Title: Biodegradation of selected phenolic compounds in a simulated sandy surficial Florida aquifer /
Full Citation
Permanent Link: http://ufdc.ufl.edu/UF00097396/00001
 Material Information
Title: Biodegradation of selected phenolic compounds in a simulated sandy surficial Florida aquifer /
Physical Description: vii, 182 leaves : ill. ; 28 cm.
Language: English
Creator: Lin, Chen Hsin, 1953-
Publication Date: 1988
Copyright Date: 1988
Subject: Phenols -- Biodegradaton   ( lcsh )
Pentachlorophenol -- Biodegradation   ( lcsh )
Aquifers   ( lcsh )
Environmental Engineering Sciences thesis Ph. D
Dissertations, Academic -- Environmental Engineering Sciences -- UF
Genre: bibliography   ( marcgt )
non-fiction   ( marcgt )
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
Bibliographic ID: UF00097396
Volume ID: VID00001
Source Institution: University of Florida
Holding Location: University of Florida
Rights Management: All rights reserved by the source institution and holding location.
Resource Identifier: alephbibnum - 001128606
oclc - 20117325
notis - AFM5810


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7_ 1. 77
E J-

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


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


This work could not have been completed without the

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


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

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

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

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


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


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
REFERENCES...... ........ ............................... 171

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


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



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


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




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


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.


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)

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

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


With the data from this research, the behavior and fate

of these important phenolic compounds in a shallow Florida


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

development of treatment methods for remediating aquifers

contaminated by such phenolic compounds.


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


(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.


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


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
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
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)

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
and WS which incorporated a melting point correction term as

log K = 6.5 0.89 ( log WS ) 0.015 ( MP ) (3-6)
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
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

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


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


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,


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
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


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
-dC/dt = k [C) = C/L [1/k1 + R T/H k ]1 (3-10)
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)

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:
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


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


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.,


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


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



c t l







Figure 3-1.

Structural formulas for some common PCP
degradation products.



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.


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


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


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


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


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


Over fiow

K= QL / AhA

Figure 4-1.

Experimental setup for hydraulic conductivity

A h



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
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,


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


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


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


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


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


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



Figure 4-2.

Experimental setup for conservative tracer



Mixture solution '
circulation line

Pum r


--i r

v1iii iiii

Sampling .
point, o1 |i^ft l

\ T

A ----

Figure 4-3. Experimental setup for column degradation

. I

"i r


- 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


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


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


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.


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.


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

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
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


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).

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


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

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

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
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


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
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

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

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

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
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


+ 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?-





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

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)]




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

3 --I -

o i
+ 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-
Id --


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






0 -_



-1.2 -0.5 -0.4 0

Log [Ce(ug/mI)]



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


--, --"

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

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

SMix-Adsp -
Mix-Desp -



-4 T,--- "--

SSng- p --
SM-s - -


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
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
a complete mixing mode was assumed, to 6.0 mg/l if a plug

flow mode was assumed. However, the highest concentration


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


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)

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.





0.7 /
0.5 -

0.4 -


0.2 -


0 0.4 0.8 1.2

Pore Volume
0 chloride + phenol

1.6 2 2.4 2,8


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

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


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

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

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

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 [

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).










0.1 -


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


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)

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.

\ "

Added 0.5 n



__ __ __ __ft __ _

0 4

I 4z1ni

8 12

4 f#20Z

5 20 24

CO .4-0

Time (daye)

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


I / I





E 0.6 -



S 0.4 -

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).

0 20D


o #301


Time (ds)
+4- 302



, -------

O- i ",

I- 1

I '
0 \
| \

0. i---



r_ -_I

,/ I.. .i

' .. -

Added 0.5 ml

ge Supernatant


40 6o

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

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



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






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