600284017
BIOLOGICAL COUNTERMEASURES FOR THE
CONTROL OF HAZARDOUS MATERIAL SPILLS
by
Neal E. Armstrong
Ernest F. Gloyna
Orville Wyss
The University of Texas at Austin
Austin., Texas 78712
Grant No. R 802207
Project flfficer '
Joseph P. Lafornara,
Oil and Hazardous Materials Spills Branch
Municipal Environmental Research Ldbqrat^ (Cincinnati)
'"
REARCH LABORATORY
OFFlCt';OF '
U.S. EWiRONMEtiTAl: PROTECTION AGENCY
CINCINNATI, OHIO 45268
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DISCLAIMER
The information in this document has been funded wholly or in part by the
United States Environmental Protection Agency under Grant No. R 802207 to
the University of Texas. It has been subject to the Agency's peer and
administrative review, and it has been approved for publication as an EPA;
document. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
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FOREWORD
The U.S. Environmental Protection Agency was created because of
increasing public and government concern about the dangers of pollution to
the health and welfare of the American people. Noxious air, foul water,
and spoiled land are tragic testimonies to the deterioration of our natural
environment. The complexity of that environment and the interplay of its
components require a concentrated and integrated attack on the problem.
Research and development is that necessary first step in problem
solution; it involves defining the problem, measuring its impact, and
searching for solutions. The Municipal Environmental Research Laboratory
develops new and improved technology and systems to prevent, treat, and
manage wastewater and solid and hazardous waste pollutant discharges from
municipal and community sources, to preserve and treat public drinking
water supplies and to minimize the adverse economic, social, health, and
aesthetic effects of pollution. This publication is one of the products of
that research and provides a most vital communications link between the
researcher and the user community.
A number of methods, including biological countermeasures, have been
considered for the control of hazardous material spills. Biological
degradation, while attractive in some respects, suffers from several
difficulties: the necessity of having on hand large quantities of
acclimated cultures; problems associated with stockpiling many such
cultures, each of which is specific to a particular substance; and the
apparent resistance of many hazardous materials to biological degradation.
This report summarizes an investigation on the feasibility of using
microbiological processes to mitigate hazarous material spills in
watercourses and should be of interest to all those concerned with building
up an arsenal of countermeasures for hazardous material incidents.
Francis T. Mayo, Director
Municipal Environmental Research Laboratory
ill
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ABSTRACT
The possibility of accidental spills of hazardous substances poses a constant
threat to the waters of the nation. Effective ways to control such spills and to
mitigate their effects include physical and chemical techniques, but biological
countermeasures have not been considered feasible to date. Determining the
feasibility of this countermeasure was the primary focus of this study.
Using the hazardous compounds, phenol and methanol, as test substances,
treatability studies were performed using acclimated bacteria to estimate their
growth kinetic and substrate removal rates and the effects of those coefficients of
environmental variables such as temperature, pH, and salinity in ranges found typically
in fresh and estuarine waters. Numerical and graphical methods were developed to
select the required amount of bacterial solids to remove some intial amount of phenol
and methanol within a selected period of time in situations approximating a contained
spill. The biological countermeasure's effectiveness was tested in simulated spill
situations in lentic and lotic environment laboratory systems, and the deleterious
effects of applying the countermeasure were examined through tests involving oxygen
depeltion and alterations in primary production.
Biological countermeasures were shown to be a feasible method for hazardous
material spill removal within certain limitations imposed by the toxicity of the
material to bacteria and its initial concentration.
This report was submitted in partial fulfillment of Grant No. R802207 by the
University of Texas at Austin under the sponsorship of the U.S. Environmental
Protection Agency. This report covers the period March, 1973 to
March, 1975.
IV
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CONTENTS
Foreword iii
Abstract . iv
Figures vii
Tables . . . xiii
Abbreviations and Symbols J - . xvi
Acknowledgment xviii
1. Introduction 1
Need for Study 1
Objectives of Study 2
Scope of Study 3
2. Conclusions 4
3. Recommendations 8
4. Development of Information for Biological Countermeasure
Feasibility Determination 9
Element of Spill Control 9
Requirements of Countermeasure 9
Information Needed 10
5. Selection of Test Materials 18
Introduction 18
Initial Selection 18
Final Selection 22
6. Literature Review 23
Introduction 23
Acetone Cyanohydrin 23
Acrylonitrile 25
Aldrin 27
Benzene 29
Isoprene 33
Methanol 33
Nitrophenol 35
Noryl phenol 36
Phenol 37
Styrene 40
Toxaphene 41
Xylene 43
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7. Screening Treatability Tests 46
Introduction 46
Procedures 46
Results 47
8. Biological Countermeasure Treatment System 64
Introduction 64
Substrate Removal and Bacterial Growth Kinetics . . 64
Experimental Methods 66
Test Results 70
9. Simulated Spill Tests 118
Introduction 118
Aquarian Tests 118
Pond Tests 135
Model Lake Tests 138
10. Countermeasure Storage 160
Introduction 160
Preservation and Recovery of a Mixed Bacteria
Culture 160
Preservation and Recovery of a Yeast Culture .... 163
11. Countermeasure Application 165
Introduction 165
Experimented Methods 165
Organization of Biological Treatability Data for
Countermeasure Design 170
Portable Treatment Systems 171
In Situ Applications of the Biological
Countermeasure 187
References 226
Appendices 241
A. Hypothesis Tests 241
B. Methanol Stripping 245
C. Computation Results of Cloth Bag Efficiency, Bacterial
Growth, and Aeration Time for the Cloth Bag
Application Tests in Batch Reactor (Reactor #1) . 250
D. Application of Cloth Bags in One-Dimensional System . . 252
VI
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FIGURES
Number
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
Microograni sin-substrate relationship
Optimum acrylonitrile concentration for detoxification by
whole cells of Azobacter vinelandii
Relationship between p-nitrophenol concentration and light
adsorption
p-Nitrophenol intensive sampling - July 8 -10, 1974 . . .
Degradation of phenol measured as total organic carbon . .
Degradation of phenol measured as total organic carbon . .
Degradation of phenol using various inoculums
Degradation of phenol using various inoculums
Effect of phenol/inoculum mass ratio on phenol removal . .
pH and salinity effects on the decomposition of phenol by
acclimated sludge
pH and salinity effects on the decomposition of phenol by
acclimated sludge
pH and salinity effects on the decomposition of phenol by
acclimated sludge
Temperature coefficient, $, for the decomposition of phenol
by acclimated sludge
The relationship between the substrate removal rate
coefficient and the cell decay coefficient for phenol
The relationship between the substrate removal rate
coefficient and the cell decay coefficient for phenol
acclimated sludge
Nutrient (N+P) effects on the decomposition of phenol
by acclimated sludge
Nutrient (N+P) effects on the decomposition of phenol
by acclimated sludqe
Page
15
50
54
56
57
58
60
61
62
78
79
80
82
83
84
85
86
VII
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FIGURES (Continued)
Number Page
18 Cell synthesis without nitrogen and phosphorous
(I)'pH buffered 88
19 Cell synthesis without nitrogen and phosphorous
(II) pH not buffered 89
20 pH variations owing to the decomposition of phenol
by acclimated activated sludge 92
21 Initial lag phase in the decomposition of phenol
by acclimated activated sludge 94
22 Comparison of theoretical and observed oxygen uptake
rates in the decomposition of phenol by acclimated
activated sludge 96
23 pH and salinity effects on the decomposition of methanol
by acclimated sludge 101
24 pH and salinity effects on the decomposition of methanol
by acclimated sludge 102
25 pH and salinity effects on the decomposition of methanol
by acclimated sludge 103
26 Temperature coefficient, 6, for the decomposition of
methanol by acclimated sludge 104
27 Temperature coefficient, 0, for the decomposition of
methanol by acclimated sludge 105
28 The relationship between the substrate removal rate
coefficient and the cell decay coefficient for methanol
acclimated sludge 107
29 The relationship between the substrate removal rate
coefficient and the cell decay coefficient for methanol
acclimated sludge 108
30 Nutrient effects on the decomposition of methanol by
acclimated activated sludge 109
31 pH variations owing to the decomposition of methanol by
acclimated activated sludge 110
vm
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FIGURES (Continued)
Number
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
The relationship between substrate removal rate and
oxygen uptake rate
Initial lag phase in the decomposition of methanol by
acclimated activated sludge
Comparison of theoretical and observed oxygen uptake
rates in methanol decomposition
Configuration of aquaria in laboratory
Analytical results of removal test No.l, Aquarium 2
(phenol) only .
Analytical results of removal test No.l, Aquarium 3
(phenol + bacteria)
Dissolved oxygen values during removal test No.l
Gross production (I) and total respiration (II) in
aquarium 1 (control) removal test No.l
Gross production (I) and total respiration (II) in
aquarium 2 (phenol) removal test No.l
Gross production (I)and total respiration (II) in
aquarium 3 (phenol + bacteria)
Schematic diagram of model lake
Diagram of sampling stations
Standard curves for dye concentration
Pattern of dye dispersion
Diagram of sampling stations, phenol dye spill with sludge
and without barrier
Barrier for phenol/dye spill
Sampling stations and barrier positions
TOC measurements during phenol/dye spill with acclimated
sludge
Page
-
112
113
116
120
123
124
126
127
128
129
140
141
143
144
146
150
151
153
IX
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FIGURES (Continued)
Number
50
51
52
53
54
55
56
57
58
59-1
59-2
59-3
59-4
59-5
59-6
59-7
59-8
Dissolved oxygen levels following phenol /dye spill with
acclimated sludge in model lake
Diagram of barrier with semi-rigid sides
Phenol/dye disappearance in model lake tests with
unacclimated sludge
Dissolved oxygen concentrations following phenol/dye spill
in model lake with unacclimated sludge
Model river system
Details of cloth bag
Details of confining barrier
Essential parts of a CSTR system
The relationship between waste sludge and total biomass
in the CSTR
Batch kinetic diagram for phenol for dilute VSS
at k=0. 01892 hr -T
Batch kinetic diagram for phenol for concentrated VSS
at k=0. 01892 hr -T
Batch kinetic diagram for phenol for dilute VSS
at k=0. 02729 hr-1
Batch kinetic diagram for phenol for concentrated VSS
at k=0. 02729 hr -1
Batch kinetic diagram for phenol for dilute VSS
at k=0. 03934 hr-1
Batch kinetic diagram for phenol for concentrated VSS
at k=0. 03934 hr -'
Batch kinetic diagram for phenol for dilute VSS
at k=0. 05674 hr~l
Batch kinetic diagram for phenol for concentrated VSS
at k=0. 05674 hr -T
155
156
157
158
166
168
169
172
173A
175
176
177
178
179
180
181
182
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FIGURES (Continued)
Number Page
59-9 Batch kinetic diagram for phenol for dilute VSS
at k=0.07239 hr~l 183
59-10 Batch kinetic diagram for phenol for concentrated VSS
at k=0.07239 hr -" 184
60 Relative efficiency of CSTR in phenol removal 186
61 Phenol transportation in the model river 189
62 Methane"! transportation in the model river 190
63 Methanol removal in one dimensional dispersion system ... 192
64 The relationship between stream velocity and material
exchange coefficient for methanol sludge 195
65-1 Efficiency of cloth bag filled with phenol sludge 199
65-2 Efficiency of cloth bag filled with phenol sludge 200
65-3 Efficiency of cloth bag filled with phenol sludge 201
65-4 Efficiency of cloth bag filled with phenol sludge 202
65-5 Efficiency of cloth bag filled with phenol sludge 203
65-6 Efficiency of cloth bag filled with phenol sludge 204
65-7 Efficiency of cloth bag filled with phenol sludge 205
65-8 Efficiency of cloth bag filled with phenol sludge 206
65-9 Efficiency of cloth bag filled with phenol sludge 207
66-1 Efficiency of cloth bag filled with methanol sludge .... 208
66-2 Efficiency of cloth bag filled with methanol sludge .... 209
66-3 Efficiency of cloth bag filled with methanol sludge .... 210
66-4 Efficiency of cloth bag filled with methanol sludge .... 211
66-5 Efficiency of cloth bag filled with methanol sludge .... 212
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FIGURES (Continued)
Number
66-6
66-7
66-8
66-9
67
68
69
B-l
B-2
Efficiency of cloth bag filled with methanol sludge ....
Efficiency of cloth bag filled with methanol sludge ....
Efficiency of cloth bag filled with methanol sludge ....
Efficiency of cloth bag filled with methanol sludge . . .
Methanol removal using cloth bags in batch reactors . . .
i . •
Phenol removal using a fixed barrier with sludge
Methanol removal using a fixed barrier with sludge
containing cloth bags
The relationship between the air stripping rate and the
reactor depth
Temperature effect on the volatilization rate coefficient .
Page
213
214
215
216
218
224
225
247
249
XII
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TABLES
Number
1
2
3
4
5
6
7
Contract List of Hazardous Materials
Physical/Chemical Characteristics of Contract Compounds. .
Results of Literature Review and Screening Tests
Air Stripping Constants for Selected Chemicals
Volatilization of Benzene in Stirred and Unstirred
Containers
Estimated Kinetic Parameters for Phenol
Kinetic Parameters for Phenol Corresponding
to kc=236 ma/1 and a=l .21
Page
19
20
21
51
52
75
77
The Relationship Between Substrate Removal Rate
Coefficient and Cell Decay Coefficient for
9
10
11
12
13
14
15
Nutrient (N,P, and Minerals) Effects on the Decomposition
of Phenol by Acclimated Sludge
Phenol Decomposition by Acclimated Activated Sludge in
Initial Lag Phase in Phenol Decomposition by Acclimated
Sludge
Comparison of Theoretical and Observed Oxygen Uptake
Rates in Phenol Decomposition by Acclimated Sludge ....
Estimated Kinetic-Parameters for Methanol
Kinetic Parameters for Methanol
The Relationship Between Substrate Removal Rate
Coefficient and Cell Decay Coefficient for
Methanol -Acclimated Sludqe
90
91
93
97
99
100
106
xm
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TABLES (Continued)
Number
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30A
30B
31
32
33
34
Methanol Decomposition by Acclimated Activated
Sludge in Natural Systems Without Chemical Aids
Initial Lag Phase in Methanol Decomposition by
Acclimated Activated Sludge
Estimated Kinetic Parameters for Paranitrophenol
Water Depths and Volumes of Aquaria
Simulated Spill of Phenol in Aquaria Removal Test No.l . .
Simulated Spill of Phenol in Aquaria Removal Test No. 2 . .
Effects of Phenol/VSS Mass Ratio on Spill
Removal Rate
Effects of Nutrient (Nitrogen and Phosphorus)
Addition With Bacteria for Control of Phenol Spill
Effects of Methanol/VSS Mass Ratio on Methanol Removal Rates
in Aquaria Test No. 1
Effects of Methanol/VSS Mass Ratio on Methanol Removal
in Aquaria Test No. 2
Effects of Nutrient Additions on Methanol Removal
Results of Methanol Spill into Ponds
Phenol Spill into Ponds
Model Lake Spill Tests
Sludge Feeding/Acclimation Schedules: Nutrients Fed
Daily to Sludge
Sludge Feeding/Acclimation Schedule
Buffer-Salts Medium
Survivors of Freeze-Storage
Survivors of Freeze-Storage Quantitated by ATP Determinations
Survivors of Lyophilization as Determined by Plate
Counts
Page
111
114
115
119
122
130
131
132
133
134
135
137
138
138
148
148
160
161
161
162
XIV
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TABLES (Continued)
Number Page
35 Survivors of Lyophilization as Quantitated by 162
ATP Determination
36 The Effects of Various Rehydration Fluids on Recovery
37
38
39
40
from Lyophil ization by Phenol -Utilizing Mixed Cultures . .
Survival of Phenol-Utilizing Geotrichum
Material Exchange Coefficient Related to the Turbulence
in Confined Reactors
Material Exchange Coefficient Related to the Stream
Velocity in Model River (Methanol Sludge)
Experimental Conditions and Kinetic Information for
the Methanol Removal Batch Tests Using Cloth Bags
163
164
196
196
217
41 Experimental Conditions and Kinetic Information for the
Phenol and Methanol Removal Batch Tests Using Fixed
Confining Barriers and Sludge Containing Cloth Bags .... 223
xv
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ABBREVIATIONS AND SYMBOLS
a --Cell yield coefficient (mg/1 VSS produced)/(mg/1 TOG utilized)
a1 —Oxygen required per substrate utilized (mg/1 O2 required)/(mg/l TOC utilized)
Aa —Air-water interface area (L2)
Ac —Surface area of cloth bag (L2)
Ar —Cross-sectional area of river (L2)
As —Surface area of water body at quiescent condition (L2)
b' —Oxygen consumption rate for endogeneous respiration (1/T), (mg/1 02
required)/(mg/l VSSVhr
C —Material exchange coefficient (L/T), cm/hr
C»C1>C2 —constants
D —Dispersion coefficient for dissolved pollutants (L2/T)
D1 —Dispersion coefficient for cloth bags (L2/T)
Da —Average diameter of air bubble
E —Cloth bag efficiency
e —Surface area expansion coefficient caused by turbulence
f —Correction factor for the substrate removal rate coefficient, k
fCe —Cloth bag distribution function in river (1/L)
—Free energy of oxidation at standard condition, Kcal/mole
G —Mean temporal velocity gradient (1/T), sec~l
g —gravity constant (L/T2)
H —Hydraulic mean depth of river (L), or reactor depth (L)f t,
hf —Head loss (L)
Hmin —Reactor depth that provides the minimum stripping rate
I —Energy gradient
k —Substrate removal rate coefficient for biological decomposition (1/T), hr~l
kT —k at temperature T°C (1/T), hr"1
ka —reaeration rate coefficient (1/T), hr~* or day-1
k
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Qw —Waste sludge flow rate (L3/T)
R —Hydraulic radius (L), m
r —Relative efficiency of CSTR system
R! —Residual in the multinomial regression for the evaluation of bacterial growth
kinetics
R2 —Residual in the multinomial regression for the evaluation of substrate removal
kinetics
Rr —Oxygen utilization rate (M/L3T), (mg/1 C>2)hr
Rri —Observed Rr
Rr2 —Theoretically computed Rr
S —Substrate concentration (M/L3), mg/1 as TOC
Sc —Substrate concentration inside a cloth bag (M/L3), mg/1 as TOC
Se —Effluent substrate concentration (M/L3), mg/1 as TOC
So —Influent substrate concentration or substrate concentration at time to
(M/L3), mg/1 as TOC
Ss —Saturated volatile substance concentration for a given partial vapor pressure
(M/L3)
T —Temperature, °C
t -Time (T)
ta —Sludge application time measured from the pollutant spill (T)
t
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ACKNOWLEDGMENTS
The authors gratefully acknowledge the contributions of the following
research associates and assistants who took part in the project: Ronald D.
Humphrey, Curtis E. Eklund, Melvin 0. Hinson, Or., H. William Hoffman,
David Peter, Alan L. Goldstein, Jim C. Spain, Robert Wetegrove, V. Nadine
Gordon, Jung W. Kim, George Oliver, Jr., And Kenneth Aicklen. Much of the
experimental work was done at the Center for Research in Water Resources
and special thanks are due to Leo R. Beard, Technical Director of the
Center, and Frank R. Husley, Technical Assistant, for their assistance.
The authors also wish to thank Mr. Thomas H. Roush and Dr. Joseph P.
Lafornara of EPA's Municipal Environmental Research Laboratory - Ci.,
Edison, NJ 08837.
xvm
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SECTION 1
INTRODUCTION
NEED FOR STUDY
The possibility of accidental spills of an ever increasing volume and array of
hazardous substances produced and shipped by industry poses a constant threat to the
waters of the nation. These spills cause varying degree of hazard and damage to a
watercourse, depending on the nature and uses of the watercourse and the type and
quantity of material spilled. The Federal Water Pollution Control Act of 1972
declared that the policy of the United States was to prohibit the discharge of
hazardous substances into navigable waters of the United States, and in Section
311(c)(2) of the Act provisions were made for the preparation of a national contin-
gency plan for the removal of oil and hazardous substances. This plan, published
August 13, 1973 in the Federal Register as the National Oil and Hazardous Substances
Pollution Contingency Plan (40 CFR 1510), delineates the actions that may be taken to
respond to a spill of hazardous materials. These actions are: Phase I - Discovery and
Notification, Phase II - Evaluation and Initiation of Action, Phase III - Containment
and Countermeasures, Phase IV - Removal, Mitigation, and Disposal, and Phase V -
Documentation and Cost Recovery.
Containment and removal or mitigation of the spilled hazardous substance are
part of Phases III and IV. In Phase III, Containment and Countermeasures, defensive
actions are to be initiated as soon as possible after discovery and notification of a
discharge. These actions may include, among other things, the placement of physical
barriers to halt or slow the spread of the pollutant and its effects on water-related
resources. In Phase IV, Cleanup, Mitigation, and Disposal, actions are taken to recover
the pollutant from the water and affected public and private shoreline areas, and mon-
itoring activities are initiated to determine the scope and effectiveness of removal
actions. Actions that may be taken include: (1) the use of sorbers, skimmers, and
other collection devices for floating pollutants, (2) the use of vacuum dredges or other
devices for sucking pollutants, (3) the use of reaeration or other methods to minimize
or mitigate damage resulting from dissolved, suspended, or emulsified pollutants, or
(4) special treatment techniques to protect water supplies and wildlife resources from
continuing damage (including biological Countermeasures).
A number of methods, including biological Countermeasures, has been considered
for the control of hazardous material spills. These methods have been reviewed by
Dawson et al. (1972), who concluded that:
"biological degradation, while attractive in some respects, suffers from
several difficulties. In order for degradation to proceed at a rapid rate, it
would be necessary to have on hand large quantities of acclimated cultures.
The problems associated with stockpiling many such cultures, each of
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which is specific to a particular substance, are obvious. Also, many
hazardous materials are apparently resistant to biological degradation."
While this evaluation points out the difficulties of biological countermeasures, it does
not rule out their use. The extent of the difficulties may not be as great as first
thought and the possibility of overcoming them has not been specifically investigated
with regard to spilled hazardous substances.
A large body of literature and experience exists in the waste treatment field
that could contribute to a further evaluation of biological countermeasures. For
example, most or all of the hazardous materials shipped and spilled are products of
some manufacturing process that produces a waste residual. This waste may contain
varying amounts of the chemical manufactured and is usually treated before being
discharged to the environment to remove the hazardous substances from the waste
stream or to reduce their concentration to the extent that they would no longer be
considered hazardous. As a result, waste treatment technology offers possible
solutions to the control of hazardous material spills. In particular, the use of
microorganisms in biological waste treatment has been developed to a sophisticated
state and has been applied to most types of industrial wastes. Biological waste
treatment has shown to be highly effective in removing a large number of hazardous
substances from waste streams .as long as these toxic substances are used by the •
microorganisms as a food substrate (Dawson et al., 1972). In addition, it is often
possible to acclimate bacteria to seemingly nonbiodegradable materials. The feasi-
bility of biodegradation has also been demonstrated by recent reports (Chears and
Myers 1973; and A.P.I., EPA, and U.S.C.G 1973) that indicate that bacteria may be
viable countermeasures for the mitigation of oil spills.
Bacterial cultures may be stored for long periods of time in a dormant state and
then later constituted for use. In the dormant state, the bacteria are in a frozen or
powder form, amenable to storage and to rapid, easy deployment without highly
specialized equipment.
Because the use of microorganisms as a countermeasure for hazardous material
spills appears to fit the criteria for potential countermeasures suggested by Dawson et
al., (1972) and because little information is available to adequately assess the
feasibility of using microorganisms as a biological countermeasure, the need for such a
study became evident.
OBJECTIVES OF STUDY
This investigation, entitled "Biological Countermeasures to Mitigate the Effects
of Hazardous Material Spills," was initiated and funded by the Environmental Protec-
tion Agency (Grant //R802207). The overall objective of the study was to investigate
the feasibility of using microbiological processes to mitigate hazardous material spills
in watercourses. Several more specific objectives were defined as follows:
1. Investigate the response requirements for any hazardous material spill and
determine the response requirements using microorganisms as a biological counter-
measure.
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2. Conduct screening tests to determine candidate microorganisms for miti-
gating the effects of certain of the Environmental Protection Agency's priority-ranked
soluble hazardous polluting substances.
3. Conduct small-scale ecological system studies to assess total ecosystem
response to these hazardous materials and their decomposition products.
4. Select candidate microorganisms for the priority list of hazardous materials
and conduct simulated spill experiments on a laboratory scale, deploying these
microorganisms.
5. Develop production, storage, reculture, and deployment methods for the
microorganisms selected.
6. Evaluate the practical feasibility of biological countermeasures for miti-
gation of hazardous material spills.
It was the intent of this work to emphasize the evaluation of the feasibility of
using microorganisms as a countermeasure to mitigate the effects of hazardous
material spills in the environment. Development of data on: (1) growth requirements
and environmental factors affecting growth of microorganisms found to successfully
break down the hazardous materials, (2) the fate of these materials and their by-
products in ecological systems, and (3) small spills into small ecological systems was
considered necessary for evaluation of the feasibility of using biological counter-
measures for mitigating effects of hazardous materials.
SCOPE OF STUDY
To accomplish the stated objectives, it was necessary to conduct literature
surveys to determine the experience of other investigators in using microorganisms to
break down hazardous materials and to determine the information available on the
effects of these hazardous materials in ecological systems.
The experimental work included laboratory culture of microorganisms, starting
with enrichment cultures to assess the types of microorganisms available to break
down various hazardous materials, followed by the determination of growth rate
characteristics of these microorganisms, including effects of environmental factors.
Other laboratory tests were conducted using small, contained, aquatic ecological
systems to determine the fate of the selected hazardous materials and of their
breakdown products in the water, sediment, and biota of natural systems. Finally,
laboratory tests were also conducted on simulated spills of these hazardous materials
in contained, aquatic ecological systems. These systems were large enough to
represent portions of several types of environmental systems so that the results could
be realistically applicable to natural systems, and the feasibility of using micro-
organisms as a countermeasure could be rationally assessed.
Investigations were made into existing programs for responding to spills of
hazardous materials, the response requirements of such programs, and the require-
ments of such programs were they to use biological countermeasures instead of
chemical or other measures. In the latter stages of this project an evaluation was
made of the tactical feasibility of using microorganisms as a countermeasure for the
spill of hazardous materials.
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SECTION 2
CONCLUSIONS
The conclusions of this study are as follows:
1. Based on criteria for assessing potential countermeasures for mitigating
hazardous material spills, biological countermeasures appear useful because:
a. Microorganisms are highly effective in removing certain hazardous
materials.
b. Microorganisms that attack a variety of hazardous materials exist.
(e.g., the Pseudo monads).
c. It should be possible to easily and rapidly deploy microorganisms in
situ or in a portable treatment system in a fresh liquid state, a
powdered state, or a freshly reconstituted state.
d. Potentially harmful secondary effects should be minor because
microorganisms are a natural part of the aquatic environment,
pathogenic bacteria will not likely constitute a significant part, if
any, of the countermeasure, noxious sludge should not be formed, and
microorganisms should not persist since they should metabolize their
own protoplasm following consumption of the hazardous material and
disappearance of the food source.
2. Previous investigations of some of the most well-known hazardous materials
show that the majority are biodegradable.
3. Based on treatability tests for phenol, methanol, and nitrophenol:
a. The following kinetic equations:
dS- kdX and dS _ _
a -
dt
satisfactorily described the bacterial growth and substrate removal
kinetics using phenol and methanol, where X is biomass concentration
(mg/1), S is substrate concentration (mg/1), t is time, a is cell yield
coefficient (biomass produced/substrate utilized), k
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b. The cell yield coefficient, a, and the Michaelis-Menten constant, Ks,
changed insignificantly with temperature.
c. The Michaelis-Menten constant, Ks, was estimated to be 236 mg/1
with a standard deviation of 70 mg/1 for phenol and 2,330 mg/1 with
a standard deviation of 1,410 mg/1 for methanol, based on total
organic carbon (TOC).
d. The cell yield coefficient, a, was estimated to be 1.21 with a
standard deviation of 0.06 for phenol and 1.25 with a standard
deviation of 0.45 for methanol, based on TOC (mg/1) and volatile
suspended solids (VSS)(mg/l).
e. The endogeneous respiration rate was closely related to the substrate
utilization rate coefficient. Thus, kj = 0.066 K°-87 and kj = 0.0115
[<0.634 (k and ^^ are based on the unit of hr~l) are proposed for the
prediction of cell decay coefficients for phenol- and methanol-
acciimated activated sludges, respectively.
f. The oxygen utilization rate can be formulated as Rr = -a'(ds/dt) + b'X,
where Rr is the oxygen utilization rate in mg/1 02/hr, a1 is a
coefficient designating oxygen requirement per substrate utilized, b1
is a coefficient designating oxygen requirement per biomass for
endogenous respiration, S is substrate concentration, X is biomass
concentration, and t is time. Based on substrate concentration as
TOC (mg/1) and biomass concentration as volatile suspended solids
(mg/1), the values a1 = 1.39 for phenol, a' = 2.23 for methanol, and b1 =
1.42 kd (b1 and k
-------
k. The absence of extra-cellular nitrogen and phosphorus resulted in a
greater cell yield coefficient, a, in phenol tests. However, the cells
grown in this condition were found to decay more rapidly than normal
cells.
k. Based on the treatability tests, the following conclusions are made pertaining
to the use of batch treatment systems as a countermeasure:
a. Batch treatment systems are preferred over continuous-stirred tank-
reactor (CSTR) systems for spills of phenol and methanol, especially
when the spill concentrations are high. Batch systems require much
less aeration time to achieve a certain effluent quality and produce
more acclimated sludge than CSTR systems.
b. Batch systems can be designed using numerical methods or using
batch kinetic diagrams.
c. Sludge-containing cloth bags were found useful for easy containment
of the sludge when consecutive batch treatments were required.
Floating cloth bags can be used for in situ treatment methods to
prevent sludge settling.
d. The material transport rate through sludge-clogged, cloth bags can be
expressed as:
vc J|c = CAC(S-SC),
dt
where Vc is the volume of cloth bag, Sc is the substance concentra-
tion inside the cloth bag, S is the concentration outside the bag, t is
time, C is the material exchange coefficient (L/T), and Ac is the
surface area of the bag.
e. Between methanol concentrations of 4,000 mg/1 and 6,000 mg/1
(VSS), the material exchange coefficient, C, remained the same. In
the turbulent flow regime, C showed a linear relationship with
velocity. The value of C for a phenol sludge concentration of 25,000
mg/1 (VSS) was approximately one-half of that for a methanol sludge.
f. For phenol sludge (15,000 mg/1 to 30,000 mg/1 as VSS). C = 0.0710
cm/hr in a laminar flow regime and C = 0.159 G^/^cm/hr in a
turbulent flow regime. For methanol sludge (4,000 mg/1 to 6,000
mg/1 as VSS), c = 0.141 cm/hr in a laminar flow regime and C = 0.317
G^/3 cm/hr in a turbulent flow regime. These values or equations are
proposed for the prediction of C where G is the mean temporal
velocity gradient (sec -1).
g. When sludge-containing cloth bags are used for the removal of spills,
the substrate removal rate by organisms can be expressed as:
ds EfcXS
dt" "
-------
where E is the cloth bag efficiency, which is obtained from
cloth bag efficiency diagrams.
h. When sludge-containing cloth bags are used in a batch treatment
system the system can be designed in the same manner as a regular
batch system except that Ek instead of k is used for the substrate
removal rate coefficient. The aeration time required to achieve a
given removal was observed to be slightly less than the theoretically
computed time, probably owing to organisms that escaped from the
cloth bags.
5. Based on spill control tests in a model river, the following conclusions can be
made:
a. Application of bulk sludge in streams is not an efficient method for
phenol and methanol removal because of sludge settling. Floating
cloth bags may be used to prevent this problem; however, this method
is highly restricted by the reaeration capacity of streams and the
large amount of acclimated sludge required.
b. Fixed, confining barriers may be used to prevent the dispersion of
spills. Once pollutants are contained within barriers they may be
treated in a batch manner. Cloth bags may be employed when the
mixing intensity is not sufficient for complete suspension of sludge.
Oxygen need only be supplied within or near the cloth bags.
6. Based on model lake tests, the following conclusions can be made:
a. Phenol spills contained by a barrier can be removed using unaccli-
mated sludge from a local activated sludge domestic waste treatment
plant.
b. Use of biological countermeasures will result in a significant impact
on the dissolved oxygen resources in the aquatic system. However,
this impact can be reduced by mechanical aeration.
-------
SECTION 3
RECOMMENDATIONS
It is recommended that development of biological countermeasures should
continue. Major research needs include:
1. Studies on counter measure storage and reconstitution to determine the shelf -
life of the stored material, the need for additions of mineral salts, and the amount of
material needed for spills of a given chemical.
2. Development of techniques for countermeasure application in quiescent and
flowing systems.
3. Determination of additional candidate chemicals for application of biological
countermeasures.
4. Further confirmation of the methods described in this report for calculating
amounts of the countermeasure needed for a. given volume spill.
-------
SECTION *
DEVELOPMENT OF INFORMATION FOR BIOLOGICAL
COUNTERMEASURE FEASIBILITY DETERMINATION
ELEMENTS OF SPILL CONTROL
The National Oil and Hazardous Substances Pollution Contingency Plan (40 CFR
1510) delineates five classes of actions that comprise the elements of spill control.
These actions are: Phase I - Discovery and Notification (discovery of a spill by the
discharger, patrol vessels, or incidental observation and the reporting of that discovery
to the proper agency), Phase II - Evaluation and Initiation of Action (evaluation of the
magnitude and severity of the spill, the feasibility of removing it, and the effective-
ness of removal actions), Phase III - Containment and Countermeasures (actions taken
to restrain the movement of the spilled material and to minimize its effects on water-
related resources), Phase IV - Cleanup, Mitigation, and Disposal (actions taken to
recover the spilled material and to monitor the scope and effectiveness of removal
actions), and Phase V - Documentation and Cost Recovery. The time needed to
implement any of these phases will depend on the location of the spill, the material
spilled, the magnitude of the spill, and so forth. Employment of a biological
countermeasure imposes special constraints on the activities in Phases III and IV and
requires that its use be carefully considered in Phase II. To understand these special
constraints, the requirements of a general countermeasure and the information needed
to judge the suitability of biological countermeasures will be discussed in this section.
REQUIREMENTS OF COUNTERMEASURE
Dawson et al., (1972) suggested the following criteria for evaluating potential
countermeasures:
1. Countermeasures should be highly effective.
2. Countermeasures should be applicable to a large number of substances.
3. Countermeasures should be amenable to rapid, easy deployment. (Highly
specialized equipment and/or chemicals that require extensive stockpiling prior to a
pollution incident or that cannot be rapidly conveyed to the scene of an accident are
undesirable.)
4. Countermeasures should be free from potentially harmful secondary effects
in the aquatic environment, including production of noxious sludges.
5. Countermeasures developed to combat spills of hazardous polluting sub-
stances should take advantage of available technology, particularly that developed to
combat oil spills.
-------
Several physical and chemical countermeasures were evaluated and the difficul-
ties of biological countermeasures were discussed. The authors also discussed the
dynamics of a spill and the problems of containment and mitigation given the type of
material spilled and the nature of the receiving water. The most important
parameter, they concluded, was the time lag between the spill and the initiation of
treatment because effects on organisms and, in many cases, process removal effici-
encies, are functions of the spilled material concentration. One could add to this
response time the time required for removal of the spilled material to safe levels.
Huibregtse et al., (1976) incorporated such requirements into a user's manual for
hazardous material spills but did not include biological countermeasures.
For a biological countermeasure to be considered feasible, the first four criteria
should be satisfied to the greatest extent possible and the biological countermeasure
should be competitive with, or at least complimentary to, the physical/chemical
countermeasures available. The experimental program developed in this project had as
its primary focus the test of feasibility of biological countermeasures using the above
criteria. The information needed to demonstrate such feasibility is discussed below.
INFORMATION NEEDED
The following information was considered essential to assess the feasibility of
the biological countermeasure: (1) screening tests to determine the general effective-
ness and applicability of the countermeasure to hazardous materials, (2) growth
kinetics tests to determine removal rates, growth rates, application rates, etc. so that
the requirements and logistics of countermeasure deployment could be determined,
and (3) simulated spills to demonstrate (under near field conditions) the feasibility and
effectiveness of the countermeasure. Each of these items is discussed more fully
below.
Screening Tests
Screening tests are simplified versions of growth kinetics tests and are con-
ducted as a short-term, batch test. The purpose is to show in a short time, with little
effort, whether the hazardous material being tested is biodegradable and, if so,
identify the organisms that are involved. In this study, screening tests were performed
after the literature survey and were used primarily to determine the candidate
hazardous materials for further testing in the growth kinetics and simulated spills
experiments.
Growth Kinetics Tests
The attractiveness of biological countermeasures for hazardous material spills is
twofold: (a) bacteria are natural components of ecological systems and their use as a
countermeasure will not constitute the introduction of a "foreign" material and (b)
bacteria will metabolize organic hazardous materials to the principal end products
carbon dioxide and water, according to the general equation:
CXHYOZ + 02 enzyme)r.n2 + H2O (1)
The microbial utilization of a hazardous material in a finite volume, mixed
reactor is described by the following equation (Pearson, 1968):
10
-------
= QS0- QS-, - qXaV, (2)
where:
V = change in hazardous material mass in system
QS0 = influent hazardous material mass
= effluent hazardous material mass
qXaV = hazardous material mass removed by cells
and:
SQ = influent hazardous material concentration (mg/1),
5} = effluent hazardous material concentration (mg/1)
Q = flow into and out of the reactor (1/d),
*a = average microorganism concentration in reactor (mg/1),
V = reactor volume (1),
t = time, and
q = hazardous material removal rate =
mg hazardous material removed/day
mg microbes
For a system with no flow (e.g., a batch reactor), Equation 2 reduces to:
which has the solution:
•qXa*> (4)
where:
q = S0-S
X t . (5)
a.
Eckenfelder (1970) has described this same process by the equation:
-rr =-kXaS, (6)
11
-------
where:
S = hazardous material concentration (mg/1),
Xa = average microorganism cell concentration in reactor (mg/1), and
k = removal rate (mg S remaining/day/mg 5/mg Xa).
In Equation 6, the product kS is equivalent to the term q in Equation 3 and in fact,
Eckenf elder (1970) found that the solution to Equation 4 may be expressed as:
*at (7)
These equations represent the overall biodegradation process, however, there are
usually a number of biochemical reactions that take place in the microorganism as the
hazardous material is reduced to elemental forms. This series of reactions may be
referred to as the breakdown pathway.
The growth of bacteria in a reactor may be expressed by the following equation
(Pearson, 1968):
v-kxv, (8)
change in cell mass = in - out + growth - decay in system
where:
XQ = influent microorganism cell mass concentration (mg/1),
X, = effluent cell mass concentration (mg/1),
Q = flow into and out of system (1/d),
X = average cell mass concentration in system (mg/1),
cl
ju = microorganism growth rate (mg cells produced/day/ mg cells),
k , = microorganism death rate (mg cells removed by death/day/mg cells),
t = time, and
V = volume.
For a system with no flow, Equation 8 reduces to:
dXa
ar 4<-kd> xa- w
It is important to note the nature of the relationship between the hazardous material
concentration and bacterial growth rate. This relationship has been shown (Pearson,
L2
-------
1968) to be very similar to the Michaeiis-Menton kinetic model for enzymatic action
and may be expressed as:
' (10)
where:
It = maximum growth rate (mg cells produced/day /mg cells),
S = hazardous material (haz. mat.) concentration (mg/1), and
K = hazardous material concentration at one-half the maximum growth
s rate (mg/1).
The techniques for deriving the maximum growth rate and the Michaelis Menton
constant, Ks, have been given by Pearson (1968).
The substrate removal rate, q, may be transformed to the growth rate,jU, by
multiplying by the yield coefficient, Y, as follows:
mg hazardous material remaining mg cells produced
mg cells - day mg hazardous material remaining
mg cells produced
mg cells - day
thus:
A
^cT • . (12)
Under steady state conditions, it may be shown from Equation 2 that q may be
determined in a continuously stirred reactor by:
QCSp-Sp
q=—UY ' (13)
In a batch reactor, q may be calculated from Equation 5.
It is well known in microbiological research that the introduction of a small
inoculum of bacteria into a medium with useable substrate initially results in growth
of the bacteria at a maximum rate,JLl, with concurrent reduction of the substrate
concentration (see Figure 1) according to Equation 3. This period is termed the
maximum growth phase. After a short time, the substrate concentration is reduced to
a level that becomes limiting to the growth of bacteria and the bacterial concentra-
tion quickly reaches a maximum. This period is termed the declining growth phase.
Following the peak, a decline in concentration occurs due to auto-oxidation and death;
this is the death phase, or the often-called endogenous respiration phase.
Initially a delay in growth, called the lag phase, may occur. The extent of the
lag phase is a function of the physiological condition of the bacteria, the size of the
13
-------
inoculum, the state of acclimation of the bacteria to the substrate, and
perhaps other effects. Once growth begins, the maximum growth rate is
characteristic of the bacteria, the other nutrients required by the
bacteria for growth, the temperature of the medium, and the toxicity of the
substrate. If the lag phase is very long, the rate at which the substrate
is consumed also lags. Since one of the most important requirements of a
countermeasure (Dawson et al., 1972) is that it be capable of immediate use
and application, the lag time must be minimized and the bacterial and
substrate characteristics that influence the lag time must be defined.
Equations 3 and 9 may be used to describe the mitigation of a hazardous
material spill and the increase in bacterial concentration, respectively,
in a batch system or in a spill situation in which the spill occurred
instantaneously (or over a very short time) and onto which bacteria were
deployed. The reduction of the hazardous material and the growth of the
bacteria would approximately follow the curve shown in Figure 1.
Equations 2 and 8, would be applicable to a continuous-flow biological
treatment system that is operated such that high removal of the hazardous
material and high bacterial retention in the system are achieved.
Operational parameters for these treatment systems have been developed in
practice for wastes containing hazardous substances (Eckenfelder, 1970).
Equations may also be developed to describe the transport of hazardous
materials in flowing systems, including biodegradation as well as other
sink terms.
In order to apply these equations, their terms must first be obtained
by experimental means. The constant for the growth rate-substrate
relationship are especially important, as is the substrate removal rate
determination.
Simulated Spills
Biological countermeasures may be employed in one of two ways: they
may be applied in situ to a spill of a hazardous material, that is, in the
receiving water itself, or they may be employed by pumping the spilled
hazardous material to a portable biological treatment system brought to the
site. For both treatment techniques, the nature and amount of the
hazardous material spilled should be determined so that appropriate
bacteria may be used and the proper amount of baterial culture applied.
Containment of the spilled hazardous material is also desirable in order to
avoid diluting effects of the natural system and to provide a controlled
environment for the bacteria. Once the nature and amount of the hazardous
material spilled have been determined and once containment has been
achieved or at least ambient concentrations determined, the bacterial
countermeasures may then be deployed. The amounts of countermeasure to be
used, the time required for action, and the secondary effects must be
determined experimentally.
The principal constraints of the biological countermeasure approach
emerge in the production, storage, and deployment system and in the
introduction of undesireable bacteria to a natural system. Production of a
14
-------
CO
c
o
tQ
S_
-M
C
-------
single bacterial species in large quantities (for use on a specific
hazardous material) would impose a serious constraint on the feasibility of
biological countermeasures. Thus, it is desirable to find bacteria that
break down a wide variety of organic substrates and that are produced
easily in large volume, such as the Pseudomonads. Pseudomonas fluorescens
for example, may be grown on sugars, amino acids, organic acids, alcohols,
aromatic compounds, and other cyclic organic compounds (Stanier 1950).
Several species of the genus Pseudomonas other than P. fluorescens have
been shown to use a variety of aromatic compounds (Gibson 1972, Chapman
1972).
Large quantities of bacterial material may potentially be stored in
several forms: (1) as a liquid culture in which substrate is supplied
continually and bacteria are produced continually, (2) in a frozen form in
which large quantities of bacterial culture must be continually
refrigerated, or (3) in a lyophilized, powdered form in which the bacteria
are stored in large quantities in airtight containers. Storage in the
liquid form is desirable in the sense that the bacteria are ready for
immediate application, but the expense of maintaining these cultures at a
number of points around the country near potential spill areas may be
prohibitive. Storage in the frozen form is more desirable because less
volume is needed, but continual refrigeration is required and a lag time is
needed to reconstitute the bacteria for application. The lyophilized form
may be the most desirable in terms of storage because the bacteria are in a
powdered form, may be stored at room temperature, and may be maintained at
many points around the country or even shipped with the hazardous material
for which bacteria have been cultured. One possible disadvantage of this
storage form is the time required (a few hours) for reconstitution of the
bacteria to an active state. Because the storage method must be determined
for each culture, investigation of this phase of deployment is required.
The methods for in situ deployment of the bacteria would depend on the
physical/chemical state of the hazardous material in the receiving water
and the storage mode of the bacterial culture. Spraying from a helicopter,
boat, or from shore should be adequate. Physical deployent of the portable
treatment plant and start-up of the bacterial culture appear to be the
critical steps in the use of a biological waste treatment countermeasure.
Use of biological countermeasures may result in the addition of
undesirable bacteria to a aquatic system. Application of pathogenic
bacteria, for example, to a spill of a hazardous material could result in
the proliferation of these bacteria as long as the hazardous material
remains. Use of activated sludge from treatment plants for in-situ
treatment or for the portable treatment plant may also result in the
application of undesirable bacteria. However,this potential hazard may not
be serious because the bacterial groups that break down hazardous materials
will not likely be pathogenic and disinfection of the effluent from the
portable treatment system should remove any undesirable bacteria from the
waste stream before discharge to the aquatic system.
An undesirable effect of biological countermeasures in the consumption
16
-------
of dissolved oxygen by the bacteria during the breakdown of the hazardous
material. Low levels of dissolved oxygen are typically found downstream of
domestic and industrial waste discharges. This problem may be avoided by
artificially aerating the receiving water after applying the bacteria or by
using small amounts of bacteria such that excessive oxygen consumption does
not occur.
Cautions about biological countermeasures are expressed in Annex X of
the National Oil and Hazardous Substances Pollution Contingency Plan. The
Plan states that biological countermeasures:
"may be used only when such use is the most desirable technique for
removing oils or hazardous substances and only after obtaining approval
from the appropriate state and local public health and water pollution
control officials. Biological agents may be used only when a listing
of organisms or other ingredients contained in the agent is provided to
EPA in sufficient time for review before its use."
17
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SECTION 5
SELECTION OF TEST MATERIALS
At the start of this project, 20 hazardous materials (individual compounds or
groups of compounds) were selected for testing. By the end of the project, two
compounds—phenol and methanol—had received most of the attention of the growth
kinetics and spill simulation tests. The process by which the initial and subsequent
selection of test compounds was accomplished is described in this section.
INITIAL SELECTION
Contract List
In conjunction with the Project Officers for the Environmental Protection
Agency, 14 compounds were selected from a list of 20 hazardous materials (Dawson et
al., 1970). Of the 14 selected (Table 1), four were given a low experimental priority.
From the physical/chemical characteristics shown in Table 2, it is apparent that
several compounds are quite soluble in water and would be dispersed by natural mixing
processes upon spillage. Several, like xylene and benzene, are relatively insoluble and
would tend to float. Others, like the pesticides, are insoluble but sorb rapidly to
suspended particles and thus would remain in suspension. Thus, from an experimental
viewpoint a variety of test hazardous materials was used with respect to physical/
chemical nature after spillage. Likewise, the chemical structure of the compounds
was variable enough to provide an adequate test of the biological countermeasure.
Selection Based on Literature Review
The literature was reviewed to gather information on the physical/chemical
characteristics of these compounds, previous biodegradation and biological treatment
investigations, and the toxicity of the compounds to organisms in fresh and marine
waters (see Section 6). The literature survey revealed that most of the contract
compounds have been shown to be biodegradable and that the breakdown pathway was
completely or partially known (Table 3). Of the 20 top-ranked hazardous materials,
most were amenable to mitigation by biological counter measures. Furthermore, the
nature of the materials' toxicity to organisms in natural systems had also been defined
to some extent, thus the concentrations that must be achieved to mitigate the effects
of hazardous materials were also known.
No hazardous materials were deleted following the literature review, but it was
evident that certain of the compounds would require special apparatus or special
handling according to the University of Texas at Austin Safety Office. Thus, those
compounds that could be tested most easily were elevated in priority.
18
-------
TABLE 1. CONTRACT LIST OF HAZARDOUS MATERIALS
Hazardous materials
Pheno 1
Methyl alcohol
Cyclic rodenticides^
Acrylonitrile
Benzene
Misc. cyclic insecticides
Styrene
Acetone cyanohydrin
Nonyl phenol
DDT
Isoprene
Xylenes
Nitrophenol
Aldrin-Toxaphene group
Total Number
Ranked Contract priority
priority! High Low
1 X
2 x
3 x
4 x
6 x
8 x
10 x
11 x
13 x
14 x
15 x
16 x
17 x
18 x
14 10 4
1Dawson, et. al. (1970).
^includes Dicetel, Endosulfon, Methoxychlor, Parathion, Methyl
Parathion, Chlordane, Dieldrin, Endrin, Heptachlor, Terpene,
Polychlorinates, Carbopheno Thion, Coumaphos Diazonon, Dioxathion
Ronnel, Chlorobenzilate, ODD, and others.
19
-------
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-------
TABLE 3. RESULTS OF LITERATURE REVIEW AND SCREENING TESTS
Biodegra-
dation Pathway
Hazardous material shown? known?
Acetone cyanohydrin
Acrylonitrile
Aldrin
Benzene
Cyclic rodent & insect.
DDT
Isoprene
Methyl alcohol
Nitrophenol
Non ly pheno 1
Phenol
Styrene
Xylene
Toxaphene
Yes1
Yes
Yes
Yes
Yes2
Yes
No3
Yes
Yes
No
Yes
No3
Yes
No
Partially
Partially
i
Partially
Yes
Partially2
Partially
No3
Yes
Partially
No
Yes
No3
Partially
No
Biodegradation
Toxicity screening
defined? tests
Yes, as
CN
Yes
Yes x
Yes x
Yes
Yes
Partially x
Yes x
Yes x
Partially x
Yes x
Partially x
Yes x
Yes x
1 Following chemical dissociation.
2 For a few compounds.
3 No direct evidence, but should be biodegradable; probable pathway known.
21
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Initial Screening Tests
The screening tests performed in this study were designed to show that
the compounds could be biodegraded with organisms from various sources. As
shown in Table 3, screening tests were performed on most of the contract
compounds and either biodegradation or volatilization was demonstrated.
Methyl alcohol, nitrophenol, and phenol were all biodegradable by the
organisms used, while isoprene, benzene, styrene and xylene volatilized
rapidly from the test containers. Volatilization tests for benzene were
conducted under various mixing conditions ranging from aerated and highly
mixed to unstirred. For the other compounds, no biodecomposition was shown
or the results were uncertain.
FINAL SELECTION
Following the screening tests and the initiation of growth kinetics
tests, it became apparent that only a few compounds could be tested as
extensively as desired in the spill simulation tests. Thus, phenol and
methanol were selected because they had been shown to be decomposable,
their breakdown pathways differed (methanol was a linear chain
decomposition, while phenol was ringed), both were soluble in water, and
both could be handled safely by laboratory personnel with reasonable
precautions. It was concluded that if the methodology of the biological
countermeasure could be demonstrated with these compounds, the feasibility
of the countermeasure would be demonstrated and the test procedures could
then be applied to other, more difficult-to-handle compounds.
22
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SECTION 6
LITERATURE REVIEW
Engineering and scientific literature was surveyed for information on the
physical and chemical properties, biodegradability, and toxicity of the contract list of
hazardous materials. Though some of this information, such as physical and chemical
properties and toxicity, was readily available in handbooks or in previous reports on
hazardous materials spills, information on microbial decomposition of these compounds
was variable in quantity and was scattered throughout the literature. For the purposes
of this study, a compilation of this information was considered necessary to determine
which of the contract compounds would be most amenable to microbial decomposition
and hence experimentation.
Following are the results of the literature survey for each contract compound
and for the specific topics: general physical/chemical properties, microbial decompo-
sition, and toxicity. Other information has been compiled through the literature
survey on other topics and is presented elsewhere in the report.
ACETONE CYANOHYDRIN
Description
Acetone cyanohydrin (CH2)2C(OH)CN (hydroxyisobutyronitrile, 2-hydroxy-3-
methylpropanenitrile, isopropylcyanohydrin) is formed by the reaction of acetone plus
hydrogen cyanide and may contain 0.2 percent free hydrogen. It is very soluble in
water, alcohol, and ether. Acetone cyanohydrin decomposes rapidly in alkali, releasing
HCN.
Physical/Chemical Properties
Physical state: colorless liquid
Molecular weight: 85.10
Melting point: -19°C
Boiling point: 82°C at 23 mm Hg
Refractive index: 1.3996
Density: 0.932 (19°C)
Vapor density: 2.95 (air = 1)
Vapor pressure: 0.88 mm Hg at 20°C
Microbial Decomposition
Little direct information could be found regarding the microbial decomposition
of acetone cyanohydrin. However, investigations of the toxic cyanide component have
been conducted. Lutin (1970) found that unacclimated activated sludge microorgan-
23
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isms were totally inhibited by 500 ppm cyanide but the effects on bacterial respiration
at lower levels of this toxicant were not analyzed.
Moore and Kin (1969) reported that decomposition probably occurred following
the spill of acetone cyanohydrin from a train wreck at Dunreith, Indiana. Five tank
cars carrying chemicals, two with acetone cyanohydrin, one with ethylene oxide, one
with vinyl chloride, and one with methyl methacrylate were involved in the wreck.
On-site inspection determined that the ethylene oxide had vaporized and also burned.
One tank load of acetone cyanohydrin was primarily intact except for a small loss of
the chemical due to volatilization and subsequent burning at the pressure relief valve.
The suspected source of stream pollution from the wreck was the 4,560 liters (1,200
gal) of acetone cyanohydrin lost from the other tank car. Estimates of the toxic
potential from the 4,560 liters (1,200 gal) of spilled material showed that it contained
1,260 kg (2,800 Ib) of cyanide and most of this toxic material found its way to the
river. Measured concentrations as high as 405 mg/1 cyanide were found in Blue Creek
immediately downstream from the site of the wreck. It was determined that aquatic
cyanide concentrations were being reduced by decomposition and dilution, but for
further protection of fish and other aquatic life, additional chemical oxidation by
calcium hypochlorite was employed.
Transformation of cyanide by biochemical means may be estimated as conver-
sion of HCN into formic acid following the reactions suggested by McKinney (1962).
The chemical oxidation of cyanide is given by Sawyer and McCarty (1967). The other
product of hydrolysis of acetone cyanohydrin, presumably a polyhydroxy alcohol, is
attacked fairly readily. Sawyer and McCarty (1967) state that acetone cyanohydrin
can be completely degraded by microbial action to form carbon dioxide and water with
aldehydes and ketones produced as intermediates.
Toxicity
Acetone cyanohydrin is very soluble in water, but decomposes to acetone and
hydrogen cyanide under alkaline conditions. Since the liquid is stable under acid
conditions, it is primarily transported in the acid state. Hydrolysis of this compound
into hydrogen cyanide represents the predominant toxic mechanism, with the acetone
by-product exhibiting a toxicity several orders of magnitude lower than cyanide.
Because of this probable environmental behavior, the toxicity of acetone cyanohydrin
components rather than of the whole compound will be discussed.
The toxicity of acetone is indicated in the work of Patrick et al., (1968) who
found that approximately 11,500 ppm acetone was the 120-hour median tolerance level
(TLm) for diatoms. The growth rate of Nitzschia linear is, a common diatom in
unpolluted waters, was reduced fifty percent at this high acetone concentration.
Anderson (1944) reported immobilization of Daphnia magna at 9,280 ppm acetone over
16 hrs, with marcosis being reversible. However, Dowden and Bennett (1965) found
that 10 ppm acetone was the median tolerance limit for Daphnia exposed to this
solvent for 24 to 48 hours, this value appears to be a gross underestimate.
Since the more actuely toxic component of acetone cyanohydrin is cyanide, it is
instructive to examine data describing-SH toxicity. Adverse effects of cyanide on
aquatic invertebrates occur at much lower cyanide concentrations than the 500 ppm
concentration found to inhibit bacterial respiration (Lutin 1970). Experiments by
Reich (1955) showed that amoeba respiration decreases with an increase in cyanide
24
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concentration. A 0.005 molar cyanide solution was lethal to these protozoans. In
addition, cyanide stimulated the utilization of glucose by a soil amoeba, but inhibited a
peptone-utilization enzymatic pathway. For the aquatic snail Physa heterostropha, a
0.432 ppm cyanide concentration was acutely toxic to half of the organisms exposed
over a 96-hour period (Patrick et ah, 1968).
Wallen et al. (1957) separately tested both components of acetone cyanohydrin
hydrolysis (acetone and CN~) on Gambusia affinis, the mosquitofish. The TLm value
for a 96-hour exposure to mosquitofish was 1.6 ppm potassium cyanide. This
concentration decreased to 0.28 ppm for bluegill sunfish (Lepomis machrochirus)
exposed for shorter periods (up to 48 hours). The equivalent toxicity for acetone over
a 96-hour test exceeded 13,000 ppm for mosquitofish and 14,000 ppm for orange-
spotted sunfish (Lepomis humilis). However, the vigorous aeration used to maintain
high turbidity as a secondary experimental parameter may have enhanced solvent
evaporation from the system and inflated the actual toxic level. Since Cairns and
Scheier (1968) showed that 50 percent survival occurred at 8,300 ppm acetone for
bluegills and because the system concentration decreased by one-half over the 96-hour
test period, loss of the volatile solvent may indeed have been a significant factor.
Freshwater fish succumbed to 0.05 to 0.10 ppm cyanide following an Indiana
railroad derailment and subsequent spillage of acetone cyanohydrin into local water-
ways (Moore and Kin, 1969). Cairns and Scheier (1968) reported that 0.07 ppm cyanide
(from KCN) allowed 100% survival of bluegills, but 0.18 ppm killed one-half the test
fish, and 0.2* ppm killed all sunfish. Previous experimentation on bluegills had
established that variations in fish size did not result in significant toxicity differences
with cyanide (Cairns and Scheier 1968). However, Turnbull et al., (1954) noticed that 1
ppm cyanide killed bluegills within an hour after violent spasms and loss of equili-
brium, but a concentration of 0.21 ppm did not evoke behavioral response or death.
Guppies (Lebistes reticulatus) required 0.26 ppm (43-hour exposure) to 0.42 ppm
cyanide (20 hours) to kill 50 percent of the fish (Chen and Selleck 1968).
Renn (1955) compared the cyanogenic properties of lactonitrile, an organic
nitrile that dissociates more readily than acrylonitrile, and potassium cyanide. Both
cyanide-producing compounds yielded similar median tolerance values for centrarchid
fishes. Yellow-breasted sunfish (Lepomis autritus) and largemouth bass (Micropterus
salmoides) partially succumbed to 0.06 ppm cyanide while the median tolerance levels
for bluegills was 0.01 to 0.06 ppm and for crappies (Pomoxis annularis) was 0.05 to 0.07
ppm cyanide. These TLm values decreased for crappies and increased for bluegills
under continuous flow bioassay techniques. Burdick et al., (1958) also used a
continuous flow apparatus for testing cyanide effects on coldwater fishes, such as
brown trout (Salmo trutta) and smallmouth bass (Micropterus dolomieui). However,
they used minimum cyanide concentrations of 0.32 ppm for trout and 0.175 ppm for
bass and lethal effects on all fishes were evident within minutes, rather than hours.
ACRYLONITRILE
Description
Acrylonitrile is the compound H2C = CHCN (vinyl cyanide, 2-propenenitrile,
cyanoethylene). Acrylonitrile is miscible with most organic solvents and exhibits a
water solubility of 5.35 percent at 20°C.
25
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Physical Properties
Physical state: colorless liquid with fairly pungent odor
Molecular weight: 53.06
Boiling point: 77.3°C
Freezing point: -83.55°C
Specific gravity: 0.8060 at 20°C
Vapor density: 1.83 (air = 1.0)
Viscosity, liquid: 0.34 cp at 24°
Refractive index: 1.3888 (9 25°C
Microbial Decomposition
Only five percent of acrylonitrile by volume is soluble in water, but, because
acrylonitrile may undergo hydrolysis upon water contact, toxic action can result from
both the compound itself and from the liberated cyanide. Bacterial oxidation of this
organic nitrile occurs through the enzymatic hydrolysis of the nitrile to a carboxyl
group, producing ammonia and organic acids (Mills and Stack 1955, Buzzell et al.,
1968). Ludzack et al., (1959) found that microorganisms capable of nitrile oxidation
are common in surface waters. Cocci and rod-shaped bacteria, usually gram-negative,
were commonly found; Pseudomonas aerogenes was especially abundant. Buzzell et
al., (1968) concluded that 400 ppm acrylonitrile constitutes the mean tolerance level
{50 percent inhibition of growth) for mixed bacterial cultures (sewage seed). Cherry et
al., (1956) found zoogleal and algal growth in natural river water with 10 and 25 ppm
acrylonitrile added. However, at 50 ppm fungi were the major component of the biota
in microcosm simulations.
Lank and Wallace (1970) have tested acrylonitrile under anaerobic conditions
and claim that a digester can receive 20 mg/1 without adversely affecting its
performance.
One important factor affecting degradability is the known volatility of acrylon-
itrile. Buzzell et al., (1969) subjected a sample of acrylonitrile to a stripping test and
found that 50 percent of the carbon was removed after 2 hours and 88 percent was
removed after 6 hours. Accordingly, they reasoned that physical stripping was
responsible for the carbon loss when experimental samples were subjected to shade-
table tests.
Toxicity
Mammals—
Acrylonitrile is toxic to humans and other mammals by inhalation of vapors,
ingestion, or skin contact. However, specific expsosure levels for humans have not
been documented. Current occupational safety levels are based on analogy to
complete hydrolysis of acrylonitrile to HCN, but cyanide formation as the predominant
mode of toxicity is still a debatable point. Brieger et al., (1952) found small quantities
of HCN in the blood of rats expossed to 100 ppm acrylonitrile vapors, but detected
none at lower atmospheric concentrations. However, other investigators (Barnes and
Cerna, 1959; Paulet and Desnos 1961, and Paulet et al., 1966) conclude that the whole
molecule, not liberated HCN, is responsible for the toxic action.
26
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Excretion of slight amounts of acrylonitrile in the urine has been noted
(Czaikowska, 1971), but higher thiocyanate levels in blood serum and urine suggested
this route for elimination of partially-metabolized acrylonitrile (Lawton et al., 1943;
Paulet and Desnos, 1961).
Acrylonitrile forms conjugates with protein and non-protein sulfhydryl groups,
decreasing tissue levels of essential amino acids, such as L-cysteine and L-glutathione
(Hashimoto and Kanai, 1965 and 1972). Inhibition of cytochrome oxidase in liver,
kidney, and brain tissues results from acrylonitrile poisoning (Tarkowski, 1968) and this
blockage of cellular metabolism may explain the accumulation of pyruvate and lactate
in acutely intoxicated animals (Hashimoto and Kani, 1972). Acrylonitrile may damage
the synthesizing function of the liver, thereby upsetting the balance of available amino
acids for growth and maintenance (Movsumzade, 1970).
Aquatic Organisms-
Several investigations of acrylonitrile toxicity to fishes have been reported.
For the marine pinfish (Lagodon rhomboides), the 24 hour-TLm was found to be 24.5
ppm (Garrett, 1957). Renn (1955) had previously determined that pinfish could
tolerate 10 to 18 ppm acrylonitrile (as acrylonitrile-N) without adverse effects.
Buzzell et al., (1968) found that 10 ppm represented the maximum concentration for
100 percent survival of bluegills. Bandt (1953) reported 100 ppm acrylonitrile as the
TLm value for fish, a significantly higher figure than previously determined.
Henderson et al., (1961) found that median tolerance values decreased consid-
erably with length of exposure for fathead minnows (Pimephales promelas), bluegills
(Lepomis macrochirus), and guppies (Lebistes reticulatus). Acrylohitrile levels of 34.3
ppm (fatheads), 25.5 ppm (bluegills), and 44.6 ppm (guppies) were acutely toxic to one-
half the test fishes within 24 hours and thus demonstrated interspecific differences in
toxic susceptibility. The same authors also noted a high cumulative (chronic) toxicity
for fathead minnows in a continuous flow system. The TLm threshold decreased from
33.5 ppm after a one-day exposure to 2.6 ppm for fish exposed to acrylonitrile for 30
days. Tainting of fish flesh by acrylonitrile did not occur at sublethal concentrations.
ALDRIN
Description
Aldrin is a broad-spectrum cyclodiene insecticide of the group that includes
chlordane, endrin, dieldrin, and heptachlor. This compound (1,2,3,4,10,10-hexachloro-
1,4,4,1,5,8,8a-hexahydro-l, 4-endo, exo-5, 8-dimethanonapthalene) possesses a melting
point of 104°C and is most soluble in aromatic hydrocarbons and carbon tetrachloride.
Aldrin has unsubstituted double bonds that readily add oxygen to form epoxy
derivatives. Upon epoxidation in sunlight, dieldrin is formed. Epoxides are also
formed in tissues (both plant and animals) and are preferentially concentrated and
stored in fats.
Microbial Decomposition
Aldrin has been shown to be degradable by several mechanisms: microbial
decomposition, photochemical oxidation, and volatilization. Kearney and Kaufman
(1972) reported that aldrin was oxidized to dieldrin by a number of soil microorganisms
27
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including species of Trichderna, Fusarian, and Penicillum. Similar microbial conver-
sions have been reported by Gakstatter (1968) and Krieger and Lee (1973).
Adsorption of aldrin by floe-forming bacteria isolated from Lake Erie has been
reported by Leshniowsky et al., (1970). A gram-positive Bacillus species and gram-
negative Flavobacterium concentrated aldrin from colloidal suspension and removed it
from the water phase upon settling. Apparently, loss of aldrin from the test solution
(1 ppm aldrin) resulted solely from adsorption of the pesticide to the organic floe and
not from any microbial metabolism. This transport mechanism appears to be highly
significant since most chlorinated pesticide biodegration occurs under anaerobic
conditions, which are likely to occur in the sediments. Hill and McCarty (1967)
reported on the anaerobic degradation of aldrin (where, by definition, degradation
referred to any measureable change in the pesticide concentration). They reported
that the anaerobic degradation of aldrin by both thick and diluted, digested waste-
water sludge suggested first-order kinetics.
Crosby (1972) stated that epoxidation in air can result in the formation of
dieldrin when aldrin is exposed to the ultraviolet component of sunlight. Similar
photoconversions have been demonstrated by Rose and Sutherland (1967) and
Georgacakis and Khan (1971).
Lichtenstein (1972) stated that volatilization was directly responsible for the
loss of pesticide residues from a given substrate. In particular, aldrin was the most
volatile of eight different insecticides tested.
Toxicity to Aquatic Organisms
Because the decomposition products of aldrin are highly toxic, the overall
problem of aldrin toxicity to aquatic life also includes the toxic metabolites and
photodecomposition products. Thus, aldrin and its conversion products have been
tested for their toxicity effects on numerous fishes and food-chain invertebrates.
Relatively few studies of aldrin toxicity to microorganisms and algae have been
conducted in comparison to the number of studies on invertebrates and fishes.
Poorman (1973) showed that aldrin levels of 50 to 100 ppm reduced cell numbers of the
photosynthetic Euglena gracilis by 12 to 17 percent, but this species encysted and
recovered when high pesticide levels were significantly diluted .
Adverse effects on invertebrates occur at much lower aldrin concentrations
than noted for planktonic species. Anderson (1960) noted that 29.2 ppb immobilized
Daphnia magna, while Sanders and Cope (1966) verified this by demonstrating that
Daphnia swimming ability was indeed inhibited at 30 ppb. Only 1 ppm aldrin induced
total mortality in lumnacid snails (Batte et al., 1951). Freshwater amphipods
(Gammarus lacustris) exhibited a 48-hour TLm value of 38.5 ppm aldrin (Gaufin et al.,
1965; Nebeker and Gaufin, 1964), although McDonald (1962) had found that 50 percent
mortality for this species exposed to aldrin for 3 to 4 hours occurred at only 0.5 ppm.
Paleomonetes kadiakensis exhibited differential toxicity resistance to aldrin, depend-
ing on whether these small freshwater shrimp were collected from agricultural areas
(185 ppm) or from wildlife refuge areas free from extensive pesticide usage (85 ppm).
Aquatic insect naiads (Acroneuria pacifica, Pteronarcys californica) appeared to be
susceptible to less than 0.2 ppm aldrin (Jensen and Gaufin, 1964; Moye and Luckmann,
1964; Gaufin et al., 1965; Jessen and Gaufin, 1966). Because of its economic
28
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importance as a food organism in Louisiana, crayfish (Procambarus clarkii) are often
grown in flooded rice fields subject to frequent pesticide applications. Since crayfish
feed predominantly on organic detritus, pesticides adsorbed to this organic matter may
be ingested and reach toxic levels in the organism. Aldrin concentrations of 0.038 ppm
were shown to be toxic to juvenile crayfish, while 0.6 ppm caused mortality in the
adult crustaceans (Hendrick and Everett, 1965). Other commercially important
organisms, especially molluscan filter-feeders (oysters, clams, mussels) have been
shown to concentrate aldrin hundreds of times over the ambient water level (Butler,
1967; Casper, 1967; Bedford et al., 1968).
The list of fishes that have been bioassayed for aldrin toxicity is quite large.
Aldrin levels of 5.2 to 60 ppb have been shown to be toxic to bluegill sunfish in 96-hour
toxicity bioassays (Tarzwell, 1959; Henderson et al., 1959; Weiss, 1964; Ferguson et
al., 1964; Profitt, 1966). However, Cope (1963T~claimed a 10 ppm TLm for bluegills
exposed only 24 hours. A decrease in median tolerance limit from 9.7 ppm to 5.6 ppm
accompanied a temperature increase of 45° to 85°F for bluegills (Cope, 1965). Other
centrarchids, such as the green sunfish (Lepomis cyanellus) and the largemouth bass
(Micropterus salmoides), exhibited identical TLm values of 0.4 ppm (Profitt, 1966).
Tarzwell (1959) and Henderson et al. (1959) showed that aldrin toxicity, as
indicated by 96-hour TLm values, was 32 to 33 ppb for fathead minnows, 28 to 30 ppb
for goldfish, and 30 to 33 ppb for guppies. Gakstatter (1968) found that radio-labeled
aldrin (14c~aldrin) was toxic to goldfish at 0.05 ppm and that conversion of this
pesticide to dieldrin proceeded rapidly in the body tissues, but more slowly in the
visceral fat. The mosquitofish, Gambusia affinis, was adversely affected by aldrin
levels of 0.05 to 2.1 ppm (Boyd and Ferguson, 1964), 0,5 ppm (Mulla et al., 1963), and
0.02 to 0.06 ppm (Ferguson, Cully et al., 1965).
Profitt (1966) demonstrated interspecific variation in susceptibility to aldrin for
Notropis minnows, with TLm values of 0.02 to 0.4 ppm for N. umbratilis, 0.02 to 0.08
ppm for P^. cornutus, and 0.6 ppm for N. blennius. The 50 percent survival level was
0.08 ppm for Notemigoneus chrysoleucas, (common golden shiner) and 0.013 to 0.185
ppm for Ictaluras melas (bullhead catfish) (Ferguson, Cully et al., 1965). Katz (1961)
reported that chinook salmon (Oncorhychus tshawytsha) found 7.5 ppb toxic, while
toxicity effects occurred at 17.7 ppb for rainbow trout ( Salmo gairdnerii) and 45.0 ppb
for Coho Salmon ( 0. kisutch). Cope (1965) determined a slightly higher TLm of 31 ppb
to rainbow trout.
BENZENE
Description
Benzene is a cyclic hydrocarbon (CgHg) that is colorless, volatile, and possesses
a distinct aromatic odor. The chemical structure consists of alternating unsaturated
bonds that create a stable resonating ring. Because it undergoes numerous substitution
reactions without breaking the ring structure, benzene serves as the parent compound
for many other industrially important aromatic hydrocarbons.
Benezene is miscible with most organic solvents, particularly alcohol and ether.
Water solubility is fairly low—0.057 percent by weight at 20°C. However, benzene
forms a two-phase azeotropic mixture with water which boils at 69.25°C compared to
80.103°C for the pure compound. Using gas chromatographic techniques, McAuliffe
29
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(1963) found that benzene solubility in water was 1,780 ppm at 25°C, highest among all
C^-Cg hydrocarbons.
Physical/Chemical Properties
Molecular weight: 78.11
Freezing point: 5.506°C
Boiling point: 80.103°C
Density: 0.87903 at 20°C
Microbial Decomposition
The microbial degradation pathway of benzene has been the subject of
extensive investigation in recent years. Gibson (1972) has proposed the following
reaction sequence for the initial formation of catechol from benzene by the bacteria
Pseudomonas putida.
^ ^^ *~ft ^-^ Vr
benzene hypothetical cis-l-2-dihydroxy-
dioxetane 1,2-dihydro-benzene
According to Dagley (1972), the unsubstituted benzene nucleus is an inert
resource structure suitably substituted by two hydroxyl groups. The catechol produced
during microbial metabolism is available for enzymic ring fission because of the two
hydroxyl groups in the ortho position.
Catechol and its carboxylated derivative, protocatechuic acid, can be catabo-
lized by either meta or ortho cleavage (Chapman, 1972). Ortho cleavage results in
fission of the bond between the two carbon atoms bearing the hydroxyl groups. In
meta cleavage, fission, occurs between a carbon attached to a hydroxyl group and a
carbon atom bonded to hydrogen. In ortho cleavage, the benzene derivative is
converted into dicarboxylic acid, whereas meta cleavage results in either an aldehydo-
acid or a keto-acid.
Gibson (1968) listed the following sequence as typifying the ortho cleavage of
catechol by Pseudomonas putida.
^•^ ^-^ <"). s~^
GOGH
catechol cis, cis- + muco- /3-ketodipate B-ketoadipate
muconic acid lactone enol lactone
30
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The B-ketoadipate thus formed is readily available for further decomposition
into succinate and acetyl-CoA, both of which can enter into the tricarboxylic acid
cycle with production of CO2 and H2O.
Ater adding benzene to a Warburg respirometer, Marr and Stone (1961) detected
catechol chromatographically in culture filtrates of Pseudomonas aeruginosa and
Mycobacterium rhodochrous adsorbed onto silica gel. The amount of benzene added
could be determined by weighing the silica gel before and after adsorption. Gibson
(1972), as stated earlier, found that Pseudomonas putida was capable of attacking
benzene. Pseudomonas putida would not grow when benzene was added directly to the
growth medium. However, good growth did occur when benzene was added in the form
of a vapor.
Malaney and McKinney (1966), in their studies of the oxidative abilities of
benzene-acclimated activated sludge, found that in 20 days it was possible to
acclimate normal activated sludge to utilize 250 mg/1 benzene as the sole carbon
source. By observing pressure changes in the absence of sludge, they, concluded that a
change in benzene concentration did not occur as a result of volatility. Presumably,
conditions did not include aeration. The bacterial species isolated from the sludge
were tentatively identified as Flavobacterium lactis, Achromiter sulfurem, Achromo-
bacter superficiale, Alcaligenes marshal!! and Rhizobium lupin!.
Toxicity
Plants—
Kauss et al., (1973) observed the reduction in growth of Chlorella vulgaris when
the organism was exposed to various concentrations of benzene. At 25 ppm benzene, a
15 percent reduction in algal growth was noted after one day, but recovery was
complete within six days. A pattern of growth depression in proportion to benzene
concentration and the recovery to nearly normal growth rates within 6-10 days was
observed at all benzene concentrations below 500 ppm. When benzene concentrations
reached 1,000 ppm or more, algal populations did not recover from the benzene
addition. Volatilization of benzene likely occurred from the unstoppered flasks used in
this experiment, thereby inflating the amounts needed to actually inhibit growth.
Photosynthetic activity of the giant kelp (Macrocystis pyrifera) is reportedly
not affected by exposure to 10 ppm benzene over a 96-hour period (North et al., 1959).
The herbicidal effects of benzene and methylated derivatives on terrestrial plants
were examined by Currier (1951). From his work with partition coefficients (water to
paraffin oil) of benzene, xylene, and other simple aromatics, an inverse relationship
between toxicity and compound solubility could be drawn. Currier concluded that
cellular penetration is facilitated by increasing the number of methyl groups on the
benzene ring, thus, xylene or toluene might be expected to be more toxic to plants
than benzene.
Aquatic Organisms-
Distinct chemoreception abnormalities in marine crabs have resulted from
exposure to sublethal concentrations of benzene (Kittredge, 1971). Feeding behavior
in the lined, shore crab (Pachygrapsus crassipes) was inhibited by benzene; however,
concentrations required to initiate disruption of chemoreception were not reported.
31
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Benzene inhibition of crab feeding behavior tended to be more transient than that
observed for more complex aromatics, such as napthalene.
Shelford (1917) evaluated the effects of benzene from waste coal tar on
freshwater fishes. The small, orange-spotted sunfish (Lepomis humilis) used in
Shelford's experiments died within one hour of exposure to concentrations of 35 to 37
ppm benzene, preceded by behavioral abnormalities such as erratic movement and
intoxication. Median tolerance limits for bluegill sunfish were 20 ppm benzene over
2k- and 48-hour exposures (Turnbull et al., 1954). Benzene at 60 ppm killed all test
fishes within two hours and was lethal to bluegills within 24 hours at 34 ppm. These
investigators calculated that 6 ppm is a safe concentration in regard to acute toxicity
for bluegills.
The data of Wallen et al., (1957) appears grossly distorted due to air stripping of
benzene from the experimental tanks. All Gambusia affinis survived in benzene
concentrations of 300 ppm, 10 times the lethal threshold described by other investi-
gators. TLm values of 395 ppm (24- or 48-hr exposure) and 386 ppm (96-hour exposure)
for the mosquito-fish were reported, with a dose of 1,000 ppm benzene killing all fish
within 16 minutes. Black bullheads (Ictalurus melas) maintained a 1,580 ppm median
tolerance limit (48-hour test), but this value decreased to 780 ppm when a 40 percent
benzene-acetone mixture was introduced to insure greater benzene solubility.
Acute toxicity of benzene to several freshwater fishes was investigated by
Pickering and Henderson (1966). Bluegills exhibited a TLm of 22.5 ppm for all
exposure intervals, almost precisely the same toxic concentration previously derived
by Turnbull et al. (1954). Fathead minnows (Pimephales promelas) tolerated benzene
slightly better in hard water (34.4 ppm TLm) than soft water (35.5 ppm TLm).
Goldfish (Carassius auritus) and guppies (Lebistes reticulatus) exhibited 24-hour TLm
values of 34.4 ppm and 36.6 ppm, respectively. The volatile nature of benzene may be
reflected in the fact that TLm values for each fish species remained the same
throughout 24-, 48- and 96-hour exposure periods.
Measurements of the respiratory stress exhibited by chinook salmon
(Oncorhynchus tshawytacha) and striped bass (Morone saxatilis) when exposed to
sublethal benzene concentrations were recorded by Brocksen and Bailey (1973). Since
15 ppm benzene initiated mortality among these juvenile fishes, levels of 5 and 10 ppm
were utilized as experimental conditions. For the salmon, 5 ppm benzene increased
respiration by 90 percent after 48 hours of exposure, while 10 ppm caused a mean
respiration rate increase of 115 percent. The respiration pattern differed significantly
for the striped bass, with increases reaching 50 percent at 5 ppm and only 15 percent
at 10 ppm over a 96-hour test period. Narcosis and related depression of respiration
below normal levels was evident, but recovery proceeded fairly rapidly after current
velocities in the respirometer were increased.
Brocksen and Bailey (1973) postulated that the differences in the lipid content
of the fish body fat were responsible for the different respiratory responses observed
for the two species. After benzene is absorbed across the gill surface into the blood,
it becomes attached to erythrocytes and lipoprotein for transport to lipid-rich tissues.
Metabolism by the liver, kidney, and various body tissues converts some benzene to
phenol, while smaller portions are excreted across the gills unchanged. Overloading of
the transport and breakdown mechanisms leads to a buildup of benzene, which
accumulates in the lipid rich nervous tissue and induces narcosis.
32
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Benzene at concentrations of 400 ppm did not produce avoidance responses in
green sunfish (Lepomis cyanellus) (Summerfelt and Lewis, 1967). A benzeneacetone
mixture failed to repel the sunfish from the area where the aromatic hydrocarbon
mixture was added. Benzene had initially been selected for study due to Shelford's
observations (1917) that it was avoided by freshwater fish.
ISOPRENE
Description
Isoprene is a colorless, volatile liquid that is a basic component of natural
rubber polymers. The chemical structure is CH2=C(CH3)CH=CH2- Although isoprene
is soluble in most common hydrocarbons and forms azeotropes with various organic
solvents, the substance is considered insoluble in water.
Isoprene can be dispersed in water by the action of soap and emulsif iers and the
resultant emulsion can be polymerized by the use of free radical initiators. Isoprene
forms peroxides when exposed to air in the absence of inhibitor substances (e.g.,
hydroquinone, tert-butyl catechol). Polymerized peroxides result in a "gummy" mass.
Physical/Chemical Properties
Boiling point: 34.067°C
Freezing point: -145.95°C
Density: 0.68095 g/ml at 20°C
Refractive index: 1.42194 at 20°C
Microbial Decomposition
No specific reports on the degradation of isoprene were located. The volatile
nature of the pure liquid and the "gummy" peroxides produced by aeration would surely
hinder the application of most biodegradation techniques.
Toxicity
Pickering and Henderson (1966) evaluated the toxicity of practical grade
isoprene to four species of freshwater fish. Bluegill sunfish with a 42.5 ppm TLm were
the most susceptible, while the TLm was 86.5 ppm for fathead minnows, 180 ppm for
goldfish, and 240 ppm for guppies. No variation between exposures of 24, 48 and 96
hours existed. In addition, isoprene toxicity to fathead minnows of different ages was
analyzed. Median tolerance limits of 75 to 85 ppm isoprene were found for minnow fry
(1 and 2 days old), juveniles (10 days old), and adults. No significant variation between
age and toxic susceptibility was indicated.
METHANOL
General Description
At room temperature, methanol (CH3OH) is a colorless neutral liquid. It is
infinitely soluble in water, although the density and viscosity constants for a
methanol-water mixture change with temperature and proportion of alcohol (Wood-
ward, 1967).
33
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Physical/Chemical Properties
Boiling point: 64.7°C
Freezing Point: -97.8°C
Density: 0.79609 g/ml at 15°C
Refractive index: 1.3287 at 20°C
Viscosity: 0.5945 cP at 20°C
Microbial Decomposition
Alcaligenes faecalis, a bacterium isolated from activated sludge, has demon-
strated the ability to metabolize methyl alcohol (500 ppm) and numerous other
aliphatic organic compounds (Marion and Malaney, 1963). An anaerobic microorga-
nism, Methanosarcina barker i, ferments methanol to methane and carbon dioxide
(Toenniessen and Mah, 1971; Bryant et ai., 1971), but such mechanisms may be of little
importance in natural systems. Numerous investigators have reported rapid degrada-
tion of methanol by microorganisms in activated sewage sludge (McKinney and Oeris,
1955; Dickerson et al., 1955; Hatfield, 1957). Previously, Placak and Ruchhoft (1947)
noted a decrease in activated sludge exposed to methanol and minimal degradation of
the compound.
The metabolic pathways for methanol proposed by McKinney and 3eris (1955)
provide for conversion of alcohol to aldehyde and then to organic acid. The organic
acid is introduced into the tricarboxylic acid cycle and is oxidized to carbon dioxide
and water. Stanier et al., (1970) reported that methanol can serve as a growth
substrate for Methanomonas methanica, producing allulose phosphate as the principal
intermediate. Bacteria that can utilize methanol and not methane include various
Pseudomonas and Hyphomicrobium species. An analogous pathway to that used by
Methanomonas methanica was utilized with pyruvic acid produced as the intermediate.
Although there were questions regarding whether methanol follows the oxida-
tion pathway of other primary alcohols (McKinney and 3eris, 1955; Marion and
Malaney, 1963), numerous recent works confirm that bacteria that grow on methane or
methanol as a source of carbon and energy, oxidize methane, methanol, formaldehyde,
and formate to carbon dioxide as follows (Patel and Hoare, 1971; Brown et al., 1964;
Dworkin and Forster, 1956; Anthony and Zatman, 1965; Anthony and Zatman, 1967a,
1967b; Hepstinsall and Quayle, 1970):
OH O O
H— C-H "2H
-)H-C-H
H
methanol formaldehyde formic acid
Oxidation of methanol and formaldehyde is known to be catalyzed by a nonspecific
primary alcohol dehydrogenase (PAD) that is activated by ammonium ions. Bacteria
reported to have this ability include Alcaligenes faecalis (Marion and Malaney, 1963),
Methylococcus capulatus (Patel and Hoare, 1971), Methanomonas methanooxidans
34
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(Brown et al., 1964) and various Pseudomonas species. (Dworkin and Forster, 1956;
Anthony and Zatman, 1965, 1967a, 19675; Heptinstall and Quayle, 1970).
Toxicity
Mammalian Toxicity—
Methyl alcohol exerts its toxic action on mammalian systems primarily upon
ingestion. Only 9.1 mg methanol/kg of body weight has been reported as the acute
oral toxicity level for rats (Welch and Slocum, 1943). Ingestion of this toxic alcohol by
humans may lead to blindness and even death. Metabolism of methanol by microbes
did not appear to follow the normal oxidation pattern of alcohol to aldehyde to organic
acid (McKinney and Jeris, 1955), but if mammalian systems follow the normal
pathway, the resultant metabolite, formaldehyde, may be the toxic agent.
Aquatic Organisms—
Although methanol exhibits an almost infinite solubility in water, aquatic
organisms appear extremely tolerant of this low molecular weight alcohol. Anderson
(1944) reported on the threshold concentrations of methanol required to narcotize or
immobilize Daphnia magna. A one-molar solution of methanol (32,000 ppm) inhibited
swimming of these cladocerans in Lake Erie water. Methanol exhibited the highest
threshold concentration of all salts and organic compounds tested. Stimulation of
movement in water beetles (Laccophilus maculosus) exposed to various anions, cations,
and simple organics was the reaction threshold tested by Hodgson (1951). Methanol
concentrations of 3.6 moles (115,000 ppm) elicited movement by 50 percent of the
beetles. This extremely high chemotaxic threshold is 190 times higher than the
detection threshold for the barnacle Balanus (Cole and Allison, 1930), but three times
less than the sensitivity level for the blowfly Phormia (Dethier and Chadwick, 1947).
The creek chub (Semotilus atromaculatus) in Detroit River water exhibited a
critical toxicity range of 8,000 to 17,000 ppm for methanol (Gillette et al., 1952). This
toxicity level indicated the acute threshold above which all test fishes died and below
which all survived for 24 hours. A typical TLm value for this type of short-term
experimental test should fall near the middle of this critical range. McKee and Wolfe
(1963) cited data previously reported showing that 8,100 ppm methanol did not injure
fingerling trout from natural waters in 24 hours. In addition, adult trout tolerated
10,000 ppm for two hours without adverse effects.
NITROPHENOL
General Description
Nitrophenol (ortho, meta, and paraisomers) occurs in phenol-acclimated cul-
tures, but at a substantially reduced rate in comparison to most other phenol and
benzene derivatives (Chambers et al., 1963). Oxygen uptake by microorganisms
exposed to 100 ppm nitrophenol was greatest for m-nitrophenol, followed by o-
nitrophenol, and then p-nitrophenol at slightly lower rates. Metabolic pathways
suggesting the removal of the nitro group as nitrite before ring cleavage have been
postulated (Simpson and Evans, 1953), but some investigators believe that benzene ring
cleavage occurs before the nitro group is eliminated. Still, the abilities of natural
-------
microbial populations to metabolize the nitrophenols and the enzymatic pathways by
which this is accomplished are largely unknown.
Toxicity to Aquatic Organisms
Applegate et al., (1957) reported that 5 ppm o-nitrophenol and p-nitrophenol
were not lethal to~~!reshwater trout, bluegills, or lamprey larvae. The 24-hour TLm
value for bluegills exposed to o-nitrophenol was 66.9 ppm, while 46.3 to 51.6 ppm o-
nitrophenol caused 50 percent mortality in 48-hour tests (Lammering and Burbank,
1961). Chronic toxicity may be significant below 50 ppm for longer exposure periods;
in fact, Applegate et al., (1957) implied that 37 ppm might be a typical 96-hour TLm
value. Loss of equilibrium by the bluegills preceeded death for most nitrophenol
concentrations but some recovery was noted at sublethal concentrations.
Bringmann and Kuhn (1959) tested several food-chain organisms from the Havel
River (Germany) for their toxicity thresholds to the nitrophenols. Daphnia exhibited
adverse effects to 14 ppm p-nitrophenol, 24 ppm m-nitrophenoi, and 60 ppm o-
nitrophenol. Microregma, a protozoan, reacted equally towards m-nitrophenol and o-
nitrophenol (20 ppm) and did not exhibit negative effects until 40 ppm o-nitrophenol
was reached. The green algae, Scenedesmus, was most tolerant to p-nitrophenol (72
ppm) and less tolerant to o-nitrophenol (36 ppm) and m-nitrophenol (28 ppm). More
than 1,000 ppm o-nitrophenol was required to inhibit Escherischia coli, although the
bacteria responded to 300 ppm of m-nitrophenol and 100 ppm of p-nitrophenol.
NONYL PHENOL
General .Description
Nonyl phenol — CgH^CgH^OH — is a clear, yellow, viscous liquid with a
slightly phenolic odor and is used in chemical manufacturing detergents, oil additives,
and rubber chemicals. It is very slightly soluble in water, but it is soluble in benzene,
ether, carbon tetrachloride, alcohol, and acetone.
Physical Characteristics
Molecular weight: 220.3
Specific gravity: 0.94 at 20°C
Distillation range: 279- 301 °C
Flash point: 140°F
Boiling point: 300°C
Microbial Decomposition
Nonyl phenol is a phenol ring attached to a chain of 9 carbon atoms. The
literature review did not reveal any specific articles on microbial decomposition of
this material; however, McKenna (1972) reported on microbial metabolism of a
benzene ring attached to a 10-carbon chain. Apparently, good growth was obtained in
a number of bacterial strains; however, it was not obvious whether the chain or the
ring compound was attacked first. Chapman (1972) stated that a chain containing 3 or
more carbon atoms can undergo oxidation provided that extensive branching is not
present.
36
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Toxicity to Aquatic Organisms
No information on the toxicity of nonyl phenol to aquatic microorganisms and
invertebrates was found. However, experimentation with larval sea lampreys (Petro-
myzon marinus), rainbow trout (Salmo gairdnerii), and bluegills has shown that 5 ppm
of this alkylated phenol causes death or "obvious distress" to the fishes (Applegate et
al., 1957). This experimental level of nonylphenol caused trout to succumb within four
Hours, while bluegills and lamprey larvae survived for 14 hours before dying.
Marchetti (1965) showed that 5.2 ppm of nonyl phenol ethoxylate, a surfactant
derivative of nonyl phenol, was toxic to juvenile rainbow trout (40 and 210 days old)
within 6 hours. Newly hatched trout alevins could tolerate 42 ppm, but after
absorption of the yolk sac the fry found 2.5 ppm to be lethal.
PHENOL
Description
Phenol (hydroxy-benzene) has a single hydroxyl group attached to a benzene
ring. Phenol is crystalline when pure and stored at room temperature. However, the
commercial grade ("liquified phenol") is either 80 to 82 percent pure phenol, with the
remainder being cresols and water. Phenol is water soluble (82 gm/1 at 15°C),
although its solubility varies with temperature up to 65.3°C. Above this temperature,
it is miscible in water in all proportions.
Physical/Chemical Properties
Molecular weight: 94.11
Boiling point: 181.75°C
Freezing point: 40.9°C
Specific Gravity: 1.07 at 20°C
Microbial Decomposition
Phenol is thought to be transformed directly into catechol (dihydroxybenzene).
The transformation of catechol into CO2 and H2O follows the ring cleavage scheme
described previously in the section on benzene. According to Chapman (1972), the
microorganism that has been utilized to demonstrate this particular pathway is
Pseudomonas putida.
Hermann (1959) found that 1,600 ppm phenol was inhibitory to sewage micro-
organisms. At this level, oxygen utilization by the sewage bacteria was halved. Over
1,000 ppm phenol was necessary to produce threshold toxicity effects on E. coli, the
common sewage bacterium (Bringmann and Kuhn, 1959). Lutin (1970) demonstrated
that 500 ppm phenol exhibited toxic action on some unacclimated sewage sludge;
however, mixed bacterial cultures that had been acclimated to 250 to 500 ppm phenol
were shown to degrade phenol as well as numerous phenol and benzene derivatives
(McKinney et al., 1956; Chambers et al., 1963). Previously, Evans (1947) isolated
numerous soil microorganisms that possessed phenol-metabolizing abilities, further
indicating the widespread environmental distribution of phenol-tolerant micro-
organisms.
37
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Evans (1947) proposed that dihydroxyphenol (catechol) was the metabolic
intermediate in the biological decomposition of phenol because it was detected in the
culture fluid. Based on the concept of simultaneous adaptation, which means that
bacterial adaptation to a single compound results in adaptation to its metabolic
intermediates, Stanier (1950) reported a widely accepted phenol decomposition
pattern. The first .step of phenol oxidation is catechol formation by tyrosinase.
Catechol is further converted to either aldehydo-acid or ketoacid, which is converted
to carbon dioxide and cell protoplasm:
short chain
•f n f*"i *\« . —. ^^^9
phenol catechol
meta
fission
Bacteria reported to have this ability include: Mycobacterium crystallophagum,
Micrococcus sphaeroides, Vibrio cuneata (Evans, 1947) and various Pseudomonas
species (Evans, 1947; Stanier, 1950; McKinney et al., 1956; Chambers etal., 1963).
Toxicity to Aquatic Organisms
Phenol and a host of phenol derivates have been extensively studied as major
pollutants with adverse effects on aquatic organisms. Since toxicity of these phenolic
compounds is generally cumulative and often synergistic, the observed environmental
disruption from phenolic waste discharge cannot be ascribed solely to phenol itself.
Biorefractory (taste and odor-causing) problems are an important sublethal effect of
phenolic compounds, especially for freshwater fish in which partitioning through the
skin and the resultant tainting of fish flesh commonly occur in phenol-contaminated
waters.
Phenol-containing wastewater was apparently responsible for alteration of the
"flora and fauna" in a Luxembourg river (Krombach and Barthel, 1963). Destruction of
the aquatic community was apparent in areas containing 10 ppm phenol. However,
presence of other toxicants in the waste may have contributed to the toxic effects
observed. When phenol was applied as a mosquito larvacide, differential survival of
fish and Anopheles (mosquito) larvae revealed that vertebrates may be more suscep-
tible to phenol than lower food-chain organisms (Knowles et al., 1941).
The green alga Scenedesmus reached its toxicity threshold concentration at 40
ppm phenol (Bringmann and Kuhn, 1959), while the diatom Navicula linearis required
258 ppm phenol to show a 50 percent reduction in growth rate over 120 hours (Patrick
et al., 1968). The phycotoxic action of phenol has been attributed to its tendency to
form compounds within the cytoplasm of Ankistrodesmus and Scenedesmus (Keyna,
1940). For the giant marine kelp Macrocystis pyrifera, one ppm phenol caused no
adverse effects, but 10 ppm produced a 50 percent reduction in photosynthetic activity
over a four-day exposure (Clendenning and North, 1960).
38
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Ukeles (1962) showed that phenol concentrations above 10 ppm were inhibitory
to growth of several algal species, especially flagellated forms (Monochrysis lutheri
and Dunaliella) and non-motile coccoid types (Protococcus and Chlorella).Proto-
coccus, Chlorella, and Dunaliella maintained approximately 50 percent of the control
growth rate at 300 ppm phenol, but populations died off at 500 ppm phenol. The
diatom, Nitzchia closterium, and Monochrysis lutheri remained viable but exhibited no
growth at 100 ppm.
Phenol toxicity to aquatic organisms has been documented by several investi-
gators. Bringmann and Kuhn (1959) showed that 30 ppm phenol is the toxicity
threshold for the protozoan Microregma. Daphnia magna were immobilized at 94 ppm
phenol in Lake Erie water (Anderson, 1944), but the crustaceans (adult) reached a 24-
hour TLm at 61 ppm and a 48-hour TLm at 21 ppm (Dowden and Bennett, 1965).
Juvenile daphnids responded negatively at lower phenol concentrations (17 ppm TLm
over 24 hours and 7 ppm TLm for a 48-hour exposure). Sollman (1949) calculated an
intermediate TLm of 28.9 ppm phenol for a 48-hour Daphnia exposure. Gammarus
pulex, Tubifex worms, and chironomid larvae were reportedly not affected by a 4.3
ppm phenol level (Liepolt, 1953). Patrick et al., (1968) demonstrated that the
freshwater snail, Physa heterostropha, achieved 50 percent survival over 96 hours at
94 ppm phenol.
Goldfish (Carrassius auritus) have been used frequently as experimental subjects
for phenol toxicity experiments. Powers (1918) observed that goldfish survived for
only 104 minutes in 259 ppm phenol. Twenty-five ppm phenol was apparently not
lethal to goldfish, whereas 41.6 ppm resulted in two-thirds mortality and 125 ppm
killed all fishes in 8 hours during experimentation by Gersdorff and Smith (1940).
Gersdorff (1943) also determined acute phenol toxicity at concentrations of 199 to
1,460 ppm, where survival times for goldfish ranged from 12 to 85 minutes. A miminal
lethal dose of phenol, when injected into goldfish tail muscle, was 230 mg phenol/kg of
fish body weight (Boni, 1965). Boni also determined that goldfish actively excreted
phenol directly without first conjugating the compound. A 44.5 ppm TLm for goldfish
exposed to phenol (96 hours) provides the most reasonable estimate of this compound's
acute toxicity (Pickering and Henderson, 1966).
Another fish used extensively for phenol toxicity bioassay is the bluegill,
Lepomis macrochirus. Trama (1955) determined that 20.5 ppm constituted the 96-hour
TLm for this sunfish. Comparable bioassay investigations of this species determined
96-hour TLm values of 11.5 ppm (Cairns and Scheier, 1959), 13.5 ppm (Patrick et al.,
1968), and 26 ppm (Pickering and Henderson, 1966). Shorter-term tests by Dowden and
Bennett (1965) found that 10 to 15 ppm phenol was an adequate 24-hour TLm for
bluegills, while Lammering and Burbank (1961) calculated 22.2 ppm phenol as the
48-hour TLm. Based on these previous bioassays, phenol concentrations of 10 to 25
ppm are acutely toxic to bluegill sunfish, indicating that these organisms have a
relatively higher sensitivity to phenol than do goldfish.
The effects of major environmental parameters (salinity, dissolved oxygen
concentration and temperature) on the toxicity of phenol to rainbow trout (Salmo
gairdnerii) have been studied by several British investigators. Brown, Jordan and Tiller
(1967) found that the median tolerance limit for phenol doubled with a temperature
increase from 6°C to 18°C. Trout acclimated to 20 percent salt water showed a 5.2
ppm TLm (48-hour test), whereas those trout held in fresh water exhibited a higher
TLm (9.3 ppm) (Brown, Shurben and Fawell, 1967). Lloyd (1961) found that reduction
39
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in dissolved oxygen concentration below the air saturation value led to increased
phenol toxicity. Phenol in concentrations of 6.5 to 9.6 ppm severely damaged the gill
filaments of trout in hard water and resulted in extensive damage to numerous organ
systems, particularly the liver, kidney, spleen, skin, small intestine, and ovary
(Mitrovic et al., 1968).
Bioassays using freshwater fishes frequently demonstrate interspecific vari-
ability with regard to phenol toxicity. During their search for an effective lamprey
larvacide, Applegate et al. (1957) noted that 5 ppm phenol was lethal to rainbow trout
in 10 hours, but was not toxic to bluegill sunfish or sea lamprey larvae. Ten to twenty
ppm phenol was indicated as the critical toxicity range (0 to 100 percent morality) for
the creek chub, Semotilus atromaculatus (Gillette et al., 1952). The green sunfish,
Lepomis cyanellus, was not adversely affected or even repulsed by phenol levels of 20
ppm (Summerfelt and Lewis, 1967). However, the minnow, Phoximus phoxinus, could
not discriminate between phenol concentrations of 4 to 400 ppm and became poisoned
by the higher doses (Jones, 1951). Forty-eight hour TLm tests produced values of 40.6
ppm for fathead minnows (Pickering and Henderson, 1966), 56 ppm for mosquitofish
(Wallen et al., 1957), 16.7 ppm for fingerling channel catfish (Clemens and Sneed,
1959), and 49.9 ppm for guppies (Pickering and Henderson, 1966).
STYRENE
Description
Styrene (vinyl benzene) is an unsaturated aromatic compound (CgH^CHsC^)
used extensively in plastics production (Coulter et al., 1967). This colorless,
flammable liquid undergoes the reactions typical of an unsaturated compound and has
been utilized in a host of plastics polymer formulations for this reason. Its water
solubility is only 0.032 percent at 25°C, although it is infinitely soluble in most organic
solvents. Pure styrene polymerizes slowly at room temperature and more rapidly
under warmer conditions. Also, oxygen tends to degrade styrene. Polymerization in
storage is retarded by 10 to 15 ppm TBC (tert-butyl catechol).
Physical/Chemical Properties
Boiling point: 145°C
Freezing point: -30.6°C
Viscosity: 0.763 cP at 20°C
Density: 0.9059 g/cm3 at 20°C
Refractive index: 1.5467 nrj at 20°C
Microbial Decomposition
The number of investigations in the literature on the microbial degradation of
phenol are minimal.This may be due to the almost total consumption of phenol for
polystyrene plastics products and the lack of unpolymerized styrene in the waste
streams of this industry. Ludzack and Ettinger (1960) referred to experimentation by
Pahren on the oxidation of styrene by unacclimated activated sludge. Only eighteen
percent of the theoretical oxidation was achieved with 10 ppm vinyl benzene, implying
relatively slow biodegradation over the 10-hour test period.
40
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Toxicity to Aquatic Organisms
Literature references to the aquatic toxicity of styrene are minimal. Bioassays
of styrene toxicity to fathead minnows, bluegills, goldfish, and guppies were performed
by Pickering and Henderson (1966). Ninety-six-hour TLm values were 25 ppm for
bluegills, 64.7 ppm for goldfish, and 74.8 ppm for guppies. Median toxicity was greater
for fathead minnows in soft water (46.4 ppm) than in hard water (59.3 ppm).
Generally, no variation between 24-, 48- and 96-hour tests occurred, except for small
TLm decreases observed in fathead minnows exposed over longer time periods.
TOXAPHENE
Description
Rather than being a distinct compound, toxaphene (CjoHioClg) is actually a
mixture containing polychloro-bicyclic terpenes. This insecticide is made by chlori-
nating camphene and its chemical structure is:
C.I-C-
The technical product is a yellowish, semi-crystalline gum that contains 67 to 69
percent chlorine and has a melting point of 65° to 90°C. However, it dehy-
drochlorinates in the presence of alkali, prolonged exposure to sunlight, and at
temperatures above 155°C. Toxapheneis generally classified with aldrin as an
organochloride pesticide. Like aldrin, it possesses a very low vapor pressure, very low
water solubility, and resistance to biological degradation.
The strong tendency of toxaphene to adsorb to various surfaces has been
reported in the laboratory as well as in natural systems. Courtenay and Roberts (1973)
reported that during their bioassay tests using plastic-lined vessels a substantial
portion of the toxaphene was sorbed on the walls of the vessels. In natural systems,
Hughes and Lee (1973) reported that environmental persistence is rather complex,
since sorption and desorption mechanisms control the presence of this material.
Toxaphene transport into lake sediments appears to be a major mechanism for
detoxification (Veith and Lee, 1971). This pesticide penetrated lake sediments as deep
as 20 cm and resisted subsequent attempts in the laboratory to be leached from the
sediments. Also, toxaphene sorbed onto suspended algae and organic particulates and
was transported to the sediment surface when this material sank.
Toxicity to Aquatic Organisms
Toxaphene is extremely phytotoxic, with phytoplankton productivity reduced 91
percent when exposed to 1 ppm for 4 hours (Butler, 1963). Concentrations below 0.01
ppm were determined to be sublethal for the green algae Scenedesmus incrassatulus
(Schoettger and Olive, 1961). However, Monochrysis lutheri (green marine flagellate)
did not grow in O.00015 ppm toxaphene and growth of Nitzchia closterium (diatom)
stopped at 0.04 ppm, but cells remained viable (Ukeles, 1962). Other algal bioassays
showed inhibited growth at 0.07 ppm for Chlorella and 0.15 ppm for Protococcus and
41
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Dunaiiella. Needham (1966) reported slight declines in blue-green algal species
following application of 90 ppb toxaphene to North Dakota lakes.
Application of toxaphene at 100 ppb caused reduction of Tendipedidae on the
lake bottom and repopulation took almost a year (Gushing and Olive, 1957; Needham,
1966). Schoettger and Olive (1961) revealed a 0.03 ppm toxicity threshold for both
Daphnia pulex and D. magna, while the Daphnia pulex tested by Cope (1966)
succumbed to one-half this level (0.015 ppm). O~ther zooplankton populations in
freshwater lakes, especially entomostracans (Cyclops, Ceriodaphnia, Diaptomus),
rotarians (Polyarthia, Keratella, Asplancha, Brachionus) and protozoans (Ceratium,
Difflugia), declined when exposed to toxaphene, (Hoffman and Olive, 1961; Needham,
1966). However, 100 ppb toxaphene (used to eliminate a fish population) did not
permanently alter the lake's invertebrate populations of physid snails, microcrusta-
ceans, and Chironomus or Chaeborus larval stages (Hilsenhoff, 1965).
The insect naiads (Pteronarcys californica, Pteronarcella badia, Claasemia
sabulosa) reacted negatively to 1.3 to 3.0 ppb over a 4 day exposure (Sanders and Cope,
1968). The amphipod Gammarus lacustris maintained 50 percent survival in 0.5 ppm
toxaphene for only 96 minutes (McDonald, 1962). In comparison, Paleomonetes
kadiakensis appeared extremely tolerant to toxaphene with 57 to 180 ppm necessary to
kill this shrimp (Ferguson, Culley et al., 1965). Toxaphene toxicity to larval
crustaceans was greatest during the molting stages. Courtenay and Roberts (1973)
used penaeid shrimp and blue crab larval stages to demonstrate maximum toxaphene
stress.
Numerous accounts of toxaphene used to eradicate the entire fish population
from a lake have been reported. Lawrence (1950) noted that 0.02 ppm was lethal to
bluegills, golden shiners, and bass fingerlings in an Alabama pond after two days.
Carp, golden shiners, bluegills, yellow perch, bonytail chub, and brown trout were
eliminated from several Arizona lakes with 0.1 ppm toxaphene. Restocking was
accomplished successfully 9 to 10 months later with rainbow trout (Hemphill, 1954).
Fukano and Hooper (1958) stated that a 5 ppb toxaphene concentration could be used
to selectively reduce small fish populations (bluegill and pumpkinseed sunfish) and not
affect larger fishes (yellow perch, largemouth bass, and rock bass). Roughfish, such as
carp and bullheads, could be controlled by 25 ppb toxaphene in Iowa lakes, although
excessive turbidity may remove some pesticide by adsorption (Rose, 1958). Stringer
and McMynn's (1958) experiments with various British Columbian lakes demonstrated
considerable interlake variation of toxaphene toxicity (10-100 ppb) to similar fish
fauna. Other investigations of roughfish control by toxaphene application have been
conducted in Florida (Huish, 1961), New Mexico (Kallman et al., 1962), Montana
(Wollitz, 1963), Alaska (Meehan and Sheridan, 1966), and North Dakota (Warnick, 1966;
Henegar, 1966).
Forty-eight-hour TLm values for bluegills ranged from 3.5 to 4 ppb toxaphene
(Tarzweil, 1959; Henderson et al., 1959; Weiss, 1964; Cope, 1966). Ferguson et al.
(1964) reported pesticide-resistant bluegills that could tolerate 1.6 ppm during short-
term bioassays. Fathead minnows exhibited median tolerance limits of 7.5 to 13 ppb
(Henderson et al., 1959; Cohen et al., 1961). Earlier, Hooper and Grzenda (1955) had
found that toxaphene toxicity was greatest to fathead minnows in hard water, while
Schaumberg et al. (1967) determined that both high temperature (39°C) and near-
freezing levels increased toxaphene toxicity to fatheads.
42
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Gambusia affinis demonstrated that resistance acquired after a single sublethal
dose can raise the toxicity level from 0.01 to 0.48 ppm (Boyd and Ferguson, 1964).
Bioassays revealed that 0.005 to 0.059 ppm is a normal toxicity range for the
mosquitofish (Workman and Neuhold, 1963; Ferguson, Cotton et al., 1965). Guppies
tested by Henderson et al. (1959) required 0.02 ppm as the toxic level, but Royer
(1966) found only 1 ppb to be lethal to this species. Mahdi (1966) ran bioassays on
Notemigoneus crysoleucas, Carassius auratus, Ictaluras melas, and Campostoma
anomalum to find median tolerance limits of 12.5-ppb and 94-ppb. The organisms
survived up to 2.5 ppb and 14 ppb under the 12.5 and 94 ppb exposure conditions,
respectively. Channel catfish fingerlings supposedly survived up to 2.5 ppm toxaphene
(Clemens and Sneed, 1959), but bullheads found 0.004 to 0.05 ppm toxic (Ferguson,
Cotton et al., 1965). According to Johnson (1966), goldfish required 0.03 to 0.100 ppm
of the toxicant before dying.
Rainbow trout and other cold-water fishes have often been used as bioassay
organisms, since restocking of these gamefish have generally followed roughfish
eradication using toxaphene in suitable lakes. Salmo gairdnerii reacted adversely to
levels as low as 4 ppb (Cope, 1966) and as high as 54 ppb (Workman and Newhold,
1963). However, most reports show a toxicity range of 8.4 to 16.5 ppb (Katz, 1961;
Webb, 1961; Mahdi, 1966). Cope (1965) indicated that temperature increases (45 to
65°F) decrease the amount of toxicant needed to cause mortality. Coho salmon
(Oncorhynchus kisutch) found 9.4 ppb as the median tolerance limit, but sockeye
salmon (O. nerka) and chinook salmon (O. tschawytscha) reacted to only 3.6 ppb and
2.5 ppb, respectively (Schoettger and Olive, 1961; Katz, 1961).
Courtenay and Roberts (1973) encountered rapid adsorption of toxaphene to
plastic-lined bioassay containers, which distorted their 1X50 values for a group of
freshwater and estuarine fishes. They found however, that toxaphene toxicity varied
directly with increasing salinity and they postulated that this pesticide interferes with
basic osmoregulatory metabolism. For fishes, the most pesticide-sensitive stage was
during development of gills in newly-hatched fry.
XYLENE
Description
Xylene, an 8-carbon aromatic hydrocarbon, is a dimethyl derivative of benzene.
Xylene occurs in three isometric ring forms, as shown below:
CH,
m-xylene
Xylenes are generally considered insoluble in water; however, McAuliffe (1963)
found that 175 ppm is the approximate water solubility of o-xylene at room
temperature. Although this degree of water solubility is only onetenth that of
benzene, it is still substantially higher than other alkylated benzenes and aliphatic
43
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hydrocarbons. m-Xylene forms an azeotrope with water, which contains 64.2 percent
m-xylene.
Physical/Chemical Properties
o-Xylene m-Xylene p-Xylene
Boiling point: 144.41°C 139.103°C 138.35°C
Freezing point: -25.182°C -47.872°C -25.182°C
Density (at 20<>C): 0.8802 g/ml 0.86417 g/ml 0.86105 g/ml
Refractive index (at 20°C) 1.50545 nD 1.49722nD 1.49582 nD
Microbial Decomposition
The degradation pathway of the three xylene (dimethyl-benzene) isomers is
similar to that of benzene and phenol. In mammalian systems, Laham (1970) has shown
that various xylenols are produced during metabolism of the xylenes. Chapman (1972)
found that Pseudomonas putida demonstrated a similar reaction sequence. He
reported that 2, 4-xylenol possesses two methyl groups that are oxidized in succession,
yielding protocatechuic acid. The latter compound, which is also a benzene degrada-
tion product, is transformed into -ketoadipic acid, an acceptable substrate for the
tricarboxylic acid cycle. For the other isomers, 2, 3-xylenol and 3, 4-xylenol, the
methyl groups are kept intact to form catechol, which can be broken up via meta
cleavage.
Microorganisms in activated sludge have demonstrated degradation of xylene
isomers, but the degradation rates differ. When acclimated to aniline (aminobenzene)
before experimentation, sludge bacteria metabolized p-xylene the most and m-xylene
the least (Malaney, 1960). Activated sludge acclimateid to large doses of benzene
exhibited significant oxidation of the xylenes. o-Xylenes were degraded the most
rapidly and m-xylenes oxidized the most slowly (Malaney and McKinney, 1966).
Zoogloea bacteria commonly isolated from polluted natural waters and wastewater
streams utilized m-toluate and p-toluate, which are metabolic derivatives of the
xylene isomers (Unz and Farrah, 1972).
Nozaka and Kusunose (1968) also found that microbial metabolism of pxylene
paralleled mammalian pathways. They determined that Pseudomonas aeruginosa
oxidized l^C-labeled p-xylene to p-toluic acid via the corresponding alcohol.
Mammalian Toxicity
None of the three xylene isomers (ortho-, meta-, para-) exhibit the longterm
depressive effects on production of red blood cells that results from benzene poison-
ing. Although experiments with mammals have demonstrated a higher acute toxicity
for xylene than benzene, inhibition of the hemopoietic system and resultant aplastic
anemia is unproved. Acute effects from inhalation of xyiene vapors generally consist
of narcosis and unconciousness rather than death (Browning, 1950).
Xylene's role as a commercial solvent has increased substantially due to
minimal occupational health dangers in comparison to the other aromatic solvents,
benzene and toluene (Browning, 1959). The neurotoxic effect of xylene is greater than
44
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benzene or toluene when considering injury to nervous tissue and activity of brain
cells, but is less than that of the longer chain alkylated benzenes.
Metabolism of xylene isomers by organisms results in oxidation to toluic acid
and subsequent combination with the amino acid glycine to form toluric acids.
Another mechanism for detoxification involves the hydroxylation of xylene to produce
xylenols. These metabolites conjugate with proteins and sulfhydryl groups in the blood
to produce glucuronides and some ester sulfates, which are easily excreted. The
metabolites of xylene are much less toxic to mammalian systems than the biotransfor-
mation products of benzene, especially catechol (Laham, 1970).
Toxicity to Aquatic Organisms
o-Xylene in concentrations between 25 to 100 ppm exerted short-term toxicity
to Chlorella vulgar is (Kauss et al., 1973). Recovery appeared to occur as a result of
volatilization of this aromatic hydrocarbon, but at concentrations near its water
solubility (171 ppm), algal growth was completely inhibited.
The acute toxicity of xylene to aquatic organisms varies considerably. Dowden
and Bennett (1965) reported a median tolerance range of more than 100 ppm but less
than 1,000 ppm for Daphnia magna exposed to xylene for 2k hours. Acute toxicity of
the xylene isomers to fish is variable, with m-xylene typically the most toxic and o-
xylene the least toxic. The TLm values for xylene are slightly lower than those for
benzene, indicating higher toxicity for these less water-soluble aromatics (Pickering
and Henderson, 1966). Pickering and Henderson determined that 50 percent survival
over 96 hours occurred at 26.7 ppm for fathead minnows, 20.9 ppm for bluegill sunfish,
36.8 ppm for goldfish, and 34.7 ppm for guppies. These values compare favorably with
previously reported toxic concentrations for short-term exposures to freshwater fishes
(Shelford, 1917; Hubault, 1936). Applegate et al. (1957) discovered that rainbow trout
and larval sea lampreys were not affected by m-xylene at concentrations of 5 ppm
during a 24-hour exposure, but bluegills exhibited sublethal effects after 10-hr
exposure. These tests (which were used to evaluate potential larvacides for control of
sea lampreys) suggested differences in toxic susceptibility between coldwater and
warm-water fishes.
45
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SECTION 7
TREATABILITY SCREENING TESTS
Screening tests were performed early in the project to determine whether the
contract compounds were biodegradable and thus candidates for the biological
countermeasure. At the same time, growth kinetics data were obtained if possible, to
better define the methodology for the subsequent growth kinetics tests reported in
Section 8. The procedures and results for the screening tests are described below.
PROCEDURES
Individual screening tests were made utilizing a two-liter sample in a batch,
aerated plexiglass reactor 8.75 cm wide, 12 cm long, and 32 cm high. At the start of
each test, each reactor was filled with 500 ml of sludge from a previous test or 500 mi
of mixed liquor from a local sewage treatment plant. Two sources were available for
mixed liquor—Friendswood, a combined industrial-municipal plant near Houston, and
Govalle, a domestic waste treatment plant in Austin. At the start of this work, it
seemed advisable to obtain sludge from sewage containing a heavy load of industrial
wastewater. However, this precaution soon proved unnecessary.
The remaining volume of the reactor was filled with distilled water to which
nitrogen, phosphorus, and an alkalizing agent, as well as the desired dosage of the test
chemical, had been added. Initially, 150 mg/1 glucose and 150 mg/1 of glutamic acid
were added to supplement the substrate for the microorganisms used. Later, however,
acclimated sludge was used and glucose and glutamic acid were deleted. In one test,
the glucose and glutamic acid were replaced by a 5% solution of yeast extract, which
provided a source of vitamins, but did not add an additional carbon source.
The length of each test was normally 24 hours, although some tests were
extended to 72 hours. Temperature was controlled at 24°C.
Suspended solids (total and volatile), oxygen consumption, pH, and total organic
carbon (TOC) were monitored throughout the tests. Normally, samples were taken at
frequent intervals over the 24-hour test period with the heaviest sampling density
placed early in the test. When glucose and glutamic acid were added, a control
reactor was used to subtract their contribution to the TOC measurments. Because
glucose (CgH^Og) has a molecular weight of 180.16, of which 40 percent is carbon, its
initial contribution to the TOC measurements was 150 mg/1 X 0.4 or 60 mg/1.
Likewise,the glutamic acid (C^HgO^N), with a molecular weight of 147.13 and a carbon
content of 41 percent, contributed 62 mg carbon/1 to the TOC measurements.
Occassionally the TOC value for the control was greater than that for a chemical;
such values were recorded as zero. A gas chromatograph was used to detect specific
chemicals in a number of the treated effluents. Correlation between the gas
chromatograph and the TOC results for phenols proved useful. Because of limitations
of the instrument, however, the results with other chemicals proved less satisfactory.
46
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Wet mounts and gram stains were made from the contents of the reaction
chambers to determine the types of microorganisms adapted to each chemical. Gram-
negative and gram-positive bacilli, and gram-positive cocci were seen. Selection of a
predominant organism was difficult with the gram stains, nevertheless, wet mounts did
provide useful information.
RESULTS
The results of the screening tests are given below for each of the chemicals
tested.
Aldrin
Four screening tests for aldrin using 9, 18, 27, and 36 mg/1 aldrin as initial
concentrations were performed on November 12-13, 14-15, 19-20, and 20-21, 1973. In
none of the tests could aldrin be detected over and above the glucose and glutamic
acid concentrations. It was concluded that aidrin either sorbed onto the reactor walls
or sludge solids, or could not be detected by TOC or the flame ionization detector
available on the gas chromatograph. Further tests to overcome these difficulties were
to be performed as time permitted; however, none were continued.
Acetone Cyanohydrin
Acetone cyanohydrin ((CH3)2C(OH)C = N, MW=85) contains a cyanide group and
has toxic effects similar to those of acrylonitrile. The same enzyme system that
detoxifies acrylonitrile (discussed below) also splits the analogous bond in cyanide
( —C=N) to ammonia and methane. This suggests that massive inoculation with
Azotobacter or Clostridium, in any area when a spill of compounds that decompose to
cyanide occurs, would be beneficial for mitigating the hazardous effects of the
compound. The organisms used in these tests do not attack acetone cyanohydrin
because the spatial arrangement of the two methyl groups prevents the enzyme from
acting on the C SN groups. However, acetone cyanohydrin decomposes readily to
form cyanide and thus is susceptible to the enzyme system under study.
Acrylonitrile
Acrylonitrile (H2C=CH-C = N, MW=53.06) is among the most dangerous of the
high priority compounds investigated in this study. It is flammable at high concentra-
tions, forms explosive mixtures with air, and even at low concentrations is highly toxic
through the effect of the -C=N moiety (less than 1 gram being a potentially fatal dose
for humans).
No bacteria were isolated that were capable of using acrylonitrile as a growth
substrate. However, it was possible to detoxify acrylonitrile by use of an enzyme
cross-reaction. The ordinary substrate of the enzyme nitrogenase is atmospheric
nitrogen, N2 or N=N, which is reduced in certain microbes to two ammonia (NH3)
molecules that are readily assimilated into amino acids and proteins. Because of the
similarity in bond structure between the natural substrate, N=N, and the cyanide
portion of acrylonitrile, R-CBN, nitrogenase will also reduce acrylonitrile by the
irreversible reacton illustrated below:
47
-------
1.0 IMH3 ammonia
nitrogenase
1.0 H2C = HC-C=N ^ 0.9 CH2=CH-CH3 propene
acrylonitrile /^ 0.1 CH3-CH2-CH3 propane
electrons
ATP
Not only is the lethal cyanide group reduced in this reaction, but the ammonia
produced is a beneficial nutrient. Propene and propane are inocuous compounds that
are readily oxidized to CO2 by a wide variety of soil bacteria.
Bacteria were available in the laboratory that were capable of producing the
acrylonitrile-detoxifying enzyme in large amounts. The aerobic Azotobacter vine-
landii and the anaerobic Clostridium pasteurianum are non-pathogenic soil saprophytes
that may be conveniently stored for long periods then grown to large numbers, with
high nitrogenase activity maintained. Although the enzyme can be purified and
concentrated and these experiments were designed to illustrate the practical use of
enzyme preparations to detoxify hazardous materials, whole cells can also be used.
When properly prepared they maintain the enzyme in a stable condition over long
periods. In addition, maintaining the enzyme in the bacteria permits the living cells to
furnish the necessary metabolites to drive the reaction forward. With enzyme
extracts it is necessary to add such auxiliary compounds as ATP, creatinine phosphate
dithionite, and others, thus complicating the problem. Azotobacter and Clostridium
cells are readily produced in massive quantities, the former being produced in the
U.S.S.R. for many years as a soil inoculant and strains of the latter being cultured
world-wide for industrial fermentations. Many strains and mutants of Azotobacter are
maintained in the authors' laboratories and methods for readily securing additional
mutants from this genus, which was formerly regarded as difficult to mutagenize, have
been perfected.
Experiments in this laboratory showed the following:
1. It was necessary to exclude or at least reduce the natural enzyme substrate,
nitrogen, if a significant rate of acrylonitrile reduction was to take place.
2 It was theoretically possible to reduce acrylonitrile at a rate equal to 20% of
the nitrogen reduction rate.
3 Enzyme saturation by acrylonitrile was achieved at 50 mM (2.65 mg/1) in cell
free systems, but in whole cell experiments optimum activity took place at about 1 mM
(0.053mg/l) acrylonitrile - this difference was probably due to the toxicity of
acrylonitrile to living cells.
4 Large-scale acrylonitrile reduction with cell-free enzyme preparations was
not practical because: (a) the cell-free enzyme was extremely oxygen sensitive, (b) a
large flow of high-energy electrons was required for activity, and (c) ATP was required
in such large amounts (about 25 ATP's per acrylonitrile reduced) that adding it was
impractical.
A useful method for the study of acrylonitrile degradation by nitrogenase-
containing bacteria was developed. In this procedure, bacteria were grown to late log
phase (about 16 hours) with ammonium in the medium, allowing rapid growth and a
48
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large crop. These bacteria were pelleted by centrifugation then resuspended in a
nitrogen-free medium in a closed vessel having an air dispersion tube and a septum-
stoppered sample port. The suspension of bacteria was then gassed continuously with
non-nitrogen inert gas. Pure argon was used for anaerobes and a mixture of 80:20
(argon: oxygen) was necessary for aerobes. Maximum amounts of nitrogenase activity
were present after approximately two hours under these conditions. When maximum
nitrogenase activity was attained, 10 ml of the culture were dis-tributed by syringe to
each of several 60-ml stoppered serum bottles containing a non-nitrogen atmosphere.
Additions of acrylonitrile or other substances were then made and acrylonitrile
decrease or propene and propane increase was determined gas chromatographicaily.
The results of a typical experiment are shown in Figure 2. The Y-axis shows relative
propene production after 120 minutes incubation at 33°C.
The 60-ml rubber-stoppered serum bottles contained an atmosphere of 80:20
(argomoxygen), approximately 1x10^ Azotobacter vinelandii cells in 10 ml of nitrogen-
free growth medium, and the specified concentration of acrylonitrile. A control bottle
containing A. vinelandii but no acrylonitrile.as well as a bottle containing 1.0 mM
acrylonitrile in sterile growth medium, showed no detectable propene.
These experiments demonstrated an application of the use of microbial enzyme
systems to detoxify compounds. The potential usefulness of this procedure, at the
present time, was predicated on the premise that the spill could be covered with a
plastic film after inoculation with the active bacteria. With further study it may be
possible to secure strains that produce an enzyme with greater preference for the
topic compound than for the natural substrate, but such study was outside the scope of
this project.
The screening tests demonstrated that acrylonitrile could be decomposed
biochemically, but the conditions for using this countermeasure in the field would be
quite restrictive.
Benzene
Seven screening tests were conducted on benzene to determine the practicality
of its removal under spill conditions. Previous work cited in the literature had shown
benzene to be biodegradable only if presented to the organisms in the vapor phase. It
was not known whether a more dilute solution could be decomposed than those used in
the screening tests.
The screening tests of August 16-17 and 20-21, 1973 were carried out with the
addition of glucose and glutamic acid to benzene concentrations of 400 and 800 mg/1.
Detection of benzene through TOC measurement was not consistent because of
masking by the glucose and glutamic acid. No benzene was detected on the gas
chromatograph. Subsequent screening tests on November 1-2, 5-6, 7-8, and 19-20, 1973
using benzene at 100, 200, 400, and 800 mg/1 without glucose and glutamic acid were
carried out. In three of the four tests, methanol (500 mg/1) was added to keep the
benzene in solution. In all these tests, benzene could not be detected on the TOC or
the gas chromatograph instrument.
While it was known that benzene was highly volatile, the rate of volatilization
from a reactor such as was used in these tests was thought to be low enough to be
measured. Yang (1968), for example, measured air stripping rates for several
chemicals and obtained the results shown in Table 4, assuming first order decay. For
49
-------
1.0
- O
•a
o
cu
c
-------
benzene, the rate of 1.71 d~l would allow 18 mg/1 of benzene to remain after one day
when the initial concentration of benzene was 100mg/l. A final screening test was
performed to determine the volatilization rate of benzene from a 200 ml volume
contained in a 30 cm2 surface area with several rates of stirring and aeration. An
initial concentration of 1,000 mg/1 was used.
TABLE 4. STRIPPING CONSTANTS FOR SELECTED CHEMICALS,
ASSUMING C=C0 exp ( Ktft
Material
Nitrobenzene
Methanol
Ethanol
Refinery waste A
Refinery waste B
Benzene
Monochlorobenzene
Aniline
K(d-l)
0.843
0.263
0.302
0.332
0.345
1.71
0.969
0.198
C(mg/l)
250
1,360
2,220
475
782
100
100
100
The results are shown in Table 5. It is apparent that with aeration, benzene will
volatilize rapidly from the reactor and from sample vessels unless special precautions
are taken.
It was concluded that the high volatility rate would restrict the use of a
biological countermeasure for benzene. The engineering feasibility of covering a spill
with an impermeable material and decomposing the spill material from the vapor
phase was not evaluated, but is a possibility.
51
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Table 5. Volatilization of Benzene in
Stirred and Unstirred Containers
Removal rate
Mixing conditions
Aerated
Stirred-large vortex
Stirred-small vortex
Stirred-no vortex
Unstirred
mgBr/dl
(mgB) (cm2)
23.*
8.7
3.3
1.2
0.22
Tl/22
(hr)
0.71
1.9
5.0
13.*
75.0
1 Mg benezene removed (Br) per hour per mg benezene
(B) per cm2 of surface area at 23°C.
2 Time for 50% decrease in concentration at 23°C.
Isoprene
Four screening tests were performed with isoprene on December 11-12 and 12-13,
1973 and on November 2 and 17, 197*. The first two tests were designed to evaluate
decomposition and the last two volatilization. Because of Isoprene's low density
(0.6810 at 20°C) and insolubility in water, it was not anticipated that the decomposi-
tion studies would be successful and they were not. The volatilization studies were
qualitative in nature because of the tendency of isoprene to polymerize and resist
measurement. Observations of the thickness of an isoprene layer on water over time
and of the presence of isoprene by smell were used to show that, with aeration,
isoprene vaporized within two hours after tests were initiated.
Methyl alcohol
Methyl alcohol (or methanol) was used in six screening tests on August 13-16 and
22-23, September *-6, and October 29 to November 1, 1973, and on August 21-23 and
29-31, 197*. In the August 1973 tests, methanol plus glucose and glutamic acid were
used. Initial methanol concentrations were 1,000, 2,000, 5,000, and 10,000 mg/1. The
results showed that methanol was removed at an average rate of 0.51 d"1, assuming
first-order kinetics, or 0.0079 mg methanol removed per day per mg mixed liquor
suspended solids (MLSS). It was assumed that the glucose and glutamic acid utilization
did not interfere or compete with methanol decomposition.
A subsequent study was performed in September 197* to determine whether
glucose and glutamic acid were needed. The results showed that the first order
removal rate for 5,000 mg/1 methanol was 0.69 d-1 without glucose and glutamic acid,
compared to 0.60 d"1 with glucose and glutamic acid. It was clear from this study that
sludge from a local treatment plant could be used to decompose methanol with only
mineral salts added.
52
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Another screening test was designed to develop initial data on growth kinetics
from batch tests. A 72-hour batch test was performed using initial concentrations of
1,000, 2,000, 5,000, and 10,000 mg/1 methanol and activated sludge from the local
Govalle Wastewater Treatment Plant. The sludge had been acclimated over a two-day
period so that methanol was the sole carbon source. Using the calculation methods of
Ford and Eckenfelder (1970), the removal rate, k, was found to be 0.005 mg methanol
removed per day per mg MLSS. The microorganism yield was estimated to be 0.45 mg
MLSS produced per mg methanol removed, while the decay rate, b, was found to be
very close to zero.
The last two screening tests, carried out in August 1974, were designed to
provide information for continuous-flow bioassays, which will be discussed in Section
8. These screening tests were "fill-and-draw" batch tests in which the reactor was
filled with a mixture of activated sludge, mineral salts, water, and the desired
concentration of substrate. After a specified period, a portion of the reactor contents
(especially the MLSS) was removed and a fresh solution of test chemical,mineral salts,
and water was added to bring the mineral salts and chemical concentrations back to
their original levels. After operating for some time, the MLSS concentrations
stabilized for the cell residence times produced in each reactor, as they would in a
continuousflow reactor, and better estimates of removal efficiency, cell yield, and
decay rates could be obtained. These tests yielded the desirable information for the
continuous-flow tests, the results of which will be presented in Section 8.
At least five distinct cultures of bacteria were isolated on methanol agar from
a methanol-adapted sludge. The cultures undoubtedly included the classical
Methanomonas methanica (an organism that grows only on methane or methanol), as
well as a Pseudomonas species, which has less restricted substrate capabilities.
Nitrophenol
Three chemical forms of nitrophenol were tested: o-nitrophenol (or 2-nitro-
phenol, melting point 44.9°C), m-nitrophenol (or 3-nitrophenol, melting point 97°C),
and p-nitrophenol (or 4-nitrophenol, melting point 114°C). Each form is only slightly
soluble in water and imparts a distinctly yellow color to the solution.
The initial screening on each of the three nitrophenol isomers was carried out
on July 30-31, August 1-2, 2-3, 7-10, and 22-23, 1974. Twenty three tests were
performed with glucose and glutamic acid. Separating the decomposition of nitro-
phenol from the glucose and glutamic acid was difficult, although there was evidence
that decomposition of the nitrophenol was occurring. Subsequent screening tests on
September 25-26, 26-27, and 27-28, 1973 were performed to elucidate the importance
of the initial form of the nitrophenol -solid or liquid. Spills of nitrophenol in the solid
form would require time for dissolution before decomposition could occur. Again,
glucose and glutamic acid masked the results. The last four screening tests were
conducted on nitrophenol on January 3-7 and 7-10, May 22-30 and June 11 to July 8,
1974. The first two of these tests used a pure culture of bacteria isolated by the
microbiological group at the University of Texas. p-Nitrophenol concentration was
determined by a correlation between concentration and light absorption (see Figure 3).
It had been noted in earlier tests that the yellow color present in each reactor at the
start of the tests would persist for some time then suddenly disappear. Disappearance
of the color was considered to be tentative evidence of p-nitrophenol decomposition.
The later of the two screening tests used acclimated activated sludge and followed the
disappearance of p-nitrophenol by TOC measurements. An example of these results is
53
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a.
H
o
•a
10 12 14 16 18 20
Concentration (mg/1)
22 24 26 28
30
Figure 3. Relationship between p-nitrophenol
concentration and light adsorption.
-------
shown in Figure 4, in which the decrease in TOC after an initial 24- to 30-hour period,
as well as the time at which the yellow color disappeared, is shown. Apparently, more
time was required for the bacteria to attack the nitrophenol molecule than to attack
the color producing group (probably the nitrogen group).
Further studies by the microbiology group showed that:
1 Two pure bacterial cultures could be isolated that decompose nitrophenol.
From cytological evidence these appear to be a Micrococcus and a Pseudomonas. Both
organisms decomposed p-nitrophenol in concentrations up to 160 ppm within 48 hours.
Higher concentrations were attacked, but the time for decomposition was prolonged.
2. The optimum pH for bacterial activity was 6.8 to 7.6. Ammonia and mineral
salts had to be added to the medium and at least 10 ppm glucose and 10 ppm yeast
extract were needed.
3. The carbon in p-nitrophenol was released as CO2- This was demonstrated by
synthesizing p-nitrophenol uniformly labeled with radioactive carbon and by trapping
the CO2 that resulted from the bacterial decomposition of the radioactive compound.
Measurement of the radioactivity showed that the carbon in p-nitrophenol was
converted almost quantitatively to CO2.
4. The fate of the nitro group on the p-nitrophenol was not determined, but it
was not used to satisfy the nitrogen requirement of the organism. Other workers have
shown that when pure cultures are used, the nitrogen appears in the medium as nitrite.
5. Initial studies indicated that the nitro group is cleaved from the aromatic
ring and is not immediately reduced. The phenol that remains is oxidized to catechol
by some cultures.
Nonyl phenol
Because of the chemical nature of nonyl phenol (water insoluble, adhesive), it
was difficult to perform screening tests in batch reactors. The three tests performed
on August 7-8, 9-10, and 22-23, 1973 did not yield conclusive evidence of decomposi-
tion. However, the microbiology group demonstrated that nonyi phenol was readily
degraded by bacteria that line up on the medium nonyl phenol interface, disperse the
nonyl phenol in smaller droplets, and eventually consume it. A mixed population was
observed, but pure cultures were not isolated.
Phenol
Anticipating that phenol might be used to demonstrate spill control through
biological countermeasures in pilot-scale studies later in the project, a substantial
number of screening tests were performed.
The earliest screening tests on phenol were carried out on July 10-18, 23-24, 24-
25, 25-26, and 26-27, 1973 using concentrations of 600 and 800 mg/1 phenol (460 and
626 mg/1 as TOC). The results, shown in Figures 5 and 6, show that over 90 percent
removal was obtained in every case within 24 hours. Subsequent tests on August 16-17
and 22-26, 1973 were designed to show the importance of acclimation and the ability of
pure cultures of phenol-decomposing microorganisms to degrade phenol. The pure
55
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io°-
Q.
O
10
"260 mg/1 MLVSS
• 80 mg/1 MLVSS
- 30 mg/1 MLVSS
- 50 mg/1 MLVSS
' 40 mg/1 MLVSS
-Disappearance of yellow color
_,
10
20
30
40
50
Time (hr)
Figure 4. p-Nitrophenol intensive sampling - July 8-10, 1974.
56
-------
600
60
O
o
H
500
400
300
200
100
x:
600 mg/liter phenol
° July 23-24
A July 24-25
n July 25-26
X July 26-27
10
Time (hr)
15
20
25
Figure 5. Degradation of phenol measured as total organic
carbon.
57
-------
600
500
c
o
S-
u
o
c
-------
cultures of bacteria and yeastlike materials had been developed by the microbiology
group of the University of Texas.
The results of the later tests are depicted in Figures 7 and 8. Both of the tests
showed that an activated sludge developed in the laboratory consistently gave the best
results. The experiment depicted in Figure 7, which like the previous runs utilized
glucose and glutamic acid, showed that these two nutrients apparently did not make a
suitable source of carbon for the yeast-like material provided by the microbiology
group. The test depicted in Figure 8 was different in two respects. First, it was run
for three days instead of one day. Secondly, the glucose and glutamic acid were
withheld and the chambers with the yeast extract were closed . Under these
circumstances the yeast culture, the bacterial culture, and the un-treated sludge took
approximately three days to decompose the phenol. The treated sludge in the presence
of glucose and glutamic acid reduced the phenol in five hours, but it took approxi-
mately 24 hours to accomplish the same removal in the absence of glucose and
glutamic acid. The extremely low value for Figure 8 was a result of the high
suspended solids level (greater than 2,000 mg/1).
Subsequent screening tests were carried out to : (1) show the importance of
supplemental nutrients (glucose and glutamic acid) (August 28-30, 1973); supplemental
nutrients apparently did not aid removal, (2) evaluate the ability of bacteria to attack
high phenol concentrations (September 4-6, 1973); greater than 94% phenol removal
was observed at initial concentrations of 900, 1,200, 1,500, and 2,000 mg/1 in one day,
(3) obtain a more precise determination of the effects of additional nutrients
(September 18-19, 19-20, and 20-21, 1973), and (4) determine the effects of micro-
organism inoculum size (January 15-16 and 22-27 and March 5-6 and 21-22, 1974);
increasing the inoculum by decreasing the phenol mass to microorganism mass ratio
increased the rate of phenol removal as shown in Figure 9. These studies provided
valuable guidance for the more detailed studies subsequently carried out.
Organisms identified in the cultures used to decompose phenol were generally
of the genus Pseudomonas.
Styrene
Screening tests were performed on August 20-21 and 21-22, 1973 using concen-
trations ranging from 100 to 1,000 mg/1 of styrene. The results of these tests indicated
that little, if any, decomposition of styrene occurred. Later tests on October 2, 1974
showed that styrene volatilized within a few hours and, like benzene and isoprene, may
not be a suitable candidate for the biological countermeasure.
Toxaphene
While a number of tests were performed on toxaphene on November 5-6, 7-8,
12-13, 19-20, and 20-21, 1973, gas chromotographic analysis of samples taken from the
reactors could not detect toxaphene. Through TOC analysis, it was not possible to
detect decomposition above the level due to addition of glucose and glutamic acid.
Apparently, the toxaphene was adsorbed onto the activated sludge solids and did not
appear in the filtered samples that were analyzed. These screening tests were
generally inconclusive regarding toxaphene decomposition.
59
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600
500 -
400 -
300
CT>
E
o
en
s-
n3
O
O
(O
cn
s-
o
rtj
£ 200
Untreated yeast culture
*•
100 -
Yeast culture
600 mg/1 phenol
Aug. 16-17, 1973
O Treated sludge
A Untreated sludge
D Untreated yeast culture
x Yeast culture
Treated sludge
V j
10 15
Time (hr)
20
25
Figure 7. Degradation of phenol using various inoculums.
60
-------
600
o
.a
fO
o
o
ea
o
id
-(->
o
500
C- 400 -
300
200
100 ~
20
600 mg/liter phenol
Aug. 23-26, 1973
O Treated sludge
A Yeast culture
D Bacteriological culture
X Untreated sludge
east culture
n
Untreated sludge
Bacteriological culture
40
Time (hr)
60
80
Figure 8. Degradation of phenol using various inoculums.
61
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iiuod >fB3jq oq.
s
Q I
f -
0\
-0
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62
-------
Xylene
Results of the seven screening tests conducted on ortho-, meta-, and para-
xylene on July 18-20, 23-24, 24-25, and 25-26, and on September 13-14, 17-18, and 18-19,
1973 indicated that xylene volatilized rapidly. As with other chemicals tested, the
results for the early screening tests for xylene were confused because of the added
glucose and glutamic acid. The later tests, which confirmed volatilization, also
showed that xylene may not be amenable to the biological countermeasure unless
special precautions can be taken to keep it in solution.
Summary
At the onset of studying this series of chemicals, it was decided that their
behavior would be sought as a contaminant in domestic sewage. Accordingly, their
behavior with time was defined in terms of the TOC content of a volume of liquid in
the batch aerator. Of the 11 different compounds studied, phenol, methyl alcohol, and
nitrophenol responded best in terms of removal. It was also shown that acrylonitrile,
acetone cyanohydrin, and nonyl phenol could be decomposed. Isoprene, benzene,
styrene, and the three isomers of xylene, because of their volatility, may not be
candidates for biological countermeasures unless the chemical can be contained in the
gaseous phase or kept in solution. Aldrin and toxaphene could not be shown to be
decomposable in these tests. Of the contract list of chemicals, those not tested were
the chemicals listed as groups of chemicals, for example, cyclic insecticides.
63
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SECTION 8
BIOLOGICAL COUNTER/MEASURE TREATMENT SYSTEM
Following the screening tests, the three chemicals that appeared to be most
amenable to biological counter measures (phenol, methanol, and p-nitrophenol) were
subjected to more intense studies designed to delineate growth kinetic coefficients,
the effects of environmental variables, and other information necessary to conduct
pilot scale countermeasure tests. The methods used and the results obtained are
described in this section.
SUBSTRATE REMOVAL AND BACTERIAL GROWTH KINETICS
The substrate removal rate by microorganisms is frequently approximated by the
following expression (Metcalf and Eddy) 1972, which is similar to the one developed by
Monod (1949) to describe the relationship between the concentration of a limiting
nutrient and the concentration of enzyme; that is:
dS = - kSX
dt Ks+S
where: S = concentration of substrate
t = time (T)
k = substrate utilization rate coefficient (1/T),
X = concentration of microbial mass (M/L3), and
Ks= substrate concentration at which the substrate removal
rate is one half of the maximum rate (Michaelis-Menten
constant) (M/L3).
McCarty (1964) and Servizi and Bogan (1963) pointed out that the ratio of the cell
mass produced to the free energy released by the oxidation of substance was almost
constant and that chemical oxygen demand (COD) or biochemical oxygen demand
(BOD) had a rough linear relationship with free energy. The following approximate
relationship between bacterial growth and substrate utilization is commonly used
(Stanier et al., 1970; Heukelekian et al., 1951):
dX = -adS - k.X
dT dt d
64
-------
where: a = growth yield coefficient (biomass produced (M/L3)/substrate utilized
(M/L3) and
kj = cell decay coefficient (1/T).
Factors Affecting Microbial Activity
Temperature--
The modified Arrhenius equation is widely used to describe the temperature
effect on the substrate removal rate (Eckenfelder, 1967; Carpenter et_al., 1968); that
is:
where: kji j 2 = substrate removal rate coefficient at temperature Tl or
T2°C and •&= temperature coefficient.
Reported values for the temperature coefficient, -fy vary widely starting from 1.0
(Eckenfelder, 1967; Carpenter et al., 1968; Wuhrmann, 1955; Eckenfelder and O'Connor,
1961; Howland, 1958; Schroepfer et al., 1960; Eckenfelder, 1966; Phelps, 1944; Zanoni,
1969; Novak, 1974), depending on the substance, temperature range, type of treatment
facility, and the procedure for evaluation of the reaction rate coefficient. The
variation of £ with temperature and substance concentration raised questions about
the validity of Equation 16 for representing the effect of temperature on microbial
activity (Zanoni, 1969; Novak, 1974).
pH and Salinity —
The pH of the internal environment of all living cells is believed to be
approximately neutral and most organisms cannot tolerate pH levels below 1.0 or above
9.5. At low or high pH, acids (which tend to exist in undissociated forms) can
penetrate into cells more easily because electrostatic forces cannot prevent them
from entering. The permeated substances can upset the internal pH balance. As pH
deviates from neutral, bacterial activity decreases (Metcalf and Eddy, 1972; Stanier et
al., 1970; McKinney 1962; Randall et al., 1972).
Bacterial cells maintain an internal osmotic pressure equal to about a 0.85%
solution of NaCl. When the environment has a lower osmotic pressure than the cell
(hypotonic), water tends to permeate into the cell. Higher extracellular osmotic
pressures (hypertonic) cause the contraction of the protoplasm as a result of water loss
through the semi-permeable cell wall. A hypotonic environment (fresh water) is the
normal condition for most bacteria and they tend to exist in a distended form,
maintaining their shape within the cell wall. Bacteria can grow in media with salt
concentrations ranging from less than 0.1% to about 10%, but their activities are
impaired with increasing salinity (Stainer et al., 1970; Burnett, 1975).
Nutrients —
When abundant nutrients are supplied, bacterial growth results in a constant
chemical composition. Bacteria consist of approximately 80% water and 20% dry
matter, the latter being approximately 90% organic and 10% inorganic. The approxi-
mate formulation of the organic fraction is C5H7N02, while the inorganic fraction is
65
-------
approximately: 50% P205, 6% K20, 11% Na20, 8% MgO, 9% CaO, 15% S03, and 1%
Fe203. It is generally agreed that nutritionally balanced wastes result in unrestricted
bacterial growth and rapid decomposition of pollutants.
Helmers et al., (1951, 1952) reported that for successful removal of waste by the
activated sludge process, nitrogen and phosphorus must be supplied at the proportion
of BOD^ : N : P = 100 : 5 : 1. Sawyer (1955) observed that the nitrogen requirement
decreased with increasing biomass as a result of decreased cell production.
Wilkinson (1958) studied the influence of nitrogen, phosphorus, and sulfur on the
production of polysaccharides by Klebsiella aerogenes. When the concentrations of
these nutrients were lowered until they became growth limiting, the amount of
polysaccharides produced per cell increased to a maximum level. At this point, the
polysaccharide to nitrogen, phosphorus, and sulfur ratios were about 32, 40, and 17,
respectively.
Symons and McKinney (J958) reported that the conventionaJ activated siudge
system functioned satisfactorily even in the absence of nitrogen for three to four
weeks. More solids were produced in the absence of nitrogen and the solids produced
were high in polysaccharides. The authors concluded that partial satisfaction of the
nitrogen requirement could result in a stable system.
Gaudy and Engelbrecht (1960, I960 reported that regardless of the presence of
extracellular nitrogen, the organic load was removed at the same rate suggesting the
possibility of an amino acid pool within the bacterial cell.
Despite the fact that trace elements are required in minute amounts, they
exhibit a pronounced effect on bacterial activity. Mg++, Mn++, Fe++, Ca++, etc., act
as cofactors for respiratory processes (Stanier et al., 1970; Oginsky and Umbreit, 1954).
Oxygen Requirement --
In the aerobic process of organic substance stabilization, the molecular form of
oxygen is the only final hydrogen acceptor. Thus, oxygen demand is a direct function
of biological metabolism and the oxygen requirement is directly related to the amount
of organic matter decomposed and rate of endogenous respiration. The oxygen
utilization rate can be formulated as follows (Eckenfelder and O'Connor, 1961):
Rr = - a' dj> + b1 X, (17)
dt
where: Rr = oxygen utilization rate ((mg/1 02)/time),
a1 = oxygen required per substrate utilized, and
b1 = oxygen required per biomass for endogeneous respiration
02)/(mg/l biomass)/time).
EXPERIMENTAL METHODS
Equipment and Reagents
The analytical instruments used for these tests included an Expanded-Scale pH
Meter from Beckman, a Salinity Refractometer (automatic temperature-compensated)
from American Optical Co., a Galvanic Cell Oxygen Analyzer from Precision
66
-------
Scientific, a 15-liter-per-minute capacity air flow meter from Gelman Instrument Co.,
and a Beckman Model 915 Total Organic Carbon Analyzer. Aerated batch reactors
made of plexiglass units 8.75 cm wide, 12 cm long, and 32 cm high were graduated at
25-ml intervals so that water loss by evaporation could be easily compensated. The
reactor volume was either 2 or 3 liters.
The basic substrates used were liquified phenol, methanoi, and glucose.
Nutrient Solution I was prepared as a nitrogen and phosphorus source and as a pH
buffer for phenol- and methanol-acclimated activated sludge and consisted of: 320 g/1
K2HPO^, 160 g/1 KH2PO^, and 120 g/1 NrtyCl.
Nutrient Solution II was prepared as a nitrogen and phosphorus source for phenol-
acclimated activated sludge and consisted of: 120 g/1 NH^Cl and 29 g/1 KH2PO^.
Nutrient Solution III was prepared as a nitrogen and phosphorus source for
methanol-acclimated activated sludge and consisted of: 91 g/1 NH^Cl and 21 g/1
A mineral solution to supply trace elements for phenol- and methanol-acclimated
activated sludge was made of: 15 g/1 MgSO^.Jh^O, 0.5 g/1 FeSO^.ZI-^O, 0.5 g/1
ZnS0^.7H20, 0.4 g/1 MnSO^.H20, and 2 g/1 CaCl2.
A silver nitrate solution (1,000 mg/1 as Ag) was used as an enzyme inhibitor.
For pH adjustment, NaHCC>3 solution (100 g/1), hydrochloric acid (1 N), or sodium
hydroxide solution (1 and 6 N) was added and for salinity adjustment, "Synthetic Sea
Salts" from Aquarium System, Inc., Eastlake, Ohio, was used.
For the determination of hardness, a buffer solution of 4.716 g/1 disodium salt of
EDTA, 3.12 g/1 MgSO^.7H20, 67.6 g/1 NrtyCl, and 572 ml/1 NrtyOH (cone.) was made.
The inhibitor was 50 g/1 NaS.9H20, the dry powder indicator was 0.5 g Eriochrome
Black T and 100 g NaCl, and the titrant was a 0.01 M EDTA solution.
All chemicals used were reagent grade.
Acclimation and Feeding of Activated Sludge
Mixed liquor from the Govalle domestic wastewater treatment plant in Austin,
Texas was used as seed. Supernatant from the mixed liquor was decanted after sludge
settlement, and substrate and nutrients were fed to the sludge. Tap water was used as
dilution water. After one to three days of aeration, ten percent of the total mixed
liquor was wasted. The above procedure was repeated throughout the test period. Ten
ml each of Nutrient Solution I and the mineral solution were fed each time. The
substrate feeding schedule for phenol acclimated sludge was:
first day .......... phenol 0.1 ml/1, glucose 940 mg/1
second day ......... phenol 0.4 ml/1, glucose 750 mg/1
third day .......... phenol 0.8 ml/1, glucose 500 mg/1
fourth day ......... phenol 1.2 ml/1, glucose 250 mg/1
fifth day to
completion ......... phenol 1.6 ml/1 (1,200 mg/1 TOC).
67
-------
The feeding schedule for methanol-acclimated sludge was:
first day methanol 2.0 ml/1, glucose 940 mg/1
second day methanol 3.5 ml/1, glucose 750 mg/1
fourth day methanol 5.0 ml/1, glucose 500 mg/1
seventh day to
completion methanol 7.0 ml/1 (2,000 mg/1 TOG).
After the initial acclimation stage, phenol sludge was fed every day and
methanol sludge was fed once every three days. The sludge was kept in the
temperature range from 17° to 21°C.
The biological treatability and countermeasure application tests were done after
at least three weeks of acclimation, when the sludge production reached a constant
level. Sludge production was checked by measuring total suspended solids (TSS) and
volatile suspended solids (VSS) before and after feeding the sludge. At steady state,
the ratio of VSS to TSS ranged from 85% to 92%.
Substrate Removal and Bacterial Growth Kinetic Study
Substrate removal and bacterial growth kinetics tests were conducted at
temperatures of 5°, 20°, and 28°C. For the tests at 5° and 28°C, the sludge cultures
were acclimated to the new temperatures by placing the sludge at the respective
temperature until 1,200 mg/1 of phenol or 2,000 mg/1 of methanol, as TOC, was
completely removed. One day was adequate for the phenol sludge to remove the
substrate. For the methanol sludge, three days were required to decompose the
substrate completely at 28°C, while one week was required at 5°C.
The acclimated sludge cultures were elutriated with distilled water several times
to reduce the residual organic carbon contents. The washed sludge was transferred to
2-liter reactors and fed with the phenol or methanol substrate and 10 ml/1 each of
Nutrient Solution I and mineral solution. The reactors were filled with distilled water
up to the 2-liter mark and aerated.
The initial concentrations of VSS and TOC for the phenol study ranged from 300
to 4,000 mg/1 and from 160 to 800 mg/1, respectively, and those for the methanol study
ranged from 200 to 1,600 mg/1 and from 500 to 1,000 mg/1, respectively. To calibrate
the methanol loss by stripping, one reactor was not supplied with sludge in methanol
tests at each temperature. Twenty-five ml of mixed liquor samples were withdrawn at
certain time intervals after aeration for the determination of MLVSS and filtrate
TOC. One drop of silver nitrate solution was added to each of the samples during
filtration to inhibit further enzymatic activity and to reduce time measurement error
caused by duration of filtration. For the determination of TOC, total carbon and
inorganic carbon were analyzed. The residual TOC contributed by the mixed liquor,
excluding substrate, was measured and subtracted from the sample TOC. The reactors
were graduated and water loss by evaporation was made up with distilled water. pH
and temperature were measured throughout the reaction periods. Ail other procedures
were performed according to Standard Methods (APHA et_al., 1970).
pH and Salinity Effects Study
Tests to evaluate the effects of pH and salinity on phenol and methanol removal
were conducted at all combinations of temperature (5°, 20°, and 28°C), pH (5,6,7,8,
68
-------
and 9), and salinity (0, 10, and 35 ppt). Two additional salinities (5 and 20 ppt) were
tested at 5°C to determine their effects on phenol removal. Salinity measurements
refer to the dilution water salinities.
The activated sludge was prepared in the same manner described in the previous
section and was acclimated to each test temperature but not to the various pH levels
and salinities. The phenol sludge was fed with 5 ml/1 of Nutrient Solution n, ten ml/1
of mineral solution, and 720 mg/1 of phenol as TOC. The methanol sludge was fed with
10 ml/1 each of Nutrient Solution III and mineral solution, and 1,050 mg/1 of methanol
as TOC. The initial MLVSS ranged from 1,000 to 1,500 mg/1 and from 450 to 700 mg/1
for phenol and methanol studies, respectively. At a given temperature, the initial
MLVSS was equalized for all pH and salinity conditions. All sampling and analytical
procedures were the same as described in the previous section. Temperature and pH
were measured and pH was adjusted frequently throughout the reaction periods.
Nutrient Studies
Nutrient studies were conducted at room temperature, which was about 20°C.
The nutrient requirements were estimated in the following manner.
From the kinetic study tests, it was concluded that approximately 65% of the
substrate utilized resulted in cell synthesis. The carbon to nitrogen to phosphorus
ratio of bacterial cells is approximately 100 : 23 : 4.6 (Novak, 1974). Thus, 720 mg/1 of
phenol as TOC feed will require 108 mg/1 (720 x 0.65 x 0.23) of nitrogen and 22 mg/1
(720 x 0.65 x 0.046) of phosphorus. In order to prevent nitrogen and phosphorus from
becoming growth limiting, a 50% excess supply (162 mg/1 of nitrogen and 33 mg/1 of
phosphorus) was considered to be adequate in the phenol tests. Similarly, 236 mg/1 of
nitrogen and 47 mg/1 of phosphorus were adequate for 1,050 mg/1 of methanol as TOC.
The effects of limiting nutrients were studied using concentrations of nitrogen and
phosphorus equal to, one-half, or one-fourth of the concentrations determined above,
as well as a zero concentration. Other conditions included tests with and without
minerals and tests with and without pH adjustment to the neutral range. For mineral
supply studies, ten ml/1 of the mineral solution was fed to the sludge. pH was adjusted
using bicarbonate.
To observe substrate removal in natural systems, the activated sludge was fed
720 mg/1 of phenol or 1,050 mg/1 of methanol as TOC, using distilled water, tap water,
ground water, and synthetic sea water as dilution water, without any other chemical
aids.
Sampling and analytical procedures were the same as those described in the
previous section. pH and temperature were measured periodically throughout the
reaction periods.
Oxygen Requirement
In addition to all the procedures described above, oxygen consumption rates (as
mg/1 02/min) were measured. BOD bottles (300 ml) were filled with mixed liquor
samples and dissolved oxygen changes with time were measured using an oxygen probe,
which was designed to seal the BOD bottle during the measurement.
69
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TEST RESULTS
Substrate removal and bacterial growth kinetics theory is well established and
numerous papers in the literature describe how the kinetic coefficients can be
evaluated using linear graphical methods. However, there appears to be a dearth of
information on kinetic coefficients for specific substances. Moreover, use of the
linear graphical methods leads one to wonder how reliable the reported evaluations
are.
The shortcomings of the linear graphical methods are reviewed briefly and
statistical methods are developed for evaluation of kinetic coefficients. Using the
statistical methods, pH, salinity, temperature, and nutrient effects on substrate
removal and bacterial growth kinetics are evaluated numerically and confidence
intervals are given for the more important parameters evaluated. Oxygen require-
ments are evaluated based on the substrate removal and bacterial growth kinetics
study and checked against the experimental results, because the linear graphical
method is apt to produce biased results. The reliability of substrate removal and
bacterial growth kinetics data are discussed in detail by comparing the experimental
results and the kinetic models.
Phenol
Kinetic Parameters —
The most widely used method for evaluating kinetic parameters (the substrate
utilization rate coefficient (k), the Michaeiis-Menten constant (KS) the cell yield
coefficient (a), and the cell decay coefficient (k^)) is a linear graphical method in
which Equations 1^ and 15 are arranged in linear forms (Eckenfelder and Ford, 1970).
Arrangement of Equation W with substitution oftS/eA for dS/dt yields:
k. (18)
p. ... X vs. L k and K are obtained from the slope and Y intercept
g AS/At S s
of the best fit straight line. Similarly, Equation 15 yields:
Then, a and kj are evaluated in the same manner as k and Kg.
The shortcomings of the above method are as follows. First, there are inevitable
experimental errors in the measurements of X (MLVSS) and S (TOC). Second, these
errors are enlarged by estimatingAX and AS. Third, the errors are further increased by
replacing dX/dt and dS/dt with &X/&t and AS/At. Last, variable transformations
(X/(As/&t), 1/S, (AX/At)/X) result in further enlargement of error and cause non-
symmetric error distributions whose variance changes with the magnitudes of the
corresponding terms. As a result, without highly accurate measurements of X and S, it
is hard to expect straight lines and the best fit straight lines do not necessarily mean
the best interpretation of the experimental results.
70
-------
Under the experimental conditions described above, the expected errors of VSS
measurements were almost constant in the VSS range of 200 to ^,000 mg/1 and were 67
mg/1 for phenol-acclimated sludge and 126 mg/1 for methanol-acclimated sludge
(estimated from Equation 28 below). Analysis of TOG was conducted in such a manner
(mainly by repeating the sample injection into the analyzer) that the errors fell within
10 mg/1. With these errors, the plots of data according to the above equation were
spread widely and there seemed to be little meaning in finding the best fit straight
lines. An example that deals with the transformation of Equation 17 to the linear form
Rr/X = - a' ((As/At)/X) + b',wm be presented in Figure 34.
The following method is developed for the evaluation of kinetic parameters from
batch test data. Rearrangement of Equation 15 yields:
dX = - a dS - kdX dt . (20)
Integration of Equation 20 from time to to t results in:
X - X0 . a(S0 - 9 - kd Xdt. (21)
If the time interval (t - to) is not big enough to allow a dramatic change jn x
during the interval, kd JfoX'dt carrbel substituted by kdX(t-t0), where X denotes the
mean biomass concentration during time to to t. Then, Equation 21 becomes:
X - XQ = a(SQ - S) - kd"x(t-t0) (21')
Under the previously described experimental conditions, the error caused by the left
hand side of the equation (X-XO) is much greater than that by the right hand side of
the equation. Thus, the terms in the right hand side of the equation may be regarded
as exact variables. Assuming the observed values of (X-XO) are normally distributed
around the true values of (X-XO) with a variance of O 2, we obtain the likelihood
function for Equation 2l':
exp ^ <22)
where: L = likelihood,
(J= standard deviation of (X-XO),
R1= (X-X0) - a(S0 - S) - kdX"(t-t0), and (23)
N = number of observations.
The values of a and kj, which provide the maximum likelihood values, are obtained
from differentiation of Equation 22 with respect to kj and a. FromoL/<3a=0 and
C)L/3kd«0,we get:
k =
d
71
-------
and
zoc-ocso-s)- kjigg-tocso-sj (25)
Since kd and a are linear combinations of (X-XO), which is assumed to be normally
distributed, they are also normally distributed. Their variances are given by:
V
where: ^""X-X )
' ~ o o
fexlt-tt>(S0-S)l2
<27'
(X-XO)O = observed value,
(X-Xo)e = estimated value from Equation 2l', and
N = number of observations.
Combining Equations 14 and 20 and eliminating dt, we obtain:
dX = -a dS + kd(^£_)ds. (29)
Integration of Equation 29 from time to to t yields:
x-x _»(a-J_)(So-S>- mr>/s). (30)
K
72
-------
If (X-X0) previously estimated from Equation 31 is used instead of (X-XO) observed, the
variance is greatly reduced and the estimation of the coef-ficients, k and K$, is
significantly improved. The residual R2 is expressed as:
R2 = (X-XQ)e - A (SQ -S) + B ln(SQ/S), (31)
where: A = a - k )
where — ^ » e/
j% ^ 41 4
(X-Xo)e = estimated value from Equation 30.
73
-------
The variance of Ks is given by:
and the variance of kj/k is:
The 80% confidence interval of k^/k is:
* U9
(M)
Therefore, the 80% confidence interval of k is:
k 6~ U2B s or
* 1-28
, or
if
In order to reduce the error caused by the term ln(So/S) in Equation 30, the data
from the reaction stage in which endogeneous respiration prevails over other activities
are not used in this estimation. Estimated kinetic parameters using the above method
are shown in Table 6. The standard deviation and confidence interval includes
experimental errors and the variability of bacterial activities with biomass concentra-
tion, substrate concentration, and other effects. As shown in Table 6, Ks and a change
with temperature insignificantly. Thus, it may be possible to conclude that the
average values of Ks and a are the inherent characteristics of phenol waste that are
not affected by temperature. Since hypothesis tests fail to disprove this assumption
74
-------
(see Appendix I, 1 and 2), all kinetic parameters can be re-evaluated based on the
above theory.
TABLE 6. ESTIMATED KINETIC PARAMETERS FOR PHENOL
Temp
5
23
28
Parameter Expected
value
Kg1 260.2
k 2 0.02026
kd 2 0.002179
a 1.258
K l 240.9
s
k 2 0.07177
kd 2 0.006755
a 1.326
K 1 206.3
s
k 2 0.06680
kd 2 0.005920
a 1 . 048
Standard
deviation
107.9
0.000838
0.077
171.2
0.001129
0.143
55.6
0.000999
0.069
80% Confi- X range
dence (mg/1)
interval
0.00630 460
0.01671 to
4,100
0.01860 470
0.03863 to
4,100
0.027 300
0.255 to
3,000
S range N
(mg/1)
up to
65
650
up to
22
720
up to
29
790
Substrate concentrations are expressed as TOG (mg/1) throughout this
study, if not specified otherwise.
Expressed in units of hr~ .
75
-------
The mean K and a vaiues are:
K = 236 mg/l,dfe =70 mg/1 and
S 3
a = 1.21,0: =0.06
cl
Corresponding kd values are obtained fromdL/dkd = 0, where L is the function defined
in Equation 14; thus,
k - a
d
For a given value of cell yield coefficient, a, the variance of k
-------
TABLE 7. KINETIC PARAMETERS FOR PHENOL CORRESPONDING TO
K
236 mg/1 AND a = 1.21
Temperature Parameter
(°C) (hr -1)
Expected
value
Standard
deviation
80% Confidence
interval
k
k.
0.01892
0.002083 0.000591
0.01636
-------
PH
"*" Observed point
20
Salinity (ppt)
30
40
Figure 10. pH and salinity effects on the decomposition of phenol
by acclimated sludge. Iso-f lines at 5 °C. *
* k
(PH=7, sal=0 ppt, temp.=5
= °'0189
78
-------
PH
10
20 30
Salinity (ppt)
40
Figure 11. pH and salinity effects on the decomposition of phenol
by acclimated sludge Iso-f linesat 21 °C.*
* k (pH=7, sal=0 ppt, temp.=21°C)+0.0610 hr"1
79
-------
pH
8 -
7 -
5 -
10
20
Salinity (ppt)
30
Figure 12. pH and salinity effects on the decomposition of phenol
by acclimated sludge Iso-f lines at 28 C.*
*k(PH=7, sal=0 ppt, temp.=28 °C) = °'0724 hr
-1
80
-------
effects are conspicuously decreased compared to those at high temperatures. This
means that the temperature effect is decreased as pH and salinity become unfavorable
to organisms. Iso-£ lines in the pH and salinity coordinates are presented in Figure 13.
The calculated Q values ranged from 1.0145 at pH 5 and salinity 35 ppt to 1.0760 at pH 7
and salinity 0 ppt. In the pH and salinity range in which the temperature coefficient
was greater than 1.06, the substrate removal rate coefficients at 23.5°C (computed
from the modified Arrhenius equation) reached those values observed at 28°C,
meaning that the temperature reached the optimum range for the organisms. In the
range with Q-less than 1.05, the Rvalues could be used to predict all k values in the test
temperature range (5 to 28 °C). Thus, it may be concluded that as pH and salinity
become more favorable to organisms, the temperature coefficient increases and the
lower limit of the optimum temperature range is lowered.
Endogeneous Respiration —
The endogeneous respiration activity appeared to be affected by all the tested
environmental factors, including temperature, pH, and salinity. With a limited
accuracy in VSS analysis, a single batch reactor could not provide a reliable estimation
of the cell decay coefficient, k
-------
PH
20 30
Salinity (ppt)
40
Effective temperature range for &:
5.0 - 23.5 UC
5.0 - 28.0
Figure 13. Temperature coefficient, 9, for the decomposition of
phenol by acclimated sludge. Iso-6 lines.
82
-------
0.007
O)
o
O
(J
O
O)
T3
O)
0.006
0.005
0.004
0.003
0.002
0.001
i i
O
O
O
kd = 0.066 k
O
0.87
I I
I
I
I i
I I
0 0.02 0.04 0.06 0.08
Substrate removal rate coefficient, k, (hr"1)
Figure 14. The relationship between the substrate removal
rate coefficient and the cell decay coefficient
for phenol acclimated sludge.
83
-------
-a
-C
a>
-o
O>
-5
-6
-7
= 0.066 k
0.87
O
-5 -4 -3
Substrate removal rate coefficient, In k (hr"1)
Figure 15. The relationship between the substrate removal rate
coefficient and the cell decay coefficient for phenol
acclimated sludge.
84
-------
700
600
500
400
TOC
(mg/1)
300
200
100
0
0
Temperature = 20 C
Initial VSS = 1,500-2,000 mg/1
0 N & P Fully supplied
H of N & P supplied
0 h of N & P supplied
x N & P not supplied
pH not buffered
IT = 0.0166/hr
T = 0.292
pH buffered
IT = 0.0551.hr
T = 0.972
ptt = 7.85
8
246
Aeration time (hr)
Figure 16. Nutrient (N & P) effects on the decomposition of phenol
by acclimated sludge.
* Minerals supplied.
85
10
-------
700
600 _
500 -
400 -
TOC
(mg/1)
300 -
200 -
100
Temperature = 20 C
Initial VSS = 1,500-2,000 mg/i
& P fully supplied
of N & P supplied
of N & P supplied
& P not supplied
not buffered
IT = 0.0154/hr
f = 0.271
= 4.5
pH buffered
k" = 0.0530/hr
f = 0.934
pH = 7.85
0
Figure 17.
Aeration Time
o
;i '
hr.
10
Nutrient (N Si P) effects on the decomposition of
phenol by acclimated sludge- Minerals not supplied.
86
-------
TABLE 8. THE RELATIONSHIP BETWEEN SUBSTRATE REMOVAL RATE COEFFICIENT
AND CELL DECAY COEFFICIENT FOR PHENOL-ACCLIMATED SLUDGE
Range of k Number of Average k
(hr~l) observations (hr ~1)
0.00 -
0.01 -
0.02 -
0.03 -
0.04 -
0.05 -
0.06 -
0.07 -
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
5
41
8
6
5
9
2
15
0.00863
0.01649
0.02619
0.03560
0.04410
0.05364
0.06100
0.07148
Average k
(hr -I)
0.00145
0.00186
0.00541
0.00358
0.00676
0.00276
0.00582
0.00671
Nitrogen and Phosphorus Effects on Cell Synthesis and Decay —
As shown in Equation 30, the net gain of cell mass is a function of the initial
substrate concentration and the substrate concentration remaining. Therefore, if the
values of kinetic coefficients (k, Ks, a, and kj) are given, the net cell production can
be predicted from the initial substrate concentration and the substrate concentration
remaining, regardless of the initial biomass concentration.
In a normal growth pattern with nitrogen and phosphorus fully supplied, the cell
decay coefficient, k^, was found to be related to the substrate removal rate
coefficient, k, as kj = 0.066 kO-87 based on the data from 91 batch reactor tests.
These estimated values of kinetic coefficients predicted the biomass growth with an
expected error of 20 to 30 mg/1 (based on Equation 40). When the biomass growth
without nitrogen and phosphorus was compared to the normal growth, the following
aspects, which are beyond the above error range, were observed (see Figures 18 and 19).
First, there was more cell production at the initial growth phase (the biomass growth
phase at substrate concentrations above 100 mg/1 as TOC). This means that the cell
yield coefficient is larger than 1.21 without nitrogen and phosphorus.
Wilkinson (1958) and Symons and McKinney (1958) reported the same result.
Second, there was rapid cell decay at the end of the growth phase (substrate
concentrations below 100 mg/1 as TOC). It may be assumed that the cells that had
been synthesized without extra-cellular nitrogen and phosphorus are more easily
decayable than the normal cells that are synthesized with abundant nutrients. This
cell decay rate appeared to increase when minerals were not supplied and when a more
optimal pH condition was provided.
,87
-------
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-------
TABLE 10. PHENOL DECOMPOSITION BY ACCLIMATED ACTIVATED SLUDGE
IN NATURAL SYSTEMS WITHOUT CHEMICAL AIDS 1
^^-\^ Test waters
Parameter •s"\^^
Initial pH
Alkalinity2
2
Hardness
TDS (mg/1)
pH Range
Average k
f
Equivalent pH
Distilled
water
6.9
4.5-7.3
0.0154
0.271
4.5
Tap
water
9.6
45
97
380
4.6-9.6
0.0186
0.328
5.0
Ground-
water
8.6
115
206
462
5.4-8.6
0.0592
1.045
7.0
Sea
water
8.3
164
7,810
38,300
4.95-8.3
0.0232
0.409
7.0
Temperature = 20 °C; k.no,, = 0.0567 hr~
/u \j
2
Expressed as mg/1 CaCO,
3 _i
Average k (hr ) means the resultant k when approximately 90% of TOC
removal was achieved.
Phenol Decomposition in Natural Systems without Chemical Aids —
Phenol decomposition resulted in a considerable decrease in pH. When the
acclimated sludge was fed with 720 mg/1 of phenol (as TOC) with no other chemicals,
the pH dropped down from 6.9 to 4.5 in distilled water, from 9.6 to 4.6 in tap water
with an alkalinity of 45 mg/1 as CaC03, from 8.6 to 5.4 in groundwater with an
alkalinity of 115 mg/1, and from 8.3 to 4.95 in sea water with an alkalinity of 164 mg/1.
These unfavorable pH conditions lasted until nearly 90% TOC had disappeared, the
removal results were equivalent to k = 0.0154 hr~l in distilled water and k = 0.0186 hr~l
in tap water. These results are approximately the same as the results obtained at
pH=4.5 and pH = 5, respectively, in a pH-regulated system. However, in groundwater
and sea water, these temporary pH changes did not interfere with the phenol
decomposition rates. The results were equivalent to those at pH 7 (see Figure 20 and
Table 10).
Initial Lag Phase ~
Even though the sludges were consistently fed with approximately 1,200 mg/1 of
TOC, less than 720 mg/1 of TOC feeding caused initial lag phases in some cases. The
lag phase was determined in the following way.
91
-------
800
TOC
« Distilled Water
•• Tap water
4. Ground water
Sea water
PH
Q: -O
D- D
A A
tf V
10
PH
0 5 10 15 20 25 30
Aeration time (hr)
Figure 20. pH variations owing to the decomposition of phenol by
acclimated activated sludge.
92
-------
If the biomass concentration, X, is considered as a constant, X, between
sampling times, then the integration of Equation 14 from time to to time t yields:
K ln(S/S + fi
xit-to)
(55)
When the k values at initial stages were noticeably smaller than those at any
other time intervals in a given reactor, these initial stages were considered to be in a
lag phase. The above method was used to prevent experimental errors from affecting
designation of the initial lag phase. Whether this lag phase in TOC removal really
means a lag phase of bacterial activities, or is merely caused by decomposition of
organics into other intermediates, will be discussed in the methanol study section.
Eight of 115 reactors displayed initial lag phases. In an extraordinary case,
duration of the lag phase was about 32 hours, which caused a 15-hour aeration lag time.
In the other seven cases, average duration of the lag phase was about four hours and
the average aeration lag time was about three hours. To achieve a certain substrate
concentration level at which organisms have recovered from the initial lag phase, a
supplementary aeration time is required in addition to the theoretically calculated
aeration time. This supplementary aeration time is designated to be the aeration lag
time and is illustrated in Figure 21. Although the phenol-acclimated sludge, prior to
the removal experiments, never experienced high salinities and high or low pH's, these
new environments did not cause any particularly detectable lag phase. The
experimental conditions and the initial lag phases are presented in Table 11.
TABLE 11. INITIAL LAG PHASE IN PHENOL DECOMPOSITION BY ACCLIMATED SLUDGE
Type of Test
Kinetic study
pH, salinity
study
Nutrient
study
Countermeasure
application
Total
Temp.
(°C)
5
23
28
5
21
28
20
28
-
Number
of
tests
10
8
6
40
15
15
20
1
115
Cases
of lag
phase
5
1
1
1
8
Average
duration
(hrs)
5
32
2
2
8
Average
lag time
(hrs)
4
15
1
2
5
Remarks
pH=7
sal=0 ppt
pH=5 . 1
sal=0 ppt
Ground-
water
Ground-
water
-
93
-------
700
500
till
.Aeration lag time = 15 hr
J — - \
1
\ \ Temperature = 5 C •
\ \ pH
\ x Salinity
= 5,1
= O.pp.t
\ \ Initial VSS = 2,000 mg/1 _
h\
o 400
s-
-------
Oxygen Requirement—
Previously, it was pointed out that removal of one gram of phenolic carbon
resulted in production of 1.21 grams of organic solids (a = 1.21). Based on the
approximate formulation of a bacterial cell (McKinney, 1962), C^H/NC^, one gram of
carbon produces 1.88 grams of organic cellular material, if it is used for synthesis.
Therefore, 64.4% (1.21/1.88) of the phenol removed is calculated to be used for cell
synthesis and the rest is completely oxidized to carbon dioxide and water.
Complete oxidation of phenol requires 3.11 grams of oxygen per gram of carbon
according to:
= 6C02 + 3H20. (56)
(72) (224)
Complete oxidation of the cell requires 2.67 grams of oxygen per gram of cell
carbon (or 1.42 grams of oxygen per gram of organic fraction of cell) according to:
C5H7N02 + 5 02 = 5 C02 + NH3 + 2 H2O . (57)
(60) (160)
Therefore, one gram of carbon will require 0.44 grams (3.11-2.67 grams) of
oxygen when it is utilized for cell synthesis. Thus, per gram of carbon, 64.4% of TOC
removed requires 0.44 grams of oxygen and 35.6% of TOC removed requires 3.11 grams
of oxygen.
The following estimation of the coefficients, a' and b', in Equation 17 are
possible:
a1 = (0.644 x 0.44) + (0.356 x 3.11) = 1.39 and b1 = 1.42 x k ..
d
Then, the oxygen utilization rate is expressed as:
Rr = -1-39^ + 1.42kdX, (58)
where dS/dt and k^ are given in Equations 14 and 54, respectively.
The oxygen uptake rates observed are compared with those computed from
Equation 58 in Table 12. Through a hypothesis test, it may be proved that a1 = 1.39 and
b1 = 1.42 k
-------
0.14
0.12 -
0.10 -
0.08
Rr
(hr")
0.06
0.04 ~
0.02 -
Rr = - 1.39 + 0.00772 X
o = Observed point
j i
I i
0 0.02 0.04
0.06
1 t
0.08 0.10 0.12
Figure 22. Comparison of theoretical and observed oxygen uptake
rates in the decomposition of phenol by acclimated
activated sludge.
95
-------
TABLE 12. COMPARISON OF THEORETICAL AND OBSERVED OXYGEN UPTAKE
RATES IN PHENOL DECOMPOSITION BY ACCLIMATED SLUDGE l
02 Uptake rate 02 Uptake rate
observed estimated Rr^ -
Rrxdng/l/hr) Rr2(mg/l/hr)
22.5
27.6
34.1
46.1
8.7
16.7
17.3
13.8
22.8
0
15.9
16.0
43.4
50.4
6.4
15.0
16.0 '
14.6
23.5
1.1
6.6
11.6
-9.3
-4.3
2.3
1.7
1.3
-0.8
-0.7
-1.1
^•Number of observations = 10.
2Mean =0.73; Standard deviation = 5.68.
Methanol
Kinetic Parameters —
Aeration of methanol resulted in a considerable amount of methanol loss by
stripping. The rate of methanol stripping may be assumed to be a first order reaction,
expressed as:
where ke is the stripping rate coefficient (time~l). (Further discussion of methanol
stripping is contained in Appendix II). When methanol removal is accomplished by both
stripping and biological decomposition, the removal rate is expressed as:
(60,
and the bacterial growth rate is:
f = -a(f t+ keS) - kdX . (61)
Integration of Equation 61 from time t to time t yields:
X-XQ = a(SQ - S) - a ke [^ S dt - k £ X dt. (62)
*° o
97
-------
When the time interval (t-to) is too short for large changes to occur in S and X, Jt0 S dt
and J£ X dt can be replaced by s(t-to) and X(t-to), respectively, where "5 and X denote
the mean concentrations of S and X between times to and t, and the following equation
results:
( - ^ _
X-XQ = a <(SQ-S) - keS(t-t0)j - kdX(t-t0). (62 )
Equation 62* is the same as Equation 21* except that (So -S) is replaced by|(s _s) _ [
9£- = - a dU + -^| - dU (66)
dU + jA ( dS - k dt) (67)
~
k e
Integration of Eq. 67 from time to to t yields:
X-Xn = (a- d) (U -U) -iln(S , /S) -k (t-t ) (68)
o k o k I ° e o
In a short time interval, UO-U can be expressed as:
Uo-U = (So-S)-keS(t-tQ), (69)
where (Uo - U) is the amount of substrate removed by biological decomposition. Then,
Equation 68 becomes:
X-XQ = (a-lfc) {(SQ -S) -ke S (t-t0)} - ln(SQ/S) -ke(t-tQ) (691)
Equations 34 through 39, with substitutions of (So - S) and Infect by
{(SQ - S) - kj (t-t0)} and {In(So/S) - ke(t-tQ)}
provide the solution for the expected values and variances of k and Ks.
When stripping is the only cause of methanol removal, the concentration at a
given time is obtained by integrating Equation 59, which gives
S = Soexp(-ket).9g (70)
-------
Plotting In S as the ordinate and t as the abscissa, ke is obtained from the slope of the
straight line.
When 1 liter of air per minute was supplied per liter of water, the stripping rate
coefficients were estimated to be 0.00330, 0.00948, and 0.0277 hr -1 at 5, 22, and 28
°C, respectively. At these stripping conditions, the estimated kinetic parameters are
shown in Table 13.
TABLE 13. ESTIMATED KINETIC PARAMETERS FOR METHANOL
Temp.
5
22
28
Parameter
K
s
k
kd
a
K
s
k
kd
a
K
s
k
l
2
2
1
2
,2
1
2
Expected
value
2,587
0.03754
-0.001981
1.013
-696
-0.05376
0.0038338
1.279
2,080
0.1902
Standard
deviation
2,747
0.007850
1.313
1,025
0.003074
0.158
674
80% Confidence
interval
^0.00.09
60.00116
*0. 00815
<-0. 01169
2-0.0162
£0.0195
X Range
(mg/1)
200
to
1,300
450
to
1,400
200
to
S Range N
(mg/1)
up to
1,000 21
up to
920 8
up to
1,150 40
0.006758 0.005910
1,600
1.444
0.291
Expressed in units of mg/1.
n
Expressed in units of hr~l.
It is likely that the distribution of the estimated parameters was spread widely
because of inconsistent methanol stripping conditions from reactor to reactor. At 22
°C, k and Ks were estimated to be negative; however, negative values for k and Ks are
contradictory in Equation 13 and it is believed that not enough observations and some
unfavorable experimental errors are the causes. Excluding these erroneous estima-
tions, the kinetic coefficients, Ks and a, changed with temperature insignificantly.
Thus, all kinetic parameters can be re-evaluated based on the assumption that the
average values of Ks and a are the inherent characteristics of methanol that are not
99
-------
affected by temperature (cf. Appendix I, 1 and 2). The mean Ks and a that
characterize methanol wastes are: Ks = 2,330 mg/1, Jks = 1,410 mg/1 and a = 1.25,
da=0.45. Re-estimated kinetic parameters corresponding to Ks = 2,330 mg/1 and a =
1.25, using equations 47 through 51 with the previously described substitutions, are
given in Table 14.
TABLE 14. KINETIC PARAMETERS FOR METHANOL CORRESPONDING TO
K = 2,330 mg/1 AND a - 1.25
Temperature
(°c)
5
22
Parameter
(hr"1)
k
kd
k
kd
Expected
value
0.04407
-0.000901
0.2805
0.003617
Standard
deviation
0.001524
0.002221
80% Confidence
interval
0.040346k&0
0.2213Sk^O.
.04854
3830
28
0.2656
0.005373 0.004642
0.2340*k&0.3057
pH and Salinity Effects —
Methanol-acclimated activated sludge was insensitive to high pH in fresh water
(see Figures 23, 24, and 25). At 5 °C, methanol removal rates were higher at pH 8 and
9 than at pH 7. At higher temperatures, the removal rate was still higher at pH 9 than
at a pH less than 6.0. In sea water (35 ppt salinity), the maximum achievable f value
was 0.7 at 5 °C, 0.1 at 22 °C, and 0.85 at 28 °C. Increasing salinity narrowed the
tolerable pH range and reduced the f values correspondingly.
Temperature Effects —
Methanol-acclimated sludge was sensitive to temperature. Between 5 and 22
°C, the substrate removal rate tripled for a 10 °C increase (0 = 1.115) at pH 7 in fresh
water, while there was no significant increase (0- = 1.0) in sea water. Between 22 and
28 °C, the temperature coefficient, -Q-, was 1.0 at pH 7 in fresh water, while it was 1.4
in sea water (see Figures 26 and 27). Thus, it may be deduced that first, the preferred
temperature range for methanol sludge is very limited. Second, bacterial activity
rapidly decreases at temperatures below the optimum temperature range. Third,
bacterial activity is not greatly affected by temperature change within the optimum
temperature range. Fourth, the lower limit of the optimum temperature range is
lowered as other environmental factors (pH and salinity) become more favorable for
the organisms. Similar results were observed for the phenol studies, but to a lesser
extent.
100
-------
PH .
9 -
8
6 ~
10
20 30
Salinity (ppt)
Figure 23. pH and salinity effects on the decomposition of methanol
by acclimated sludge. Iso-f lines at 5 C.
(PH=7, sal=0 ppt, temp.=5
= °-04407 hr
-1
101
-------
pH
8
10
20 30
Salinity (ppt)
074
064
40
Figure 24. pH and salinity effects on the decomposition of
methanol by acclimated sludge Iso-f lines at 22 °C.
(PH=7, sal=0 ppt, temp.=22
= °-2805
102
-------
8
k .891
I
.587
10
20 30
Salinity (ppt)
40
Figure 25. pH and salinity effects on the decomposition of
methanol by acclimated sludge. Iso -f lines at 28 °C.
k(PH=7, sal=0 ppt, temp.=28 °C) = °'2656 hr
-1
103
-------
-1.0711 1.0617
PH
1.0598
10 20 30
Salinity ijppt)
40
Figure 26. Temperature coefficient, 0, for the decomposition
of methanol by acclimated sludge Iso-0 Lines at
5-22 °C.
104
-------
PH
9 -
8 -
7 :
5 F
1199
1.1612 \
1.J192
4394
3939 -
1.2911
10 20 30
Salinity (ppt)
40
Figure 27. Temperature coefficient, 0, for the decomposition
of methanol by acclimated sludge Iso-0 lines at
22 - 28 °C.
105
-------
Endogeneous Respiration —
The cell decay coefficient, kj, showed a close relationship with the substrate
removal rate coefficient, k, when tabulated according to the ranges of k (see Table 15
and Figure 28). The following relationship between kj and k was established from the
In k
-------
0.006
0.005
r-^ 0.004
!s-
0.003
o
OJ
o
o
0.002
0)
•o
O)
o
o.ooi ••
kd = 0.0115 k
0.634
0.1
0.2
-1
0.3
Substrate removal rate coefficient, k (hr )
Figure 28. The relationship between the substrate removal rate
coefficient and the cell decay coefficient for methane!
acclimated sludge.
107
-------
-5.0
-5.5
-6.0
OJ
o
o
o
o
O)
O
-6.5
-7.5
kd = 0.0115 k
0.634
-a.5
I O
-3.0
-2.5
-2.0
-1.5
Substrate removal rate coefficient, in k
Figure 29. The relationship between.the substrate removal rate
coefficient and the cell decay coefficient for methanol
acclimated sludge.
108
-------
1,000s
Methanol stripping
° N&P and minerals supplied
A No nutrients
10 15 20
Aeration time (hr)
Figure 30. Nutrient effects on the decomposition of methanol by
acclimated activated sludge.*
* Temperature = 22 °C, pH = 6.9, initial VSS = 450 mg/1
109
-------
1000
900 -
800
700
600
TOC
(mg/1)
500
400
300
200
Temperature = 22 C
Initial VSS = 450 mg/1
Methanol stripping
TOC
• Distilled Water
A Ground water
Sea water
a -a
10
20 30
Aeration Time (hr)
40
50
Figure 31. pH variations owing to the decomposition of methanol
by acclimated activated sludge.
110
-------
that oxygen uptake was nil when there was no TOC removal by organisms at initial
stages and that it increased as the TOC removal rate increased (see Figure 32). Thus,
it is believed that the TOC removal rate is directly related to the bacterial activity.
Thirty-one of 66 tests showed an initial lag phase. The average duration of the
initial lag phase in the 31 cases was 9 hours and the average aeration lag time was 6
hours (see Table 17). The longest lag phase was observed at pH = 5.6, salinity =
0 ppt,and temperature = 22°C and it lasted for about 37 hours, causing a 30-hour
aeration lag time. The patterns of the inital lag phase are shown in Figure 33. It is
not certain which factors were responsible for the lag phase and the duration of the
lag phase. However, the duration of the lag phase generally appeared to decrease with
increasing temperature, except for the above extreme case.
TABLE 16. METHANOL DECOMPOSITION BY ACCLIMATED ACTIVATED SLUDGE
IN NATURAL SYSTEMS WITHOUT CHEMICAL AIDS
Test waters
Parameters
Distilled
water
Ground-
water
Sea
water
Initial pH
9
Alkalinity
2
Hardness
TDS (mg/1)
pH Range
Average pH
Average k (hr
f
Efficiency of
dilution water
7.6
7.1-7.6
7.3
0.181
0.646
0.813
8.6
115
206
462
7.5-8.6
7.8
0.0884
0.315
0.473
8.4
164
7,810
38,300
8.2-8.4
8.3
0.0315
0.112
1.0
1
2
Temperature = 22 °C; k9.
Expressed as mg/1 Ca CO,
= 0.2805 hr
Ill
-------
2200 r
2000
1800
1600 -
1400
1200 -
1000 -
800 -
600
400 -
200 -
Methanol stripping
20 30
Aeration time (hr)
Figure 32. The relationship between substrate removal rate and
oxygen uptake rate.
50
112
-------
1200
1100 Duration = 37 hrs.
1000
900
800
700
600
TOO
(mg/1)
500
400
300
200
100
0
Lag time = SOhrs
Temp. = 22 °C
pH = 5.6
sal = 0 ppt
Temp. = 22 °C
pH = 7.4
sal = 10 ppt
10 20
Methanol Stripping
(ke = 0.00540/hr)
30 40 50
Aeration time (hr)
60
70
80
Figure 33. Initial lag phase in the decomposition of methanol by
acclimated activated sludge.
113
-------
TABLE 17. INITIAL LAG PHASE IN METHANOL DECOMPOSITION BY
ACCLIMATED ACTIVATED SLUDGE
Type of test
Kinetic study
pH, salinity
study
Nutrients
study
Countermeasure
application
Total
Temp.
(°C>
5
22
28
5
22
23
22
24
28
-
Number
of
tests
5
2
5
15
15
15
5
3
1
66
Cases
of
lag phase
4
-
9
7
8
-
3
-
31
Average
duration
(hr)
12.5
-
9
13.5
6
-
7
-
9
Average
lag time
(hr)
9.5
^
8
8
2.5 i ,•
-
4
-
6
Oxygen Requirement —
Complete oxidation of methanol requires 4 grams of oxygen per gram of carbon
as indicated in the following equation:
CH3OH + § O
(12) (48)
= C02 + 2 H2O
(72)
Complete oxidation of bacterial cell requires 2.67 grams of oxygen per gram of
cell carbon (Equation 57). Thus, 1 gralri of methanol carbon requires 1.33 grams (4.0 -
2.67 grams) of oxygen when it is utilized for cell synthesis.
One gram of methanol carbon removed by organisms results in 1.25 grams of
organic fraction of cell production (a = 1.25), while theoretically, it will make 1.88
grams of organic solids if it is used entirely for cell synthesis. Therefore, 66.5%
(1.25/1.88) is utilized for synthesis. Per gram of carbon, 66.5% of TOC decomposed by
organisms will require 1.33 grams of oxygen and the remaining 33.5% of TOC removed
by organisms will require 4 grams of oxygen. Thus, a' = (0.665 x 1.33) + (0.335 x 4) =
2.23 and b' = 1.42 kd .
Then, the oxygen utilization rate (Equation 17) is expressed as:
where -4-r and
at
(73)
are calculated in Equations 63, 64, and 71.
114
-------
The oxygen requirement tests were conducted independently from the kinetic
study tests on which Equation 73 is based. When 81 observations of oxygen uptake rate
measurements were compared to the results from Equation 73, the mean difference
between the observed and theoretical oxygen requirements was 0.317 mg/1 02/hr ± 13.0
mg/1 02/hr. Statistically, it was shown that this difference in oxygen utilization rate
was not significant (see Appendix 1,4). Therefore, it was concluded that Equation 73
could be used to predict the oxygen requirements.
In Figure 34, the theoretically evaluated values of a1 and b' are justified through
a linear graphical method. When Rr/X is plotted against - (AU/At)/X, the plotted
points should fall on a straight line, the slope and Y intercept of which are a' and b1
respectively (Equation 17*). But, the experimental errors and variable transformations
hardly allow any straight lines to be drawn. However, it is reasonable to assume that
the probabilities of overmeasurement and undermeasurement of all the variables
involved are the same. Therefore, the straight line must roughly divide the plotted
points equally. Equation 73 fulfilled this requirement by dividing the observed points
38 to 43.
Through the above two justifications, it was concluded that a1 = 2.23 (based on
a = .25) and b1 = 1.42 k,j (kj is based on k Equation 71) provided a
reliable estimation of oxygen demand.
p-Nitrophenol
Results of the p-nitrophenol screening tests were subjected to the same growth
kinetic coefficient estimation procedures as were used for phenol and methanol. The
estimated kinetic parameters are given in Table 18. The values for k and kj are very
similar to those for phenol, but the ks concentration is substantially less, as is the
value of a.
TABLE 18. ESTIMATED KINETIC PARAMETERS FOR PARA-NITROPHENOL
Expected Standard 80% Confidence X Range S Range N
Parameter value deviation interval (mg/1) (mg/1)
K8(mg/l)
k(hr~1)
k.dir'1)
a
a
42.39 88.14
0.08509 £.0.01785
^-0.03075
0.004058 0.003963
0.4300 0.1125
20
to
240
up to
220 34
Temp. = 20 C
115
-------
0.10- 38 points
0.8
0.6
Rr/X,
0.4
0.2
43 points
o
o o
o-
- O O v-er Ov.
°0oo /p° oo Rr = - 2.23 ^ + 0.00635 X
(at 20 °C)
0.02 0.04 0.06 0.08
Figure 34. Comparison of theoretical and observed oxygen uptake
rates in methanol decomposition.
116
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SECTION 9
SIMULATED SPILL TESTS
To determine toxicity of the test compounds in natural systems and to test
deployment methods for the biological countermeasures, a number of model systems
were established. Fifteen aquaria, of 57-liter -capacity each, were established with
sediments, plants, and Crustacea to simulate a portion of a slow-moving stream or a
pond. These aquaria received a continual flow of aged tap water initially, and, later,
groundwater, with a residence time of about 2 days. Because these aquaria could be
used, cleaned, and reestablished rather quickly, they were used initially to determine
the response of natural systems to spillage of the contract list of hazardous materials.
Total system productivity was monitored to determine this response. Then, hazardous
materials were spilled into the aquaria to test the in situ and portable treatment
system countermeasure techniques. Ten small pond ecosystems were also established.
Each was 1.8 meters in diameter, contained 800 liters of water, was stirred continu-
ously by a pump, and was furnished with sediments and plants like the aquaria. These
systems were used to simulate a portion of a pond and their size diminished some of
the wall effects characteristic of the aquaria.
The results of the aquaria and of the pond tests were used to develop
experimental methods for large-scale biological countermeasure tests in a 30.5-m
diameter, 3.0-m deep model lake. Spills of hazardous materials were made in the tank
and the in situ countermeasure techniques used to mitigate the material were
evaluated. Flowing-system tests were carried out in a model river, which had been
used previously for radioactive material transport studies. The model river is 61
meters in length and consists of two parallel channels, each equipped with flow and
water level regulation devices and established with sediments, plants, and small
aquatic organisms. The experiments carried out in the model river system are
described in Section 11 along with the countermeasure application techniques.
The aquaria and ponds were located in temperature-regulated rooms with
controlled fluorescent lighting, while the model river and model lake were located
outside.
While the primary purpose of these studies was to investigate the feasibility of
using a biological countermeasure in a near real-life system, secondary objectives
were: (1) to examine hazardous material removal rates as a function of bacterial
culture, amounts of bacterial mass (specifically, the ratio of mass of hazardous
material spilled to the bacterial mass applied), the addition of nutrient salts, the
method of bacteria application, and the effects of dilution in flowing systems and (2)
to examine the effects of the biological countermeasures on the quality and eco-
systems of the receiving water.
117,
-------
AQUARIA TESTS
Rationale
As mentioned above, the aquaria tests were designed to provide initial
experience in the use of biological countermeasures for spilled hazardous materials in
ecosystems, as well as to show the effects of the spills. It was assumed that removal
of spilled materials in the aquaria could be monitored as it had been in the treatment
reactors. Therefore, the general rationale of the experiments was to "spill" one of the
test chemicals, follow its removal over time through sampling and analysis, and
determine the effects of the countermeasure on the system. Rates of removal,
effects of the ratio of material mass spilled to bacterial mass added on the removal
rates, and effects of nutrient additions were investigated.
The basic measurement parameter, removal rate of the spilled material, was
calculated in the following way.
Because the aquaria used in these studies were continuously supplied with fresh
water and were mechanically mixed by stirrers, the aquaria could be treated as
continuous-flow, stirred reactors. The concentrations resulting from spills into such
systems may be described by the following equation:
(i+k)t (74)
5 - 3Qe ,
Where: S = concentration of spilled material at time t after spillage
,
S = initial concentration of spilled material = M/V (M/L ),
$ = hydraulic residence time = V/Q(T),
k = biological decomposition rate (M/(TM)),
t = time (T), ,
V = system volume (L ), _
Q = flow into or out of system (L /T), and
M = mass spilled (M).
Thus, in these spill control tests, it was desired that k be calculated by measuring S at
various times during the test, determining the overall removal rate of the spilled
material, and then solving for k knowing the hydraulic dilution rate, 1/6-.
Procedures
Apparatus —
The 57-liter capacity aquaria used were 60 cm long, 31 cm wide, and 29 cm
deep. The water level was maintained so that with sediments in the bottom, the
working water volume was about 45 liters. Aged tap water and, later, groundwater
were metered into each aquarium through tygon tubing constricted by a screw clamp
valve. Flow rates were measured twice daily and adjusted when necessary (because of
their tendency to change over time). At the outlet end of each aquarium, a constant
level siphon was used, which had to be serviced periodically to remove attached algae
and bacteria. The water in each aquarium was gently stirred by high torque stirrers
equipped with a stirring rod to which was attached a smaller cylinder perforated with
holes. Dye tests showed that at most, complete mixing occurred within a few minutes.
118
-------
Dissolved oxygen measurements were made with a Precision Scientific Co.
probe and temperature was measured by a thermocouple on the probe or by a mercury
thermometer.
The aquaria were illuminated by banks of 40-watt fluorescent lights placed
directly overhead and were put on a 12-hour "on", 12-hour "off" cycle. Light intensities
were in the range of 300 to 500 foot-candles at the level of the plants in the aquaria.
Nine of the aquaria (Aquaria 1 through 9) were operated individually, that is,
water was metered into and drained from each aquarium . Six of the aquaria (Aquaria
10 through 15) were operated in series. Water was metered into Aquarium 15 and was
conveyed to Aquarium 14, then Aquarium 13, and so forth through siphons; water was
drained from Aquarium 10. The configuration of these aquaria in the laboratory is
shown in Figure 35 and the water depths and volumes of each aquarium are given in
Table 19.
TABLE 19. WATER DEPTHS AND VOLUMES OF AQUARIA
Aquarium
number
1
2
3
4
5
6
7
8
Water
depth
(cm)
25.4
24.8
23.8
23.8
24.8
23.8
24.4
23.8
Water
volume
U)
46.3
45.2
43.4
43.4
45.2
43.4
44.6
43.4
Aquarium
number
9
10
11
12
13
14
15
Water
depth
(cm)
25.4
22.2
22.5
24.8
24.1
23.5
24.1
Water
volume
a)
46.3
40.6
41.2
45.2
44.0
42.9
44.0
Experimental Procedures —
The general procedure for all tests conducted was to permit the aquaria
ecosystems to stabilize for several weeks before a test was begun. The systems were
considered to be stable when the production and respiration values and ratios were
relatively constant.
After the material to be spilled was selected, the amounts to be spilled and the
amounts of acclimated or unacclimated bacteria to be added were determined. Flow
rates, thus residence times, were set and a sampling schedule was established.
A culture of bacteria was fed the test chemical in a reactor for several weeks
prior to the spill tests. Just prior to the spill tests, the culture was allowed to settle,
the supernatent was decanted, and a measured volume (about 10jP) of the sludge was
readied for addition to the aquaria. The TSS and VSS content of the bacteria culture
were measured after settling.
119
-------
S-
o
fO
i_
o
c
«3
cr
(T3
O
c
o
fO
s_
CT>
in
CO
O)
s_
3
cn
120
-------
3ust prior to spillage, a volume of water equal to the bacterial culture volume
was removed from the aquarium to receive the bacteria so that the aquarium volume
returned to a desired level upon addition of the culture. This procedure was not done
for test chemical additions since the volume added was usually only a few milliliters.
Spilling was accomplished by simply pouring the test chemical or bacterial
culture into the appropriate aquarium near the stirrer to insure mixing.
Generally, a water sample was taken prior to spillage and then after spillage
with decreasing frequency over a period of a few days to several months. These
samples were routinely analyzed for TOG, total oxygen demand (TOD), pH, TSS, VSS,
and the compound spilled (gas chromatography). Occasionally, plate counts were
made.
Daily dissolved oxygen and temperature measurements were made immediately
after the lights came on and after they were turned off so that production and
respiration values could be calculated using the three-point method. Hourly samples
for dissolved oxygen were taken on several occasions to verify that the three-point
method could be applied.
Data Analysis ~
After each test, the TOC or gas chromatograph results of sample analyses were
plotted on semi-logarithmic paper to permit calculation of the slope of the curve.
From this slope value, the hydraulic dilution rate was subtracted to obtain the
biological decomposition rate, which was then related to the experimental conditions
imposed. VSS and plate count data were plotted on arithmetic paper to follow the net
growth of the bacteria added over the period of the experiment.
The dissolved oxygen and temperature data were used to calculate the
production and respiration rates and their ratios. These data were then plotted with
time to show the effects of the material spilled and of the countermeasure on the
ecosystem.
Results
Two hazardous materials were used in the aquaria spill tests -- phenol and
methanol. The results of the tests using these two compounds are described below.
Phenol —
A total of six sets of experiments were conducted in the aquaria using phenol.
Removal test No. 1 — The first of these tests was performed during the period
from March 17 to May 1, 1974 with spill monitoring occurring on April 10-13, 1974. Five
aquaria were used in the initial experiment (Aquaria 1,2,3,4, and 6); three of these
aquaria (Aquaria 1, 2, and 3) contained the rooted plant Vallisneria and the other two
(Aquaria 4 and 6) contained only phytoplankton. One aquarium in each set was used as
a control (Aquaria 1 and 6) and no phenol was spilled into these aquaria. Also in each
set, one aquarium received phenol (theoretical initial concentration of 95 mg/1) but no
bacteria (Aquaria 2 and 4). In the aquaria with the Vallisneria, one aquarium received
phenol plus bacteria (Aquarium 3).
121
-------
Following the "spill" of phenol into the aquaria, the concentrations of phenol
and bacteria were monitored using gas chromatography, TOC, TOD, and VSS analysis
and plate counts. The results of the gas chromatography and bacterial concentration
analyses are presented in Figures 36 and 37. The results for Aquarium 2 given in
Figure 36 show a decrease in phenol due strictly to washout of phenol by water flowing
through the system. Note that the VSS concentrations were very small.
In contrast, the decrease in phenol concentration in Aquarium 3, which received
phenol and bacteria, was very rapid after the first 24 hours. Note in the lower portion
of Figure 37 that the VSS concentration representing the bacterial mass was high
initially and dropped off to a rather stable concentration of about 12 mg/JL The
bacteria added to this aquarium settled rapidly onto the sediments and grew there for
several days before disappearing through self-oxidation. Also given in the lower
portion of Figure 37 are plate count concentrations of bacteria that are able to use
phenol as a substrate.
The observed rates of decay of phenol in these aquaria are given in Table 20.
When the dilution rates are subtracted from the observed decay rates, one may obtain
the rate of biological decay of phenol. In the two aquaria that received only phenol
(Aquaria 2 and 4), the biological decay rate was larger than the dilution rate. The net
effect of adding bacteria was to reduce phenol concentrations to less than detectable
levels within three days.
TABLE 20. SIMULATED SPILL OF PHENOL IN AQUARIA, REMOVAL TEST NO. 1
(APRIL 1970 to MAY 3, 1974)
Disappearance rates (
-------
100
90
80
70
60
50
40
I
o.
30
20
10
Dilution rate:
Observed de-cay:
0.48 d
0.62 d
Biological decay: 0.14 d
o—o Observed
Theoretical
-1
-1
I i I I
Phenol only
(Aq. 2)
Washout
1 I I i i I i i i
10
20 30
Time (hr)
40
50
100
50 -
1
1 1 1 1 1 1 I 1
1
10
40
20 30
Time (hr)
Figure 36. Analytical results of removal test Mo. I, Aquarium
2 (phenol only).
50
123
-------
Phenol -t- bacteria
(Aq.3)
Washout
I 30
o
I
Bu
Observed
Theoretical
20
Dilution rate :
Observed decay:
0.50 d
1.12 d
Biological decay: 0.62 d
10
20 30
Time (hr)
40
50
10
20 30
Time (hr)
40
Figure 37. Analytical results of removal test Mo. 1, Aquarium
3 (phenol + bacteria).
50
4.0
- 3.5
3.0
124
-------
receiving no phenol dropped, apparently as a result of bacterial growth and phenol
decomposition. Results of the production and respiration analyses are given in Figures
38, 39, and 40. The gross production (?G) and total respiration (Rj) in the control
aquarium (Aquarium 1) are given in Figure 39. As indicated in the figure, production
and respiration averaged about 7 mg/l/d. There was a slight increase in PG and R_
from day 0 to day 10, which occurred as the aquarium was changed from a static to a
flow-through system. Stabilization of this aquarium following the change in the nature
of the system is also shown in the ratio of gross production to total respiration, which
is given in the lower portion of Figure 38.
Once phenol was spilled, the ecosystems in the aquaria responded rapidly. In
Aquarium 2, which received only phenol, there was a two-day delay in response to the
spill as shown in Figure 40. On the second day following the spill, gross production of
the system dropped to near zero, while total respiration increased to over 9 mg/l/d.
However, respiration suddenly dropped to near zero, but five days after the spill,
production and respiration began to recover to values higher than pre-spill levels.
Some cycling of production and respiration levels is evident between days 20 and 40 in
Figure 40. These overall changes are also reflected in the ratio of production to
respiration, which is given in the lower portion of Figure 40.
A similar response was given by the ecosystem in Aquarium 4; however, because
this aquarium received bacteria and phenol, the response was more immediate. Within
a few hours following the addition of bacteria, the dissolved oxygen dropped to zero in
the aquarium. The gross production dropped to near zero and respiration increased
rapidly for a day then dropped to just under 5 mg/l/d. By the fourth day after the
phenol spill, the ecosystem was recovering and production and respiration levels far
exceeding the prespill levels were measured as shown in Figure 41. Maximum
production and respiration values of about 19 mg/l/d were reached. These overall
changes are also reflected in the ratio of production to respiration, which is given in
the lower portion of Figure 41.
It is obvious that the ecosystems in these aquaria responded dramatically to the
phenol spill. The aquaria were monitored almost continuously following the spill and
they returned to essentially pre-spill conditions.
The bacterial concentrations present were measured as volatile suspended solids
(VSS) or were determined from a plate count (Figures 35 and 36). In Aquarium 2,
essentially background levels of 3 to 4 mg/1 VSS were found, whereas much higher
levels were found in Aquarium 3, to which bacteria had been added. However, these
added bacteria settled to the bottom of the aquarium and remained there as a white,
fluffy mass until long after the phenol had been completely removed. After 15 to 20
days, this bacterial layer disappeared.
Removal test No.2 —A second test was performed during the period April 19 to
June 3,1974 to confirm the results of the first test. Using Aquaria 7, 8, and 9, enough
phenol was spilled to produce an initial concentration of approximately 100 mg/1 in
Aquaria 8 and 9, while Aquarium 7 was kept as a control. The results of this test are
given in Table 21.
The effects of phenol addition on dissolved oxygen levels and on production and
respiration were essentially the same as in the first test. The only difference was that
the production values in Aquarium 8, receiving only phenol, did not drop as much as in
Aquarium 2 in the earlier test.
125
-------
10 !—
s r
Control (Aquarium 1)
.0 -
10
20
30
40
50
60
L
Phenol only
(Aquarium 2)
Phenol + bacteria
(Aquarium 3)
I I
20 30 40
Time (days after 26 Mar 74)
Figure 38. Dissolved oxygen values during removal test No. 1.
126
-------
12
cn CD
CD
Q.
C O
O ••-
o s_
3 •<-
•O Q.
O
Q. S_
in i—
(/) (O
O 4->
i- O
C3 I—
8
0
0
10 20
Time (days)
30
40
1.4
1.0
O
•M
I-
o;
c
Q.
0.6
0.2
0
II
A r\f
30
40
10 20
Time (days)
Figure 39. Gross production (I) and total respiration (II) in
aquarium I (control) Removal test No. 1.
127
-------
12
o
Phenol spill
10 20
Time (days)
30
40
1J4
20
Time (days)
Figure 40. Gross production (I) and total respiration (II)
in aquarium 2 (phenol) removal test No. 1.
128
-------
iU
1 - 15
>> "O
IO -v.
•o <—
^. -^
11
•S „ 10
a at
a.
M
• c
e o
O i-
0 i.
•o "3. 5
O (/» J
o. s-
at i—
I/I re
O •!->
u o
CO 1—
0
1.5
.2 1.0
2
i—
as.
a.
0.5
0
i i i
Phenol //I
/» i
R , / \ ,'A '\
K«r ' •' ^ X-!— ' s / ' \*
T\. y \ /; ^
• { -1
". J^ij
r\j " I v'i
i /
,1
i \A/ ,
0 10 20 30 4C
Ti'me'^days) -
II.
r
: \
~ \ , ~ ~ \ • "\^. " '^' \ — ~T£.
U | V
j
j i
[ I
m. ' t —
i
', 1
, JV" ,
0 10 20 30 40
Time (days)
Figure 41. Gross production (I) and total respiration (II) in
Aquarium 3 (phenol + bacteria). Removal test No. 1,
-------
TABLE 21. SIMULATED SPILL OF PHENOL IN AQUARIA, REMOVAL TEST NO. 2
(APRIL 19 to MAY 3, 1974)
Disappearance rates (d~l) Environmental Conditions
Aquarium Conditions Initial Obsd. Dilution Biol. Avg.
Number phenol decay rate decay pH
cone. rate rate
Avg. Avg.
DO te
(mg/1) (
1
2
7 Control 0
8 Phenol 92.2
9 Phenol + 86.5
bact.
Phenol/VSS ratio =1.1
Lowest DO value.
0.485 _- 8.2
(7.5T
1.83 0.51 1.32 8.2
(0.0)
3.30 0.49 2.81 6.7
mg/mg.
9.1 20.5
8.0 20.6
1. 20.9
Effect of phenol/VSS mass ratio test — To determine how much bacterial mass
must be added to a known mass of spilled chemical, a test was carried out during May
1974. Phenol and bacteria were added in such proportions that their ratio ranged from
12.9 to more than 4,300 mg/mg. Also tested was a bacterial culture, developed by the
microbiology group, that was acclimated to phenol and that would stay in suspension.
The initial phenol concentration in each of the three aquaria (Aquaria 4, 5, and
6) was 485 ppm. Aquaria 4 received a phenol and bacterial culture developed from
activated sludge, while Aquaria 5 and 6 received phenol and two different volumes of
an unacclimated bacterial culture developed by the microbiology group.
The results of these tests are given in Table 22 and show the importance of the
phenol/VSS mass ratio. At the lowest mass ratio, the biological decay rate was 3.1 d ,
while it was less than 0.4 at higher ratios. It should be noted that, in each case, a
48-hour time lag was observed before significant decomposition of phenol occurred.
Because adding large amounts of bacteria can lower dissolved oxygen concen-
trations to zero, the removal rate can be inhibited. Thus, one must strike a
compromise between the amount of bacteria added and the risk of lowering dissolved
oxygen to inhibiting levels. In this particular experiment, Aquarium 4 was aerated,
dissolved oxygen levels remained above 5.0 mg/1, and a very high removal rate was
observed. Aquaria 5 and 6 were not aerated, dissolved oxygen levels dropped to near
zero, and lower removal rates were observed. When the results for Aquarium 3 in the
previous experiment (Table 20) are compared with those of Aquaria 5 and 6, it is still
apparent that a low phenol/VSS mass ratio is desirable, but aeration can greatly
enhance the removal rate, as in the case presented here, by a factor of 5.
Nutrient Addition Test -- The next test was performed to examine the
importance of nutrient additives to the biological countermeasure. In the treatability
studies, mineral salts had to be added to provide the necessary amounts of nitrogen,
phosphorus, and other nutrients for bacterial growth. The nutrient addition test was
carried out in January 1975 and involved three test conditions: (1) Aquarium 1 received
phenol only (500 ppm), (2) Aquarium 2 received phenol (500 ppm) plus bacteria, and
130
-------
(3) Aquarium 3 received phenol (500 ppm) plus bacteria and NaHCO,,
(NHJoSCX in amounts sufficient to provide the concentrations .of
phospnorus used in the treatability studies.
CL, and
nitrogen and
The results of these tests (Table 23) show that nutrient additions produced a
two-fold increase in the decay rate as compared to Aquarium 2, even though the
dissolved oxygen level fell to zero near the end of the test. The importance of
nutrients was reflected in the VSS concentrations, which are indicative of bacterial
concentrations. In Aquarium 1, the VSS concentration never exceeded 12 mg/1. In
Aquarium 2, which received bacteria but no nutrients, the VSS dropped from an initial
concentration of 214 mg/1 to less than 10 mg/1 within six hours and stayed below 6 mg/1
for the rest of the test. In contrast, the VSS concentration in Aquarium 3 dropped
initially (as in Aquarium 2) but then rose after 10 hours following the spill to 38 mg/1
before decreasing again.
TABLE 22. EFFECTS OF PHENOL/ VSS MASS RATIO ON
SPILL REMOVAL RATE (MAY 1974)
Disappearance rates (d~*) Environmental Conditions
Aquarium
number
Phenol/ Obs. Decay Dilution Biol.
VSS mass rate
ratio
(mg/mg)
rate decay
rate
Avg. Avg.
temp . D . 0 .
(°C) (mg/1)
Avg.
PH
4-
5
12.9
43.7C
4367.4'
3.58
0.88
0.76
0.5
0.5
0.5
3.1
0.38
0.26
21.5
21.5
22.0
8.0 .
(7.5)'
5.0 ,
(o.o)-
5.0 ,
(o.o)-
7.3
7.4
7.5
Phenol added to each aquarium was 21,400 mg.
Rates observed over 48-hour time lag.
Aerated.
Activated sludge-adapted bacteria.
Lowest D.O. value.
Laboratory culture of adapted bacteria.
131
-------
TABLE 23. EFFECTS OF NUTRIENT (NITROGEN AND PHOSPHORUS) ADDITION
WITH BACTERIA FOR CONTROL OF PHENOL SPILL (JANUARY 1975)
— 1 2
Disappearance rates (d ) Environmental Conditions
Aquarium Conditions Phenol/ Obsd. decay Dilution Biol. Avg. Avg. Avg.
number VSS rate rate Decomp. pH D.O. temp.
mass ratio rate (mg/1) (°C)
(mg/mg)
1 Phenol
2 Phenol + 1.9
bacteria
3 Phenol + 1.9
bacteria +
nutrients
0.576
0.984
1.656
0.528
0.446
0.514
0.048
0.538
1.142
7.4
7.5
7.6
8.8
(8.5)2
7.2
(4.5T
5.6
(0.3T
20.5
18.6
20.3
1 Phenol added = 21,440
2
Average over a 30-day period.
3
Lowest D.O. value
Methanol —
Following the aquaria tests with phenol, the methanol tests could be designed
with some confidence to provide information on decay rates and effects. Tests were
conducted to determine the influence of methanol/VSS mass ratios and the influence
of nutrients on removal rates.
Effects of methanol/VSS mass ratio test No. 1—In June 1974, the first methanol
test in the aquaria was carried out. It was designed to examine the effects of the
methanol/VSS mass ratio on methanol removal rates with aeration. Six aquaria were
involved in the test; the initial conditions and results are summarized in Table 2k.
Initial concentrations of methanol ranged from 1,000 to 10,000 ppm and the mass
ratios ranged from 38 to 1934 mg/mg. While the methanol removal rates ranged from
0.02 to 0.48 d , there appears to be little correlation between the mass ratios and
the removal rates. As with phenol removal, low dissolved oxygen conditions may
inhibit the bacteria and lower the removal rates. Such could have been the case for
methanol removal in Aquaria 7, 8, and 9. It is interesting to note that methanol
removal in Aquarium 3, to which no bacteria were added, was very high; apparently,
bacteria in the aquarium were readily able to break down the methanol.
132
-------
TABLE 24. EFFECTS OF METHANOL/VSS MASS RATIO ON METHANOL REMOVAL RATES,
TEST NO. 1 (JUNE 1974)
Disappearance
; rates (d"1)
Aquarium Initial Methanol Obsvd. Dilu.Biol.
number methanol VSS decay .rate decomp.
cone. mass ratio rate rate
Environmental
conditions^
D. 0.
(mg/1)
Temp.
(ppm)
(mg/mg)
1
2
3
7
8
9
1
2
5,000
1,000
10,000
5,000
1,000
5,000
Average over
Lowest D.O.
193
38
379
1,934
379
-
a 30-day
value.
0.653
0.684
0.635
0.898
0.510
0.904
period.
0.465
0.423
0.459
0.408
0.493
0.446
0
0
0
0
0
0
.188
.261
.176
.490
.017
.458
8.1
8.3
8.4
8.9
9.2
9.0
8
8
8
7
(1
8
(2
6
(1
.8
.8
.8
•2 o
.O)2
o
!o)2
.5
.2)
21
20
21
21
21
21
.3
.9
.5
.4
.3
.5
Effects of methanol/VSS mass ratio test No. 2 — The second test to examine
the effects of the methanol/VSS mass ratio was conducted in February 1975. Three
different mass ratios ranging from 1.75 to 349 mg/mg were used with initial methanol
concentrations ranging from 5,000 to 100,000 ppm. Aquarium 1 received only
methanol, while Aquaria 4, 5, and 6 each received equal amounts of bacteria and
different amounts of methanol. To avoid inorganic nutrient limitations, NaHCO,,
JoSO* were each added twice a day to Aquaria 4, 5, and 6.
KH-PCX, and
The results of the test (Table 25) indicate some biological removal in each
aquarium, with the lowest rate in Aquarium 1 as expected. The removal rates in
Aquaria 4 and 5 were relatively low, but apparently the dissolved oxygen
concentration, which dropped to near zero in both aquaria within 24 hours after the
spill, inhibited removal. Somewhat surprising was the removal of methanol at the
initial concentration of 100,000 ppm. Biological activity was present as the dissolved
oxygen dropped to near zero in Aquarium 6 after 72 hours, but removal was apparent,
perhaps through air stripping, immediately after the spill of methanol.
It is possible that some removal was a result of aeration and, as shown in
Section 8, this could account for a substantial portion, 0.23d~ , of the observed rates
in Aquaria 1 and 5. However, the aeration rate was not of the magnitude used in the
experiments described in. Section 8. Thus, it is assumed that loss through aeration
would be less than 0.23 d and perhaps negligible.
133
-------
TABLE 25. EFFECTS OF METHANOL/VSS MASS RATIO ON METHANOL REMOVAL
NO. 2 (FEBRUARY 1975)
Disappearance
rates (cH)
Environmental
conditionsi
Aquarium Initial Methanol Obsvd. Dilu. Biol.
number methanol VSS decay rate decomp.
cone. mass ratio rate rate
(ppm) (mg/mg)
pH D. 0. Temp.
(mg/1) (°C)
1
4
5
6
5,000
5,000
10,000
100,000
-
1.75
34.9
349.4
0.70
1.14
0.77
0.96
0.53
0.52
0.48
0.50
0.17
0.62
0.29
0.46
9.4
8.7
7.3
7.7
9.0
1.0
1.0
4.6
19.1
19.1
19.8
20.3
1Average over a 30-day period.
Effects of nutrient additions — The effects of nutrient additions following a
spill of methanol were studied in a test carried out in January 1975. Using initial
methanol concentrations of 5,000 ppm and a methanol/VSS mass ratio of 26.6 mg/mg,
Aquarium 1 received only methanol, Aquarium 7 received methanol plus acclimated
bacteria, Aquarium 8 received methanol plus bacteria plus nutrients, and Aquarium 9
received methanol plus bacteria plus twice the mass of nutrients placed in Aquarium 8.
The nutrients were added twice each day to Aquarium 8 and 9 to produce the indicated
concentration upon each addition. The aquaria were also aerated continuously.
From the results given in Table 26, it is apparent that nutrient additions aided
biological removal since the biological removal rate in Aquarium 8 (with nutrients) was
twice that in Aquarium 7 (without nutrients). However, doubling the nutrient
concentrations did not produce an increase in the decay rate, because either: (1)
nutrient concentrations were at saturation levels for growth or (2) low dissolved
oxygen levels, (even with aeration) were inhibitory.
134
-------
TABLE 26. EFFECTS OF NUTRIENT ADDITIONS ON
METHANOL REMOVAL (JANUARY 1975)
Disappearance Environmental
rates (d ) conditions
Aquarium Conditions Initial MeOH/VSS Obsvd. Dil. Biol. pH D.O. Temp.
number methanol mass ratio decay rate decomp. (mg/1) (°C)
cone. (mg/mg) rate rate
(ppm)
1
7
8
9
Methanol
Methanol+
bacteria
Methanol+
bacteria*
nutrients
Methanol+
bacteria*
2x nutrients
5,000
5,000
5,000
5,000
26
26
26
26
.6
.6
.6
.6
0
0
1
0
.64
.72
.03
.94
0.48
0.52
0.53
0.52
0.16
0.20
0.50
0.42
8.2
7.3
7.2
(
7.3
(
8.9 .
(8.4)2
6.3
(3.4)2
3.0
:«i.o)2
1.8
:
-------
Procedures
The ponds were rigid, plastic wading pools arranged in two rows and equipped
with banks of fluorescent light to provide 300 to 400 foot-candles at the water
surface. They were filled to a depth of about 10 cm with sediment and to a depth of
about 12 cm with water and were planied with Vallisneria, a rooted aquatic plant. The
surface area of each pond was 2.0 m (diameter of 1.6 m) and the average volume was
312 1. Mixing in each pond was achieved by submersible pumps, resulting in circular
water movement. Vigorous aeration was also used in several ponds.
The general experimental procedure was to spill methanol or phenol, followed
immediately by the bacteria culture into a pond. The bacterial culture was acclimated
to phenol over a one-week period in a batch reactor then transferred to a 200-liter
container for further growth. After stirring to ensure initial mixing, the concentration
of the spilled chemical was monitored in the pond by TOC, TOD, and gas
chromatography analyses.. TSS, VSS, pH, temperature, and dissolved oxygen were also
monitored. A variation of this pattern was the use of a portable treatment unit to
which pond water was pumped, treated by an acclimated culture, and returned to the
pond by gravity flow. In this case the spilled chemical concentration was also
monitored in the treatment unit effluent.
At the end of each test, the chemical concentration data were plotted and the
removal rate calculated using Equation 75. Averages of pH, temperature, and
dissolved oxygen were computed.
Results
Methanol Spill Tests —
A simulated methanol spill was carried out in 3une 1974 in Ponds 1, 2, 3, and 5.
Two concentrations of methanol (1,000 and 5,000 ppm) and four bacterial culture
concentrations (0, 1.0, 3.0, and 5.0 liters of culture of 665 mg/1 VSS) were used. The
initial conditions of the test are given in Table 27.
The study was conducted over an eleven-day period, and the results are given in
Table 27. Because of the relative equality of the decay rates, it is difficult to
determine if biological decay, volatilization, or sorption processes were important.
There were increases in VSS concentrations on the third day in Ponds 1 and 2 and on
the fifth day in Ponds 3 and 5, indicating bacterial growth. Also, decreases in
dissolved oxygen in Pond 5 after day 5 indicated continued growth. The decay rates
were substantially less than those measured in the aquaria, but initial VSS
concentrations were apparently too small to provide an inoculum larger than that
already existing in the ponds.
That minimum inocula are required for methanol removal in aquatic systems
was shown in this test.
136
-------
TABLE 27. RESULTS OF METHANOL SPILL INTO PONDS
(JUNE 1974)
Pond Initial Methanol Initial Bacteria MeOH/ Decay Environmental Conditions
number Methanol added culture culture VSS rate
cone. (kg) cone. added ratio (d~l) pH D.O. Temp.
(ppm) (mg/1) (mg) (mg/mg) (mg/1) ( C)
1
2
3
5
1
2
1,000
1,000
5,000
5,000
Average
Lowest DO
0.29
0.29
1.45
1.45
over a
value.
0 0
5.4 1995
1.8 665
9.0 3325
30-day period.
145
2180
436
0.15
0.13
0.11
0.10
8
8
8
8
.1
.2
.2
.0
8.0(5
7.0(6
6.8(5
4.0(0
.9)
-0)?
.4)?
.2)2
25
25
26
26
.2
.7
.0
.3
Phenol Spill Tests —
The overall purpose of the phenol spill experiment conducted in November 1974
was to test the biological countermeasure on phenol in a larger container than the
aquaria. Secondary objectives were to examine the effects of using acclimated as
opposed to unacclimated bacteria and to evaluate use of a portable treatment unit
containing acclimated bacteria.
The ponds used, the treatment technique applied, and the initial phenol and
bacteria concentrations are given in Table 28. Pond 6 was used as a control, receiving
phenol but no bacteria, while Pond 7 received phenol and acclimated bacteria. Ponds 8
and 9 received two levels of unacclimated bacteria acquired just prior to the test from
the Govalle Treatment Plant. Initial phenol concentrations ranging from 148 to 170
ppm and initial bacteria concentrations from 128 to 224 mg VSS/1 were used. Phenol to
bacteria VSS ratios ranged from 0.71 to 1.4 mg/mg.
The observed decay rates are given in. Table 28 and show that essentially no
removal occurred in the control pond (0.02 d~ ), while rapid removal took place in the
pond with acclimated bacteria (0.76 d ), even though the dissolved oxygen levels
dropped to zero for a short period. The two ponds receiving the two levels of
unacclimated bacteria had low removal rates (0.08 and 0.15 d~ , respectively), but the
rates were in proportion to the amount of bacteria added. The removal rate in Pond 3,
treated by the portable treatment unit, was 0.10 d~ . Since this removal rate was in
part a function of the pumping rate of pond water through the treatment system, one
cannot directly compare this rate with those from the other ponds. Removal in the
treatment unit itself was as high as 90 percent initially, but decreased to about 20%
after 10 days when the influent concentration decreased substantially.
In summary, this experiment confirmed the feasibility of the countermeasure
with acclimated sludge, the possible use of unacclimated sludge, and the use of a
portable treatment unit under proper conditions. The use of the latter unit is
discussed in more detail in Section 11.
137
-------
TABLE 28. PHENOL SPILL INTO PONDS
(NOVEMBER-DECEMBER 1974)
Pond Conditions
number
Initial Initial Phenol/
phenol bacteria VSS mass
cone. cone. ratio
(ppm) (mg/1) (mg/mg)
Phenol Environmental Cond.
decay
rate
PH
DO
Temp.
(mg/1) (UC)
3 Port. trtmt. unit
6 Control
7 Accl. culture
8 Unaccl. culture
9 Unaccl. culture
Calculated.
2
Lowest D.O. value.
MODEL LAKE TESTS
165
166
153
170
148 ;
0
0
137 1.2
128 1.4
224 0.71
0.10
0.02
0.76
0.08
0.15
7.4
8.0
7.1
7.8
7.7
5.0(0.5)?
9.5(8.8),
6.0(0.0),
8.5(7.5),
s. 8(7. or
17.9
16.0
15.6
16.0
16.6
In order to assess, on a larger scale, the candidate techniques for hazardous
material spill cleanup developed in laboratory tests, model lake tests were performed.
These tests included dye tests to determine general mixing patterns of the model lake
and simulated phenol spills with or without barriers, but with bacteria. The tests
performed are summarized in Table 29, and are described below.
TABLE 29. MODEL LAKE SPILL TESTS (1976)
Spill material
Sludge
Barrier
TEST
Dye Phenol Acclimated Unacclimated with without
Spill I
(June 16-17)
Spill II
(June 22-24)
Spill III
(July 6-9)
Spill IV
(July 20-Aug.2)
x
X
Dye Spills
Rationale —
As a prelude to a phenol spill, dye was spilled in the model lake to determine
the spreading action or dispersion rate and to determine whether stratifications
existed. The bottom slope and the circular shape of the model lake perimeter, as well
as wind action on the surface of the water, could influence dispersion, while the
density of the substance spilled could cause stratification.
138
-------
Before phenol could be spilled, the best method for simulation of a spill needed
to be determined. A dye that could be detected easily and that could be used in small
enough concentrations so that it would not interfere analytically with other param-
eters was needed. The volume of dye necessary for detection had to be determined
and a technique had to be found for dispensing the dye.
To evaluate data from future phenol spills, a method for calculating mass
balance was developed using the dye tests. This permitted the recovery of the dye and
the spilled material mass to be calculated, which, in turn, enabled calculation of the
decomposition rate of spilled phenol.
Procedures --
Apparatus — The pond (Figure 42) used for the model lake is a 30.5rm diameter,
brick-lined, circular structure that was formerly a clarifier for a magnesium plant.
Along the outer perimeter, it is about 2.1 m deep, sloping to a depth of about 3.6 m
near the center. Immediately around the center pole is a trench, 0.9 to 1.2 m wide and
1 m deep, making an approximate depth at the pose of 4.6 m. The model lake had a
volume of about 1.78 x 10 liters.
A sampling grid (Figure 43) was formed using 0.64-cm (1/4-inch) sash cord
stretched across the pond, forming a 9.1-m x 9.1-m grid. Three cords were stretched 3
m apart across the pond in a north-south direction. Three more cords were stretched
in an east-west direction. Sample stations were marked with tape at the intersection
of the cords.
To sample the various stations, a 3.6-m, flat-bottomed, aluminum boat was
used. It was propelled by pulling along the sash cord grid. Water samples were taken
with a 1.0-liter Van Dorn sampler at various depths. To determine the temperature
and dissolved oxygen, a Yellow Spring Instrument Oxygen meter, Model 51A, was used.
A Turner Fluorometer was used to detect the concentration of dye present in
the water samples.
Methods —
Spill—In order to introduce the dye into the water column, a toy balloon was
filled with 300 ml of liquid Rhodamine B dye. Rhodamine B dye was used because of
the low concentrations (ppb) that could be detected on the fluorometer.
The dye-filled balloon was tied to a rope and lowered to a depth of 1.2 m at
station No. 22. The balloon was then burst at 0800 hours on the day of the test with a
broken piece of glass fastened on the end of a pole.
Sampling—Water sampling was begun one hour (0900) after dispensing the dye
and additional samples were taken at 1200, 1600, and 2200 on the day of the spill and at
0800 on the following day.
The water samples were taken from a boat at 3 depths-0.15 m and 1.2 m below
the surface and at 0.3 m off the bottom. The stations sampled at these depths were 11,
139
-------
Cross section
Top view
Figure 42. Schematic diagram of model lake.
140
-------
Figure 43. Diagram of sampling stations
141
-------
12, 13, 21, 22, 23, 31, 32, and 33. Stations A and C were each sampled from the shore at
0.15 m and 1.2 m below the surface.
The dissolved oxygen and temperature were determined at all three depths only
at station 22.
Weather observations were made at each sampling time. Approximate wind
velocity and direction and cloud cover were noted.
Plastic bottles (125 ml) were filled from the Van Dorn sampler. A total of 31
samples was taken at each sampling time, one representing each depth and station
described previously. The excess water from the Van Dorn sampler was emptied into a
large plastic can for disposal on shore away from the pond.
Analysis of samples— Before the sample could be analyzed on the fluorometer,
standard curves had to be constructed relating meter fluorescence units and dye
concentration. To do this, the following volume to volume concentrations of dye were
made: 25, 50, 100, 200, 300, 400, 500, 600, 800, 1,000, 1,200, and 1,500 (ul/1). These
were read on the fluorometer at the proper range, and a curve was plotted for each
range as shown in Figure 44. These curves were then used for determining the
concentration of dye present in each water sample.
The samples that had been collected were stored in the dark and were
temperature stabilized. At completion of the sampling, the fluorescence of the
samples was measured with the Turner Fluorometer. The concentrations (ppb) of the
samples were determined from the previously prepared standard curves.
Data analysis—Concentrations of the samples from each station and depth were
calculated. Cross-sectional drawings were made in the east-west vertical plane, the
north-south vertical plane, and the horizontal plane. On these drawings were recorded
the concentration at all stations and depths. There was a set of drawings for each
sample time. Contour lines were then drawn to indicate the various concentration
patterns.
To determine the percent recovery of spilled dye, the average concentration of
the dye in the pond at the last sample period was calculated. Next, the mass of spilled
dye was calculated, and finally, using the volume of the pond and the average final dye
concentration, the mass of recovered dye was caluclated.
Results-
Dispersion of the dye—It was possible to visually observe initial dispersion of
the dye. At approximately two hours after the spill, the dye appeared to cover the
entire pond. Based on the data plots, mixing in the pond was complete within 24 hours,
except for a slightly higher concentrated area of dye in the deepest part of the pond
(Figure 45).
Recovery of the dye— Some 300 ml of dye were initally spilled into the pond.
The average concentration of the dye at the final sampling period was found to be 195
ppb (or 0.2 iil/l). The mass remaining was 356 ml (mass (ul)= V(l) x ul/1 = 1.78 x 10 Ix
0.2 ul/1 x 10 ul/ ul). It can be concluded that little, if any, of the dye was lost during
the spill. Therefore, the use of Rhodamine B dye appeared to be a satisfactory method
of establishing a dilution baseline and of "tagging" the phenol during spills.
142
-------
100
2. 75
-------
A __
|ol54
I
11
12
13
d1 177
127
o 150
o 172
o!64
o!73
Q154
21
22
23
o200
31
32
o291
Figure 45. Pattern of. dye dispersion
33
°177
o 184
?nn
0168
o!77
-^
I o200_
0159
°159
h- 10' H
144
-------
Phenol/Dye Spill with Sludge and Without Barrier
Rationale—
In the combination phenol/dye test, dye was used to "tag" phenol (estimate
dispersion) and to establish a dilution baseline. Visual and analytical tracing of dye
dispersion indicated the location of phenol. The establishment of a dilution baseline
provided a rate with which to compare the degradation of phenol. It was assumed that
phenol would be diluted at a similar rate as the dye.
Acclimated sludge was used in the test, but without a barrier. The purpose of
the test was to simulate a natural spill situation in which no barrier was available or in
which use of such was not feasible. It was assumed that the spreading action of the
phenol/dye and of the bacteria would progress at the same rate.
Observations were made of the effect of the spill on the aquatic organisms in
the pond. The acute toxic effect on the fish could be readily observed.
Procedures-
Apparatus-- Following the previous dye test, the pond was pumped out to a
depth of approximately 0.3 m and then refilled to the original level with groundwater.
The same grid as before was used, but an additional sash cord was added in the east-
west direction 9.1 m south of the other ropes (Figure 46). A larger sampling area was
made.
The other apparatus — boat, temperature and dissolved oxygen meter, and
fluorometer — were the same as in the previous test. To detect phenol (76% of which
is carbon), the Beckman 915 Total Organic Carbon Analyzer was used. With this
instrument, total carbon, inorganic carbon, and total organic carbon present in the
samples were determined.
To determine the total and volatile suspended solids in each sample, a Millipore
vacuum filtering apparatus with Grade 934AH glass-fiber filter papers was used.
A 104 C drying oven was used to dry the samples and a 600 C muffle furnace
was used to drive off the volatile solids.
Methods —
Approximately 190 liters of activated sludge was obtained from the aeration
basin at the Govalle Sewage Treatment Plant in Austin, Texas. The sludge was
acclimated with aeration in a 208-1 drum. During the period of acclimation, nutrients
needed to sustain bacterial growth, as shown in Table 30 A and B, were fed to the
sludge daily. Phenol was fed in increasing amounts and glucose was fed in decreasing
amounts (Table 30 B) during acclimation, furnishing the carbon source necessary for
maintaining growth. Immediately before the spill, the total suspended solids and the
volatile suspended solids were determined on the acclimated sludge. Also, 17 liters of
phenol were mixed with approximately 80 liters of water in a 208-1 drum. It was
assumed that the phenol would go into solution with this water and thus go into
145
-------
Figure 46. Diagram of sampling stations, phenol dye spill with
sludge and without barrier
146
-------
solution more readily with the water in the pond. To the water-phenol mixture, 250 ml
of Rhodamine B dye were added.
The drum of phenol/dye solution and the drum of acclimated sludge were
positioned beside the pond. Using a Little Giant model 3E-12R submersible pump with
15.2 m of gar-den hose attached, the phenol/dye solution was pumped into the pond at
station 0608. The end of the garden hose was barely submerged.
After spilling the phenol, the sludge was pumped in a similar manner, spreading
it across the surface in an approximate 3.0-m radius around the boat.
Water sampling was begun one hour (0930) after the spill. Additional samples
were taken at 1230, 1630, and 2200 on the day of the spill, at 1000 on the first day after
the spill, and 0900 on the second day after the spill.
Samples were taken as before at 0.15 m below the surface and at 0.30 m from
the bottom at stations 0809, 0609, 0409, 0207, 0601, 1007, and 0611. Samples were also
taken at 0.15 m below the surface, 1.2 m below the surface, and 0.3 m off the bottom
at stations 0608, 0407, 0607, 0807, 0804, 0604, and 0404. All parameters were
measured and samples were taken at each depth and station.
Weather observations were made at each sampling time as before. The analysis
for determining the concentration of dye in each sample was done as in the previous
test. Determination of total suspended solids (TSS) and of volatile suspended solids
(VSS) was accomplished by filtering 50 ml of sample through pre-weighed, glass-fiber
filter papers with a Millipore vacuum filter apparatus. The filter papers were then
dried for one hour at 104 C and weighed to determine the TSS (mg/1). The filter paper
was then ignited at 600 C for 15 minutes and reweighed to determine the VSS (mg/1).
Filtrate from the VSS and TSS tests was used to determine Total Carbon (TC),
Inorganic Carbon (1C), and Total Organic Carbon (TOC). A 20-ul sample was injected
into the appropriate furnace port on the Total Organic Carbon Analyzer. From these
injections, TC and 1C were obtained. The difference between the two was TOC.
In order to detect toxicity effects on the organisms in the pond, visual
observations were made at the time of the first appearance of affected fish and at the
time when the last appearance of dead fish occurred.
Results —
From cross-sectional plots of the dye data, it can be seen that within six hours,
the upper portion of the ponds was uniformly mixed. However, higher levels of dye
were found near the bottom, with the deepest area concentrating the dye. Within 24
hours, mixing in the pond was complete, apparently due to a brisk wind. The dye
maintained a fairly constant concentration throughout the test.
From the background samples taken before the spill it was determined that
there were 137.2 ppb (or 137 ul/1) of dye in the pond. This represents a total dye
volume of 244 ml in the pond. Some 250 ml were added with the spill of phenol
bringing the total volume to 494 ml.
147
-------
Calculations of the average concentration of dye in the pond at the last
sampling period showed that 861 ml were recovered. This apparent excess of dye at
the test end could not be accounted for.
At the end of the first hour, the phenol settled to the bottom below the spill
station. By the 24th hour after the spill, the phenol was dispersed completely. Up to
this point the mass of phenol as TOC had shown no significant change. However, at 36
hours there was a significant increase. This was perhaps due to decomposition of
aquatic organisms killed by the toxicity of the phenol, resulting in the release of
organic carbon into the water column. Because of this analytical problem, the results
of the test were considered inconclusive. The acute toxic effect of phenol on fish in
the pond was evident. Fish began floating by the end of the first hour and affected
fish died within the first 24 hours of the test.
TABLE 30 A. SLUDGE FEEDING/ACCLIMATION SCHEDULES:
NUTRIENTS FED DAILY TO SLUDGE
Nutrient Concentration (g/JL) Amount per 190 1
K.HPO.
2 4
KH.PO.
2 4
(NH4)2S04
MgSO. . 7H-0
s 4 2
FeSO. . 7H00
4 2
ZnSO. . 7H-0
4 2
MnSO. . H_0
4 2
CaCl7
3.2
1.6
1.5
0.15
0.005
0.005
0.004
0.02
608
304
285
28.5
0.95
0.95
0.76
3.8
TABLE 30 B. SLUDGE FEEDING/ACCLIMATION SCHEDULE
Day
1
2
3
4
5
Phenol (ml/ 190 -O
19
76
152
228
300
Glucose (g/190 Ji)
178.6
142.5
95
47.5
—
Phenol/Dye Spill With Acclimated Sludge and Barrier
Rationale --
To determine the feasibility of using a barrier coupled with the biological
countermeasure, this test was conducted. First, it was necessary to know whether a
barrier would actually contain the spilled material, thus, the design features for a
148
-------
suitable barrier were considered. It was assumed that since the specific gravity of
phenol and of the bacteria was greater than water, the barrier needed to extend from
the surface to the bottom of the water column. Second, it was necessary to simulate
actual spill and counter measure application conditions, and third, it was necessary to
sample in such a way as to detect phenol decomposition, bacteria growth, and leakage
under the barrier.
Procedure —
Apparatus— The barrier was made of a polyethylene sheet (6.1 m x 10.7 m), nine
styrofoam floats (0.9 m x 0.2 m x 0.08 m), 10.7 m of 9.5-mm steel chain, and duct
tape.
The floats were spaced along the 10.7-m edge of the polyethylene sheet and
taped in place. The sheet was folded over the floats and taped to itself, the chain
was encased in the same manner along the other 10.7-m edge (bottom). The ends of
the polyethylene sheet were overlapped and taped with duct tape forming a cylinder
approximately 3 m in diameter with floats along the top and a chain (for weight) along
the bottom (Figure 47). The remaining apparatus was the same as for the previous
spills.
Methods —
Sludge was again obtained from Govalle Sewage Treatment Plant and
acclimated as in the previous test. The model lake was pumped down to within 0.3 to
0.6 m of the bottom and refilled with clean groundwater. The barrier was positioned
in the pond with the center at station 0608 (Figure
At 0830 hours, 10 ml of Rhodamine B dye was poured inside the barrier,
followed by 10 liters of liquid phenol. The water was stirred with a stirring rod during
the addition of phenol. The acclimated sludge was pumped in as before, but with
stirring as during the phenol addition. The volume of the expanded cylinder was
approximately 22.2 x 10 1. However, stirring resulted in the barrier walls collapsing
inward, making the volume inside the barrier substantially less. Pond water was
pumped into the barrier in an attempt to re-expand the walls, but was only partially
successful.
Water samples were taken using the same methods as before at 1000, 1300, and
1700 hours on the day of the spill, at 0830 and 1600 on the second day, at 0830 on the
third day, and at 0830 and 1600 on the fourth day at three depths — 0.15 m, 1.2 m, and
0.15 m from the bottom. The stations sampled were 0607, 0608, and 0609.
The results for dye, phenol, and V5S concentrations were analyzed in the same
manner as for the previous phenol/dye spill.
Results —
A large proportion of the dye remained inside the barrier throughout the test
and the dispersion pattern of the dye was very similar to that of phenol. Phenol was
retained within the barrier with some leakage under the barrier on the deep side. The
TOC data indicated a decrease from 90 ppm to 35 ppm (61%) inside the barrier within
four days (Figure 49). Since the TOC outside the barrier maintained a constant
background level except at the bottom at Station 0607, where leakage under the
149
-------
s
on
Vol. ~2.22 x 10 1
Figure hj. Barrier for phenol/dye spill.
150
-------
0607/ 0608 \0609
Figure 48. Sampling stations and barrier position.
151
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barrier occurred (see Figure 49), it was assumed that the organisms present in the
acclimated sludge were in fact decomposing phenol. The rate of decomposition
calculated from Figure 47 was 0.24 d~ . A small portion of this rate may be attributed
to dilution according to the dye results, although it is possible that the dye was also
decomposed or sorbed onto suspended solids.
Twenty-four hours after the spill, the dissolved oxygen at all depths at station
0608 dropped significantly. The dissolved oxygen at mid and bottom depths fell to
below 1 mg/1 (Figure 50), indicating bacterial utilization of the oxygen. The dissolved
oxygen at stations 0607 and 0609 maintained a normal level throughout the test.
Immediately after the spill, TSS and VSS levels in the barrier were high ( 20-50
mg/1) because of the bacteria added. After 24 hours however, only very small ( 0 mg/1)
concentrations were detected. Over the next 24- to 48- hour period, there was an
increase in TSS and VSS concentrations (to 22 mg/1) within the barrier, compared to
those levels observed outside the barrier, thereby indicating bacterial growth.
The primary problem encountered in this test was a mechanical one involving
the barrier. Stirring action caused the sides to collapse inward. However, even with
this problem, the barrier apparently did contain the phenol and bacteria.
Phenol/Dye Spill with Unacclimated Sludge and Barrier
Rationale —
Sludge that had not been acclimated was used in this spill test to determine the
feasibility of using readily available sludge directly from the sewage plant, thus
reducing the time needed for acclimation. It was assumed that there would be a slight
lag in decomposition while in situ acclimation took place, but it was thought that this
would still result in rapid decomposition of phenol.
An improved barrier with semi-rigid sides was designed and built. It was
assumed that this barrier would maintain its shape but still be maneuverable.
Procedures —
Apparatus —
The polyethylene sheets and floats on the barrier used in the previous test were
added to a skeleton constructed of 1.3-cm PVC pipe and T's. Onto each of two 4.9-m
by 1.3-cm PVC pipes, were loosely slid two T'S (1.3 cm x 1.9 cm x 1.9 cm). The ends of
these two pipes were then connected with 1.3 cm T's, thus forming a circular structure.
A second circular structure exactly like the above was made. Six 1.8-m lengths of 1.3-
cm PVC pipe were then connected to the T's of each circular structure, becoming the
rigs of the cylindrical skeleton (Figure 51). The skeleton was slid inside the poly-
ethylene barrier, positioned with the top skeletal ring just under the floats, and then
taped into place. The barrier dimensions were approximately: diameter3.0 m,
circumference-9.8 m, and height- 2.4 m. This barrier was then located with its center
at station 0608.
The remaining apparatus was the same as for the previous spills.
154
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Figure 51. Diagram of barrier with semi-rigid sides,
156
-------
Station 0608
* v Dye (ppb)
Phenol (ppm)
1 2 3 4 5 6 7 8 9 10 11 12 13 14
Figure 52. Phenol/dye disappearance in model lake tests with unacclimated sludge.
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Methods —
Since there was no acclimation of bacteria, on the day prior to the spill
approximately 190 1 (50 gal) of activated sludge were obtained from the aeration basin
at the Govalle Sewage Treatment Plant. The sludge was returned to the laboratory
and aerated overnight.
The water in the pond was pumped down and refilled with groundwater before
the spill and the improved barrier was put into place.
Immediately before the spill, 10 ml of Rhodamine B dye were mixed with 10
liters of liquid phenol and at 0830 hours this mixture was poured into the center of the
barrier. The unacclimated sludge was pumped in as before, but there was no stirring
during the spill. The sludge was spread about the surface with a hose to insure
coverage of the spill area.
Water samples were taken on July 20, 1975 at 0930,1200, and 1600 hours, on July
21 at 0830 and 1600 hours, and on July 2 through August 2 (omitting July 31 and August
1) once daily at about 0900 hours. Samples were taken at the same depths and stations
as in the previous test.
Results —
The dye remained primarily within the barrier as in the previous test and it was
assumed that the phenol was also contained. The rate of decrease of phenol (measured
as TOC) was substantially less than in the previous test. At the end of the eighth day,
the concentration of phenol (TOC) had decreased by about 50% (Figure 52) and by the
end of the fourteenth day there was a total decrease of 67%. In contrast, in the
previous test using acclimated sludge, it took four days to obtain 50% reduction in the
phenol concentration. The rate of phenol reduction was d based on the data shown in
Figure 51. It was also apparent that the dye decreased at approximately the same
rate. Based on the sampling data at stations 0607 and 0609 outside the barrier, no
significant amount of phenol or dye was lost under the barrier, thus the decrease in
dye concentration cannot be attributed to dilution. With this conclusion about the dye,
it was also concluded that the phenol was not diluted, but was decomposed.
By the third day of the spill, the dissolved oxygen at all depths of station 0608
reached a concentration of less than 1 mg/i (Figure 53) and this level was maintained
for a period of 48 hours, indicating a high level of bacterial action. With the exception
of the eighth and eleventh days, there was a steady increase in DO from the fifth day
to the end of the test. The two exceptions could be the result of a response to cloud
cover, which would decrease the production of oxygen by algae in the system. The
steady increase in DO indicates recovery of the system from the effects of the spill.
Stations 0607 and 0608 maintained a high DO concentration as would be expected in a
normal system.
The barrier used in this test proved very satisfactory. It maintained its shape
well and, as previously discussed, apparently contained the spilled material.
159
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SECTION 10
COUNTERMEASURE STORAGE
One of the deficiencies of the biological countermeasure often mentioned is the
problem of storage. It had been assumed (Dawson et al., 1972) that liquid storage
would be necessary, but storage in a dry, powdered form is also possible. The
experimental work described in this section focused on ^showing the feasibility of
storage methods using freezing and lyophilization as two techniques of phenol storage.
PRESERVATION AND RECOVERY OF A MIXED BACTERIA CULTURE
When sewage sludge was inoculated into a buffer (Table 31) supplemented with
. 500 ppm of phenol, a mixed culture developed, which after several transfers, consisted
largely of a Pseudomonas sp. .The preservation of this phenolutilizing culture was
tested by inoculating 2.8 x 10 colony-forming units (CPU) of culture into 10-ml
solutions of buffer/salts medium with the following additions: (1) no addition, (2) 5%
glucose, (3) 5% glutamate, and (4) 5% peptone. Each solution was then divided into
two 5.0-ml samples, one for preservation by freezing, the other for preservation by
lyophilization. One percent cellulose was added as a stabilizing agent to the cultures
to be lyophilized.
TABLE 31. BUFFER-SALTS MEDIUM
Salt
MgS04.7H20
NaMo 04
tfTf r\f\
KH2 ^4
Na2H P04
Ca C12
Fe Cl-
NH4 Cl
Cone, (mg/1)
112.5
5.0
2.5
680.0
700.0
27.5
0.5
2600.0
All samples were frozen over a period of 30 minutes in a - 20° C freezer. The
samples to be test-preserved by freezing were left at - 20 C; the samples to be
lyophilized were removed from the freezer and placed in dry ice-acetone for 30
seconds, then vacuum-dried overnight, flame sealed, and stored at room temperature.
After one week the 5.0-ml frozen samples were thawed and diluted by the addition of
5.0 ml of 0.5% peptone in buffer-salts solution. The freeze-dried samples were
rehydrated by the addition of 10.0 ml of 0.5% peptone in buffer-salts. ATP
160
-------
determinations were made at one and at four hours after rehydration or thawing; plate
counts of colony-forming units were done at 30 min after removal of the cultures from
storage conditions. The results obtained are recorded in Tables 32 through 35.
TABLE 32. SURVIVORS OF FREEZE-STORAGE
QUANTITATED BY PLATE COUNTS
Additions
None
5% glucose
5% glutamate
5% peptone
Remaining
organisms
(CFU/ml)
3.4 x 108
3.8 x 108
2.6 x 108
3.4 x 105
% Survival
CFU/ml
survived 1
24.0
27.0
19.0
0.024
9
Before freezing, 1.4 x 10 colony forming units per ml (CFU/ml)
were present.
TABLE 33. SURVIVORS OF FREEZE-STORAGE QUANTITATED
BY ATP DETERMINATIONS
Fg ATP/ml remaining Fg ATP/ml remaining
Additions 1 hr. after thawing^>2 4 hr. after thawing!>
None
5% glucose
5% glutamate
5% peptone
5.35 x 108
5.84 x 108
1.91 x 109
6.60 x 107
4.71 x 108
3.64 x 108
9.17 x 108
4.68 x 107
1 9
Before freezing 1.4 x 10 Fg ATP/ml were present.
2 —12
Fg = femtogram (10 g).
161
-------
TABLE 34. SURVIVORS OF LYOPHILIZATION AS DETERMINED
BY PLATE COUNTS
Concentration
Additions
None
5% glucose
5% glutamate
5% peptone
(CPU/ml)
<102
2.3 x 108
1.7 x 108
-------
Colony-forming units were determined by dilution of the seven test cultures in buffer-
salts medium and plating in triplicate on nutrient agar plates supplemented with 500
ppm phenol. The results obtained are presented in Table 36. These data show little
significant difference among the rehydration media tested and indicate that
lyophilized phenol-utilizing mixed cultures may be successfully rehydrated in phenol-
containing medium.
It was also of interest to investigate the recovery of energy-generating ability
by a lyophilized mixed culture of phenol-utilizing microorganisms. A culture that
contained 4.7 x 10 Fg* ATP/ml was lyophilized and stored at room temperature as
previously described. After one week the vial was opened and the culture rehydrated
with 10 ml of 500 opm phenol in buffer salts. The 10-ml fluid sample was incubated
with shaking at 25 C in a capped 18-cm (7-in) test tube and 0.1 ml samples were
removed immediately and at intervals of 1.5, 3, 4, and 8 hours for ATP determination
in a DuPont Luminescence Biometer. The results of this experiment show an initial
drop in ATP concentration followed by a steady rise from 1.5 to 8 hrs. This may
indicate that after an early period of recovery, during which energy generation is
reduced, ATP formation or cell number rapidly increases.
TABLE 36. THE EFFECTS OF VARIOUS REHYDRATION FLUIDS ON RECOVERY
FROM LYOPHILIZATION BY PHENOL-UTILIZING MIXED CULTURES
Concentration
Additions (CFU/ml)
9
Demineralized water 1.3 x 10
Buffer-salts with:
a
No additions 1.3 x 10
100 ppm phenol 1.2 x 109
9
500 ppm phenol 1.4 x 10
9
1% peptone 1.7 x 10
9
1% glutamate 1.0 x 10
9
1% glucose 1.4 x 10
PRESERVATION AND RECOVERY OF A YEAST CULTURE
A phenol-utilizing yeast strain, Geotrichum candidum, was grown in a defined
medium consisting of 0.02 M phosphate, 1.0 g/1 NhL Cl, and small amounts of other
inorganic salts supplemented with 500 ppm phenol as the sole carbon source. After 24
hours of incubation, 0.01% yeast extract was added to the culture and incubation was
continued for another 24 hours. Cells were harvested by centrifugation, washed, and
* Fg= femptogram (10 gm)
163
-------
suspended in a 50% dilution of the mineral-salts growth medium or dilute mineral-salts
medium supplemented with 5.0% glucose. Samples of each suspension were frozen and
lyophilized. Another set of 0.5-ml aliquots were placed in petri dishes and dried under
vacuum without prior freezing. Plate counts were made on samples of the suspension
before preservation.
Samples were rehydrated in mineral salts medium supplemented with 500 ppm
phenol, incubated with shaking for 45 minutes, diluted, and plated on mineralsalts agar
supplemented with 500 ppm phenol. Table 37 shows the CPU's in the two
circumstances and confirms the suspicion that yeast do not survive lyophilization as
well as bacteria.
TABLE 37. SURVIVAL OF PHENOL-UTILIZING GEOTRICHUM1
Lyophilization Dried from liquid
No addition 5% glucose No addition 5% Glucose
CFU/ml
% Survival
2.7 x 106
0.81
7.3 x 107
22
2.1 x 106
0.64
1.3 x 108
39
1 8
Initial count was 3.3 x 10
164
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SECTION 11
COUNTERMEASURE APPLICATION
The feasibility of biological countermeasure methods for the removal of phenol
and methanol is largely dependent on the amount of acclimated cultures at hand.
Phenol and methanol are products or by-products of some manufacturing processes and
are inevitably discharged in waste streams. If these wastes are biologically treated,
the treatment plants are sources of acclimated cultures. Acclimated bacterial
cultures may be stored in a frozen or powdered form in a dormant state for later use
(Armstrong et al., 1974). However, these storage problems are beyond the scope of
this study and this chapter only investigates the use of acclimated cultures for the
actual abatement of phenol and methanol spills. To improve usefulness of the
cultures, some auxiliary devices are employed and their mechanisms are studied.
Owing to the experimental and time limitations, some of this work is largely theo-
retical. However, all of these theoretical developments are soundly based on the
experimental results or accepted theories, or are directly or indirectly proved by the
laboratory-scale tests. For example, the material exchange mechanism through cloth
bags is investigated based on the experimental results. Then, the cloth bag efficiency
and the substrate removal kinetics using cloth bags are theoretically developed from
the material exchange mechanism. Finally, theoretically developed substrate removal
kinetics are used to predict the experimental results within a five percent error.
EXPERIMENTAL METHODS
Equipment and Reagents
In addition to the equipment and reagents used in the biological treatability
tests, the following chemicals and devices were employed.
Rhodamine B dye was used for the purpose of visual tagging of the material
being transported.
A model river was used for material transportation and bulk sludge application
tests, tests for material exchange through cloth bags, and confining-barrier application
tests. The model river consists of two parallel channels (Figure 54); each is 38.1 cm
wide, 45.72 cm deep, and 60.95 m long and has a 10- to 15-cm layer of sediment on the
bottom. The bottom sediment was transplanted from Lake Austin in 1962 and has
grains with a geometric mean diameter of 0.173 mm and a geometric standard
deviation of 1.70 mm (Kludo and Gloyna, 1969). The flow is controlled using a
constant-head tank, inlet valves, and V-notched weirs. Water depth is regulated with
an adjustable outlet weir. Groundwater is supplied to the model river from a 1.9x10^-1
supply tank.
165
-------
-p
w
>.
CQ
!H
-------
Cloth bags were employed as a means of sludge containment. They were made
of cotton cloth with an average pore diameter of about 0.2 mm and were cylindrical in
shape with dimensions of 3.18 cm (diameter) and 25 cm (length) and were kept fully
expanded with wire-cloth frames. Styrofoam was attached to the top of the bag as a
floating aid and had a small hole (about 1 cm diameter) in the middle for the purpose of
sludge injection and sampling. Details are shown in Figure 55.
Confining barriers were used to prevent dispersion of spilled materials. They
were made of vinyl, were cylindrical in shape, and had a 27-cm diameter and a 40-cm
height with an open top and bottom. A wire frame support maintained the cylindrical
shape. Styrofoam was attached to the top of the barrier when flotation was desired
and weight was applied to the bottom when a fixed barrier was desired. Details are
shown in Figure 56.
i
Procedures
Material Transportation and Application of Bulk Sludge in the Model River —
Phenol and methanol were spilled in the model river with and without
acclimated activated sludge. The sludge was sonicated for several seconds to reduce
floe size for the purpose of improving sludge floatability. To detect material
transport visually, a minute amount of Rhodamine B dye was added to the phenol and
methanol. Samples were taken at various distances downstream from the point of
spillage at certain time intervals and were analyzed for TOC and VSS. TOC and VSS
contributed by the flume water was subtracted from the TOC and VSS measurements
to delineate the TOC due to phenol or methanol alone and the VSS contributed by
acclimated sludge alone. TOC due to the Rhodamine B dye was negligible. The model
river flow was 10.0 liter/min and the water depth was 15.8 cm, giving a 603-cm cross-
sectional area and a 0.276-cm/sec velocity.
Material Exchange through Cloth Bags —
Material exchange rates through the cloth bags were measured in batch
reactors and in the model river. Elutriated and settled sludge was in-activated with 10
mg/1 of silver, using a silver nitrate solution, and was mixed with known amounts of
phenol or methanol. A small amount of Rhodamine B dye was added to the sludge to
detect the material transport visually and to plan the sampling times.
In batch reactor tests, cloth bags were filled with sludge while slowly immersed
into 3-liter reactors. Five reactors were aerated at an air flow of 1 liter air/min/liter
water to provide turbulence and five other reactors were maintained in a quiescent
condition. These tests were carried out for both phenol and methanol sludge. The
water outside the bags was sampled at pre-set time intervals for TOC analysis. The
water volume decrease outside the bags caused by sampling was accounted for in the
data analysis.
In the model river, the cloth bags were filled with methanol-acclimated sludge
prepared in the previously described manner while they were gradually immersed in
such a way that the hydrostatic pressure inside the bags did not allow the flume water
to flow into the bags. The mixed liquor inside the bags was analyzed for TOC at the
end of each test. The test flume velocities were 0.106,0.191, 0.575, 1.551,2.250, and
2.731 cm/sec. Three or four bags were used for each test velocity. Rhodamine B dye
solution was used to visually observe the flow regimes.
167
-------
1 cm
Hole
Floating aid
(Styrofoam)
Wire-cloth
frame
Cotton cloth
(Pore diameter
= 0.2 mm)
Figure 55. Details of cloth bag.
168
-------
Floating aid
(Styrofoam)
27 cm
cm
Wire frame
Weight
for fixed
barrier
Figure 56. Details of confining barrier.
16.9
-------
Cloth Bag Application —
Three-liter batch reactors were filled with 30 ml each of Nutrient Solution I
and the Mineral Solution. Cloth bags with elutriated settled sludge were immersed in
the reactors, which were aerated at room temperature. They were fed with 700 mg/1
of methanol (as TOC) and waters outside the bags were sampled at pre-set time
intervals for TOC analysis. To calibrate methanol loss by stripping, one reactor was
operated without a cloth bag.
Confining Barrier Application
Floating barrier— The model river velocity and depth were adjusted to 0.85 cm/sec
and 35.2 cm, respectively. The floating barrier was immersed in the flume so that the
clearance between the barrier and the bottom of the flume was less than 3 cm.
Rhodamine B dye solution was spilled inside the barrier and color intensity was
observed visually. Thirty minutes after the dye spill, the color intensity inside the
barrier approached the same level as that outside the barrier. Because of this poor
containment efficiency, no further tests were done with the floating barrier.
Fixed barrier— Under the flume conditions described above, phenol and
methanol were spilled in the model river. The fixed barriers were dropped at the spill
site. Cloth bags with known amounts of acclimated activated sludge were applied
inside the barriers and air was supplied. The water outside the bags was sampled and
analyzed for TOC. TOC contributed by the flume water was subtracted from the TOC
measurements. For the barrier efficiency study, cloth bags and air were not supplied.
ORGANIZATION OF BIOLOGICAL TREATABILITY DATA FOR COUNTERMEASURE
DESIGN
The procedure for obtaining information that is needed in setting up counter-
measure plans for biological treatability can be summarized as follows.
1 Examine the characteristics of water to be treated (spill concentration, pH,
alkalinity, salinity, temperature etc.).
2. From Figure 20 and Table 10, or Figure 31 and Table 16, estimate the
equivalent pH for treatment or average pH and efficiency of dilution water if no
chemical aids will be used.
3. From iso-f diagrams at the nearest temperature, Figures (10, 11, 12, 23, 24, or
25 ) evaluate f corresponding to the given pH and salinity. Then,
kT0 = f k(pH_7, sal- ppt)*
is the efficiency of dilution water.
4. From iso-9 diagrams (Figs, 13, 26, or 27), find 9. Then, k-r = kj 0(T-TO).
5. Evaluate kj from kj = 0.066 k^-87 for phenol and from ^ . _ Q.0115 k^'634 for
methanol.
170
-------
6. Evaluate b1 from b1 = 1.42 k
where: Q = flow rate (L3/T),
V = reactor volume
S0= influent substrate concentration (M/L^), and
Se= effluent substrate concentration (M/L^).
At steady state (ds/dt=0) the total biomass required to achieve substrate reduction
from S0 to Se may be found from:
VX = fo(S0-SP)-( kgVSeXKs+Sg)] (77)
kSe
V f = QXQ - (Q-Qw)Xe+ aQ(So-Se- k^ V ) . kjx . Q^, (78)
where: XQ= influent biomass concentration (M/L ),
X = effluent biomass concentration (M/L ),
Q = sludge waste flow rate (L3/T), and
3
X = biomass concentration in sludge waste flow (M/L ).
inr
If influent and effluent solids are negligible, then:
aS(S-S)-akS-kX-wX. (78')
At steady state ( -rr- = 0) the sludge waste required to achieve the total biomass (VX) in
the reactor is:
171
-------
Influent
Q
So> Xo
Air
Aeration tank
V
Se, X
Return sludge"
Sedimentation tank
Q-Qw
Effluent
Se, Xe
Qw
Se, Xw
Figure 57. Essential parts of a CSTR system.
172
-------
QWXW= a-fo(S0-Se) - keVSeV kdVX.
J (79)
The maximum achievable total biomass in the reactor is: a/k^ {Q(SO - Se) - kgVSe}
which is obtained when no sludge is wasted. In order to prevent sludge washout, sludge
waste should be less than a£2 (So - Se)-keSeVJ(see Figure 58).
The total oxygen requirement at steady state is:
Rr V = a'Q(S0- Se) - keSeV - b'XV . (80)
Design of Batch Treatment System
Batch treatment systems do not necessarily require sophisticated facilities.
Any containers can be turned into batch reactors, if it is necessary. Another
advantage of batch treatment systems may be that the effluent quality can be
controlled very easily. Thus, batch treatment methods offer a highly promising
solution for abatement of phenol and methanol spills.
Batch systems can be designed using a numerical method. The time required to
reduce the substrate concentration from S[[_\ to Sj and the biomass concentration at
an i-th station are obtained from Equations 62 and 69,
respectively, or:
(81)
a.f
and
(82)
Superscripts and subscripts are the iteration indexes and station indexes, respectively.
In Equations 81 and 82, X.and t-l are used in calculating t.and X., respectively.^ Thus,
iteration is required to solve for these two values. In order to get started, X^°'can be
set equal to Xj_j. The iteration can be stopped when the difference between xp1"1' and
jq(n-l) is negligibly small.
When stripping loss is not involved in the reactor, Equations 81 and 82 are
simplified to:
(83)
173
-------
o
o
d)
0)
-C
4->
C
1/1
CO
ro
f Q (So - Se) - ke VSe
VX = | Q (So - Se) - ke VSe
Q (So - Se) - ke VSe
Waste sludge (Q
w
Figure 58. The relationship between waste sludge and total
biomass in the CSTR.
USA
-------
and
Since X.is directly solved from Sj without the aid of t., there is no need for iteration
when Equations 83 and 84 are concerned.
Equations 81 and 84 are based on the approximation,
k (y. + x •">
N l Ai-i t-j
tH
Therefore, the S increment, (S. - S. .), should be selected so that it produces very
small relative changes in X, calculated as (X.-X. i)/X. ..
/ I />-<• ^-* 1"™* 1"~1 \
When the S increment was satisfied(KSj_~Sj_jX/Xi-± ^ ^/^2.OC>) and
(| (Si-Si-i)/Si-i|£l/"lo)stable results were produced.
Computation results were utilized in developing the batch kinetic diagrams for
phenol for k values of 0.01892, 0.02729, 0.3934, 0.05674, and 0.07239 hr"1 at dilute and
concentrated VSS concentrations as shown in Figures 59-1 to 59-10. The following
example explains the use of these diagrams.
Example —
Problem—500 mg/1 TOC of phenol waste is to be treated in a batch reactor
under environmental conditions that give k = 0.0645 hr~ . What is the aeration time
required to obtain 99% phenol removal and what will be the final biomass concentra-
tion if 2,000 mg/1 of VSS is available initially?
Answer—Use the batch kinetic diagrams that bracket the k value to be used.
Draw VSS lines starting from 500 mg/1 TOC and 2,000 mg/1 VSS in diagrams for k =
0.05674 h?l and k = 0.07239 hr~l (Figures 59-8 and 59-10). Then, read the VSS's at 5
mg/1 TOC and the times of the intersecting points of TOC and VSS lines at the initial
and final points.
At k = 0.05674 hr"1 : (1) VSS at 5 mg/1 TOC = 2,460 mg/1, (2) time at 500 mg/1
TOC = - 6.1 hours, (3) time at 5 mg/1 TOC = 5.8 hours, and (4) time required for 99%
removal = 5.8 - (-6.1) = 11.9 hrs.
At k = 0.07239 hr -1: (1) VSS at 5 mg/1 TOC = 2,460 mg/1, (2) time at 500 mg/1
TOC = -5.0 hours, (3) time at 5 mg/1 TOC = 4.5 hours, and (4) time required for 99%
removal = 4.5 - (-5.0) = 9.5 hrs.
Then, the time required for 99% phenol removal at k = 0.0645 hr~ is, by interpolation,
9 5 i (11 ? 9 5) °-07239- °-06^. -10 7 hrs
?.;> + ui.* 7.V 0.07239 -0.05674 ~ 1U
-------
2000
1000
700
500
300
200
100
70
50
TOC
(rag/1)
30
20
10
7
5
3
2
200
400
600 800 1000
VSS (mg/1)
1200
1400
Figure 59-1. Batch kinetic diagram for phenol for dilute
VSS at k = 0.01892 hr"1 .
175
-------
2000
1500
2000
2500 3000
VSS (mg/1)
3500
4000
Figure 59-2. Batch kinetic diagram for phenol for concentrated
VSS at k = 0.01892 hr"1.
176
-------
0000
200
600 800
VSS (mg/1)
1000
1200
1400
Figure 59-3. Batch kinetic diagram for phenol for dilute
VSS at K - 0.02729 hr"1.
177
-------
2000
1500
2000
2500 3000
VSS (mg/1)
3500
4000
Figure 59-4. Batch kinetic diagram for phenol for concentrated
VSS at k = 0.02729 hr"1.
178
-------
40
200
400
600
800
VSS (rng/1)
1000
1200 1400
Figure 59-5. Batch kinetic diagram for phenol for dilute
VSS at k = 0.03934 hr"1.
179
-------
2000
1000
700
500
300
200
100
70
50
TOC
(mg/1)
30
20
10
7
5
3
2
200
400
600
800
VSS (mg/1)
1000
1200
1400
Figure 59-7. Batch kinetic diagram for phenol for dilute
VSS at k = 0.05674 hr"1.
181
-------
2000
1000
700
500 -
300
200
TOO
70
50
30
20
10
7
5 -
1500
2000
2500 3000
VSS (mg/1)
3500
4000
Figure 59-8. Batch kinetic diagram for phenol for concentrated
VSS at k = 0.05674 hr"1.
182
-------
2000
200
400
600 800
VSS (mg/1)
1000
1200
Figure 59-9. Batch kinetic diagram for phenol for dilute
VSS at k = 0.07239 hr"1.
1400
183
-------
2000
1000
700
500
300
200
100
70
50
TOC
(mg/1)
30
20
10
7
5
3
2
1500
2000
2500 3000
VSS (mg/1)
3500
4000
Figure 59-10. Batch kinetic diagram for phenol for concentrated
VSS at k = 0.07239 hr"1.
184
-------
When three hours of aeration lag time by initial lag phase are taken into consideration,
the total aeration time required will be about 14 hours and the final VSS concentration
will be 2,460 mg/1.
Oxygen requirements for this system change with time. S and X at a given time
are obtained from the batch kinetic diagrams and oxygen requirements at the time are
solved from Equation 58 or 73.
Comparison of CSTR and Batch Systems
The biological decomposition rate in the CSTR system is kXSe/(Ks+Se) and in
the batch system is kXS/(Ks+ S). Changing the fractions to a common denominator,
one gets ((fgX(Ks Se + S Se))/((Ks+Se)(Ks+ S)) for the CSTR system and ((UX(KSS +
SSe))/((Ks+ Se)(Ks+ S)) for the batch system. Since S in the batch reactor is always
greater than or equal to the design effluent quality, Se, the removal rate in the batch
reactor is always better than that in the CSTR.
When there is no stripping loss, the hydraulic detention time T , required to
achieve effluent quality, S , in the CSTR is:
sft)(K.+sp) (85)
k x^
In the batch reactor, Tbis solved from Equations 83 and 84. Let the relative
efficiency of the CSTR, r, denote the ratio of , to , or:
= t/t .
(86)
The relative efficiency of the CSTR turned out to be mainly dependent on the
influent to effluent quality ratio, S /S . As influent concentration increases the
relative efficiency of CSTR sharply decreases. If an effluent quality of 5 mg/1 TOC of
phenol is required with 3,000 mg/1 of initial VSS available, the efficiency is 0.7 for 10
mg/1 influent, 0.27 for 50 mg/1 influent, and 0.05 for 500 mg/1 influent (see Figure 60).
Thus, for highly concentrated spills, the batch systems can be considered as a
necessity rather than as a preference.
Biological countermeasures for accidental spills may require as much
acclimated activated sludge as is available. Batch systems also have an advantage
over CSTR systems in this respect. The sludge production rate in a CSTR, Pr , is:
(S- S6) - akeVSe- kaV X
and that in a batch reactor, Pr. , may be expressed as:
Pr= oV (S0-Se)-
(88)
'o
If the stripping loss and endogeneous respiration (second and third terms of the
equations) are neglected, the ratio of sludge production rate in a CSTR to that in a
batch reactor becomes Tb/Tc, which is identical to the relative efficiency of a CSTR.
185
-------
I I I I I I I I
3 5 10 30 50 100
Influent concentration (mg/1 TOG)
Effluent concentration (mg/1 TOC)
300
Figure 60. Relative efficiency of CSTR in phenol removal
(for effluent quality of 5 mg/1 TOC).
186
-------
Therefore, batch systems produce more sludge.
The following may be considered as disadvantages of batch systems:
1. Batch systems are more susceptible to shock loading.
2. Initial lag periods need to be considered.
3. Rapid substrate removal may cause an unfavorable pH range.
4. Sludge production is a nuisance when it is to be disposed.
5. When consecutive batch treatments are required, there are time losses
between each batch, owing to the draining and filling of the reactors and the sludge
return.
Methanol-acclimated activated sludge turned out to be highly dispersive and
this may impose a serious problem in sludge return when there is no special treatment.
Sludge-containing cloth bags can be an easy solution for the sludge return. They can
curtail considerable time loss between each batch treatment. Functions of cloth bags
are discussed in detail later in this chapter.
IN SITU APPLICATIONS OF THE BIOLOGICAL COUNTERMEASURE
Application of Bulk Sludge in One-Dimensional Dispersion System
When spills occur in rivers and the use of portable treatment facilities is
restricted, bulk sludge application methods may be a solution for the treatment of
spills. As usually practiced, the dispersion in rivers may be considered one-
dimensional.
The one-dimensional dispersion equation is expressed as:
£s.
ax
where: 5 = concentration at a distance x and time t (M/L ),
u = mean flow velocity (L/TLand
D = dispersion coefficient (L /T).
For a point source injection of a substance, the concentration distribution has been
solved to be (Taylor, 1954; Harris, 1963):
- (X-Ut
<90)
where: M = total mass of substance injected (M) and
S 2
A = the cross-sectional area of channel (L ) .
187
-------
Fischer (1966) reported that the mechanism of initial diffusion, characterized by
tail effects, was markedly different from that described by Equation 90, so that
Equation 89 was the only effective formulation after the initial period. However,
phenol and methanol transport tests did not show any evidence of Fischer's theory.
The dispersion coefficients were estimated using the Fischer's moment method:
(91)
£. «-> v
where:
1 ' (92)
* X - ut . (93)
Thus, the dispersion coefficients estimated were not affected by the initial
diffusion mechanism, if there was any. The computed concentration distributions from
Equation 90 satisfactorily agreed with those observed (Figures 61 and 62). The
skewness of the distributions is thought to be attributed to the winds rather than to
the initial mixing mechanism. The model river ran from north to south and a southerly
wind prevailed throughout the test periods. Test flow velocities could not overcome
the winds and the surface water flow was largely governed by winds. During the
phenol transport tests, there was a temporary northerly wind, which gradually changed
to a mild southerly wind. Thus, there were reverse tail effects in the beginning, which
is contrary to Fischer's theory. When the flow rates were extremely small, phenol
sank to the bottom, reproducing the so-called tail effect, and methanol floated with
arbitrary distributions governed by the winds. Therefore, Equation 90 is thought to be
effective in prototype channels on which the winds have little effect.
When dispersive acclimated activated sludge is deployed over the spills accord-
ing to the pollutants dispersion model, the distribution function of both sludge and
pollutants may be approximated as Equation 90. Then, the total mass removal rate
can be expressed as:
dx-f~Arkrsix ,
J_ '
where k is the pollutant removal rate coefficient (time" ) contributed by volatilization
and bottom sorption, etc, (other than biological reactions). Equation 94 holds true only
when there is enough oxygen supply. In most streams, aeration is restricted and the
bulk sludge application method cannot be efficient for highly concentrated spills.
When the spilled material concentration at peak point is smaller than K and when A is
constant, Equation 94 may be simplified to:
- f°AfKrScbc
~u
188
-------
II
4->
C
o
Q.
Q.
1/1
O>
-o
o
O)
ra
4J
s_
o
Q.
to
O
c
cu
-
o
-------
4-
o
O
O.
O.
oo
s-
O)
S-
o
Q.
CO
(O
s_
O
c
«3
-e
O)
O)
5-
0>
O
O
O
CTi
O
CO
O
Lf)
O
oo
O
C\J
(L/6ui) uoqjBO
190 :
-------
<95>
where M is the total mass of VSS applied. At low substrate concentrations, bacterial
growth Inters the stationary phase (see Fig. 29), so that M may be considered
constant. Then, the total mass of pollutant at time t is solved by integrating Equation
95. If tit ,
where: M = total mass spilled,
M = total mass at time t (measured from the spillage), and
t = sludge application time.
cL
And, if
A distinct disadvantage of this method can be characterized by the term^ft*
involved in Equation 96. Without enought acclimated activated sluge to compensate
for this disadvantage this method may require an unfeasibly long time, as implied by
the t within the exponential term.
Unfortunately, most of the sonicated phenol sludge applied over the spills
settled down immediately so that there was no measurable amount of suspended
organisms left and there was no measurable phenol loss during the transport (k = 0).
More than 60% of the sonicated methanol sludge settled within 45 minutes, and the
rest, 39 grams of VSS, remained suspended throughout the test time. There was
methanol removal during the transport, but, it was not possible to conclude how much
was due to biological degradation (see Figure 63).
Application of Cloth Bags
The usefulness of acclimated activated sludge can be increased by reducing the
amount of idle sludge, which is a byproduct of sludge settling and return. If sludge-
containing cloth bags are introduced into the reactors, this process is not necessary.
Floating cloth bags can also be used for in situ treatment to prevent sludge settling
when the mixing intensity is not great enough for sludge suspension. The cloth pores
are rapidly clogged with bacterial floe when the bags are filled with concentrated
sludge. Through the cloth, liquid exchange is free while the transport of bacterial floe
is greatly inhibited.
191
-------
60
CJ
o
H
60
00
c
to
0)
o
9
(II
CO
M
(0
O
H
50
30
20
10
With sludge
(39 gram VSS)
Without sludge
23
Time (hr)
Figure 63. Methanol removal in one-dimensional dispersion system
(Temp. = 22°C).
192
-------
Material Exchange through Cloth Bags—
The material transport rate into a cloth bag will be proportional to the surface
area of the bag and to the concentration gradient. Thus, the following is possible:
(s-se)
(98)
where: V = volume of cloth bag (L ),
3
S = concentration outside the bag (M/L ),
S = concentration inside the bag (M/L ),
2
A = surface area of the bag (L ), and
C = material exchange coefficient (L/T).
The material exchange coefficient, C, was estimated using the following methods.
In a confined reactor, the total mass, M, of some substance does not change if
the substance is injected inside a cloth bag with inactivated sludge. This mass balance
may be described by:
V + VcSc = M' (99)
where V is the volume outside the bag. Thus Equation 97 becomes:
.da.- _£&.( M-V0S-VCS6) . ,. .
dt V0Vt C (100)
Let W denote the difference between the equilibrium concentration and the
concentration outside the bag at time t. Then,
W = M _ S, (101)
where,
Then, Equation 99 becomes:
V = VQ + V . (102)
dW= C A W
dt V ^
Integration of Equation 103 from tirrfe i, to t_ yields:
In (W, /
ACV ( V t,
In a dispersive stream, the concentration outside a bag is negligible if a
substance is injected inside the bag, which is filled with inactivated sludge. Therefore,
193
-------
(105)
Integration of Equation 105 from time t. to t- yields:
C = V^ In ( Scl / S_2 )
Ac < t2 ~ c]?
The estimated material exchange coefficients using Equations 104 and 106 are
given in Tables 38 and 39. The coefficients for methanol sludge were twice those for
phenol sludge. Even though the concentration of methanol sludge increased from 4,200
mg/1 to 6, 760 mg/1 (as suspended solids) owing to the sludge settleability, it did not
have any perceptible effect on the material exchange rate. In the laminar flow
regime, the material exchange coefficient remained at the same level observed in the
quiescent condition. The material exchange coefficient showed a linear relationship
with velocity in the turbulent flow regime when the flume depth was kept constant
(see Figure 64).
Like other phenomena of hydraulic mixing and stirring, the rate of material
exchange through cloth bags can also be stated as a function of the mean temporal
velocity gradient. Therefore, the following formulation is possible:
C = Cj Gc2, (107)
where:
c., c- = constants,
G = mean temporal velocity gradient, or mixing intensity
p = density of water (gram/m ),
g = gravity constant (m/sec ),
LL = absolute viscosity of water (gram/m/sec),
h, = head loss in a given stretch (m), and
t , = detention time in the stretch (sec).
194
-------
o
o
o
<£
14-
O)
o
o
en
to
.c
o
X
QJ
tO
S-
OJ
0
0
1.0 2.0
Model river velocity, u (cm/sec)
Figure 64. The relationship between stream velocity and material
exchange coefficient for methanol sludge.
3.0
195
-------
TABLE 38. MATERIAL EXCHANGE COEFFICIENT RELATED TO
THE TURBULENCE IN CONFINED REACTORS
Sludge
Phenol
Methanol
Suspended
solids
(mg/1)
26,130
4,200
Volatile
S S
(mg/1)
23,720
3,860
Turbulence
Quiescent
Turbulent
Quiescent
Turbulent
Exchange
coef . ,
(cm/hr)
0.0710
0.684
0.141
1.350
C, Number of
observations
20
20
20
20
Turbulent conditions were made by aeration through glass tubes with an air
flow rate of 1 liter air/min/liter water. The turbulence caused by this air
flow rate was just enough for complete suspension of phenol-acclimated sludge
without cloth bags.
TABLE 39. MATERIAL EXCHANGE COEFFICIENT RELATED TO THE STREAM VELOCITY
IN THE MODEL RIVER (METHANOL SLUDGE )
Velocity Exchange coef. Flow ,
(cm/sec) Reynolds No. (cm/hr) regime''
Number of
observations
0.106
0.191
0.575
1.551
2.250
2.731
440
800
2,400
6,500
9,400
11,400
0.155
0.131
0.133
0.256
0.349
0.425
Laminar
Laminar
Turbulent
Turbulent
Turbulent
3
3
3
3
4
4
SS = 6,760 mg/1, VSS = 6,270 mg/1, temperature = 28 °C,
hydraulic radius = 0.124 m.
Flow regimes were visually observed using a Rhodamine B dye solution.
If Manning's formula is employed to estimate the head loss and the detention
time, c- should be 2/3 in order to describe the linear relationship between the
exchange coefficient and the velocity in the turbulent regime. Manning's formula is
expressed as:
u=il1/2R2/3,
where:u = mean stream velocity (m/sec),
n = coefficient of roughness,
I = energy gradient, and
R = hydraulic radius in m.
196
(108)
-------
Therefore,
If the coefficient of roughness, n, of 0.025 is selected for a test channel with
moderate weeds (Davis, 1952), the following is proposed for prediction of the material
exchange coefficient in the turbulent flow:
C = 0.159 G2/3 (for phenol sludge) and (110)
C = 0.317 G2/3 (for methanol sludge), (111)
where C has the units cm/hr, and G, sec" .
Reaeration in streams is also thought to be a result of hydraulic mixing. Camp
and Meserve (1974) brought together a number of observations on large rivers and on an
experimental channel and interpreted reaeration as a function of the mean temporal
velocity gradient. Fair et al, (1968) approximated the Camp and Meserve's analysis as:
k H = 29 G2/3
cl
where k is the reaeration coefficient (day ~ ), H is the hydraulic depth of the stream
(ft), andTJ is the mean temporal velocity gradient (see" ). Since k H is equivalent to C
in this study, Equations 110 and 111 agree with their analysis.
Efficiency of Cloth Bag — '
The mass transport rate into a cloth bag should equal the mass removal rate by
the organisms inside the bag. Therefore,
Vc ^Sc =cA(S-Se\
where X = V5S concentration inside the bag (mg/1).
Thus, the substrate concentration inside the bag (S ) is solved to be:
S
Let the cloth bag efficiency, E, denote the ratio of substrate removal rate
inside the bag to the removal raj:e_without the bag, or:
E= xkXcSc \ AXcS N So (K*
'
197
-------
Thus, E is a function of the substrate concentration (S), the cloth bag shape
factor (V /A ), the substrate removal rate coefficient (k), the material exchange
coefficient (<5), and the biomass concentration inside the bag (X ). Among the above
factors, Xc is the least important factor in the range of 15,000 mg/1 to 25,000 mg/1 (as
VSS) for phenol sludge and 4,000 mg/1 to 6,000 mg/1 (as VSS) for methanol sludge. The
cloth bag efficiency diagrams for phenol and methanol sludge are presented in Figures
65-1 to 65-9 and 66-1 to 66-9 for various values of k, C, and S. The main purpose of
these diagrams is for the proper design of cloth bags under various circumstances. Use
of the diagrams is as follows:
First, find a diagram for a given or desired set of k and C values. Then, from
the diagram, find the cloth bag efficiency corresponding to the cloth bag shape factor,
V /A , and the substance concentration range (between the spill concentration and the
desired concentration after treatment).
The cloth bag efficiency decreases as the cloth bag shape factor (V /A )
increases, as the substance concentration decreases, as C decreases, and as K
increases.
Application of Cloth Bags in Confined Reactors—
When the sludge-containing cloth bags are used in a confined, batch reactor, the
total mass removal rate by organisms is expressed as:
, + ^> (115)
where N is the number of cloth bags. Therefore,
dS (Ek)XS m,v
dt = KTT" U16)
where:
Y (117)
X =
_— .
Because there is no change in the bacterial growth kinetics, the only difference
between two batch treatment systems with and without cloth bags is that Ek, instead
of k, should be considered as the substrate removal rate coefficient in the cloth bag
systems. If E remains almost constant throughout the reaction period (between the
spill concentration and the desired effluent concentration), the system can be designed
using the methods described in the section entitled Design of Batch Treatment System,
with a substitution of Ek for k. If E changes with substrate concentration, the cloth
bag efficiency should be estimated at each station using Eqs. 104 and 105, before
entering the iteration process of Eqs. 72 and 73.
198
-------
1.0
0.8
u 0.6
o
<£
en
fO
0.4
0.2
0
2000 1000 500 300
' i i i i i i
i I ' i i i
I _ L
100 50 30
10
Substrate concentration, S (mg/1 TOC)
Figure 65-1. Efficiency of cloth bag filled with phenol sludge
(k = 0.01892 hr"1 , C = 0.07 cm/hr, and
X = 20,000 mg/1).
\+
192
-------
2000 1000 500 300 TOO 50 30
Substrate concentration, S (mg/1 TOC)
10
Figure 65-2. Efficiency of cloth bag filled with phenol sludge
(k = 0.01892 hr'1, C = 0.3 cm/hr, and
Xc = 20,000 mg/1).
200
-------
2000
1000 500 300 100 50 30
Substrate concentration, S (mg/1 TOC)
Figure 65-3. Efficiency of cloth bag filled with phenol sludge
(k = 0.01892 hr"1, C = 0.7 cm/hr, and
X = 20,000 mg/1).
201
-------
2000
1000 500 300 TOO 50 30
Substrate concentration, S (mg/1 TOC)
Figure 65-4. Efficiency of cloth bag filled with phenol sludge
(k = 0.05674 hr"1, C = 0.07 cm/hr, and
Xc = 20,000 mg/1).
202
-------
1.0
0.8
=; o.e
cj
CD
•r—
o
4-
en
Ja 0.4
0.2
0
I r
i i i i r
i i i i i i i i i
i i i i i
2000 1000 500 300 100 50 30 10
Substrate concentration, S (mg/1 TOC)
Figure 65-5. Efficiency of cloth bag filled with phenol sludge
(k = 0.05674 hr"1, C = 0.3 cm/hr, and
Xr = 20,000 mg/1).
203
-------
1.0
0.8
o>
•r—
0
O)
cn
(O
0.6
0.4
o
o
0.2
I I I I I
I I
III I
I I I I I
2000 1000 500 300 100 50 30
Substrate concentration, S (mg/1)
10
Figure 65-6. Efficiency of cloth bag filled with phenol sludge
(k = 0.05674 hr"1 , C = 0.7 cm/hr , and
Xr = 20,000 mg/1).
\+
204
-------
1.0
0.8 -
LU
0.6 -
O)
(O
JD
0.4 -
0.2 -
2000
1000 500 300 100 50 30
Substrate concentration, S (mg/1)
Figure 65-7. Efficiency of cloth bag filled with phenol sludge
(k = 0.07239 hr"1 , C = 0.07 cm/hr , and
Xc = 20,000 mg/1).
205
-------
1.0
0.8 -
0.6 -
O)
•r—
u
4-
O)
03
O
CJ
0.4 -
0.2 -
0
2000 1000 500 300 100 50 30
Substrate concentration, S (mg/1)
10
Figure 65-8 Efficiency of cloth bag filled with phenol sludge
(k = 0.07239 hr"1 , C = 0.3 cm/hr , and
Xr = 20,000 mg/1).
V*
2Q6.
-------
1.0
0.8 -
0.6
o
O)
CJ1
to
O
O
0.4 -
0.2 -
0
' ' i i I i—|—i 1
' ' ' ' ' I—I 1
2000 1000 500 300
TOO 50 30
10
Substrate concentration, S (mg/1)
Figure 65-9. Efficiency of cloth bag filled with phenol sludge
(k = 0.07239 hr"1 , C = 0.7 cm/hr , and
Xr = 20,000 mg/1).
\*
207
-------
111 I I I I—I
V /A = 0.1 cm
3000
1000 500 300 100 50 30
Substrate concentration, S (mg/1 TOC)
Figure 66-1. Efficiency of cloth bag filled with methanol sludge
(k = 0.04407 hr"1 , C = 0.14 cm/hr , and
Xr = 5,000 mg/1).
c
208
-------
1.0
0.8
-«0.6
u
CD
0.4
0.2
3000
i i
ill i
= 0.5 cm
LJ ' i i i i—I L
_L
J L
1000 500 300 100 50 30
Substrate concentration, S (mg/1 TOC)
10
Figure 66-2. Efficiency of cloth bag filled with methanol sludge
(k = 0.04407 hr"1 , C = 0.7 cm/hr , and
Xc = 5,000 mg/1).
209
-------
1.0
0.8
£ 0.6
u
0 cm
I I I I I
i i i i i i i i
1000 500 300 100 50 30
Substrate concentration, S (mg/1 TOC)
10
Figure 66-3. Efficiency of cloth bag filled with methanol sludge
(k = 0.04407 hr"1 , C = 1.35 cm/hr , and
Xc = 5,000 mg/1).
210
-------
1.0
0.8 -
-------
1.0 r—-r
0.8
0.6
-------
1.0
0.8
0)
•I-
u
- 0.4
o
o
0.2
3000
V _/Ar = 2.0 cm
c L.
_L
I'''''' ' L
1000 500 300 100 50 30
Substrate Concentration, S (mg/1 TOC)
10
Figure 66-6. Efficiency of cloth bag filled with methanol sludge
(k = 0.2256 hr"1 , C = 1.35 cm/hr , and
Xc = 5,000 mg/1).
213
-------
1.0
0.8
o
c
(J
•+-
4-
O)
CD
_Q
jC
O
o
0.6 "
0.4 -
0.2 -
0
3000
1000 500 300 100 50 30
Substrate concentration, S (mg/1 TOC)
10
Figure 66-7. Efficiency of cloth bag filled with methanol sludge
(k = 0.2656 hr"1 , C = 0.14 cm/hr , Qnd
X = 5,000 mg/1).
214
-------
1.0
0.8 -
>> 0.6
c
-------
i.ori—r
0.8
O)
o
en
-------
The theoretically computed substrate removal patterns using the latter method
are compared to the observed ones in Fig. 67. The tests were conducted at 24 °C with
all nutrients and pH buffer provided, distilled water as a dilution water, and methanol
as the substrate. The stripping rate coefficient of 0.0181 hr~ was estimated from a
separate reactor with the same conditions except for the acclimated activated sludge.
All other necessary kinetic coefficients were estimated from diagrams and equations
defined in this study. Detailed experimental conditions and necessary kinetic
information are tabulated in Table 40.
TABLE 40. EXPERIMENTAL CONDITIONS AND KINETIC INFORMATION FOR
THE METHANOL REMOVAL BATCH TESTS USING CLOTH BAGS
Parameter
Condition
Source
Temperature
PH
Dilution water
Nutrients (Nitrogen, Phos-
phorus and minerals)
Cloth bag shape
factor (VC/AC)
K
£
a
k
24.0 °C
6.9
Distilled water
Provided
0.672 cm (#1),
0.648 cm (#2),
0.695 cm (#3)
2,330 mg/1
1.25
0.28 hr
-1
-1
Initial X (mg/1)
(mg/1) Initial X
0.00513 hr
1.35 cm/hr
6,780 (#1), 7,040 (#2),
6,400 (#3)
380
Figs. 24, 27
and Eq. 3
Eq. 71
Table 38
A total of 1.14 grams of VSS were applied initially in 3-liter reactors (X = 380
mg/1). After 40 hours of aeration, total VSS was measured to be about 2.25 grams and
the VSS that had escaped from cloth bags was less than 90 mg (30 mg/1). Thus, less
than 10% of the initial biomass applied escaped from the bags and there was a net
increase of biomass inside the bags.
The aeration time required to obtain a given substrate removal was observed to
be about 5% less than that computed. The organisms that escaped from bags are free
from the cloth bag barrier (E = 1.0) and dilute biomass concentrations outside the bags
produced improved k values, as pointed out previously.
For detailed computation results for the cloth bag efficiency, biomass growth,
and aeration time required, see Appendix B.
217
-------
700
600
500
400
TOC
(mg/1)
300
200
100
0
0
Methanol Stripping
(k = 0.0181/hr)
Reactor #3
Reactor #2
o
A
Observed TOC removal
in reactor #1
Observed TOC removal
in reactor #2
Observed TOC removal
in reactor #3
• Theoretical TOC removal curve
30
10 20
Aeration time(hr)
Figure 67. Methanol removal using cloth bags in batch reactors.
40
218
-------
Application of Cloth Bags in One-Dimensional Dispersion
System—
When N number of sludge-containing cloth bags are deployed over the spills
according to the pollutants dispersion model, the distribution function of cloth bags
can be given by the following equation, assuming that the dispersion coefficient of
cloth bags is the same as the pollutant dispersion coefficient:
» (118)
where the dimension of f is 1/L. Then, the total mass removal rate can be expressed
as:
dx.
dt ~ ^ Ks+ s ^o
When the peak spill concentration is considerably smaller than K , the cloth bag
j frkr
J>
s;
efficiency becomes almost constant with substrate concentration (see Figures 65 and
66). Then, the total mass of the pollutant at time t is obtained by integrating Equation
>m the clo
-MJL =
119 from the cloth bag application time, t , to t, or:
— cl
(120)
where M = total biomass applied (= N V X ).
Oxygen Requirement Considerations in the Application of Cloth Bags
The oxygen transport rate into a cloth bag should balance the oxygen consump-
tion rate by the organisms inside the bag. When the dissolved oxygen (D.O.) level
inside the bag is designed to be zero, the following equation is possible:
t EkX,.S + b'Xt) = CAC0,
where 0 = DO level outside bags (mg/1).
Thus, the required DO level outside the cloth bags is solved to be:
=
The cloth bag shape factor, V /A , or the amount of sludge inside the bag, X ,
needs to be designed so that the required DO level does not exceed the saturated level
to avoid anaerobic conditions inside the bag.
* The dispersion coefficient for cloth bags was assumed to be the same as that for
pollutants because the difference does not affect the pollutant removal rate signifi-
cantly. See Appendix C for more details.
219
-------
In a confined reactor, the oxygen supply rate into the system should equal the
oxygen consumption rate by organisms.
Thus,
V k (O - O) = NCA O, (122)
a, S C
where: V = volume of reactor (L ),
k = aeration rate coefficient (1/T),
a 3
Os= saturated DO level (M/L ), and
N = number of cloth bags.
Therefore, the aeration devices should be able to provide oxygen at the aeration
rate coefficient, k . or
Cl
V Os- 0
In a stream, the oxygen supply and consumption rate balance is expressed as:
Ak (0-0) = fCA 0. (123)
i d o v. C
The number of cloth bags per unit length of streams (f ) is restricted by the
reaeration abilities of streams (k ), or:
Ac. ko, Qs-o
At TT O (123)
Therefore cloth bags applied in excess of this number will cause an aerobic
conditions within the bags.
The above gas transfer mechanisms deal with mechanical transport only. If
oxygen is supplied at the cloth site by means of air diffusers, rather than by
mechanical mixing, the above restrictions will be greatly alleviated. In addition to the
oxygen supplied from the surrounding water by mechanical mixing, the cloth bags will
receive oxygen through that portion of the bag that is in contact with air. This
mechanism has not been studied.
Application of Material Confining Barriers
Containment of a spill in a concentrated form provides better treatment
efficiency and, in addition, reduces damage to the total environment. Brown (1972)
reported that an inflatable plastic barrier was highly effective in confining phenol,
220
-------
methanol, and other soluble hazardous substances. The barrier was constructed of a
highly flexible, fiber-reinforced plastic with an air-inflated flotation collar, which
supported the barrier, and a water-inflated seal, which sealed the barrier to the
bottom of the waterway. The barrier was maintained in position by a mooring system.
As was described previously, a floating barrier with an open bottom was used in
the model river to contain spilled materials; however, it failed to confine material
long enough for effective biological treatment. Therefore, similar models were used.
The cylindrical barrier had a bottom area of 575 cm and a water depth of 35
cm. The loss of contained material was thought to be due to bottom sediment
sorption, seepage through the bottom sediment, leakage through the bottom
seal, volatilization through the air-water interface, and so forth. These
losses due to other than biological decomposition may be categorized into a
first order reaction with respect to the substance concentration. When the
water body inside a barrier is to be treated in a batch manner, the total
mass removal rate can be expressed as:
— — K XS — Kr S
where k is the pollutant removal rate coefficient (time" ) contributed by other than
biological decomposition.
Use of a 575-cm area of Lake Austin sediment, which had a geometric mean
diameter of 0.173 mm with a geometric standard deviation of 1.70 mm, resulted in a k
value of 0.00346 hr" for phenol. The value for methanol was observed to be 0.00873
hr without aeration. The difference between the two coefficient values was believed
to be due to methanol volatilization. When methanol was aerated through glass tubes
with an air flow rate of 1 liter air/min/liter water, k increased to 0.00952 hr"
_*!
When the mixing intensity within the barrier is not enough for complete
suspension of the acclimated sludge as required for in situ batch treatment, sludge-
containing cloth bags may be used. Then, Ek instead of k should be used in Equation
or:
dS _
dt ~
Then, S and X change with time and are obtained from Equations 81 and 82 with
substitutions of Ek for k and k for k .
The confining barrier application tests were carried out using sludge-containing
cloth bags as the acclimated bacterial source with phenol and methanol as a substrate,
with groundwater flowing in the model river, and without any chemical aids at 28 C.
Oxygen was supplied at the cloth bag sites by means of aeration. The equivalent pH
for phenol removal in the groundwater was 7.0 (see Table 10). Therefore, k at 28 °C is
estimated to be 0.0724 hr" for phenol (see Figure 12). The average pH for methanol
221
-------
removal in the groundwater was 7.8 with an efficiency of 0.473 (see Table 16).
Therefore, k for methanol at 28 °C is estimated to be 0.1143 hr (=0.2656 x 0.91 x
0.473) (see Figure 25). Detailed experimental conditions and information necessary for
the prediciton of S and X with aeration time are tabulated in Table 41.
At the end of two days of aeration, about 2% of the applied phenol sludge and
about 5% of the applied methanol sludge initially leaked from the cloth bags.
Theoretically computed substrate removal patterns of phenol and methanol, on the
assumption that all the sludge stayed within the cloth bags, are compared to the
observed substrate concentrations in Figures 68 and 69. The aeration times required
to obtain a given substrate removal were observed to be approximately 5% less than
those computed, probably due to organisms that escaped from the cloth bags.
222
-------
TABLE 41. EXPERIMENTAL CONDITIONS AND KINETIC INFORMATION FOR THE
PHENOL AND METHANOL REMOVAL BATCH TESTS USING FIXED CONFINING
BARRIERS AND SLUDGE-CONTAINING CLOTH BAGS
Parameters
Temperature (°C)
Dilution water
Initial pH
Equivalent pH
(or average pH)
Efficiency of
dilution water
f -Factor
k (hr"1)
K (mg/1)
s
kd (hr'1)
a
C (cm/hr)
Initial XG (mg/1 as VSS)
Initial X (mg/1 as VSS)
Cloth bag shape factor,
V /A (cm)
Test
Phenol
28
Groundwater
8.6
7.0
(Table 10)
1.0
(Table 10)
1.0
(Figure 12)
0.0724
236
0.00672
(Eq. 41)
1.21
0.684
(Table 18)
31,320
1,350
0.675
substance
Methanol
28
Groundwater
8.6
7.8
(Table 16)
0.473
(Table 16)
0.91
(Figure 25)
0.1143
2,330
0.00291
(Eq. 58)
1.25
1.35
(Table 18)
5,754
635
0.685
223
-------
oo i
00
(0
-IT -
o
o
GO
cnur
to
s-
4->
to
^ o
3 O
oo
o
o
Phenol loss by other than
biological decomposition
= 0.00346/hr)
Theoretical TOC removal
Cloth bag
efficiency
OO
•
o
o
CXJ CD
to
O -Q
O
O
O Observed TOC removal
10 20 30
Aeration time (hr)
40
50
Figure 68. Phenol removal using a fixed barrier with sludge-
containing cloth bags.
224
-------
Methanol loss by other than
biological decomposition
(k = 0.00952/hr)
Cloth bag
efficiency
Biomass
growth
Theoretical
removal
Observed TOC removal
20 30
Aeration time (hr)
Figure 69. Methanol removal using a fixed barrier with
sludge-containing cloth bags.
225
-------
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240
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APPENDIX A
HYPOTHESIS TESTS
A-l. INFERENCE ABOUT MIC H AELIS-MENTEN CONSTANT, kgt FOR PHENOL.
Ho: Ks(28°C) = Ks
Hl: Ks(28°C) XKs
mg/1
Ks(28°C) =206'3 mg/1
S,, =55.6 m g/1
Ks(28°C)
n=29
T c _ 206.3 - 236
5"5!>
t27}ag9= 2.4727
'T'S''
-------
A2. INFERENCE ABOUT CELL YIELD COEFFICIE NT, a, FOR PHENOL.
V a(28°C) =T
Hl: a(28°C) = a
a = 1.21
a(28 °C) = 1'048
Sa(28°C) = 0-069
n =29
RejectH0if |T.S.|7tn_2A9g
T S - -., . 2
(1(569
'27,0.99 = 2'4727
IT. s. Igg
HJs not rejected. Therefore, the m ean, a = 1.21, can be represented as the
ceTI yield coefficient for all test tern peratures.
242
-------
A-3. INFERENCE ABOUT a1 AND b1 FOR PHENOL.
HQ : RrrRr2 =0
H1 : Rr,-Rr2 ^0
Rr]-Rr2 =0.73
°(RrrRr2) =5.68
n = 10
R eject HQ if |T9|>tg>a75
T = - Pj73. = 0>
y 5.68AJ10
- 0.7027
H is not rejected. Therefore, there is no evidence that the estim ated oxygen uptake
rates based on a1 = 1.39 and b1 = 1.42k , are different from the actually observed rates.
243
-------
A-4. INFERENCE ABOUT a1 AND b1 FOR METHANOL.
H : Rr-| -Rr2 = 0
H-| : Rr| -Rr2 ^0
By, = fl T17
-i—1\ rp u.oi /
(Rr.,-Rr2) = 13.0
= 81
Re]ectHo if
=0.219
Qn
80 13/V^T
t80,0.75 =0.6776
|T80|
-------
APPENDIX B
METHANOL STRIPPING
The boiling point of m ethanol is 64.7° C at! atm , thus it is highly
volatile. As pointed out previously, considerable amounts of methanpl were
removed by volatilization; therefore, m ethanol stripping as well as biological
decom position, can play an im portant role in m ethanol rem oval.
Eckenf elder .etjL (1957)form ulated the stripping kinetic model as:
CS' (B-l)
where: k = volatilization rate coefficient,
3
CL = stripping airflow rate (L /T),
a
c = a constant, and «
V = reactor volume (L ).
However, there was a disagreed! ent regarding the value of the constant, c (Engelbrecht et
_al., 196^ Gaudy £t_a1., 1961). Eckenf elder et _al. (1957) reported that the c value ranged
from 0.15 to 0.35, while E ngelbrecht et _aL 0961) reported that c was about 0.85 rather
than 0.35. Since Equation Bl has no rational basis, the following air stripping kinetics
theory was developed.
The widely accepted gas transfer equation is:
O
where: V = reactor vol urn e (L ),
k = volatilization rate coefficient (L/T),
2
Aa = air- water interface area (L ), and
S_ = %aturated volatile substance concentration for a given partial pressure
(fi/L3).
The vapor pressure of m ethanol in atmosphere can.be regarded as zero. The rate
coefficient, ke, in Equation 59 is equivalent to ky a in Equation
245
-------
B-2. W hen 1 liter air/ min/liter water, was supplied in the m ethanol rem oval study, the
maximum value of k was 0.0277 hr" at 28°C. This means that the saturated m ethanol
concentration (S ) for the m ethanol vapor pressure in the stripping air bubble also is
negligible . Equation B-2 can be reduced to:
£"*vrs (B-3)
When air is supplied with an average bubble size of Dg (diameter) at an airflow rate of
Q 3, the air- water interface area is solved to be:
a
where: e = surface area expansion coefficient caused by turbulence,
2
A = surface area of water body at quiescent condition (L ),
Q = air flow rate (L3/T),
Q
D = average diam eter of air bubble (L),
a
t , = detention tim e of air bubble in the reactor (H/vb),
H = depth of reactor (L), and
vb = bouyancy velocity of air bubble (L/T), a function of D g.
Therefore, Equation B-3 becomes:
As shown in Equation B-5, the stripping rate largely depends on the
geometry of the reactor tank, air bubble size, and the air flow rate.
Equation B-5 is graphically interpreted in Figure B-l . The reactor depth
that provides the minimum stripping rate, Hmjn, is solved by setting
Equation B-5=0, or:
Da
The stripping rate increases as the reactor depth increases or decreases
from Hm-jn. As the reactor depth decreases from Hm-jn, volatilization
through the surface area of the water body dominates and as the reactor
depth increases from Hmi-n, air stripping by air bubbles dominates.
* The saturated m ethanol concentration was calculated based on the data from Lane
Handbook of Chemistry (Dean, 1973).
246
-------
H
Figure B-l. The relationship between the air stripping rate and the
reactor depth.
247
-------
At a quiescent condition (A = A ), the volatilization rate coefficient,
k , was estim ated to be 0.07T16 cm/hr It 5° C, 0.2044 cm/hr at 22° C, and
0.5973 cm/hr at 28° C (Figure B-2). As temperature neared the boiling point,
the volatilization rate coefficient sharply increased, so that the Arrhenius
equation failed to describe the tern perature effect on the rate coefficient.
When 1 liter air/min/liter water of air was supplied through glass tubes with
a dlam eter 0.3 cm into the 3-liter reactors with a surface area (A ) of 105
cm ,the air-water interface areaincreased by 32.5% (A = 1.325 M ).
248
-------
10 15 20
Temperature (°C)
25
30
Figure B - 2. Temperature effect on the volatilization rate coefficient.
249
-------
APPENDIX C
TABLE C-l COMPUTATION RESULTS OF CLOTH BAG EFFICIENCY, BACTERIAL GROWTH,
AND AERATION TINE FOR THE CLOTH BAG APPLICATION
TESTS IN BATCH REACTOR (REACTOR #1)
s
(mg/TTOC}
660
650
640
630
620
610
600
590
580
570
560
550
540
530
520
510
500
490
480
470
460
450
440
430
420
410
400
390
380
370
360
350
340
E
0.7932
0.7890
0.7848
0.7806
0.7764
0.7722
0.7679
0.7637
0.7594
0.7551
0.7508
0.7465
0. 7422
0.7379
0.7336
0.7293
0.7249
0.7206
0.7163
0.7120
0.7077
0.7034
0.6991
0.6948
0.6905
0.6862
0.6820
0.6777
0.6735
0.6692
0.6650
0.6608
Ek
(hr'1)
0.2221
0.2209
0.2198
0.2186
0.2174
0.2162
0.2150
0.2138
0.2126
0.2114
0.2102
0.2090
0.2078
0.2066
0.2054
0.2042
0.2030
0.2018
0.2006
0.1994
0.1981
0.1969
0.1957
0.1945
0.1933
0.1921
0.1910
0.1898
0.1886
0. 1 874
0.1862
0.1850
xc
(mg/rVSS)
6,780
6,905
7,031
7,157
7.284
7,411
7,538
7,667
7,795
7,924
8,053
8,183
8,313
8,443
8,574
8,705
8,838
8,967
9,098
9,229
9,361
9,492
9,624
9,755
9,887
10,018
10,149
10,280
10,411
10,541
10,671
10,801
10,930
X
(rng/1 VSS)
380.0
387.0
394.0
401.1
408.2
415.4
422.5
429.7
436.9
444.1
451.4
458.6
465.9
473.2
480.5
487.9
495.2
502.6
509.9
517.3
524.7
532.0
539.4
546.8
554.1
561.5
568.8
576.2
583.5
590.8
598.1
605.4
612.6
1
(hrs.)
0
0.327
0.656
0.987
1.320
1.655
1.992
2.332
2.674
3.019
3.368
3.719
4.074
4.432
4.793
5.159
5.529
5.903
6.282
6.666
7.055
7.449
7.850
8.256
8.669
9.089
9,516
9.950
10,393
10,845
11.306
11.776
12.257
250
-------
TABLE C-l (Continued)
s
(mg/1 TOC)
330
320
310
300
290
280
270
260
250
240
230
220
210
200
190
180
170
160
150
140
130
120
110
100
90
80
70
60
50
40
E
0.6567
0.6525
0.6484
0.6443
0.6402
0.6361
0.6321
0.6281
0.6241
0.6202
0.6163
0.6125
0.6087
0.6049
0.6012
0.5975
0.5939
0.5904
0.5869
0.5835
0.5802
0.5770
0.5739
0.5710
0.5681
0.5656
0.5632
0.5611
0.5594
0.5582
Ekl
(hr-1)
0.1839
0.1827
0.1816
0.1804
0.1793
0.1781
0.1770
0.1759
0.1748
0.1737
0.1726
0.1715
0.1704
0.1694
0.1683
0.1673
0.1663
0.1653
0.1643
0.1634
0.1625
0.1616
0.1607
0.1599
0.1591
0.1584
0.1577
0.1571
0.1566
0.1563
x.
(mcg/l VSS)
11,059
11,187
11,314
11,441
11,567
11,692
11,816
11,938
12,060
12,180
12,299
12,416
12,530
12,643
12,754
12,862
12,966
13,068
13,165
13,258
13,346
13,427
13,502
13,568
13,624
13,667
13,694
13,699
13,676
13,612
X
(mg/1 VSS)
619.8
627.0
634.1
641.2
648.3
655.3
662.2
669.1
675.9
682.6
689.3
695.9
702.3
708.6
714.8
720.9
726.7
732.4
737.9
743.1
748.0
752.6
756.8
760.5
763.6
766.0
767.5
767.8
766.5
762.9
t
(hrs.)
12.750
13.254
13.771
14.302
14.847
15.409
15.987
16.584
17.201
17.839
18.502
19.190
19.906
20.654
21.436
22.257
23.121
24.034
25.001
26.031
27.133
28.319
29.604
31.008
32.555
34.281
36.233
38.485
41.147
44.408
251
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APPENDIX D
APPLICATION OF CLOTH BAGS IN ONE-DIMENSIONAL DISPERSION SYSTEM
When the dispersion coeffi dent for cloth bags is different from the
dissolved pollutant dispersion coefficient, the total m ass of the pollutant
at tim e t is solved to be:
so
where: D' = dispersion coefficient for cloth bags.
It appeared from a visual dispersion study of twenty cloth bags in the
model river that D1 was not significantly different from D. However, if
D1 = %D, then:
s _ 1.15 -let,
~Y 6
w here:
Y = exp 4
Therefore, the difference in D and D1 does not significantly affect the
removal rate.
252
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