. ZoBell, C.E. (1969). Microbial Modification of Crude Oil in the Sea. In Proceedings of Joint Conference on Prevention and Control of Oil Spills . American Petroleum institute, Washington, D.C., pp. 317-326.

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BIOGRAPHICAL SKETCH Chen Hsin Lin was born on April 12, 1953 in Taipei, Taiwan. He graduated from Cheng-Kung University with a Bachelor of Science degree in civil engineering in 1976. He worked as an assistant engineer with Ta-Shuan Engineering, Inc., Taipei, Taiwan for one year between 1976 and 1977. He passed the Professional Environmental Engineer Examination in 1977 and joined the Taipei Water Department as an associate engineer. He came to the United States in 1983 when he was accepted for graduate study at Auburn University, Department of Civil Engineering, where he completed the Master of Science degree in June 1985. He attended the University of Florida, Department of Environmental Engineering Sciences, in August 1985, and is currently a candidate for the Doctor of Philosophy degree in environmental engineering. He was married to Lily Chen in January 1979 and they have two boys, Albert, 8 years old and Andrew, 6 years old. 182

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T certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy. Wesley/ Camar Miller, Chairman Professor of Environmental Engineering Sciences I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy. W. Emmet t Bolch Professor of Environmental Engineering Sciences I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy.
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I certify that T have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy Daniel P. Spangle, Associate Professor of Geology This dissertation was submitted to the Graduate Faculty of the College of Engineering and to the Graduate School and was accepted as partial fulfillment of the requirements for the degree of Doctor of Philosophy. December, 198 3 Dean, Graduate School

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