United States
Environmental Protection
Agency
Hazardous Waste Engineering
Research Laboratory
Cincinnati OH 45268
EPA/540/2-86/002
September 1986
vvEPA
Superfund
Systems to Accelerate
In Situ Stabilization of
Waste Deposits
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EPA/540/2-86/002
September 1986
SYSTEMS TO ACCELERATE IN SITU STABILIZATION
OF WASTE DEPOSITS
by
M. Amdurer, R. T. Fellman,
J. Roetzer, C. Russ
Envirosphere Company
Two World Trade Center
New York, NY 10048
EPA Contract Number: 68-03-3113
Task 37-2
Project Officer
W. Grube
Hazardous Waste Engineering Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
HAZARDOUS WASTE ENGINEERING RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL 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 Contract
No. 68-03-3113 (Work Order 37-2) to JRB Associates with a subcontract
(JRB No. 2-817-33-956-72-8) to Envirosphere Company. It has been
subject to the Agency's peer and administrative review and has been
approved for publication as an EPA document.
This report is intended to present information on the potential
application of a number of in situ treatment technologies for the
stabilization of deposits containing various organic waste
compounds. It is not intended to address every conceivable waste
type or all possible applications of the technologies described.
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
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FOREWORD
Today's rapidly developing and changing technologies and industrial
products and practices frequently carry with them the increased generation
of solid and hazardous wastes. These materials, if improperly dealt
with, can threaten both public health and the environment. Abandoned
waste sites and accidental releases of toxic and hazardous substances to
the environment also have important environment and public health
implications. The Hazardous Waste Engineering Research Laboratory assists
in providing an authoritative and defensible engineering basis for
assessing and solving these problems. Its products support the policies,
programs and regulations of the Environmental Protection Agency, the
permitting and other responsibilities of State and local governments and
the needs of both large and small businesses in handling their wastes
responsibly and economically.
This report describes the important elements of technology which
need to be understood and applied in order to accomplish an effective in-
situ stabilization of hazadous wastes on an uncontrolled waste site.
This will be useful to remedial action plan designers and plan approval
staffs in clarifying the advantages and disadvantages of various in-situ
stabilization technologies with respect to pollutant, site, and technology
interreactions. For further information the reader should contact the
Land Pollution Control Division of the Hazardous Waste Engineering Research
Laboratory.
Thomas R. Hauser, Director
Hazardous Waste Engineering
Research Laboratory
m
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PREFACE
This report covers a broad area of potential in situ surface and subsurface
remediation activities at uncontrolled waste sites. Conventionally,
remediation has been conceived of as containment and isolation, or as
excavation followed by secured landfill deposition or incineration. These two
"conventional" methods may not always be practicable or economically
achievable, especially where contamination of a large area has occurred.
Furthermore, the waste material is not necessarily destroyed} thus objections
to these methods have arisen based upon the long lifetimes of many hazardous
wastes in the environment and the undemonstrated longterm reliability of
engineered safety measures.
The application of in situ systems to accelerate the stabilization of waste
deposits offers a potential solution to the concerns raised with respect to
conventional approaches. The fundamental concept for in situ stabilization is
chemical treatment of the waste within the soil medium. This requires the use
of a treatment agent (reactant), a means for delivering the reactant to the
waste and usually a means for recovering the products of the reaction. This
three part concept — delivery, recovery and treatment selection — is
discussed in this report. In situ treatment may be used in concert with other
remedial actions. For example, excavation and treatment on the surface
(incineration, soil washing) may be used for the most highly concentrated
source materials, followed by in situ treatment of the plume or other
remaining, less-contaminated areas.
It is important to note that while each individual step of the basic in situ
treatment process is part of a broad body of technical experience, the
combination of •injection, reaction and recovery as a system for in situ
remediation has scarcely been practiced and is in its infancy as an integrated
technology. This finding frustrates efforts to systematize selection of
reactant, delivery and recovery systems because of the near absence of proven
field experience. One danger of which the reader must be clearly forewarned
iv
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is that a report such as this, which attempts to systematize the process of
selecting an in situ treatment system, may appear to promise too much in terms
of specificity and definition and may appear to suggest .that firm solutions
exist. To underscore this, the authors note that, at best, application of
systems to accelerate the in situ stabilization of waste deposits is in its
early stage of development. Further, the application of these systems to
uncontrolled waste sites will require a site-by-site, customized approach
since the subsurface geohydrology, waste inventory and site history will be
unique for each site, a situation defying systematization. There is also
extensive need for laboratory simulations and testing prior to implementation
of in situ systems. Data are too scarce to predict, with confidence, the
efficiency and practicality of the methods. However, the data at this stage
in the evolution of these technologies can justify the development of general
guidelines. This report provides the general guidelines and the basis for
them. It also serves as a starting point for developing the site specific
requirements for implementation of in situ systems to accelerate the
stabilization of waste deposits, and for identifying future research and
investigation needs in this area.
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ABSTRACT ; . -.-...'
In situ systems to accelerate the stabilization of waste deposits have been
presented as alternatives to containment, isolation or excavation as methods
for remediation of uncontrolled waste sites. In situ applications involve
three essential elements: selection of a chemical or biological agent
(reactant) which can react with and stabilize the waste, a method for delivery
of the reactant to the deposit and a method for recovery of the reaction
products or mobilized waste. The most promising applications for in situ
treatment methods are for spill or plume types of contamination,, where the
contaminants are relatively evenly distributed and preferably in liquid form.
Delivery of reactants to solid, heterogeneous, low permeability deposits will
be more difficult. In situ methods may find particular application when used
in combination with other remedial measures, for example, removal of the
source material and in situ treatment of the plume.
Four reactant categories have been examined: biodegradation, surfactant-
assistant flushing, hydrolysis, and oxidation. Of these, biodegradation and
surfactant-assisted flushing appear most promising as in situ treatment
techniques. For any treatment technique, the potential toxicity of the
applied reactant and any intermediate compounds or by-products must be
carefully evaluated. Furthermore, the potential for undesirable reactions
with other contaminants present must be studied (e.g., oxidation of phenol
with hydrogen peroxide may also oxidize chromium (III) to the more toxic
hexavalent chromium).
Methods of delivery of reactants based upon gravity include surface flooding,
ponding, surface spraying, ditching, and subsurface infiltration beds and
galleries. Forced injection (pumping) may also be used. Permeability is an
important consideration in selecting the delivery system. Gravity delivery
methods require a permeability of the soil/waste medium in the range 10
_q
cm/sec to 10 cm/sec (280 to 2.8 ft/day). Forced injection is most
effective at a permeability in the range of 10~ cm/sec to 10~ cm/sec
vi
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(280 to 0.28 ft/day); below this permeability limit a potential application
offorced injection for reagent delivery coupled with electro-osmosis for
recovery may exist. Additionally, gravity systems should be considered only
when the waste deposit lies in the unsaturated zone and when the depth to the
bottom of the deposit is less than 5 meters (16 feet). Otherwise, forced
injection should be considered.
Recovery systems using gravity include open ditching and buried drains, and
pumped methods include wellpoint and deep well systems. Basically, the same
limitations that apply to delivery systems are also true for recovery systems.
Gravity-induced recovery works best when the water table is within 5 meters
(16 ft) of the surface. For depths in the range of 0-8 meters (0-26 ft), well
points can also be considered. Depths greater than the suction limit (about 8
meters or 26 ft in practice) will require the use of down-hole pumps for
recovery.
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CONTENTS
PAGE
DISCLAIMER
FOREWARD
PREFACE
ABSTRACT
CONTENTS
LIST OF FIGURES
LIST OF TABLES
ACKNOWLEDGEMENTS
1.0 DELIVERY AND RECOVERY SYSTEMS FOR WASTE DEPOSIT STABILIZATION
1.1 ENGINEERING FEATURES AND GEOHYDROLOGIC PARAMETERS
1.1.1 Introduction
1.1.2 Waste Deposit Settings and Methods of Delivery
and Recovery
1.1.3 Geohydrologic Parameters
1.2 DELIVERY TECHNOLOGIES
1.2.1 Gravity Delivery Methods
1.2.2 Forced Delivery Methods
1.2.3 Summary and Example Applications
1.3 RECOVERY TECHNOLOGIES
1.3.1 Gravity Recovery Methods
1.3.2 Forced Recovery Methods
1.3.3 Summary
1.4 SPECIAL METHOD OF DELIVERY AND RECOVERY ENHANCEMENT
(ELECTRO-OSMOSIS)
1.5 COMPARATIVE ANALYSIS OF ALTERNATIVES
1.5.1 Importance of Various Parameters in Gravity
vs Forced Systems
1.5.2 Application of Various Systems
iii
iv
vi
viii
xii
xvi
1
2
2
3
4
13
15
35
39
43
45
53
57
60
63
63
63
2.0 BIODEGRADATION
2.1 INTRODUCTION
74
74
IX
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CONTENTS (Cont'd)
2.2
2.3
2.4
3.0
3.1
3.2
3.3
3.4
3.5
3.6
3.7
3.8
ANALYSIS OF DATA
2.2.1 Microbial Mechanisms of Catabolism
2.2.2 Development of Microbial Agents
2.2.3 Factors Affecting the Use of Microbial Agents
2.2.4 Susceptability of Various Chemical Classes to
Biodegradation
APPLICATION TO WASTE DEPOSITS
2.3.1 Site Assessment
2.3.2 Case Histories of In Situ Treatment of Surface
Waste Deposits or Spills
2.3.3 Case Histories of In Situ Treatment of Subsurface
Waste Deposits or Spills
2.3.4 Liquid Surface Waste Deposits
2.3.5 Renovation of Waste Disposal Sites
SUMMARY
SURFACTANT-ASSISTED FLUSHING
INTRODUCTION
BACKGROUND AND THEORY
SURFACTANT CHEMICAL CHARACTERISTICS
SURFACTANT APPLICATION TO SUBSURFACE DEPOSITS:
EXISTING INFORMATION ON SURFACTANT BEHAVIOR
3.4.1 Tertiary Oil Recovery
3.4.2 Petroleum Spills
SURFACTANT APPLICATION TO SUBSURFACE DEPOSITS:
GEOCHEMICAL AND ENVIRONMENTAL FACTORS
3.5.1 Groundwater Chemistry
3.5.2 Soil Chemistry
ENVIRONMENTAL EFFECTS
3.6.1 Biodegradability
3.6.2 Toxicity
SUMMARY
CONCLUSIONS
PAGE
75
75
76
80
80
91
92
96
99
108
109
110
126
126
127
128
136
136
138
143
143
143
147
148
150
151
154
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CONTENTS (Cont'd)
4.0
4.1
4.2
4.3
4..4
4.5
HYDROLYSIS
INTRODUCTION
HYDROLYSIS MECHANISMS AND KINETICS
4.2.1 Hydrolyzable Organic Groups
4.2.2
4.2.3
4.2.4
4.2.5
Effect of pH on Hydrolysis Rates
Effect of Temperature on Hydrolysis Rates
Effect of Solvent Composition on Hydrolysis Rates
Catalysis
ACCELERATION OF HYDROLYSIS RATES IN WASTE DEPOSITS
4.3.1 Aklyl Halides
4.3.2 Halogenated Ethers, Epoxides, and Alcohols
4.3.3 Epoxides
4.3.4 Esters ( Carboxylic Acid Esters)
4.3.5 Amides
4.3.6 Carbamates
4.3.7 Phosphoric and Phosphonic Acid Esters
4.3.8 Miscellaneous Compounds (including Alkylating
Agents and Pesticides)
CASE HISTORY OF BASE-CATALYZED HYDROLYSIS
SUMMARY
PAGE
157
157
158
159
162
164
166
166
167
168
170
170
170
173
173
176
179
179
181
5;0 CHEMICAL OXIDATION 185
5.1 HYDROGEN PEROXIDE 186
5.1.1 Properties of Hydrogen Peroxide 186
5.1.2 Oxidation of Organics by Hydrogen Peroxide 187
5.1.3 Application Potential of Hydrogen Peroxide for - 187
In Situ Treatment
5,2 OZONATION > 191
5.2.1 Properties of Ozone 191
•;•••'.. 5.2.2 Oxidation of Organics by Ozone 193
5.2.3 Applications of Ozonation 195
5.2.4 Application Potential of Ozone for In Situ Treatment 202
XI
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5.3
CONTENTS (Cont'd)
HYPOCHLORITES
5.3.1 Properties of Hypochlorites
5.3.2 Treatment Applications of Hypochlorites
5.3.3 Potential for In Situ Treatment of Waste
Deposits Using Hypochlorites
PAGE
204
204
207
208
6.0
6.1
6.2
6.3
6.4
6.5
APPLICATION AND DESIGN OF SYSTEMS TO ACCELERATE
STABLIZATION OF WASTE DEPOSITS
INTRODUCTION
REMEDIAL INVESTIGATION
FEASIBILITY STUDY
6.3.1 Evaluation of Biodegradation for In Situ
Stabilization of Waste Deposits
6.3.2 Evaluation of Flushing and Surfactants for Waste >
Deposit Stabilization
6.3.3 Evaluation of Hydrolysis for Waste Deposit
Stabilization
6.3.4 Evaluation of Oxidation for Waste Deposit Stabilization
APPLICATION AND DESIGN OF DELIVERY/RECOVERY SYSTEMS
FOR IN SITU TREATMENT
6.4.1 Determining the Requirements of a Delivery/
Recovery System
6.4.2 Site Evaluation
6.4.3 Selecting the Delivery and Recovery Methods
6.4.4 Field Demonstration Program
6.4.5 Evaluating Alternative Methods
6.4.6 Detailed Design and Implementation :
CASE HISTORY OF RI/FS AND IN SITU TREATMENT ;
OF CONTAMINATED SOIL AND GROUNDWATER i
6.5.1
6.5.2
6.5.3
6.5.4
6.5.5
Site Summary
Remedial Investigation
Feasibility Study
Description of the Treatment System
Cost Data for the Bibcraft Site
INDEX
APPENDIX A
212
212
213
214
217
223
227
232
233
234
236
236
241
241
241
242
242
243
245
248
251
257
A-l
XII
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' LIST OF FIGURES
Page
1-1 Porosity, Specific Retention and Specific Yield 14
Variations with Grain Size
1-2 General Layout of a Gravity Delivery and Recovery 18
System
1-3 Typical Plan of Ditch and Flooding Type Gravity 23
Delivery System
1-4 Gravity Delivery System Using Ponding 25
1-5 Gravity Delivery System Using Ponding in a Thin
Impervious Stratum 26
1-6 Gravity Delivery System Using Surface Spraying 28
1-7 Gravity Delivery System using Ditches 30
1-8 Subsurface Gravity Delivery System Using
Infiltration Galleries 32
1-9 Subsurface Gravity Delivery System Using an Infiltration Bed 34
1—10 Injection Pipes for Forced Delivery 36
1-11 Forced Delivery. Using Injection Wells 37
1-12 Method for Calculating Location of a Buried Pipe 46
Recovery System
1-13 Gravity Recovery Using A Ditch 49
1-14 Gravity Recovery Using Buried Perforated Pipes 51
1-15 Wellpoint System for Forced Recovery 55
1-16 Deep Well Recovery System 58
.1-17 Limits of Recovery Methods Applicable to Different Soils 59
2-1 Biocraft Site Plan 102
2-2 Biocraft Biodegradation Treatment System-Basic 104
Process Flow Diagram
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LIST OF FIGURES (Cont'd)
Page
3-1 The Effects of Solution pH, Electrolyte Composition 145
and Soil Composition on Surfactant Adsorption to Soil
4-1 pH Dependence of Hydrolysis Rate by Acid, Neutral, and 163
Base Promoted Processes
4-2 Effect of pH on Hydrolysis of Ethyl Acetate 165
5-1 Decomposition Rates of Ozone in Various Waters 192
5-2 City Waterworks, Karlsruhe, West Germany 200
5-3 Basic Flow Diagram for Ozonation of Groundwater 201
at Karlsruhe
6-1 Evaluation of Biodegradation 218
6-2 Evaluation of Flushing and Surfactants 224
6-3 Evaluation of Hydrolysis 228
6-4 Systematic Approach to Delivery/Recovery System Selection 235
6-5 Groundwater Surface Contours, Biocraft Site 244
xiv
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v LIST OF TABLES
Page
1-1 Hydraulic Conductivity, Porosity and Drainage 10
Characteristics of Materials
1-2 Design Limitation for Infiltration-Percolation Systems 22
1-3 Delivery and Recovery Systems 64
1-4 Relative Importance of Geotechnical Parameters in 65
Gravity and Forced Systems
1-5 Matrix for Delivery Methods 67
1-6 Matrix for Recovery Methods 68
2-1 Examples of Biological Renovation at Contaminated 98
Surface Sites
2-2 Summary of Microbiological Treatment Technologies 112
2-3 Summary of Organic Groups Subject to Microbial Metabolism 113
3-1 Surfactant Characteristics 130
3-2 Properties of Selected Organic Compounds which 133
Indicate the Potential Effectiveness of Surfactants
3-3 Summary of Experiments on Surfactant-Enhanced 139
Gasoline Recovery
3-4 Results of Surfactant-Flushing of Contaminants 142
from Test Soil
3-5 Environmental Chemical Properties of Selected 149
Commercial Surfactants
4-1 Groups of Organic Compounds that are Generally 160
Resistant to Hydrolysis
4-2 Groups of Organic Compounds that are Potentially 161
Treatable by Hydrolysis
4-3 Hydrolysis of Alkyl Halides 169
4-4 Hydrolysis of Halogenated Ethers, Epoxides, Alcohols 171
XV
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LIST OF TABLES (Cont'd)
4-5 Hydrolysis of Epoxides
4-6 Hydrolysis of Esters
4-7 Hydrolysis of Amides
4-8 Hydrolysis of Carbamates
4-9 Hydrolysis of Phosphoric and Phosphonic Acid Esters
4-10 Hydrolysis of Miscellaneous Compounds (Including
Pesticides)
4-11 Applicability of Base Catalyzed Hydrolysis as a
Treatment Method for Organic Chemicals
5-1 Organic Chemical Classes Ability to React with
Hydrogen Peroxide
5-2 Organic Chemical Classes Ability to React with Ozone
5-3 Ozonation of Various Compounds in Water
5-4 Organic Chemical Classes Ability to React with
Hypochlorites
6-1 Potential Applications of Treatment Methods to
Waste Contaminants
6-2 Costs of Remedial Action at the Biocraft Site
Page
172
174
175
177
178
180
183
188
194
197
205
215
252
XVI
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ACKNOWLEDGEMENTS
The authors wish to extend their thanks and appreciation to Dr. Walter Grube -
Soil Scientist & Project Officer, Land Pollution Control Division, Hazardous
Waste Engineering Research Laboratory, USEPA; Mr. Donald Banning - Research
Chemist, Hazardous Waste Engineering Research Laboratory, USEPA; and
Dr. Edward Repa - Senior Hydrologist, JRB Associates, for their committed
efforts in guiding the authors' efforts and constructively commenting upon the
authors' product over the course of this challenging project.
XV11
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SECTION 1
DELIVERY AND RECOVERY SYSTEMS FOR WASTE DEPOSIT STABILIZATION ~~
In order to remediate waste deposits in situ, reactants (chemical, biological
or both) must be delivered to the waste deposit and surrounding soil
containing the contaminants. During or following treatment, spent reactants
or stabilization by-products may require removal from the waste deposit and
the surrounding soil. This section provides information currently available
concerning the various aspects of delivery and recovery systems usable for in
situ stabilization. Sections 2-5 provide information currently available on
the biological and chemical processes and reagents that may be applicable to
waste deposit stabilization. Section 6 integrates all of the foregoing, to
establish guidance concerning the use of stabilization technologies at
specific sites.
The specific objectives of Section 1 are to:
1. Identify and review the various soils engineering parameters
affecting the selection and application of delivery and recovery
systems (Section 1.1)
2. Review the types and features of various gravity and forced delivery
systems (Section 1.2)
3. Review the types and features of various gravity and forced recovery
systems (Section 1.3)
4. Review delivery and recovery enhancement technology available through
electro—osmosis (Section 1.4)
5. Present a comparative analysis of alternative delivery and recovery
systems (Section 1.5).
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1.1 Engineering Features and Geohydrologic Parameters
1.1.1 Introduction
An evaluation of potential delivery and recovery methods which could be
employed for in situ treatment of waste deposits must consider the physical
setting of the waste deposit in terms of its relationship with the subsurface
soil and groundwater, and the geohydrologic parameters of the waste deposit
and surrounding soil media. In theory, there are numerous possible
combinations of settings and geohydrology which could arise. The following
assumptions were made in order to confine the discussion to the most common
situations likely to be encountered:
o Waste deposits are in a solid state and no immiscible flow is
expected.
o Waste deposits are located within the upper unconsolidated formation
(above or below the groundwater table) and are not present in a
confined aquifer or bedrock.
o Recovery of the reaction products and spent reactant will be
exclusively from a water table aquifer.
o Solutions applied will have physical characteristics similar to water
and precipitates which may form will not significantly affect deposit
porosity and hydraulic conductivity.
Based on an understanding of the geohydrology of the site it may be determined
that it is necessary to completely contain the waste deposit and any leachate
generated, or place a hydrologic barrier downgradient of the deposit to assist
the recovery system. Grout curtains, slurry walls, and sheet pilings have
been used for this purpose. A description of these methods is beyond the
scope of this work, but is contained in companion documents to this report,
including A W Martin Associates (1978), A D Little (1983), USEPA (1984a),
Spooner et al. (1984), and Repa and Kufs (1985).
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1.1.2 Waste Deposit Settings and Methods of Delivery and Recovery
Given the above assumptions, four basic methods (and combinations thereof)
have been investigated and conceptually applied as possible delivery and
recovery systems. These are:
b Gravity delivery by surface or subsurface means
o Forced delivery
o Gravity recovery
o Forced recovery.
These four methods are briefly discussed in the following paragraphs.
1.1.2.1 Gravity Delivery
Gravity delivery can be applied in cases where the waste deposit is present
either partially above or below the natural ground level, or completely below
the natural ground level and overlain by a permeable material. The
qualitative factors that would give preference to the use of gravity delivery
systems include:
o Shallow depth to waste deposit from the surface (less than 5 meters
or 16 feet)
o Highly permeable cover material between land surface and the waste
—3
deposit (greater than 10 cm/sec or 2.8 ft/day)
o Highly permeable waste deposit and surrounding soil media (greater
than 10~3 cm/sec or 2.8 ft/day)
o Waste deposit located above the groundwater table
o High surface soil infiltration rate (greater than 10 cm or 4 inches
per day)
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o Availability of a relatively long treatment time (i.e., months to
years).
Gravity delivery can also be applied, however, in cases where the subsurface
deposit is overlain by an impermeable cover if the impermeable layer can be
cost-effectively removed, or subsurface methods of gravity delivery (e.g.,
infiltration galleries) are employed, thus eliminating the need for extensive
excavation of the impermeable cover. ;
1.1.2.2 Forced Delivery I
Forced delivery methods can be used in all the conditions noted a.bove, but may
not be suitable in some cases of surficial or very shallow waste deposits
(forced injection of a fluid into a very shallow deposit could lead to soil
cracking caused by excessive uplift pressures, leakage to the surface, etc).
However, shallow injection well points can be used for some shallow waste
deposits.
1.1.2.3 Gravity and Forced Recovery
Recovery system feasibility is significantly affected by depth to the waste
deposit and by the geohydrologic properties of the waste deposit and
surrounding soil media. Gravity recovery is most suited to ishallow waste
deposits (less than 10 meters or 33 feet deep, and less than 5 meters or 16
—3
feet below the water table) with hydraulic conductivities of 10 cm/sec
(2.8 ft/day) or greater. Forced recovery is generally required for deeper
deposits or less permeable conditions.
1.1.3 Geohydrologic Parameters
The geohydrologic parameters of the waste deposit and surrounding soil media
which can have an effect on the selection of delivery and recovery systems are:
o Surface soil infiltration rate
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o Hydraulic conductivity of the waste deposit, cover material and
surrounding soil media
o Interrelated parameters of porosity, specific yield, specific
-. retention and grain size distribution of the host soils and waste
deposit.
Detailed descriptions of the methods for measuring and evaluating these
parameters during site investigations, and their use in evaluating site
geohydrology, are contained in Freeze and Cherry (1979), USEPA (1980a), USEPA
(1980b), and Repa and Kufs (1985). The following brief discussion of these
parameters, however, is provided.
1.1.3.1 Infiltration Rate
The infiltrative capacity of the surface soil above a waste deposit can be a
limiting parameter to the rate at which reactants can be applied by surface
gravity delivery methods. The infiltration rate of the soil is defined as the
rate (cm/min) at which water (or other fluid) enters the soil through its
surface. The infiltration rate is a function of both texture and structure of
the soil as well as moisture content. For example, the drier the soil
profile, the higher the infiltration rate. Conversely as the soil pores fill
with fluid, the infiltration rate will decrease until an approximate
steady-state condition is approached. Usually the steady-state infiltration
rate is used as the design criterion for the hydraulic loading rate of surface
delivery methods. Thus, it is necessary to know the cumulative water intake
of a soil column as a function of time to be able to calculate the fluid
application rate required for treatment of the waste deposit. If the natural
infiltration rate is too low, it can be increased by tilling the surface
(USEPA, 1984b).
The equation for expressing the short-term change in infiltration rate is
approximated by:
(1-1)
where: I = infiltration rate at time, t (length/time)
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n
a constant representing the instantaneous intake rate at
time equal 1 minute
an exponent which, for most soils, is negative with
values between 0 and -1
time
The equation for expressing the cumulative intake of fluid at time, t, is:
Y = AIt(ttfl)/(ttH)
(1-2)
where: Y - cumulative intake (length) and other parameters are as
defined above
The coefficients needed for computation in the above equations can be secured
from data obtained in field infiltrometer studies. The field determined
values for (Y) and (t) can be plotted on log-log paper with the slope of the
line of best fit equal to n+1. The coefficient A may be calculated from
equation 1-2 using any point (t,Y) on the line.
Methods and tools which can be used to determine infiltration rates in the
field are basin flooding, sprinkler infiltrometers, cylinder infiltrometers,
and lysimeters (Bouwer, 1964; Bouwer, 1966; Bouwer and Rice, 1967). Basin
22
flooding involves using an area ranging from 1m (10 ft ) to 0.1 hectare
(0.25 acre), flooding it with water, and measuring the infiltration rate
(USEPA, 1977). Sprinkler infiltrometers are used primarily to determine the
limiting application rate for systems using sprinklers. Cylindrical
infiltrometers and lysimeters are most commonly used. The cylinder
infiltrometer technique involves driving a metal cylinder into the soil to a
depth of about 15 cm (6 inches). The cylinder is usually 13-35 cm (6 to
14 inches) in diameter and 25 cm (10 inches) long. A buffer zone around the
cylinder is formed either by diking or placing another cylinder around the
first one. The buffer zone serves to prevent lateral flow of water from the
inner cylinder. Water is then introduced into the inner cylinder until a
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steady-state condition is reached, while the water level in the buffer zone is
kept at the same level. A lysimeter test involves the recovery of an
undisturbed soil core sample whose infiltration rate is measured in the
laboratory. The field implementation, methods of calculation, theoretical
background, limits of applicability and potential problems associated with
these methods are discussed in greater detail in Meinzer (1923), USEPA (1977),
USEPA (1980a), and Freeze and Cherry (1979).
In all cases the infiltration tests should be performed with the solution
which will actually be used during the in situ treatment process. This is
preferable because the physical and chemical properties of the solution
(density, viscosity, ionic strength, adsorption properties) may alter the
value of the infiltration rate compared to that determined using water alone.
1.1.3.2 Hydraulic Conductivity
Hydraulic conductivity (K) is.defined by Darcy's law in the equation:
where:
Q = KIA, or
v = KI = Q/A,
(1-3)
(1-4)
Q
K
I
A
v
Q
= flow rate (length /time)
= hydraulic conductivity (length/time)
= hydraulic gradient (dimensionless)
= cross-sectional area (length )
= specific discharge ("Darcy velocity") (length/time)
The actual flow velocity of groundwater or a conservative (i.e., not
attenuated) contaminant is given by: ;
V = KI/neff = v/neff.
(1-5)
where ngff = the effective porosity for flow of the medium
(i.e, excluding "dead end" pores through which
the fluid will not flow).
-------
In coarse-grained materials, n is close to the specific yield (see
below). However, n ff in fine materials may be much lower than the specific
yield. If either the total porosity or specific yield is used to calculate V,
erroneously low estimates of velocity may result (Bear, 1979; Gibb et al.,
1985).
Values of hydraulic conductivity (K) depend on properties of the fluid (such
as viscosity) as well as that of the porous medium. In general, the hydraulic
conductivity is anisotropic, i.e., the K for horizontal flow (K,) is not
equal to the K for vertical flow (K ). The hydraulic conductivity K can
also be expressed by (Luthin, 1957):
K = C(d5orPg/u
(1-6)
where:
C - proportionality constant for the medium, based
geometric characteristics (dimensionless)
d,._ = mean grain diameter of the medium (length) |
3
p = density of the fluid (mass/length )
u - kinetic viscosity of the fluid (mass/length x time)
2
g = acceleration of gravity (length/time )
on
In this equation Cd,-n (also defined as k, the specific or intrinsic
permeability) is a function of the medium alone, while pg/u is a function of
the fluid alone. The hydraulic conductivities for fluids other than water can
be estimated using these relationships and the other fluids' viscosities and
densities.
Hydraulic conductivity does not remain constant but may vary over time, for
example, from increased swelling of soil clay particles, change in pore size
distribution or change in the chemical nature of the soil-water interactions.
In particular, the addition of chemical reagents during in situ treatment may
significantly alter this value. It is therefore important to evaluate the
hydraulic conductivity under conditions of reactant saturation.
8
-------
Obtaining a representative value of the hydraulic conductivity at a particular
site is essential to establish the rate at which a treatment solution can move
through the waste deposit and surrounding soil media. Not only is it
important to obtain hydraulic conductivity of the medium but the hydraulic
conductivity of the waste deposit must also be determined. The relationship
between the K values of the host material and the waste material will dictate
which delivery or recovery methods may be applicable and effective. For
example, attempting to deliver a fluid by gravity into a waste and host
material both with hydraulic conductivities of lxlO~3 cm/sec (2.8 ft/day)
would normally be feasible; however, if the waste material is surrounded by a
host material with a hydraulic conductivity of 1x10 cm/sec (28 ft/day),
the fluid will try to follow the path of highest hydraulic conductivity (path
of least resistance) and travel within the host material, not effectively
entering and permeating the waste deposit.
Several methods, including laboratory and field (slug and pump) tests, that
can be employed to obtain a value for the saturated hydraulic conductivity of
a particular medium. The theory, implementation and interpretation of these
methods are described in detail in several sources, including USEPA (1977);
USEPA (1980a), Kepa and Kufs (1985), Olson and Daniel (1981), Freeze and
Cherry (1979) and Luthin (1957).
Laboratory tests using permeameters can be made on undisturbed core samples
taken in the field with an appropriate core sampler. These methods give the
vertical hydraulic conductivity. Although inexpensive and conceptually easy,
laboratory hydraulic conductivity tests require much care to achieve accuracy.
Furthermore, laboratory determined hydraulic conductivities may not agree with
field measurements from the same location (e.g., Olson and Daniel, 1981).
Hydraulic conductivity can also be crudely estimated according to the
effective grain size of a soil. The effective grain diameter (d1Q) means
that 10 percent (by weight) of the soil particles are finer than the specified
diameter. Table 1-1 shows that, for example, media with effective grain size
(d10) between 0.6 to 0.1 mm (0.02 to 0.004 inches) will have a hydraulic
conductivity between approximately lxlO~ and 1x10 cm/sec (280 and 2.8
-------
TABLE 1-1
HYDRAULIC CONDUCTIVITY, POROSITY AND DRAINAGE CHARACTERISTICS OF MATERIALS (USEPA, 1980b)
K (HYDRAULIC CONDUCTIVITY)
1
28,
D 1 KT1 ID"2 TIT3 1(T*
XX) 2800 280 28 2.8 0.28
I I I I I
10-5
0.03
1
1D-8
0.003
1
10-7 10"8 CM/SEC
0.0003 0.00003 FT/DAY
I 1
1 10-1 TO'2 lo'3 104 Iff5 IffO ID'7 10-8 CM/SEC
2 1 .6 .2 .1 .06 .02 .01 .006 .002 .001 " MM
I I I I I I I '
0.08 0.04 0.024 0.008 0.0040.0024 0.0008 0.00040.0002 0.00003 0.00004 INCHES
EFFECTIVE GRAIN DIAMETER, d10
GRAVEL CS.SAND | MED. SAND | FINE SAND | CS. SILT
CLEAN GRAVELS _, .. VERY FINE SANDS
, CLEAN SANDS , ,
1* "1 1"
COARSE FINE
. SAND-GRAVEL MIXTURES. TILL ,
MED. SILT
_^t
FINE SILT
SILTS. ORGANIC & INORGANIC
CLAY
-i
VARVED CLAYS. ETC. - i
HORIZONTAL K
,. SAND-SILT-CLAY MIXTURES. TILL
VERTICAL K
rl
K
RANGES OF
TVflCSl
REAL SOILS
RANGE OF POROSITY VALUES
FOR VARIOUS MATERIALS
Type
very fine sand
fine sand
concrete sand
fine to medium sand
medium sand
medium to coarse sand
clean gravel
concrete gravel
cm/ sec
5 x 10~3
2 x 10~2
2 x 10~2
5 x 10~2
W1
1.5 x 10"1
10
25
ft/day
14
57
57
140
280
425
2.8 x 104
7.1 x 104
Type
clay
sand
gravel
sand and gravel
municipal waste
- --
Porosity
45 -
35 -
30 -
20 -
30 -
55
40
40
35
40
-------
ft/day). Media with d,n between 0.1 to 0.005 mm (0.004 to 0.0002 inches)
_O _£
will have a hydraulic conductivity between 1x10 and approximately 1x10
cm/sec (2.8 to 0.003 ft/day).
The field bore-hole method or single well pump test is a commonly used method
for measuring the in situ hydraulic conductivity of saturated soils. It is
considered to be the simplest and most reliable of the available methods.
Studies conducted through the years demonstrate that this method primarily
measures the horizontal component of the hydraulic conductivity. The
hydraulic conductivity can also be determined in the field by pumped-well
aquifer tests using a series of wells. This method involves discharging or
recharging one well at a known rate and measuring the response of water levels
in the observation wells. Water level responses are then mathematically
related to the hydraulic conductivity (Lohman, 1979; Freeze and Cherry, 1979;
Repa and Kufs, 1985). Typical hydraulic conductivities of various soil
materials are listed in Table 1-1.
Methods are also available for measuring the vertical hydraulic conductivity
in the unsaturated zone. These include the double-tube method, the gradient
intake method, and the use of an air entry permeameter. Detailed discussions
of principles and equipment involved in each of these methods can be found in
Bouwer (1964, 1966), Bouwer and Rice (1967), and Black (1965).
1.1.3.3 Specific Yield, Specific Retention, Porosity and Grain
Size Distribution
Porosity is important in determining the quantity of fluid which can be
physically accommodated by the media during delivery, and the velocity of
groundwater (or reactant) flow through the media. Specific yield and specific
retention values are important properties for modeling groundwater Jflow. In
addition, these parameters indicate how much of the delivered reactant will
remain within the soil or waste deposit, and how much will be released. An
important consideration is that these parameters should be determined for the
waste deposit in addition to the surrounding soil, since the waste deposit may
control the maximum application rate and flow rate.
11
-------
The specific yield of soil (with respect to water) is defined as the volume of
water which will drain by gravity from a saturated soil sample, divided by the
total volume of the soil sample. For relatively coarse-grained materials
(sands and gravels), .the specific yield is approximately equal to the
effective porosity for flow (see Equation 1-5). The specific yield is
expressed as a percentage (Johnson, 1967): >
where:
Sy - 100 (VW/VS)
w
(1-7)
specific yield (dimensionless)
3
volume of liquid removed by gravity (length )
3
volume of soil (length )
The specific retention of soil with respect to water is defined as the volume
of water which will be retained in an initially saturated soil sample against
the pull of gravity, divided by the total volume of the sample. It is
expressed as a percentage (Johnson, 1967):
R - 100(Vr/Vs) (1-8)
where: R s specific retention (dimensionless)
V = volume of water retained by the soil against the
r 3
pull of gravity (length )
3
V s volume of soil (length )
s
Porosity (n) can be defined as the ratio of the aggregate volume of the
interstices (pores) of the rock or soil sample to its total volume. It can be
expressed as a percentage by the following equation (Johnson, 1967):
n - 100 (Vw/Vg) = 100 (Vs-V)/V£
(1-9)
where:
n
V
w
total porosity (dimensionless)
3
total volume of soil or rock (length )
3
volume of water required to saturate the sample (length )
aggregate volume of the solid particles that make up
3
the sample (length )
12
-------
The porosity of a sample is best measured in the laboratory. Gibb et al.
(1985) describe experiments to measure the effective porosity (n ff) in
geologic materials. It can also be estimated using a grain-size distribution
curve (e.g., USEPA, 1977). Typical porosities of various soil materials are
listed in Table 1-1. An example of the relationship between porosity,
specific yield and specific retention according to grain size is shown in
Figure 1-1.
Various laboratory and field test methods have been developed to determine
porosity, specific yield and specific retention. These are reviewed in
Johnson (1967), USEPA (1977), Bouwer (1978) and Freeze and Cherry (1979).
1.2 Delivery Technologies
Delivery techniques used for artificial groundwater recharge or wastewater
treatment by land application may be applicable to waste deposit remediation
efforts. These techniques introduce water or reactant solutions into waste
deposits in order to react with contaminants in the waste deposits or flush
contaminants from the deposits to the groundwater table. Flushed contaminants
can subsequently be collected and treated above ground. The available
delivery methods are grouped into two generic categories: gravity and
forced. Gravity methods apply the flushing or reactant solution directly over
the waste deposit (if the waste deposit is at the surface) or deliver the
solution through the surrounding soil to the waste deposit. Forced delivery
methods inject the flushing or reactant solution directly into the waste
deposit or surrounding soil through pipes by means of an applied pressure. In
both cases the solution enters the groundwater for subsequent recovery after
passing through the waste deposit. When considering any delivery (or
recovery) method, the reactant and groundwater flow should be modeled (using
conventional flow net analysis or mathematical models) so that design
parameters can be tested and proper delivery of reactant and recovery of spent
solution can be assured. Groundwater flow analysis is described in most
groundwater texts, including Freeze and Cherry (1979) and Cedergren (1977),
and mathematical modeling is reviewed in Wang and Anderson (1982).
13
-------
FIGURE 1-1
POROSITY, SPECIFIC RETENTION AND
SPECIFIC YIELD VARIATIONS WITH GRAIN SIZE
50
45
40
35
30
25
20
15
10
5
0
\ i i i i i i i \ i i i r
POROSITY
SPECIFIC RETENTION
SAND
GRAVEL
V)
cc
UJ
o
O
CD
I/I
1/16 1/8 1/4 1/21 2 4 8 16 32 64 128 256 M
0.0020.0050.010.020.040.080.160.320.631.3 2.5 5.0 10.1 INCHES
MAXIMUM 10% GRAIN SIZE
NOTE:SOIL SAMPLE FROM SOUTHERN COASTAL BASIN, CA,
CLASSIFICATION SHOWN IS THE LABORATORY CLASSIFICATION OF
THE SAMPLE.
SOURCE: US EPA, 1977
14
-------
The following sections present discussions of the design considerations of the
following delivery technologies:
o Gravi ty
- Flooding
- Ponding
- Spray Irrigation
Ditches
Subsurface Spreading (infiltration galleries and beds)
o Forced
- Injection Wells
1.2.1 Gravity Delivery Methods
Gravity delivery methods are applicable at or near the ground surface and can
be classified into two groups: surface and subsurface spreading. Surface
spreading involves the application of the flushing or reactant solution over
the waste deposit or overlying soils. Subsurface spreading requires the
distribution of the flushing or reactant solution through the use of
infiltration galleries or beds. The selection of a gravity delivery method
depends on the following four factors:
Infiltration Rate and Soil Hydraulic Conductivity - The infiltration rate of
the surface soil and hydraulic conductivity of the waste deposit and overlying
soil affects the volume of solution that can be applied by gravity delivery
methods. The infiltration rate can be increased by surface soil preparation
(i.e., tilling), but increasing the hydraulic conductivity of subsurface
materials is generally difficult or impossible.
Location of the Waste Deposit - Location of the waste deposit with respect to
the ground surface and water table affects the selection of a gravity delivery
method. If the waste deposit is exposed at the surface, then surface
application methods such as flooding, spraying and ponding can be considered.
15
-------
As the depth to the top of the waste deposit increases, the effectiveness of
surface application is reduced because of lateral spreading or attenuation of
the applied solution. If the waste deposit is located below the groundwater
table, surface application will not be successful because the applied solution
will be diluted, and not readily penetrate the water table to reach the waste
deposit (although if its density is significantly greater than that of
groundwater penetration may occur).
Topography -- The topography of the waste disposal site is a primary factor in
the selection of a delivery method. Flooding or ponding should only be
considered for terrain with slopes less than 3%. Trenching or ditching would
be more effective for topography characterized by slopes greater than 3%.
Climate - The climatological influences at a site affect the selection of
gravity delivery methods. Of particular importance are frost penetration
depth and groundwater level variations (caused by seasonal changes or tidal
effects). .
1.2.1.1 Selection of Gravity Delivery Methods
In developing the selection methodology for gravity. delivery methods, it is
assumed that the deposit will be completely saturated with the applied
solution, and the applied solution will be recovered by interception of the
water table, i.e., gravity recovery (this assumption is necessary for
calculation of required application rates).
Before selecting a gravity system,1 the following conditions must be known
about the site:
o Surface topography and area of the site
o Sustained infiltration rate (I)
o Configuration of the waste deposit
- areal dimensions (LxW)
- thickness of deposit (d)
- depth to deposit (d ) from the surface
16
-------
o Aquifer thickness (above impermeable layer) before solution
application (H,)
o Hydraulic conductivity (K) of the waste deposit and surrounding soil
media (based on minimum value)
o Depth to water table (S ) from the surface
o Porosity (n) based on maximum value.
Once these parameters are known the required application rate and the
application rate attainable by gravity methods can be calculated to determine
whether gravity delivery methods are viable.
1.2.1.2 Determination of the Required Application Rate
To satisfy the treatment objectives of complete saturation of the waste
deposit and required detention time, a solution should be applied at a certain
rate. For gravity recovery (using a drain or buried pipe placed perpendicular
to groundwater flow) this rate can be determined by using the following
equation (USEPA, 1977; Huisman and Olsthoorn, 1983) (see Figure 1-2):
K
- Hd) /L ( L + 2X)
(1-10)
where:
X =
L =
K =
Total saturated thickness above aquiclude required to meet the
assumption that the waste deposit is fully saturated (length)
Elevation of the recovery system above the impermeable layer
(i.e., the original water table elevation before solution
application) (length)
Distance from the edge of the deposit to the recovery
system (length)
Length of the deposit parallel to the groundwater flow (length)
Average hydraulic conductivity of the waste deposit and
surrounding soil (length/time)
Required application rate per unit area of the deposit to
satisfy the saturation criterion (length/time) .
17
-------
FIGURE 1-2
GENERAL LAYOUT OF A GRAVITY DELIVERY
AND RECOVERY SYSTEM
00
r
_X-J
INDUCED
WATER TABLE
9 » >»»*"""'"*
ORIGINAL
WATER TABLE
/////////////////// CLAY OR BEDROCK
-------
In this equation (USEPA, 1977) all values except q-, and X are measured. The
application rate, q,, multiplied by the area of application (L x W)
represents the total flow of solution through the saturated waste deposit and
surrounding soil to the recovery system which is located at a distance, X,
from the edge of the waste deposit. For continuous in situ treatment
operation, the unit flow rate, q,, is also the required recovery rate to
maintain the treatment processes in a stable, steady state condition.
Assuming a reasonable distance of X for the recovery system from the edge of
the waste deposit, the application rate, q,, can be estimated by solving
equation (1-10).
Another equation expresses the relationship between the application rate
(q..), time (t) required for saturation of the waste deposit and surrounding
soil media, and the location of the recovery system (X) as follows:
t = n (Hc + Hd) X/2q W
(1-11)
where:
t = time required for saturation of waste deposit and
surrounding soil media (time)
n = porosity (dimensionless)
w = width of deposit perpendicular to groundwater flow
and other parameters are as defined above.
1.2.1.3 Determination of the Maximum Application Rate by Gravity
Methods
The amount of solution that can be introduced into a deposit by gravity
delivery methods depends on the area of application (wetted area) and its
sustained infiltration rate (assuming either that the deposit is at the
surface, or the infiltration rate is less than the hydraulic conductivity of
the subsurface soil so that the hydraulic conductivity is not a limiting
factor). For spraying or surface flooding, the area of application is
19
-------
determined by the areal extent of the waste deposit. The total amount of
solution (Q0) that can infiltrate into the deposit would be:
Qo- AI8 (1-12)
3
where: Q = maximum application rate of the deposit (length /time)
2
A s area of application (length )
I - sustained infiltration rate (length/time)
s
The results of this equation, should be compared to equation (1-10) to check
that the hydraulic conductivity of any soil layers between the surface and the
waste deposit is not rate-limiting.
For application using ditches Q will depend on the wetted area (sides and
bottom) of the ditch and the sustained infiltration rate through the wetted
area at a particular level of solution in the ditch. This can be calculated
if the infiltration rates through the ditch sides and bottom are determined in
field pilot studies (or assumed to be the same as the soil infiltration
rate). For application using ponds, infiltration beds or galleries, the soil
permeability will be the rate-limiting parameter. In this case, the equation
for delivery rate of solution would be Darcy's law (Equation 1-3) where A =
area of bottom of the pond, infiltration beds or galleries.
i-
The parameter qQ (= QQ/A, the application rate per unit area of the
delivery system) represents the surface infiltration or subsurface delivery
rate (using Equation 1-3) achievable by a gravity delivery method. If q, is
set equal to q in Equation (1-10), solving for X will define the minimum
distance at which to locate the recovery system for the shortest treatment
pathway and the minimal treatment time.
To achieve saturation, which is a basic assumption for this methodology, q
must be equal to or greater than q-,. If q is less than q-,, the next
step would be to determine whether q can be increased. Such increase could
be effected by increasing applied head of the solution, varying soil and waste
properties by tilling (USEPA, 1977), or introducing forced delivery systems.
20
-------
1.2.1.4 Surface Applications
Flooding — In effecting solution application by flooding, the solution is
spread over the land surface in a thin sheet. This technique is similar to
the flooding that is practiced as an irrigation method for agricultural land.
It is also used as a method of artificial recharge for aquifers located near
the ground surface. The method is effective and of low cost in areas which
are flat or gently sloped (generally less than 3 percent slope) and uniform
(gullies or ridges are absent) (ASCE, 1972), the waste deposit is at or near
the surface, and soil and waste deposit infiltration and hydraulic
_3
conductivity are high (greater than 10 cm/sec or 2.8 ft/day). Preferred
soil and waste deposit conditions are those similar to sands (SW), loamy sands
(SM-SW), and sandy loams (SM) (ASTM, 1969).
Design parameters utilized for infiltration-percolation systems for
wastewaters can also be applied to surface flooding systems. As seen in Table
1-2, an application rate of 300 cm (120 inches) per week is possible when
waste deposit and surrounding soil characteristics are similar to those of
'."'•' ' O
sandy soils. This would be the equivalent to approximately 5200 m /per
hectare (560,000 gallons per acre) per day.
In flooding applications, ditches can be used to distribute the solution
across the up-slope end of the waste deposit area. Weirs placed at regular
intervals along the ditch divert the solution to the spreading area. The
direction of flow in the spreading area can be controlled by strategically
placing embankments. Peripheral berms and a collection ditch at the lower end
of the area are required to prevent the solution from flowing out of the
application area. A typical plan of a ditch flooding system is shown on
Figure 1-3. Information on construction parameters and costs of berm
construction is presented in A D Little (1983).
Since the liquid is applied directly to the soil surface, even a thin layer of
impermeable material between the surface and the waste deposit would impede
infiltration of the liquid and make this application method ineffective. The
natural infiltration rate can be enhanced by tilling or furrowing the surface
21
-------
TABLE 1-2
DESIGN LIMITATION FOR INFILTRATION - PERCOLATION SYSTEMS
Factor
Range of Feasible Values
Liquid loading rate
10 to 300 cm/wk, or 2 z 10~5 to
6 x 10~4 cm/sec (4 to 120 in/wk)
Annual application rate
5 to 155 m/yr (17 to 500 ft/yr)
Land required for
1-mgd (3785 m3/day)flow
1 to 25 hectares (2 to 62 acres)
plus buffer zones
Application
techniques
Usually surface
Soils
Rapidly permeable soils, such as
sands, loamy sands and sandy loams
Source: Pound and Crites, 1973
22
-------
FIGURE 1-3
TYPICAL PLAN OF DITCH AND
FLOODING TYPE GRAVITY DELIVERY SYSTEM
PREVAILING GROUND
SLOPE GENERALLY
LESS THAN 3 PERCENT
DISTRIBUTION
DITCHES
DITCH OUTLET
SURFACE
COLLECTION
DITCH
SUPPLY DITCH
HAZARDOUS WASTE
DISPOSAL BOUNDARY
23
-------
soil (USEPA, 1977; USEPA, 1984b), but deeper impermeable layers may not be
reached by this method and other delivery systems will therefore be required.
Flooding should be implemented in a uniform manner so that dry spots do not
result. Because the solution is applied in a thin sheet over the ground
surface, it is susceptible to freezing, so this method is limited to use
during non-freezing conditions. Also, because the solution is exposed to the
atmosphere, this method is not suitable for application of reactants which are
volatile or susceptible to photo-oxidation. i
Ponding — Ponding can be used to increase the infiltration rate of the
applied solution above that achieved by flooding. Ponds can be constructed
either by excavating a few feet into the ground or by constructing low berms.
The bottom of the pond is utilized as an infiltrative surface for the solution
to enter the ground and the depth of the solution in the pond becomes the
driving force to increase infiltration rates (i.e., the gradient, I, in
Equation 1-3 is greater than 1.0).
The ponding method is suitable when the deposits are of a sandy or loamy (SM
or SW, ASTM, 1969) nature and when the ground surface is relatively flat. For
irregular terrain, a large number of ponds or considerable excavation would be
required. Although there is not a specified maximum slope for pond
construction, constructing a pond would be progressively more difficult on
steeper slopes. As an example, a pond with a length of 100 meters on a 10
percent slope would require a downgradient berm with a height greater than 10
meters. If 2:1 (horizontal:vertical) side slopes were used, the width of the
base would be greater than 40 meters.
A pond can be constructed on level ground without excavating any material from
the surface. By surrounding the area with low levees or berms, the liquid can
ba contained. In this manner, the contaminated area can be treated without
having to remove any soil. Such a system is depicted on Figure 1-4.
Information on construction costs of berms is presented in A D Little (1983).
When the waste deposit is overlain by a relatively thin layer (less than 2
meters or 6 feet) of impervious material, the ponding method can be utilized
by excavating the impervious layer. Figure 1-5 depicts such a system.
24
-------
FIGURE 1-4
GRAVITY DELIVERY SYSTEM
USING PONDING
ro
on
TO TREATMENT
RECOVERY
WELL
POND
J_
INDUCED GROUNDWATER TABLE
WASTE DEPOSIT
ORIGINAL
GROUNDWATER TABLE
CLAY OR BEDROCK
TO TREATMENT
RECOVERY
WELL
-------
FIGURE 1-5
GRAVITY DELIVERY SYSTEM
USING PONDING IN A THIN
IMPERVIOUS STRATUM
ro
INDUCED
WATER
TABLE
RECOVERY
WELL
>>(\ i I Us
V 4 y
HAZARDOUS
WASTE DEPOSIT
IMPERVIOUS
STRATUM
ORIGINAL
GROUNDWATER
TABLE
RECOVERY
WELL
CLAY OR BEDROCK
-------
Surface Spraying — Sprinkler-type irrigation systems are used to deliver the
liquid directly to the ground surface (Figure 1-6). This technique has
commonly been used for land-based treatment of wastewaters (USEPA, 1977).
Sprinkler distribution simulates rainfall and is less susceptible to
topographic constraints than other surface methods. The sprinkler
distribution system can be applied on a ground slope of up to 20 percent
(USEPA, 1977). Surface spraying is most effective when the deposit is at the
ground surface and has a high infiltration rate.
Surface spraying systems consist of one or a series of sprinkler heads
connected to a header pipe. The procedure for sprinkler system design
involves the determination of the optimum rate of application, sprinkler
selection, sprinkler spacings and performance characteristics, and design of
laterals (Fry and Grey, 1971; USEPA, 1977). Surface spraying involves a
significant utilization of equipment and its capital and operating costs are
substantially higher than those for other gravity methods.
The optimum rate of application for a sprinkler system is the rate that
ensures uniform distribution under prevailing climatic conditions without
exceeding the infiltration rate of the soil (USEPA, 1977). Sprinkler
selection is based primarily on conditions of service, such as type of
distribution system, pressure limitations, application rate, clogging
potential, and effects of winds (USEPA, 1977). Sprinkler spacings and
performance characteristics are jointly analyzed to determine the most uniform
distribution pattern at the optimum rate of application. USEPA (1977) and Fry
and Grey (1971) contain detailed information on sprinkler system design.
A W Martin Associates (1978) provides order-of-magnitude cost estimates for
sprinkler systems.
Surface spraying is not recommended if volatile organics are contained in the
solution being applied, because much may be lost by evaporation, and
volatilization may create odor problems. In addition, photo-oxidation may
occur during spraying operations. Another important consideration for use of
a sprinkler system is the clogging of nozzles caused by scaling of the applied
solution.
27
-------
FIGURE 1-6
GRAVITY DELIVERY SYSTEM
USING SURFACE SPRAYING
oo
TO
TREATMENT
SPRINKLERS
TILLED SOIL * INDUCED W ATE RT ABLE
RECOVERY WELL
RECOVERY WELL-
CLAY OR BEDROCK
TREATMENT
ORIGINAL
GROUNDWATER
TABLE
-------
Ditches — The ditch method of surface spreading utilizes relatively
flat-bottomed ditches to transport the solution over the application surface
providing the opportunity for percolation. Generally, ditches are relatively
shallow and narrow (1 to 2 meters or 3-6 feet wide) and make use of both the
bottom and side surfaces for infiltration of liquid to the ground. Gradients
in the ditches should be slight for erosion prevention and maintenance of an
adequate residence time for infiltration. Ditches can be constructed by
excavating surface material or building small embankments. Figure 1-7 depicts
a typical ditching system.
Ditches would be effective for surface application in circumstances where it
is not desirable to completely cover the entire area with the reactant
solution. Runoff control is not necessary, since all of the applied solution
is contained within the ditch system. This method of application is suitable
for a subsurface deposit overlain by pervious soils. If the surface layer is
impermeable, this method may still be valid, providing that excavation of the
ditches is deep enough to penetrate into more permeable materials. Ditches
would not be suitable for sites located in areas of very irregular terrain.
Shallow ditches would be limited to use during non-freezing weather periods.
Because this method has less surface area exposed to the atmosphere than
flooding, however, it would be less susceptible to rapid freezing. If the
floor of the ditch is below the frost level, infiltration can take place even
if the ground surface is frozen.
1.2.1.5 Subsurface Delivery Systems
The infiltration gallery (or trench) and infiltration bed delivery methods are
classified as subsurface gravity systems because the direct application of the
liquid is not on the ground surface. These systems consist of excavations
filled with a porous medium (coarse sands or gravels) that aid in distributing
the liquid throughout the waste deposit. The large void spaces of the porous
medium provide for storage and easy delivery of the solution.
29
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FIGURE 1-7
GRAVITY DELIVERY SYSTEM
USING DITCHES
TO TREATMENT
00
o
RECOVERY.
WELL
INDUCED WATER TABLE
WASTE DEPOSIT
GROUNDWATER TABLE
TO TREATMENT
RECOVERY
WELL
CLAY OR BEDROCK
-------
These methods are suitable where the waste deposit is subsurface and where
ground freezing is a recurring problem. Infiltration galleries can be
installed under the freezing zone thereby permitting gravity application of
solution for year round operation. Also, where the subsurface waste deposit
is located at a depth which makes surface application impractical (i.e.,
greater than about 5 meters or 16 feet), or it is overlain by a layer of
impermeable material which is not economic or feasible to excavate, gravity
application of solution is still possible by utilizing subsurface methods.
These methods are also more suitable for application of volatile or
photo-oxidizable materials.
Infiltration Gallery — An infiltration gallery consists of a trench that is
filled with gravel or stones. The solution fills the void spaces in the
gallery and is distributed to the surrounding soils and waste deposit.
Infiltration occurs in both the horizontal and vertical directions (Figure
1-8) (USEPA, 1980b; A W Martin Associates, 1978).
This method works best in cases where the waste deposit and surrounding soils
are of a sandy or loamy (SM or SW) nature. .Hydraulic conductivities of soils
-2 —4
and wastes between 1x10 cm/sec and 1x10 cm/sec are (28 to 0.28 ft/day)
best suited for these delivery systems. This method can be installed to
penetrate an impermeable surface soil, so that the subsurface systems contact
directly with more permeable strata. If the soils are of a silty nature, with
—4 — 5
hydraulic conductivities between 1x10 cm/sec and 1x10 cm/sec (0.28 and
0.028 ft/day) (Table 1-1), these techniques may still be used but the
application rate and therefore the treatment time will be much slower. As
with ditches, the application rates of a gallery are best determined by
inflow-outflow measurements in the field. Design application rates, number,
spacing, and depth of galleries are based on such field results.
The recommended packing of fill media for use in this system is either gravel
or crushed rock sized 2 to 6 cm (0.8-2.5 inches) in diameter. Generally, the
smaller sizes are preferred because the infiltrative surface of the soil has
more direct contact with the liquid. The rock should be washed before being
put in place to remove fines that may clog the bottom infiltrative surface.
31
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FIGURE 1-8
SUBSURFACE GRAVITY DELIVERY SYSTEM
USING INFILTRATION GALLERIES
TO
TREATMENT
BACKFILL-,
TO
TREATMENT
CO
ro
'/T'e-:. _ WASHED
PERFORATED *»*#«\i
PIPE
INDUCED GROUNDWATER
WASTE DEPOSIT
ORIGINAL
GROUNDWATER TABLE
RECOVERY
WELL
RECOVERY
WELL
CLAY OR BEDROCK
-------
The solution can be introduced into the gallery by injection in different
locations along the length of the gallery or through perforated distribution
pipes. The pipe used for the distribution can be constructed of the following
materials: clay, bituminized fiber, concrete, plastic (acrylonitrile-butadiene-
styrene:ABS), polyvinyl chloride (PVC), styrene rubber plastic (SR), or
polyethylene (PE). If water is to be the application liquid, then any of
these materials will suffice. However, some of the plastic-type pipes may not
be suitable for certain organics, bases, acids or other additives. This must
be confirmed with the manufacturer before installation. The perforation size
in the pipes, spacing of holes along the pipes and spacing between galleries
will depend on site specific conditions (USEPA, 1980b).
Infiltration galleries provide effective gravity application methods in
circumstances where other methods may not be feasible, such as in areas of
steep slopes and uneven terrain. Galleries are limited to areas where
topography has slopes less than 25 percent (Pound and Crites, 1973). With
slopes steeper than 25 percent the use of construction equipment may be
difficult.
Infiltration Beds — Infiltration beds (Figure 1-9) are similar to galleries
with the exception that they are wider and contain more than one perforated
distribution pipe. The bed method depends almost entirely on infiltration
through the bottom, with little infiltration through the sidewall surfaces.
This method is suitable when the waste deposit and surrounding soil media have
characteristics like sandy (SW) or loamy (SM) soils. Because of the greater
width of the beds, they are limited to applications in which the topography is
relatively flat (slopes less than 5 percent) (USEPA, 1980b). Beds generally
are less expensive to construct than galleries per unit area because they
require a single excavation, grading and bed-laying procedure.
Typically, the perforated distribution pipes within infiltration beds would be
placed 1 to 2 meters (3-6 feet) apart. In cases of more impermeable soils,
the pipe placement could be as close as 0.5 meters (1.5 feet). The design and
the materials used for the bed system are similar to those for the gallery
system.
33
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FIGURE 1-9
SUBSURFACE GRAVITY DELIVERY USING
INFILTRATION BED SYSTEM
GROUND LEVEL
TO
REATMENT
TO
TREATMENT
INDUCEDV3ROUNDWATER LEVEL
ORIGINAL
GROUNDWATER TABLE
CLAY OR BEDROCK
34
-------
These systems would have the same limiting factors and effectiveness as
infiltration gallery systems; however, infiltration bed systems are more
limited by steep slopes and uneven terrain. Beds have the advantage of
saturating a much larger area than a single trench. A bed is also simpler to
install than a comparable multi-trench system, since the piping goes to a.
common header and the entire system is installed as one single excavation.
1.2.2
Forced Delivery Methods
Forced injection is the process in which a fluid under pressure is forced into
the waste deposit and surrounding soil through pipes which have been
strategically placed to deliver the solution to the zone requiring treatment.
This method is generally suitable for a deposit having a hydraulic
conductivity greater than 1x10 cm/sec (0.28 feet/day) which would
represent a fine sand/coarse silt material.
The injection process may be accomplished by using either an open end or
slotted pipe. An open end injection pipe as shown on Figure 1-10 consists of
an EW size (3.5 cm or 1-3/8 inch OD) pipe or equivalent. The lower element of
the pipe contains an expendable .and movable point for driving into thfe ground
without plugging. Additional lengths of pipe, are added as it is driven to the
desired depth. The slotted pipe (Figure 1-10) may be plastic (PVC) pipe, 3.8
cm (1.5 inch) in diameter placed in an 8 to 10 cm (3 to 4 inch) borehole. The
lower portion of the pipe is slotted over an interval corresponding to the
zone to be treated, and surrounded by gravel or coarse sand as shown on Figure
1-10. Above the slotted portion, a cement grout is placed around the pipe up
to the ground surface. Figure 1-11 shows the application of injection wells
in a waste deposit.
A forced delivery system, unlike the gravity systems, is conceptually
independent of surface topography and climate and can be designed to
accommodate any of the waste deposit configurations that have been discussed.
Since the applicability and design of the forced injection delivery system
depends heavily on the site geohydrological conditions, the site must be
investigated by means of test borings with field hydraulic conductivity
35
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FIGURE 1-10
INJECTION PIPES
FOR FORCED DELIVERY
—V
co
en
•v-
EW ROD OR
EQUIVALENT
EXPENDABLE
PLUG
CEMENT GROUT
GRAVEL OR COARSE
SLOTTED PIPE
OPEN END
SLOTTED
-------
FIGURE 1-11
FORCED DELIVERY USING
INJECTION WELLS
DIRECTION
OF
GROUNDWATER
FLOW
O
O
O
INJECTION WELLS
O
• RECOVERY WELLS
WASTE DEPOSIT
PLAN VIEW
INJECTION WELL
TO
TREATMENT
INDUCED
GROUNDWATER
TABLE
WASTE
DEPOSIT
ORIGINAL
GROUNDWATER
TABLE
DIRECTION OF INDUCED FLOW
CLAY OR BEDROCK
CROSS SECTION
37
-------
testing as well as laboratory geotechnical testing. Test borings serve
primarily to establish or confirm the waste deposit configuration and the
depth to the groundwater table, both of which will be essential in the layout
of a forced injection system. In situ and laboratory tests can provide data
on hydraulic conductivity, rates of groundwater flow and dispersion of
injected solutions, particle size, and porosity, which are also needed for a
system design.
The particle size analysis would give an indication of the "injectability" of
the site soil in question. In general, a soil is not considered "injectable"
if over 10 percent of the sample passes the #200 sieve. However, injection
may be successful, albeit to a lesser degree, when the soil sample exhibits
10-20 percent fines (passing the #200 sieve). The potential for injection
under this sub-optimal situation would have to be investigated for the
specific deposit. Forced injection will work best in well-sorted granular
materials having relatively high effective porosities (specific yields)
ranging from 25 to 55 percent and average particle sizes larger than that of
fine sand/coarse silt, about 0.05 to 0.1 mm (0.002 to 0.004 inch) (US Dept of
Navy, 1982). In situ pumping tests can be made in exploratory borings to
provide information on both the hydraulic conductivity and the flow rates
under different injection pressures. The in situ hydraulic conductivity of
the undisturbed soil determines the rates at which the fluid will be accepted
under varying pressures.
A maximum injection pressure must be established to prevent hydraulic
fracturing and uplift in the deposit. This fracturing might causse the fluid
to travel toward the surface rather than seeping through the formation. To
—4 2
avoid this the injection pressure should be kept below 1.5 x 10 N/m per
meter (1 psi per foot) of overburden above the injection level (Winterkorn and
Fang, 1975). If the rate of injection is kept constant, the pressure measured
at the entry of the injection hole depends on the size of the voids in the
soil (i.e., porosity), the viscosity of the solution, and the hydraulic
conductivity of the soil. These three factors acting as resistances determine
the relationship between pressure and rate of injection.
38
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Depending on the injection pressure and the corresponding flow rate (Q) that
is selected, a spacing between injection holes can be determined using the
following formula (Huisman and Olsthoorn, 1983):
1/3
where:
r = 0.62 (Qt/n)
(1-13)
r = radial distance of solution penetration (length)
o
Q = rate of solution application (length /time)
n = porosity of soil (dimensionless)
t = pumping time (time)
The pumping time (t) is determined by the configuration of the waste deposit
and delivery/ recovery system. It will be determined by dividing the measured
flow rate (Q) by the theoretical distance traveled by an element of fluid from
the injection point to the recovery point. This must all be within the time
framework set up for the clean-up operation at the site. Based on the above
equation, the well spacing should be approximately 2r. A grid work of
injection wells would be set up accordingly, making sure to cover completely
the contaminated area in plan.
It should be noted that open end pipes .would be better suited for soils or
waste deposits where it appears that the hydraulic conductivity in the
vertical direction approaches the hydraulic conductivity in the horizontal
direction (i.e., K =K.) . Slotted pipes, on the other hand, eject the
solvent in the horizontal direction along the axis of the riser pipe.
Therefore, they would be more useful in soils or waste piles where it appears
that the horizontal hydraulic conductivity exceeds the vertical hydraulic
conductivity (i.e., KV is less than Kh) .
1.2.3 Summary and Example Applications
As discussed in the foregoing sections, gravity and forced methods may be used
for delivery of solutions to waste deposits. A number of gravity delivery
methods including flooding, ponding, surface spraying, ditches, infiltration
galleries and infiltration beds may be used. The selection of a particular
39
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gravity delivery method would depend upon: surface topography, infiltration
rate, configuration of waste deposit, groundwater hydrology, hydraulic
conductivity, soil porosity and local climate. Forced delivery would be
cost-effective for a waste deposit and surrounding soil having a lower
hydraulic conductivity (down to 1 x 10 cm/sec, or 0.28 ft/day) and low
infiltration rate(below 10 cm or 4 inches per week), which would preclude the
delivery of treatment solution by gravity. Such systems would typically be
comprised of an injection pipe (open end or slotted) and pump. The
applicability of any delivery system would depend heavily on the site
geohydrological conditions. Test borings and in situ pumping tests should
therefore be conducted prior to the engineering and design of the delivery
system.
Determining the required application rate of treatment solution is the most
important engineering effort for both gravity and forced delivery systems.
The solution application rate must be established based on consideration of
various site parameters and waste deposit characteristics as well as the
location and rate of recovery system operation. An example is presented below
to demonstrate the procedures to estimate the required/allowed application
rate of treatment solution.
In this example, it is assumed that a field geohydrological survey was
conducted and the following data were generated: ;
o The waste deposits lie immediately above the water table.
o The length (L) of the waste deposit parallel to the groundwater
gradient is 30 m and its width is 40 m; i.e., area (A) = 1200 m2.
o Height of the original water table above an impermeable layer (H,)
is 3 m.
o The thickness of the waste deposit is 1.5 m; thus the . thickness of
induced saturation (H ) would need to be 1.5 m + H, = 4.5 m.
c d
40
-------
o At the site, the surrounding soil is sandy with a hydraulic
conductivity of 1 x 10 cm/sec.
o Field testing has shown that the sustained infiltration rate (I )
is approximately 3.5 x 10 cm/sec and the soil porosity (n) is 45%.
Laboratory geotechnical and waste tests have determined that the
hydraulic conductivity of the waste deposit is approximately
1 x 10 cm/sec, and the reaction time (See Section 2-5)
requires a maximum of 10 minutes.
The first step in establishing a delivery system concept is to determine the
allowable application rate of solution based on the infiltration rate of the
soil. Equation (1-12) is applied as follows:
Q = AI
os
= 1200 x 3.5 x 10 = 4.2 x 10 m /sec = 0.25 mJ/min, and
/ O
q= Q/A=I =2.1x10 m per square meter per minute.
(1-14)
Thus, the maximum rate at which solution can be introduced into the soil is
O / O
0.25 m /minute, or 2.1 x 10 m per square meter per minute.
The second step in the process is to determine the required application rate
of solution based on the hydraulic conductivity and recovery system location
(X). This application rate represents the flow required to maintain the in
situ treatment system at steady state conditions, i.e., to maintain the waste
deposit entirely under a saturated condition with a recovery system operated
at a designated distance, and at a recovery rate equal to the delivery rate.
The hydraulic conductivity of surrounding medium and waste deposit (K) is 0.06
m/minute. Equation (1-10) is then applied as follows:
qI = K(Hc2-Hd2)/L(L+2X)
= 0.06 (4.52-32)/30 (30 + 2X)
qn = 0.68/(900 + 60X)
41
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The third step in the process of establishing the conceptual design is to set
q, s q to define the minimum distance the recovery system should be
located from the edge of waste deposit (xmin)« Setting q-^ equal to 2.1 x
10 m/min (from above) and solving for X, one obtains X_. = 39 m, which
can be rounded to 40 m.
The fourth step is to determine the time required to saturate the waste
deposit and surrounding soil prior to the commencement of steady state
operation. Equation (1-11) is applied for this purpose as follows:
t = n(H - H, )X
c d
2qW
0.45 (4.5 + 3) x 40
80q
1.69
(1-11)
If the actual application rate q = q = q,, t = 8036 min = 5.6 days. It
appears that the duration of approximately 6 days to saturate the waste
deposit completely and to start the recovering operation is reasonable. Based
o
on t = 6 days, the initial application rate will be 0.27 m /min. Thereafter,
o
operated at 0.27 m /min (71 gpm) the delivery/recovery system would be at
steady state.
Another illustration for consideration is an area of clayey soil with a waste
deposit that has a low permeability similar to the surrounding clay medium.
Presented below are the basic assumptions for this example.
HC - 4.5 m
H, - 3 m
d -4 -5
K = 1 x 10 cm/sec = 6 x 10 . m/minute
L = 30 m
(Waste Deposit Area (A) = L x W = 30 x 40 m2 = 1200 m2
n - 0.3 (30%) ;
I = 7 x 10~5 cm/sec
s ,
42
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Determining q and q, yields:
o 1
7 x 10 cm/sec
-5/3
4.2 x 10 m /square meter per minute
= 6 x 10~5 (4.52 - 32)/30 (30 + 2X)
= 6.8 x 10~4/(900 + 60X)
Letting q = q1 results in X = -14.7 m. A negative solution indicates
that the soil hydraulic conductivity (as reflected by q,) rather than
infiltration (as reflected by q ) governs the application process. Thus, a
solution for this situation can be achieved only by assuming the location of a
recovery system, calculating the saturation (q,) value, checking that it is
less than q , then checking this against q and t estimated for the site.
As an example, assume a reasonable distance of 25 m for the location of the
recovery system downgradient of the waste deposit. Then,
q_ = 3.8 x 10 m/minute
t = 0.30 (4.5 + 3) 25/80q
" 0.70/q
Letting q = q-, t = 1.9 x 10 inin = 128 days.
The practical application flow rate (Equation 1-1.2) will be:
Qo = q±A = 3.8 x 10~6 x 1200 = 4.6 x 10~3 m3/min (1.2 gpm)
Delivering an approximate flow of 4 liters per minute over an area of 1200
2
m is not considered a reasonable practice. Similarly, a saturation time of
over 3 months may be considered unreasonable. Therefore, a forced method
should be applied instead of a gravity delivery method.
1.3 Recovery Technologies
The available recovery technologies can, like delivery technologies, be
grouped into two general categories: gravity and forced methods. The
43
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recovery technologies discussed herein for in situ treatment of waste deposits
are those widely used in groundwater recovery and construction dewatering
operations. Gravity recovery depends upon interception of the groundwater
downgradient from the waste deposit (i.e., down the regional groundwater
gradient or radially in the case of an induced groundwater mound). Thus,
after applied treatment solutions pass through the waste deposit and enter the
groundwater, the resultant fluid is collected in an interceptor system (i.e.,
open ditch or buried drain) by simple gravity flow. -Forced recovery systems
utilize well points, deep wells or vacuum well points located downgradient of
(or radial to) the waste deposit to remove spent solutions by mechanical
means. The primary factors affecting the application of recovery methods
are: depth to groundwater, depth to impermeable layer, and geohydrologic
properties of the waste deposit and surrounding soil.
Depth to groundwater is a constraint for gravity systems only to the extent
that' there are practical limits to which excavation can be performed
(typically less than 5 meters or 16 feet: Huisman, 1972) before costs become
excessive or related widths of excavation become impractical. For forced
systems, depth to groundwater affects system installation costs and
operational energy consumption costs. Depth of the water table aquifer (i.e.,
depth to impermeable layer) is a constraint on gravity systems and wellpoints
because it is sometimes necessary (depending on waste and local geohydrologic
conditions) for the recovery system to penetrate much of the thickness of the
aquifer to ensure complete recovery. Thus the operation limits of gravity
systems and wellpoints may preclude the use of these systems in such
circumstances.
The primary geohydrologic property affecting recovery system application and
feasibility is hydraulic conductivity. As hydraulic conductivity is reduced,
the rate of spent solution recovery is decreased, thus increasing the period
necessary for recovery. In a related manner, forced recovery systems will
require greater energy utilization as hydraulic conductivity decreases.
Hydraulic conductivity is related to the grain size distribution, within the
waste deposit and soil. Table 1-1 depicts the general relationship between
hydraulic conductivity and effective grain size (Din) distribution of a soil
44
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with water as the transported fluid (Federal Highway Administration, 1976).
For any recovery system being considered, the groundwater flow to the system
should be determined using conventional hydrologic analyses or mathematical
modeling. Hydrologic analysis of flow to recovery systems is covered in
Freeze and Cherry (1979), Cedergren (1981), Bouwer (1978) and Federal Highway
Administration (1976). Repa and Kufs (1985) provide a good practical handbook
on groundwater recovery.
1.3.1 Gravity Recovery Methods
Gravity recovery of spent solution and reaction products from a waste deposit
can be accomplished through the use of open ditches or buried perforated
pipes. The flow to the gravity recovery system is governed by the same
factors that control flow to a well (e.g., hydraulic conductivity and
hydraulic gradient), in accordance with Darcy's Law (Equation 1-3). Whereas
hydraulic conductivity is a function of the waste deposit and surrounding
soil, hydraulic gradient can be controlled by appropriate placement of the
gravity recovery system in relation to the waste deposit and groundwater
table. However, since the objective of the delivery system is to maintain
saturation throughout the depth of the deposit, attention must be given to
assure that saturation occurs. Bouwer (as presented in USEPA, 1977) has
developed an equation to determine the distance at which recovery systems
should be placed. This is the same equation introduced in Section 1.2.1 where
the recovery distance (X) from the outer edge of the infiltration area (see
Figure 1-12) can be calculated as:
X
- K (H -
c
) / 2q L - L/2
(1-15)
where: K = Hydraulic Conductivity (length/time)
qn = Solution application rate/unit area of the deposit
(length/time)
L = Length of the deposit parallel to the groundwater flow, (length)
H = Total saturated thickness required, i.e., distance from the
c
top of the waste deposit to the impermeable layer (length)
H, = Height of the recovery system above impermeable layer (length)
Q —
45
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01
FIGURE 1-12
METHOD FOR CALCULATING LOCATION
OF A BURIED PIPE RECOVERY SYSTEM
SOLUTION APPLICATION RATE (q)
RECOVERY
DRAIN
ORIGINAL
GROUNDWATER
TABLE
LENGTH OF THE DEPOSIT (L)
INDUCED
WATER
TABLE
CLAY OR BEDROCK
1) RECOVERY DISTANCE (X) = (Hc2 - Hd2)K/2q - L/2
2) MAXIMUM GRAVITY APPLICATION RATE (qo) = LI
(I =NATURAL INFILTRATION RATE)
SET a = q. TO DEFINE MINIMUM X
IF X IS TOO LARGE FOR THE SITE, THEN:
• INCREASE q0 (INCREASE HYDROSTATIC HEAD; TILL SURFACE)
• USE FORCED DELIVERY SYSTEM
-------
Therefore, X can be determined by measuring K and setting H,, measuring the
depth of the deposit, and using the q-i value determined for the delivery
system.
1.3.1.1 Open Ditch
Open ditches, consisting simply of a ditch or trench excavated into the
groundwater table, have been used successfully for the collection and
transport of groundwater from shallow aquifers. The recovered liquid is
ultimately conveyed to a sump from which it can be either returned to the
delivery system, collected for disposal or further treated. The ditch may or
may not be lined with stones or some other porous medium to maintain the
structural stability of the side slopes. Ditches and trenches work best in
permeable media such as sands, where the hydraulic conductivity is greater
than lxlO~3 cm/sec (2.8 ft/day),.
Generally, open ditches are limited to depths not exceeding 4-5 meters (13 to
16 feet) below ground level and the absolute maximum that has been recommended
for groundwater recovery is approximately 8 meters (25 feet) (Federal Highway
Administration, 1976). Since the recommended side slopes are usually within
the range of 1:1.5 to 1:2 (verticalrhorizontal), a ditch 8 meters (25 feet) in
depth could have a width at the surface of more than 32 meters (100 feet).
Clearly, the surface expression of deep trenches becomes quite large and
volumes of earth to be removed become significant. A further consideration is
that since they are open to rainfall, ditches that are deep and have side
slopes of 1:2 or greater may take in considerable amounts of direct rainfall.
If the ditches lead to a treatment system, this will put an additional load on
that system.
Ditches can be installed on moderately steep terrain (slopes less than
25 percent). For steeper slopes, the upgradient end of the excavation would
become progressively more extensive, thus increasing trench excavation
volume. Also, problems may be encountered in mobilizing excavation equipment
on very steep slopes.
47
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Ditches should be designed to a minimum depth of 1-1.5 meters (3 to 5 feet)
below the groundwater table. Upon selection of the drainage ditch, the depth
from the bottom of the ditch to the impervious layer beneath the water table
aquifer (H,) should be determined and groundwater flow from the infiltration
area to the recovery ditch should be modeled to ensure that the ditch will
recover all of the applied solution. If not, then the design of the ditch
(both horizontal and vertical extent) should be altered to assure that the
recovery will be complete and that the waste deposit remains saturated.
The design of recovery ditches is described in Federal Highway Administration
(1976) and order-of-magnitude construction costs are presented in A D Little
(1983). A typical design is shown in Figure 1-13. Because ditches .and
trenches are designed to transport the spent solution in addition to
recovering it, they should be designed with a cross section of adequate area
and relatively gentle slope (1 to 5 percent) to control water velocities,
reducing friction losses and erosion of the side slopes. For unconsolidated
deposits, the velocity in the ditches or trenches must be kept well below the
scour velocity. A porous (gravel fill) lining can be added to prevent erosion
if higher velocities are desired or required.
1.3.1.2 Buried Drains
Buried drainage pipes containing either slots, perforations, or open joints
are another type of gravity collection method similar to the infiltration
galleries described in Section 1.2.1 (gravity delivery techniques). The
drainage systems are constructed by excavating a trench and laying drainage
pipes, made of steel, concrete, asbestos-cement, clay, or plastic, at the
bottom. The trench is then backfilled with gravel or other porous material to
a designated depth (up to the saturated water level) and the rest of the
trench is backfilled with soil. Often the gravel is covered with fabric to
prevent fine soil from entering the gravel from above and clogging the drain.
An impermeable barrier (liner or slurry trench) may be required on the
down-gradient end of the trench to prevent the flowthrough of the intercepted
and contaminated groundwater if the surrounding materials have a moderately
high permeability. Detailed information on drain construction is provided in
48
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FIGURE 1-13
GRAVITY RECOVERY USING A DITCH
ID
DELIVERY
SYSTEM
11
-GROUND LEVEL
ORIGINAL GROUNDWATER
LEVEL
''MW/@//#
DITCH
DIRECTION OF FLOW
CLAY OR BEDROCK
-------
Federal Highway Administration (1976), Luthin (1957), U S Dept of Agriculture
(1972), and Repa and Kufs (1985). A buried drain collection system is
illustrated on Figure 1-14.
Buried drains are best suited for sands, with hydraulic conductivity greater
than 1.0 x 10~3 cm/sec (2.8 ft/day). They can be utilized in silty soils
but will result in long recovery times, and complete recovery of applied
solutions may be difficult to ensure. Drains can be installed in areas of
rough terrain and steep slopes, since they will be completely embedded within
the groundwater table. The only restrictions on the installation of the
drains is accessibility for excavation and pipe laying equipment to the
particular location.
It is technically feasible to excavate a trench to almost any depth desired;
however, the cost of construction could become prohibitively high. Although
hydraulic backhoes can excavate to depths of about 15 meters (50 feet), for
economic reasons, the trench depth for groundwater recovery from a -waste
disposal site should be limited to about 5 meters (16 feet) below ground level.
The same design principle discussed for open ditches will also apply for
buried drains. The location (both horizontal and vertical) should be such
that it satisfies the requirements in Equation (1-14). The velocity in the
pipe should be maintained above 0.5 meters (1.5 feet) per second to prevent
settling of any materials and should be less than 1 meter (3 feet) per second
to prevent high friction losses and uneven distribution of the drawdown over
the length of the drainage pipe. As with all delivery or recovery systems,
groundwater flow to the drain should be modeled to determine that the drain
will recover the spent solution and reaction products. - 1
The disturbed area caused by porous drains is relatively small compared with a
trench or ditch system. Because the drains are placed within the water table
and covered with earth, freezing problems do not occur during the winter
months. Furthermore, recovery of volatile or photo-oxidizable compounds is
enhanced. Porous drains can clog because of chemical precipitates, and this
may require an extensive maintenance effort to correct. The expense of
50
-------
FIGURE 1-14
GRAVITY RECOVERY WITH BURIED PIPES
DELIVERY SYSTEM
INDUCED
WATER
TABLE
J£Z_
ORIGINAL GROUNDWATER
LEVEL
-BACKFILL
-WASHED
STONE
"-PERFORATED
PIPE
DIRECTION OF FLOW
CLAY OR BEDROCK
-------
installing a porous recovery drain is high, and the volume of groundwater
recovered by this gravity system is low compared to pumping methods, although
operating costs are much lower (Federal Highway Administration, 1976;
A W Martin Associates, 1978; A D Little," 1983).
1.3.1.3 Permeable Treatment Beds «
Permeable treatment beds are a variation of recovery trenches or drains, in
which contaminated groundwater is treated as it flows through the bed.
Treatment beds may be used alone if the contamination is primarily in the
aqueous phase (e.g., a spill), or in combination with other treatment methods
which remediate the contaminant source while the permeable treatment bed
controls the downgradient plume. The groundwater may then be recovered for
further treatment or discharge; alternatively, the treated groundwater may not
be removed from the ground but simply continue its natural flow. Permeable
treatment beds are applicable only to sites with relatively shallow
groundwater tables (i.e., the limitations are similar to those for recovery
with buried drains). The bed should fully penetrate the contaminant plume and
be keyed into an impermeable stratum for maximum'"expo sure of! the contaminated
groundwater to the treatment material (this is particularly true if the
groundwater is not subsequently recovered).
To date, permeable treatment beds have not been used for in situ treatment of
contaminants, although bench- and pilot-scale tests have been performed to
determine treatment-effectiveness (Park, 1985; Repa and Kiifs, 1985). Potential
problems in using this technique include chemical saturation' of the bed
material, short effective life of the bed material and plugging of the bed
with precipitated substances.
Potential bed fill material includes limestone, activated .carbon., glauconitic
green sands, coal, fly ash, soil containing clay materials, natural (zeolites)
or synthetic ion exchange resins, and polymeric adsorbents (Park, 1985; Repa
and Kufs, 1985). Limestone would be used .primarily for neutralization ' of
acids or precipitation of metals. However, the increase in groundwater pH as
it passes through a limestone bed may increase the rates of base-catalyzed
52
-------
hydrolysis of some organic contaminants in the groundwater (see Section 5.0).
Activated carbon is commonly used as a treatment method for adsorption of
hydrophobic (non-polar) organic contaminants in water. However, significant
problems such as plugging of the bed, short lifetime or saturation of the
carbon, and desorption might occur with the use of activated carbon in
permeable treatment beds (Repa and Kufs, 1985). Glauconitic green sands are
used primarily to treat trace metal contamination. Reduction of odors
(Spoljaric and Crawford, 1978; as cited by Repa and Kufs, 1985) during
treatment with this material suggests that volatile organic compounds may also
be adsorbed. Coal and fly ash appear promising for adsorption of organics
(Park, 1985), but leaching of other contaminants (e.g., trace metals) needs to
be evaluated. Zeolites and ion-exchange resins are used primarily for
adsorption of trace metals, but synthetic polymeric adsorbents (macroreticular
resins, e.g., XAD resins) effectively adsorb a wide range of organic compounds.
The high cost of these resins, however, would severly limit their use in situ
unless a built-in regeneration system were included in the design.
In summary, permeable treatment beds may have limited application in specific
cases, particularly for temporary remedial measures, but their costs and
limitations render their use for long term in situ treatment of organic
contaminants unlikely at present.
1.3.2 Forced Recovery Methods
Forced recovery is the process by which a fluid is pumped from pipes or wells
strategically placed in the waste deposit for removal, recycle or treatment.
When employed in shallow groundwater regimes, such systems are called
wellpoint systems. When employed in deep groundwater regimes, such systems
are termed deep well systems. The design rate of removal of liquid from the
wells should be greater than the rate that the reactant solution is being
delivered (by gravity application or injection), since some surrounding
groundwater will also be drawn into the recovery system. The specific
recovery rate required will depend on hydrologic conditions at the site, and
should be determined by modeling the groundwater flow regime from the point of
delivery to the recovery system.
53
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1.3.2.1 Wellpoint Systems
Wellpolnt systems are the most commonly used dewatering! methods in
construction practice today and such systems are applicable to a wide range of
excavations and groundwater conditions. The technology can be readily adapted
for use as a recovery system in managing waste deposits. A wellpoint system
is usually the most practical method for dewatering where the site is
accessible, the groundwater is shallow and hydraulic conductivity of the waste
' ' —.1 — ^
deposit and surrounding soil media ranges between 1x10 and 1x10 cm/sec
(280 to 2.8 ft/day). For deep groundwater conditions, more than about 8
meters (25 feet), it will be necessary to use ejector wells (Repa and Kufs,
1985) or deep wells with turbine or submersible pumps.
A conventional wellpoint system consists of one or more stages of wellpoints
(wellpoints connected to a header at a common elevation) which are installed
in a line, a ring or radially around the waste deposit at spacings of from 1-5
meters (3 to 15 feet). The wellpoints are attached to 3.8 or 5 cm (1 1/2 or 2
inch) riser pipes connected to a common header pumped with one or more
wellpoint pumps as shown on Figure 1-15. The wellpoints are small well
screens composed of either brass or stainless steel mesh, slotted brass or
plastic pipe, or wire wrapped on rods to form a screen. Wellpoints generally
range in size from 5 to' 10 cm (2 to 4 inches) in diameter and 1 to 1.5 meters
(3 to 5 feet) in length, and are constructed with either closed ends or
self-jetting tips. It may be judged necessary to add a filter around the
wellpoint, depending upon the nature of the waste deposit area being drained.
A wellpoint pump is a combined vacuum and centrifugal pump which is connected
to the header and pumps water from the wellpoints. Generally, a stage of
wellpoints would be capable of draining a deposit about 5 meters (16 feet)
thick. Draining a deposit that is greater than 5 meters (16 ft) thick
generally requires a multi-stage installation of wellpoints.
A vacuum wellpoint system is essentially the same as a conventional wellpoint
system except that a partial vacuum is maintained in the sand filter around
the wellpoint and riser pipe. This vacuum increases the hydraulic gradient
54
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FIGURE 1-15
WELLPOINT SYSTEM FOR
FORCED RECOVERY
WELLPOINTS (VACUUM
OR
CONVENTIONAL)
HEADER
PLAN VIEW
DELIVERY SYSTEM
DIRECTION OF
INDUCED FLOW
COMMON HEADER
*- / WASTE
/ DEPOSIT
CLAY OR BEDROCK
CROSS SECTION
55
-------
producing larger flows to the wellpoints (Fruco and Associates, 1966). Vacuum
wellpoint systems are used in deposits with hydraulic conductivities from
IxlO"3 to as low as lxlO~5 cm/sec (2.8 to 0.03 ft/day).
Actual field pump testing at the desired recovery elevation must be performed
to determine the flow rates at which the fluid can be recovered,, This will
provide the necessary data to design the recovery well spacing and grid
pattern. The flow rates for a given wellpoint can be used to calculate a
radius of drawdown, using standard well drawdown theory (Urguhart, 1968;
Freeze and Cherry, 1979; Repa and Kufs, 1985). A typical layout for a
wellpoint recovery system, consisting of installations downgradient from a
waste deposit, is shown on Figure 1-15. Order-of-magnitude construction costs
are presented in A W Martin Associates (1978) and A D Little (1983).
The efficiency of both conventional and vacuum wellpoint recovery systems are
limited by the soil and waste deposit hydraulic conductivities. With a low
hydraulic conductivity, the pumping period required for recovery of treatment
solution may exceed the time frame established to accomplish the remediation
of the waste deposit. Under these conditions, the well spacing may also have
to be very close, resulting in an unacceptable capital and operating cost.
1.3.2.2 Deep Well Systems
Deep well systems are particularly suited for recovering groundwater from
depths below the suction limit (about 8 meters or 25 feet) or for dewatering
large areas where large volumes of fluid must be removed. This requires
higher rates of pumping than those obtained with a wellpoint recovery system.
Mixed and axial flow pumps powered by electricity, gasoline, or diesel are
available in discharge ranges from 0.3 to 1.6 m3/sec (5000 to 25,000 gpm) at
heads up to 30 meters (100 feet). Deep well turbine pumps are available in
sizes from 0.01 to 0.4 m3/sec (200 to 6000 gpm), with head capabilities up
to 180 meters (600 feet) (Fruco and Associates, 1966).
56
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Deep wells for dewatering are similiar in type and construction to commercial
water wells. They commonly have screens with a diameter of 10 to 45 cm,(4 to
18 inches) and lengths up to 90 . meters (300 feet). A filter is usually
installed around the screen to prevent the infiltration of the deposit
materials into the well and to improve the yield of the well. As in the case
of a wellpoint system, deep wells may also be used in conjunction with a
vacuum established at the recovery area. This serves to induce a larger flow
to the well. Construction details are described in Repa and Kufs (1985) and
order-of-magnitude costs are estimated in A W Martin Associates (1978) and
A D Little (1983).
Actual in situ recovery rates, radii of influence and spacing arrangement will
be arrived at in much the same way as in the wellpoint recovery systems,(Repa
and Kufs, 1985, describe the methodology for assessing the effects of pumping
from deep wells). A typical layout for a deep well recovery system is shown
on Figure 1-16. ..
1.3.3 Summary . . -
Recovery systems can use either gravity or forced methods. Gravity systems
are generally applicable in shallow groundwater regimes (less than 8 meters or
25 feet) and forced systems are applicable in deep groundwater regimes. The
hydraulic conductivity of the waste deposit and surrounding soil media and the
time required to accomplish remediation must also be considered in selecting
recovery systems. Figure 1-17 gives guidance on which methods are suitable
for recovery systems depending on the grain size of the soil. This figure
indicates that gravity recovery systems, as well as well points and deep
wells, are limited to media with .an effective grain size (d,Q) between 0.1
to 1 mm. (0.004 to. 0.04 inches) which generally have a hydraulic conductivity
between SxlO"1 to 10 cm/sec (1420 to 2.8, ft/day) (Table ,1-1). For media
with effective grain size (d,) between 0.1 to 0.01 mm (0.004 to 0.0004
,0
* _
inches) which generally have a hydraulic conductivity between 10 to 10
cm/ sec (2.8 to 0.03 ft/day) (Table 1-1), recovery of water may be possible
Media with hydraulic
10
_ t\
using well points equipped with vacuum pumps.
conductivity less than 10~5 cm/sec (0.03 ft/day) i.e., d,n less than 0.01
57
-------
FIGURE 1-16
DEEP WELL
RECOVERY SYSTEM
DEEP WELL.
o
o
PLAN VIEW
DELI VERY SYSTEM
TO TREATMENT
^IMTJAL_GROyNDWATER TABLE
•"• ——• CONE OF ~
DEPRESSION
DIRECTION OF INDUCED FLOW
TURBINIIE
PUMP
CLAY OR BEDROCK
CROSS SECTION
58
-------
cn
ID
t-
o
Ul
K
Ul
I-
UJ
u
te.
Ul
a.
FIGURE1-17
LIMITS OF RECOVERY METHODS APPLICABLE TO DIFFERENT
SOILS
U.S. STANDARD SIEVE
OPENINGS IN INCHES
U.S. STANDARD SIEVE NUMBERS
10 1416 20 30 40 50 70 100 140 200
'/Y////.
LIMITS FOR
GRAVITY SYSTEMS
INCLUDE SUMPS,
WELLPOINTS AND
DEEP WELLS
SUBAQUEOUS EXCAVATIONS
OR CUTOFF WALLS
REQUIRED
LIMITS FOR
WELLPOINT VACUUM METHOD
ELECTR
OSMOSIS
SSIBLE
0.1 0.05
0.004 0.002
I-
K
Ul
V)
K
i
100
0.001 MM
INCHES
GRAIN SIZE
GRAVEL
COARSE
FINE
SAND
COARSE
MEDIUM | FINE
SILT OR CLAY
SOURCE: US DEPARTMENT OF THE NAVY, 1982
-------
mm (0.0004 inches) can only be dewatered by means of other enhancement
techniques (e.g., electro-osmosis, which is discussed in the following section)
used in conjunction with wells or well points.
1.4 Special Method of Delivery and Recovery Enhancement (Electro-Osmosis)
When a waste deposit exhibits a hydraulic conductivity of less than 10
cm/sec (0.03 ft/day) and the fluid to be extracted is high in inorganic
constituents, electro-osmosis may be used to increase the flow rate of the
fluid through the waste deposit and surrounding media. Groundwater migration
by electro-osmosis is initiated by applying a direct electric potential to
electrodes installed in the ground at a selected spacing within the low
hydraulic conductivity media. The electric potential applied to the
electrodes causes the positive ions in the pore water to move from the anodes
(positively-charged electrodes) towards the cathodes (negatively-charged
electrodes). The movement of the positive ions develops tension in the media,
causes the chemical composition of the groundwater to change, and forces the
pore water to flow from the anodes to the cathodes. These actions result in
the reduction of the water content of the deposit (Loughney, 1973) or movement
of water from injection to recovery wells if these wells form the anodes and
cathodes, respectively. Some waste deposits that do not permit application of
standard delivery and recovery methods because of low hydraulic conductivities
might be rendered treatable when these methods are combined with electro-
osmosis. By making wellpoints, the anode and cathode, movement of the
treatment solution through the deposit may be accelerated.
Figure 1-17 (in the preceeding section) shows the grain size distribution for
which electro-osmosis should be considered. The corresponding hydraulic
conductivity for these materials ranges from 1x10
cm/sec (0.03 to 0.0003 ft/day). |
to less than 1x10
A site investigation should be performed prior to selecting any injection or
recovery system. Once the site conditions are known and it has been
determined that the hydraulic conductivity may be enhanced by electro-osmosis,
conductivity tests should be performed to determine the electro-osmotic
60
-------
transmission coefficient, k (volume of water transmitted through a unit
cross section in unit time by application of a potential of 1 volt/cm normal
to the cross section). This coefficient is determined in the laboratory or
field and used in a Darcy-type equation.
The discharge of a cathode wellpoint, Qe, may be estimated from the equation
(Fruco and Associates, 1966).
(1-15)
where:
k = coefficient of electro-osmotic permeability
e 2
(length /volts x time)
i = electrical gradient between electrodes (volts/length)
z = length of electrodes (length)
a, = effective spacing of wellpoints (length)
The current required can be estimated from the following empirical
equation developed by Maclean and Rolfa (as presented in Loughney,
1975):
= (Ac + B)/t
(1-16)
where:
1^ = current required per gram (pound) of water expelled
(amps)
t = time
c = clay content of soil, i.e., weight of soil finer than
0.002 mm (0.00008 inches)(percent)
A - constant . .
B = constant
61
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Current requirements commonly range between 15 and 30 amps per recovery well,
and power requirements are generally high. However, regardless of the expense
of installation and operation of an electro-osmotic dewatering system, it may
be the only effective means of dewatering or permeating certain fine-grained
soils.
In an electro-osmotic dewatering system, the depth of the electrodes should be
at least 1.5 meters (5 ft) below the bottom of the contaminated deposit that
is to be dewatered. The spacing and arrangement of the electrodes may vary,
depending on the configuration of the area to be dewatered and the voltage
available at the site. Cathode spacings of 8 to 12 meters (25 to 40 ft) have
been used, with the anodes installed midway between the cathodes. Electrical
gradients of 5 to 13 volts per meter (1.5 to 4 volts per ft) distance between
electrodes have been successful in electro-osmotic dewatering. The electrical
gradient should be less than about 50 volts per meter (15 volts per ft) of
distance between electrodes for long-term installations to prevent loss in
efficiency caused by heating of the ground. Applied voltages of 30 to 100
volts are usually satisfactory; a low voltage is usually sufficient if the
groundwater has a high mineral content (i.e., high conductivity) (Fruco and
Associates, 1966).
Electro-osmosis would only be cost-effective in waste sites having low
hydraulic conductivities. Hydraulic conductivities of less than 10 cm/sec
(0.0003 ft/day) prior to initiating electro-osmosis can be made to exhibit
hydraulic conductivities in the range of 10 to 10 cm/sec (0.03 to
0.003 ft/day) for a gradient of 3 volts per meter (one volt per ft) using this
method (Fruco and Associates, 1966). It should be noted that these enhanced
hydraulic conductivities are nevertheless very low, i.e., groundwater or
leachate flow velocity will still be very slow.
Power consumption is a major limitation on the economic feasibility of the
procedure. Use of long term, low power options should be considered when it
is possible to do so within the time frame established for the site
remediation.
62
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1.5 Comparative Analysis of Alternatives
The application of chemical solutions into a waste deposit to provide in situ
treatment or mobilization of contaminants from the deposit requires an
appropriate delivery and recovery method. The selection of delivery and
recovery techniques requires an understanding of the parameters governing
these systems, including site conditions, nature and configuration of the
waste deposit, geohydrologic features, and surface hydrologic characteristics.
As discussed in the previous sections, two major delivery and recovery
techniques, gravity and forced, are possible. Alternatives that can be
considered for delivery and recovery systems are given in Table 1-3. A
comparative analysis of gravity and forced systems focusing on parameters
affecting selection of these systems and an engineering judgement on the
application of each alternative system is presented in this section.
1.5.1 Importance of Various Parameters in Gravity vs Forced Systems
Gravity methods utilize natural gravity forces to effect the delivery and
recovery of solutions, while forced methods utilize mechanical mechanisms to
deliver or withdraw the solution from the waste deposit. Soil infiltration
rates and hydraulic conductivities are key parameters controlling the
effectiveness of a gravity method. Pressure head and hydraulic conductivities
are the key design criteria for pumped and vacuum type forced methods.
Table 1-4 provides a list of applicable parameters affecting the design of
gravity and forced delivery and recovery systems.
1.5.2 Application of Various Systems
The selection of delivery or recovery methods depends primarily on the
geohydrologic characteristics of the disposal site. Two matrices (one for
delivery systems and one for recovery systems) have been developed in order to
guide in the identification of feasible delivery and recovery methods for a
63
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Delivery/Recovery
Techniques
Delivery Systems
Gravity Methods
TABLE 1-3
DELIVERY AND RECOVERY SYSTEMS
Forced Methods
Alternatives
Flooding, Ponding, Ditch, Surface
Spraying, Infiltration Gallery,
Infiltration Bed
Injection Pipe (open end or slotted)
Recovery Systems
Gravity Methods
Forced Methods
Ditch, Buried Drain
Wellpoint with Vacuum
Wellpoint without Vacuum
Deep Well
Electro-Osmosis
-------
TABLE 1-4
RELATIVE IMPORTANCE OF GEOTECHNICAL PARAMETERS IN GRAVITY
AND FORCED SYSTEMS
Gravity Methods Forced Methods
1. Hydraulic Conductivity (K)
2. Infiltration Rate (I )
s
3. Application Rate (q,)
4. Configuration of Water Table and
Waste Deposit Location (H , H,)
5. Time to Reach Saturated Condition (t)
6. Homogeneity
7. Relation of Hydraulic Conductivity
between Waste Deposit and Surrounding
Medium
8. Relationship between Infiltration
Rate and Hydraulic Conductivity
NI
LI
LI
LI
LI
LI
NI
NOTE: I = Important
LI = Less Important
NI = Not Important
1 = wellpoints are limited by depth to the watertable
65
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given set of site conditions. These matrices, presented in Tables 1-5 and 1-6
are based on engineering judgement and experience. The major criteria used to
identify potentially suitable delivery and recovery systems are discussed in
the following subsections.
1.5.2.1 Hydraulic Conductivity
Gravity delivery methods would be applicable for situations in which the waste
deposit and surrounding soil media have hydraulic conductivities between
—1 —3
1 x 10 cm/sec and 1 x 10 cm/sec (280 to 2.8 ft/day). Forced delivery
methods would be applicable for situations where hydraulic conductivities are
—3 —4
between 1 x 10 cm/sec and 1 x 10 cm/sec (2.8 to 0.28 ft/day). At a
site where the hydraulic conductivity is less than 1 x 10 cm/sec (0.28
ft/day), enhancement techniques such as electro-osmosis or hydrofracturing
would be required.
In terms of recovery systems, a site where the hydraulic conductivity is
—1 -3
between 1 x 10 cm/sec and 1 x 10 cm/sec (280 to 2.8 ft/day) would be
amenable to open ditches and buried drains. Also at such a site, wellpoint
and deep well systems would be suitable. The vacuum well point recovery
system would be a feasible technique for a site having a hydraulic
—3 —5
conductivity between 1 x 10 cm/sec and 1 x 10 cm/sec (2.8 to 0.03
ft/day). For a waste deposit having a relatively low hydraulic conductivity
(below 1 x 10 cm/sec or 0.03 ft/day), electro-osmosis may be the only
effective recovery method.
1.5.2.2 Depth of Waste Deposit Cover
In general, gravity delivery and recovery methods would be more effective for
a shallow waste deposit with a thin cover, while the depth of the waste
deposit and cover thickness depth would not affect the application of forced
methods (except recovery by wellpoints since these are limited to 5 meters per
stage). In practice, gravity delivery and recovery systems are preferred for
a waste deposit site having a total depth of waste and cover of less than 5
meters (16 ft). Forced delivery and recovery systems are more suitable for
depths greater than 5 meters (16 ft).
66
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TABLE 1-5
MATRIX FOR DELIVERY METHODS
I | Thickness
Delivery (Location of the deposit In
1 of
Methods 1 relation to existing (Contanlnatlonl overlying
-
GRAVITY
1. Flooding
2. Ponding
3. Surface
Spraying
4. Ditches
5. Infiltration
Galleries
6. Infiltration
Bed
FORCKD
1. Injection
Pipes
groundwater table | starts at
Unsatu-lPartlallyl ISur-
rated ISaturatedlSaturatedlface
X
X
X
X
X
X
X
LE
LE
NA
LE
LE
LE
X
NA
NA
NA
NA
NA
NA
X
X
X
X
NA
NA
NA
Sub-
surface
X
X
X
X
X
X
X
1 Infiltration
Topography I Rate
Impermeable I (Slope) 1 cm/hr
layer 1 1 (Inches/hr)
l< 1.5ml >1.5m
0 l(>5ft)l(>5 ft)
X
X
X
X
X
X
X
NA
X
NA
X
X
X
X
1
NA
NA
NA
NA
X
X
X
FlatlO-3Zl>3Z
X
X
X
X
X
x
x
X
X
X
X
X
X
X
NA
NA
LE
X
X
NA
X
.1-.2
(.3-5)
X
X
X
X
X
X
X
.06-.! |<.06
(.15-.3) K0.15
X
X
X
X
X
X
X
NA
LE
NA
X
X
X
X
1
1 Depth to Bottom
Hydraulic Conductivity 1 of the
en/ sec (ft/day) 1 Waste Deposit
1 Meters (ft)
-1 -3
(280-2.8)
LE
X
LE
*
X
X
X
-3 -4. -4 -7
10 -10 | 10-10
(2.8-0.28)1(0.28-0.0003)
NA
LE
NA
LE
LE
LE
X
NA
NA
NA
NA
NA
NA
x(2)
<5
«16)
X
X
X
X
X
X
X
5-12
(16-40)
LE
LE
LE
LE
X
X
X
C>40)
NA
NA
NA
NA
NA
NA
X
01
X " Applicable
I.E • Leas Kffectlve
• NA • Not Applicable
(1) • May need combined gravity and forced delivery.
(2) * Applicable with electro-osmosis.
-------
TABLE 1-6
MATRIX FOR RECOVERY METHODS
Recovery
Methods
GRAVITY;
Open Ditches
and Trenches
Depth To Groundwater Hydraulic Conductivity
0-5 m
(0-16 ft)
X
5-12m
(16-40 ft)
NA
cm/ sec
>12 m (>280-2.8
(>40 ft) ft/day)
NA X
10"3-10"4
cm/ sec
(2.8-0.28
ft/day)
LE
io-4-io-7 .
cm/ sec
(0.28-0.003
ft/day)
NA
Porous Drains
NA
NA
LE
NA
FORCED:
Well point
Deep Well
Vacuum Well
Poi nt
Electro-
osmosis
X
NA
X
X
X
X
X
X
NA
X
NA
X
X
X
NA
NA
LE
LE
X
NA
NA
NA
LE
X
X = Applicable
LE = Less Effective
NA - Not Applicable
68
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Design criteria related to the configuration of the waste deposit and its
surrounding media can be summarized as follows:
1. Gravity delivery methods are most effective if the waste deposit is
situated in the unsaturated zone, and at the surface or at a
relatively shallow depth (less than 5 meters or 16 feet).
2. If the deposit is covered with a thin layer of impervious material
(less than 1.5 meters or 5 feet) gravity delivery might require
excavation but would probably still be more cost-effective than
forced delivery.
3. Open ditches are generally limited to depths not exceeding 4-5 meters
(13 to 16 feet) below ground level. The trench depth of buried
drains for groundwater recovery from a waste disposal site should
also be limited to a maximum of 16 feet below ground level.
4. A stage of well points would be capable of draining a deposit of
about 5 meters (16 feet) in depth. Draining a deposit of greater
than 5 meters (16 feet) generally requires a multi-stage installation
of wellpoints and vacuum pumps to assure the maintenance of the
maximum vacuum in the column. In this case, down-hole pumps may be
more cost-effective.
5. For deep groundwater conditions, more than 8 meters (25 feet), it may
be more practical to use deep recovery wells with turbine or
submersible pumps.
6. Deep well systems are particularly suited for dewatering large areas
at greater depth where large volumes of fluid must be removed.
1.5.2.3 Climate
The influence of climate is more significant for gravity delivery and recovery
systems than for the forced methods. Freezing and frost penetration may
69
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preclude the operation of gravity delivery and recovery systems, particularly
flooding, ponding, ditches and surface spraying methods. Subsurface spreading
methods may be suitable where ground freezing is a recurring problem.
Infiltration galleries or beds can be installed under the freiezing zone,
thereby permitting year-round application of solution. A forced delivery
system, unlike the gravity systems, is conceptually independent of surface
topography and climate and can be designed to accommodate any waste deposit
configuration.
1.5.2.4 Relationship Between Waste Deposit and Soil Medium
The most important relationship between the waste deposit and the surrounding
soil medium is that of their hydraulic conductivity values. If the waste
deposit has a lower hydraulic conductivity than the surrounding soil, gravity
delivery would not be reliable, because the solution would most probably
bypass the waste deposit. Therefore forced injection of reacta.nt solution
directly into the waste deposit would be required. Gravity delivery methods
are thus most effective in relatively homogeneous deposits where the applied
solution can be evenly distributed throughout the deposit. In a heterogeneous
environment, the waste deposit may not be effectively reached by gravity
delivery methods.
The hydraulic gradient formed by the natural conditions cannot; be easily
altered in gravity delivery methods. With forced delivery methods,, the
hydraulic gradient can be increased by increasing the injection pressure.
This pressure would increase the transmission rate of the applied solution
through the medium.
Using gravity delivery methods, the applied solution will generally have to
travel from the point of application through the overlying soil to reach a
subsurface waste deposit. Forced delivery methods allow direct application of
reactant solution into a waste deposit, eliminating attenuation or reaction of
the solution with the overlying soil.
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References
American Society of Civil Engineers (ASCE), 1972. Groundwater Management
Manual. ASCE Manual 40, ASCE, New York, NY.
American Society for Testing and Materials (ASTM), 1969. Classification of
Soils for Engineering Purposes, ASTM D2487-69. Annual Book of ASTM Standards,
ASTM, Philadelphia, PA.
Black, C. A. (ed). 1965. Methods of Soil Analysis, Part I: Physical
Properties. Agronomy 9, Amer. Soc. of Agron., Madison, WI.
Bouwer, H., 1964. Measuring Horizontal and Vertical Conductivities of Soil
with the Double Tube Method. Soil Sci. Soc. Amer. Proc. 28:19-23.
Bouwer, H., 1966. Rapid Field Measurement of Air-Entry Value and Hydraulic
Conductivity of Soil as Significant Parameters in Flow System Analyses. Water
Resources Research 2:729-738.
Bouwer, H. and R. C. Rice, 1967. Modified Tube-Diameters for the Double Tube
Apparatus. Soil Sci. Soc. Amer. Proc. 31:437-439.
Cedergren, H. R., 1977. Seepage, Drainage and Flow Nets (2nd edition).
J. Wiley and Sons, New York, NY.
Federal Highway Administration, 1976. Grouting in Soils. FHWA-RD-76-27. FHA
Office of Research and Development, Washington, DC.
Freeze, R. A. and J. A. Cherry, 1979. Groundwater. Prentice Hall, Englewood
Cliffs, NJ 604 pp.
Fruco and Associates, 1966. Dewatering and Groundwater Control for Deep
Excavation. U.S. Army Engineering Waterways Experimental Station, Vicksburg,
NJ.-
Fry, A. W. and A. S. Grey, 1971. Sprinkler Irrigation Handbook. Rain Bird
Sprinkler Mfg Corp., Glendora, CA.
Gibb, J. P., M. J. Barcelona, J. D. Ritchey and M. H. LeFaivre, 1985.
Effective Porosity of Geologic Materials. In: Land Treatment of hazardous
Wastes: Proc of the llth Annual Res. Symp. EPA/600/9/85-013. HWERL, US
Environmental Protection Agency, Cincinnati, OH. pp. 190-197.
Huisman, L., 1972. Groundwater Recovery. Winchester Press, New York, NY.
Huisman, L. and T. N. Olsthorn, 1983. Artificial Groundwater Recharge. Pitman
Advanced Publishing, New York, NY.
Johnson, A. I., 1967. Specific Yield - Compilation of Specific Yields for
Various Materials. Geological Survey Water Supply Paper 1662-D. U.S.
Geological Survey, Alexandria, VA.
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Little, A. D., 1983. Handbook for Evaluating Remedial Action Technology
Plans. EPA-600/2-83-076. MERL, U S Environmental Protection Agency,
Cincinnati, OH.
Lohman, S. W., 1979. Groundwater Hydraulics. U.S. Geological • Survey
Professional Paper 708. U.S. Geological Survey, Alexandria, VA. ?
Loughney, R. W., 1975. Construction Dewatering by Electro-Osmosis, .Educator
Wells and Deep Wells. In: Joint AEG-ASCE Symposium on Practical Construction
Dewatering, May 16, 1975.
Luthin, J. N. (ed.), 1957. Drainage of Agricultural Lands. American Society
of Agronomy, Madison, WI.
Martin Associates, A. W., 1978. Guidance Manual for Minimizing Pollution from
Waste Disposal Sites. EPA-600/2-78-142. MERL, U.S. Environmental Protection
Agency, Cincinnati, OH.
Meinzer, 0. E., 1923. Outline of Groundwater Hydrology. U.S. Geological
Survey Water Supply 494 (reprinted 1968). U.S. Geological Surveyj, Alexandria,
VA.
Olson, R. E. and D. E. Daniel, 1981. Measurement of Hydraulic Conductivity of
Fine Grained Soils. In: Permeability and Groundwater Contaminant Transport
(T. F. Zimmie and C. 0. Riggs, eds), ASTM STP 746. American Society for
Testing and Materials, Philadelphia, PA.
Park, J. E., 1985. Permeable Materials for the Removal of Pollutants from
Hazardous Waste Leachates. Proceedings of the llth Annual Research Symposium
on Land Disposal of Hazardous Wastes, HWERL, USEPA, Cincinnati, OH, p.19-26.
Pound, C. E. and R. W. Crites, 1973. Wastewater Treatment and Reuse by Land
Application, Volumes I and II. Office of Research and Development, USEPA,
Washington, DC.
Repa, E. and C. Kufs, 1985. Leachate Plume Management. Draft Report for
HWERL, U.S. Environmental Protection Agency, Cincinnati, OH.
Spooner, P. A., G. E. Hunt, V. E. Hodge and P. M. Wagner, 1984. Compatibility
of Grouts with Hazardous Wastes. EPA-600/2-84-015. MERL, U.S. Environmental
Protection Agency, Cincinnati, OH.
U.S. Department of Agriculture, 1972. Drainage of Agricultural Land. A
Practical Handbook for the Planning, Design, Construction and Maintenance of
Agricultural Drainage Systems. U.S. Dept. of Agriculture, Soil Conservation
Service, Washington, DC.
U.S. Department of the Navy, 1982. Soil Mechanics. NAVFAC DM-71. Naval
Facilities Engineering Command, Alexandria, VA.
USEPA, 1973. Wastewater Treatment and Reuse by Land Appliction. USEPA Office
of Research and Development, Washington, D.C.
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USEPA, 1976. Erosion and Sediment Control - Surface Mining in the Eastern
U.S. EPA-625/3-76-006. USEPA, Washington, DC.
USEPA, 1977. Process Design Manual for Land Treatment of Municipal Wastewater.
EPA-625/1-77-008, U.S. Environmental Protection Agency Center for Environmental
Research Information, Cincinnati, OH.
USEPA, 1980a. Procedures Manual for Ground Water Monitoring, at Solid Waste
Disposal Facilities. Manual SW-611, USEPA Office of Water and Waste
Management, Washington, DC.
USEPA, 1980b. Design Manual for Onsite Wastewater Treatment and Disposal
Systems. EPA-625/1-80-012, USEPA Office of Research and Development,
Washington, DC.
USEPA, 1980c. Lining of Waste Impoundment and Disposal Facilities. Manual
SW-870, USEPA Office of Water and Waste Management, Washington, DC.
USEPA, 1982. Remedial Action at Waste Disposal Sites. EPA-625/3-76-006.
USEPA, Washington, DC.
USEPA, 1984a. Slurry Trench Construction for Pollution Migration Control.
EPA-540/2-84-001 MERL, U.S. Environmental Protection Agency, Cincinnati, OH.
USEPA, 1984b. Review of In-Place Treatment Techniques for Contaminated
Surface Soils. EPA-540/2-84-003a. MERL, U.S. Environmental Protection
Agency, Cincinnati, OH.
USEPA, 1984c. Summary Report: Remedial Response at Hazardous Waste Sites.
EPA-540/2-84-002a. MERL, U.S. Environmental Protection Agency, Cincinnati, OH.
USEPA, 1984d. Case Studies 1-23: Remedial Responses at Hazardous Waste
Sites. EPA-540/2-84-002b. MERL, U.S. Environmental Protection Agency,
Cincinnati, OH.
Urguhart, L.C. (ed.), 1968. Civil Engineering Handbook. McGraw Hill Co, New
York, NY.
Wang, H. F. and M. P. Anderson, 1982. Introduction to Groundwater Modeling.
W. H. Freeman Co., San Francisco, CA.
Winterkorn, H. F. and H-Y. Fang, 1975. Foundation Engineering Handbook.
Van Npstrand Reinhold Co., New York, NY.
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SECTION 2
BIODEGRADATION
2,1 Introduction
Recent developments in applied microbiology (Bitton and Gerba, 1984; API,
1982; Doggett, 1983; Jhaveri and Mazzacca, 1983; Kellogg et al, 1981; Kopecy,
1983; Krupka and Thibault, 1980; Litchfield and Clark, 1973; Zitrides, 1978)
have made in situ biological treatment of hazardous organic materials in soil,
water, and groundwater a potentially cost effective alternative to chemical or
physical methods of site reclamation. Biological treatment involves the use
of native microbes, selectively adapted bacteria or genetically altered
microorganisms that have been modified through specific gene mutation or
genetically assisted molecular breeding to degrade a variety of organic
compounds. Biodegradable compounds include industrial surfactants, organic
solvents, crude and refined petroleum products, pesticides and herbicides,
polychlorinated biphenyls, polycyclic aromatic hydrocarbons and other classes
of organic compounds (Bitton and Gerba, 1984; Kobayashi and Rittman, 1982) as
discussed below. The biological treatment process usually involves the
addition of nutrients and oxygen, and may take place completely within the
deposit or (more commonly) partly above ground in environmentally-controlled
bioreactors and partly within the waste deposit. Typically, contaminated
groundwater is recirculated from downgradient recovery weills, through
bioreactors and conditioning processes (e.g. aerators) at the surface, and
reinjected at upgradient locations for further in situ degradation.
This section discusses a number of organic waste treatment methods that
involve the use of microbial agents. Processes of waste biodegradation are
identified that may be used as the sole treatment or in conjunction with
chemical or physical methodologies or both. These methods are representative
of an emerging technology and significant advances can be expected in the near
term. It should be recognized that as these existing methods are superseded
by more advanced techniques, new procedures should be considered in future
strategies concerning the in situ treatment of wastes. The information
included on biological methods of waste stabilization was obtained from
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published reviews, literature, reports on demonstration studies and personal
communications with commercial firms that are actively developing this
technology. Methodology guidelines have been developed from these sources.
2.2 Analysis of Data
2.2.1 Microbial Mechanisms of Catabolism
It has long been recognized (Atlas, 1981; Horvath, 1972; Zobell, 1946) that
many microorganisms have the ability to utilize hydrocarbons as the sole
source of carbon and energy. These microbial communities react to the a wide
range of naturally occurring and anthropogenic (synthetic) hydrocarbons.
Hydrocarbons are catabolized (broken down to simpler substances) by
microorganisms using three general mechanisms (Atlas, 1981; Focht and Chang,
1975; Sokatch, 1969; Stanier et al, 1976). These are aerobic respiration,
anaerobic respiration, and fermentation. In general, aerobic degradation
processes are more often used for biodegradation because the degradation
process is more rapid and more complete, and problematic end products
(methane, hydrogen sulfide) are not produced. However, anerobic degradation
is important for dehalogenation (Bouwer and McCarty, 1983).
In aerobic respiration, organic molecules are oxidized to carbon dioxide
(CO ) and water or other end products using molecular oxygen as the terminal
electron acceptor. Oxygen may also be incorporated into intermediate products
of microbial catabolism through the action of oxidase enzymes, making them
more susceptible to further biodegradation. Microorganisms catabolize
hydrocarbons by anaerobic respiration in the absence of molecular oxygen using
inorganic substrates as terminal electron acceptors. In
CO
anaerobic
,-2,
2 is reduced to methane (CH,), sulfate (SO, ) to
respiration,
sulfide (S *"), and nitrate (N0») to molecular nitrogen (N9) or
+
ammonium ion (NH,). Hydrocarbon sources are degraded by fermentation
using substrate level phosphorylation as the terminal electron acceptor.
Fermentation results in a wide variety of end products including carbon
dioxide, acetate, ethanol, proprionate, butyrate, etc.
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In most cases, naturally occurring microbial activity can decompose organic
materials of both natural and synthetic origin to harmless or stable forms or
both by aerobically mineralizing them to C0_ and water, or anaerobically
decomposing them to C02, CH, and water (Alexander, 1981; Atlas, 1981;
Bitton and Gerba, 1984; Boethling and Alexander, 1979a and 1979b:[ Evans, 1977;
Kobayashi and Rittman, 1982; Perry, 1979; Sokatch, 1969; Stanier et al, 1976;
Zobell, 1946). Some anthropogenic compounds can appear relatively refractory
to biodegradation by naturally occurring microbial populations because of the
interactions of environmental influences, lack of solubility, absence of
required enzymes or other factors as discussed by Alexander (1981). However,
the use of properly selected or engineered microbial populations, maintained
under environmental conditions most conducive to their metabolic activity
(including microbial growth and continued catabolic breakdown of waste
compounds), can be an important means of biologically transforming or
degrading these otherwise refractory wastes (Doggett, 1983; Evans, 1977;
Horvath, 1972; Kaplan et al., 1982; Knap and Williams, 1982; Kobayashi and
Rittman, 1982; Krupka and Thibault, 1980; Nasset, 1983; Stoddard et al, 1981;
Thibault and Elliott, 1983; Zitrides, 1978). Indeed, it has been postulated
by Horvath (1972) that the concept of molecular recalcitrance (Alexander,
1981) to degradation by microorganisms may not be valid.
2.2.2 Development of Microbial Agents
Microbial systems are available to treat a wide variety of hydrocarbons
(Bitton and Gerba, 1984; Atlas, 1981; Kobayashi and Rittman, 1982; Kopecky,
1983; Zitrides, 1978) including chlorinated and unchlorinated alkanes,
aromatics and polycyclic aromatics, nitrosamines, pesticides and herbicides,
phthalate esters, etc. Biological agents available to degrade these compounds
may occur as, or arise from, naturally occuring microorganisms (Kobayashi and
Rittman, 1982). In addition, biological agents may be acclimated to specific
organic materials or mixtures through a system adaption or mutation/adaptive
regimen (Bitton and Gerba, 1984; Kobayashi and Rittman, 1982; Kopecky, 1983;
Zitrides, 1978), or through the use of plasmid insertion. Microbial strain or
system acclimation may include enzyme induction, strain selection and
mutation. The use of specific nutrients (vitamins, nitrogen, phosphorus,
76
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trace elements, etc.) to encourage microbial growth, and surfactants to
increase substrate solubility can also produce novel biological agents or
systems for the degradation of organic pollutants (Kobayashi and Rittman,
1982; Kopecky, 1983; Zitrides, 1978).
Systems have been developed to treat subsurface soils, groundwaters, surface
spills, lagoons, ponds and other surface waters. These systems include the
use of activated sludge treatment (Kobayashi and Rittman, 1982), fixed film
reactors (Kobayashi and Rittman, 1982), subsurface injection (Kobayashi and
Rittman, 1982; Zitrides, 1978), groundwater pumping for surface treatment
(Zitrides, 1978), and surface application combined with soil turning (Kopecky,
1983; Zitrides, 1978). Several commercial firms have developed proprietary
strains of microorganisms or are capable of adapting native populations for
use in waste site renovation, and have also developed the engineering
technology and treatment systems required for these methods. The kinetics of
biodegradation, as well, as design considerations, for the implementation of
biodegradation systems (i.e., hydraulic design, aeration/oxygenation systems,
use of hydrogen peroxide, ozone and other oxygen sources, nutrient addition,
and operation and maintenance requirements) are described in Repa and Kufs
(1985), A D Little (1983), Jhaveri and Mazzacca (1983), and USEPA (1984d).
Appendix A identifies specific native microflora, microbial consortia,
laboratory derived strains and commercially available microorganisms and the
organic compounds that they are able to transform or degrade. The
environmental conditions that prevail during the course of treatment are
described when data are available. Appendix A also identifies catabolic end
products, degradation rates and treatable concentrations as reported in the
literature. Appendix A indicates that almost every class of organic compound
can be degraded by some microorganism. These microbes include representatives
from the obligate anaerobes, anaerobic bacteria, heterotrophic bacteria,
oligotrophic bacteria, phototrophic bacteria, actinomycetes and fungi.
Bacteria isolated from the environment are often identified to the genus level
only, or if speciated are assigned a strain number. This is done to avoid
confusing them with other members of the genus or species that have not
77
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demonstrated the abilities associated with the organism identified by the
strain number. In other instances, an organism may be identified by the
plasmid it carries. Plasmids are extrachromosomal, inheritable pieces of DNA
(also called episomes) which can be transmitted to other cells. These
episomes are identified by initials which may indicate the substrates they
degrade (such as TOL plasmid specifying toluene and xylene degradation or SAL
plasmid specifying salicylate degradation) or simply by laboratory code
numbers (such as pAC25 specifying 3-chlorobenzoate degradation). Bacteria
characterized by strain or plasmid code identifier in the literature as haying
the capacity to degrade specific compounds or classes of compounds are
identified in this document by that strain or plasmid code.
Obligate anaerobic bacteria are represented by hydrolytic bacteria (which
catabolize saccharides, proteins, lipids); hydrogen producing acetogenic
bacteria (which further break down the products of hydrolytic bacteria, e.g.,
fatty acids and neutral end products); homolactic bacteria (which catabolize
multicarbon compounds to acetic acid); and methanogenic bacteria (which break
down acetic acid to methane and carbon dioxide). The strict anaerobes require
anoxic environments and oxidation-reduction potential of less than -0.2 volts.
These microorganisms are commonly referred to as methanogenic consortia and
are found in anaerobic sediments or sewage sludge digesters. These organisms
play an important role in reductive dehalogenation reactions, nitrosamine
degradation, reduction of epoxides to olefins, reduction of nitro groups and
ring fission of aromatic structures (Evans, 1977; Kobayashi and Rittman, 1982).
The most commonly isolated microbes in areas contaminated with hydrocarbons
are heterotrophic bacteria (i.e., bacteria for which complex organics, rather
than inorganic materials, are the chief source of nutrients) represented by
the genera Pseudomonas, Achromobacter, Arthrobacter, Acinetobacter,
Micrococcus, Vibrio, Brevibacterium, Corynebacterium and Flavobacterium. The
first five genera are of special importance in hydrocarbon degradation
(Kobayashi and Rittman, 1982; Ornston, 1971; Rogers et al, 1981). The genus
Pseudomonas, an environmentally ubiquitous bacteria, has proven to be
especially versatile in its ability to readily adapt to a wide variety of
substrates (Ornston, 1971). Pseudomonads have been adapted and genetically
78
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engineered to degrade an expanding array of substrates including, among
others, halogenated aromatic ring structures (Evans, 1977; Furukawa and
Chakrabarty, 1982; Kellogg et al, 1981; Kilbane et al, 1982; Ornston, 1971;
Serdar et al, 1982; Zitrides, 1978). Members of this genus are able to
catabolize these compounds aerobically using oxygen as the terminal electron
acceptor or anaerobically by nitrate respiration.
Oligotrophic bacteria are defined as microbes that live under conditions of
low productivity (carbon flux of less than one mg/l/day). The Caulobacters
(Poindexter, 1981) are the best known group of obligate oligotrophics but a
number of bacteria, fungi or actinomycetes are capable of adapting to an
existence under these conditions (Kobayashi and Rittman, 1982). These
organisms are found in biofilms and appear to have multiple inducible enzyme
systems and are therefore capable of metabolizing a wide variety of substrates.
Phototrophic microorganisms (i.e., those which obtain energy from sunlight)
include algae, cyanobacteria (blue-green algae) and photosynthetic bacteria.
These organisms are involved in biological transformations rather than
degradation. They are important in that the metabolic products they form from
otherwise refractory organic compounds become the growth substrate of
heterotrophic bacteria (Kobayashi and Rittman, 1982).
Actinomycetes are morphologically similar to both bacteria and fungi and are
known to attack a wide variety of complex organic compounds including phenols,
pyridines, glycerides, sterols, halogenated and unhalogenated aromatic
compounds, paraffins, other long chain organics and lignocellulose. They are
obligate aerobes and are capable of growth under oligotrophic conditions.
They can grow under wide extremes of pH and temperature and are resistant to
desiccation (Kobayashi and Rittman, 1982). Fungi have non-specific enzyme
systems that enable them to degrade or transform hydrocarbons of complex
structure or chain length. These organisms play an important role in the
degradation of aromatic structures including polychlorinated biphenyls (Bumpus
et al., 1985). However, the metabolism of these compounds is often incomplete
and requires an association with bacterial populations to assure complete
mineralization (Kobayashi and Rittman, 1982).
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2.2.3 Factors Affecting The Use Of Microbial Agents ;.
Generally, microorganisms require adequate levels of inorganic and organic
nutrients, growth factors (vitamins, magnesium, copper, manganese, sulfur,
potassium, etc.), water, oxygen, carbon dioxide and sufficient biological
space for survival and growth. One or more of these factors is usually in
limited supply and the various microbial competitors adversely affect each
other through the struggle for these limiting factors (Rosenzweig and
Stotzkey, 1980). Additional factors which can influence microbial
biodegradation rates include microbial inhibition by the test compound, the
number and physiological state of the organisms as a function of available
nutrients, the seasonal state of microbial development, predators, pH (optimum
range is 6-8), and temperature (Fannin et al., 1981). The optimal temperature
for aerobic biodegradation processes is 68°F to 97°F (20-37°C). However,
groundwater temperatures are below this range in many parts of the United
States, leading to suboptimal biodegradation rates (Repa and Kufs, 1985
provides a map of typical groundwater temperatures in the United States).
Interactions between these and other potential factors can cause wide
variations in degradation kinetics.
The studies of Liang et al. (1982) have indicated that there may be risk
associated with the use of biological agents in organic waste treatment.
Components of this risk include the probabiliites of release, survival and
growth of non-indigenous microbes, and the consequent occurrence of some
undesirable environmental or ecological change. An analysis of risk should be
performed for these microorganisms, and any other biological agent
intentionally used for waste treatment, to insure that adverse environmental
impacts will not result from their use. |
2.2.4 Susceptibility of Various Chemical Classes to Biodegradation
This discussion evaluates the various biological treatment technologies
inducible from naturally occurring microbial ecosystems or available
commercially, that are applicable to the degradation of organic waste
materials, including halogenated and unhalogenated alkanes, aromatics and
80
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polycyclic aromatics, pesticides and herbicides, nitrosamines, phthalate
esters, and many others. The catabolic reactions leading to the
biodegradation of organic materials are fairly well known and can be found in
several reviews (Atlas, 1981; Bitton and Gerba, 1984; Evans, 1977; Kobayashi
and Rittman, 1982; Ornston, 1971). Studies by Kobayashi and Rittman (1982)
and others have demonstrated that properly selected microbial populations and
the maintenance of environmental conditions most favorable to their metabolic
activity can degrade significant quantities of organic materials. The role of
microorganisms in hydrocarbon biodegradation is so extensive that Kobayashi
and Rittman (1982) concluded that attempts to generalize the relationships
between chemical structure, substitutions, chain length or molecular size and
biodegradability have so many exceptions that they should be considered only
as broad guidelines.
The relative biodegradability of specific organic compounds can be estimated
based on the ratios of various parameters describing their oxygen requirement
for decomposition. Specifically, the ratios of the 5— day biochemical oxygen
demand (BOD5) to chemical oxygen demand (COD), or the 21-day BOD (BOD21)
to ultimate oxygen demand (UOD) , indicate what proportion of compounds can be
degraded biologically (estimated by BOD) compared to the bioref ractory portion
which would require chemical decomposition (estimated by COD or UOD).
BOD,-/COD ratios and BOD91/UOD ratios (also called refractory index, or RI)
J £• -L
for various compounds are listed in Repa and Kufs (1985). In general,
phenols, alcohols, esters, aldehydes, carboxylic acids and some simple
aromatic compounds (benzene, toluene, napthalene) appear to have relatively
high degradability using these relationships, while halogenated phenols,
aliphatics and aromatics appear to be less readily biodegraded (Repa and Kufs,
1985).
.2.2.4.1 Non-Halogenated Branched and Straight Chain Alkanes
Atlas (1981) discussed the microbial degradation of n-alkanes with chain
lengths up to C,, . The initial degradation produces a primary alcohol,
followed by an aldehyde and a monocarboxylic (fatty) acid. The carboxylic
acid is further oxidized to a shorter-chain fatty acid. The catabolism of
81
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long chain carboxylic acids can be inhibited by shorter chain carboxylic
acids, thus preventing further degradation of the longer chain carboxylic
acids (Atlas and Bartha, 1973). ;
Straight chain and branched alkanes are readily degraded by a wide variety of
soil and salt water bacteria, and activated sludge microorganisms (Hill and
McCarty, 1967; Kobayashi and Rittman, 1982; Kobayashi and Tchan, 1978; Murray
and Van der Berg, 1981; Ornston, 1971; Wolfe et al.,1980; Yordy and Alexander,
1980). Organisms identified in the literature include pseudomonads (de Smet
et al., 1981; Litchfield and Clark, 1973; Ornston, 1971; Perry, 1979),
actinomycetes (Perry, 1979), yeast microorganisms (Perry, 1979), Bacillus (de
Smet et al., 1981), Corynebacterium (de Smet et al., 1981), methylotrophic
bacteria (Mancinelli et al., 1981), and activated sludge organisms (Kobayashi
and Rittman, 1982). These alkanes, especially those with shorter chains,
often occur in nature as plant and animal byproducts. Microorganisms more
readily adapt to the catabolic use of these compounds than to more complex
structures, esentially because of their simple form and availability in the
environment. Native populations of microorganisms have been shown to degrade
aliphatic hydrocarbons ranging in concentration from 1 ppm or less up to
approximately 1,000 ppm or more. Within the concentration range specified,
the biodegradation of these aliphatics appears to be dependent on the
solubility of the hydrocarbon in the environment (Atlas, 1981). The ability
of these organisms to degrade higher concentrations of specific aliphatics can
be enhanced by the process of adaptation or genetic manipulation (Doggett,
1983; Kobayashi and Rittman, 1982; Kopecky, 1983; Ornston, 1971; Zitrides,
1978).
Litchfield and Clark (1973) found that significant populations of bacteria are
present in groundwater contaminated with hydrocarbons including gasoline, fuel
oil, and other petroleum products. They found that waters containing less
3
than 10 ppm hydrocarbons generally had populations of less than 10
organisms per ml while waters with hydrocarbon concentrations in excess of 10
ppm generally supported populations on the order of 10 organims per ml.
Species were identified as belonging mostly to the genera Pseudomonas and
Arthrobacter. ,
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Highly branched isoprenoid alkanes are degraded to dicarboxylic acids. Methyl
branching, which generally increases the resistance of hydrocarbons to
microbial attack, requires that microorganisms use additional degradational
mechanisms (Atlas, 1981). Representative acyclic hydrocarbons subject to
biodegradation are shown in Appendix A.
-'£
Mutants of Pseudomonas, Aerobacter and Micrococcus have been shown, under
laboratory conditions, to be capable of the total degradation of nitriles,
cyanides and amines at concentrations ranging from 250 ppm to 500 ppm in a
matter of hours. When used in a bioreactor renovation of waste water, these
organisms were capable of degrading acrylonitrile at concentrations ranging
from 100 to 1,000 ppm to less than 1 ppm over a period of 3 months (Krupka and
Thibault, 1980; Nassef, 1983; Zitrides, 1978). The degradation of specific
compounds within these groups, concentrations listed, degradation times and
microorganisms involved are identified in Appendix A.
2.2.4.2 Aromatic Compounds and Phenols
Extensive studies on the catabolism of aromatic compounds by microorganisms
have identified many of the pathways and mechanisms involved in their
degradation. These cyclic compounds have been reported to be substrates for
cooxidation with the formation of an alcohol or ketone (Atlas, 1981; Horvath,
1972; Jacobson et al., 1980; Perry, 1979). Substituted cyclic compounds are
more easily degraded than unsubstituted forms, particularly if the substituent
is an n-alkane of adequate chain length (Atlas, 1981; Perry, 1979). Microbial
attack in such cases usually occurs first on the substituent, producing an
intermediate product such as cyclohexane, carboxylic acid or similar compound.
Bacterial (procaryotic) degradation of aromatic compounds usually involves the
formation of a diol, followed by ring cleavage and the production of a diacid
(Atlas, 1981; Evans, 1977; Ornston, 1971). Eucaryotic organisms, in contrast,
oxidize aromatic compounds to the trans diol (Atlas, '1981). Aromatic
compounds can be degraded both aerobically and anaerobically (Atlas, 1981;
Evans, 1977; Ornston, 1971). Microbial systems "capable of degrading various
aromatic compounds are identified in Appendix A by compound.
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Degradation rates for cyclic aliphatics and aromatic hydrocarbons by native
microbes are slower than for acyclic compounds. Native microorganisms will
completely or partially degrade these compounds at concentrations below 100
rag/1 over a period of 10 to 90 days (Horvath, 1972; Rogers et al., 1981; Rubin
et al., 1982). Little data are available concerning mutant bacterial
degradation rates on unsubstituted aromatic hydrocarbons. However, genetically
altered strains of Psuedomonas, Alcaligenes or Micrococcus have been shown in
the laboratory to completely degrade substituted aromatics at concentrations
of 200 to 500 ppm over an interval of several hours to a few days (King and
Perry, 1975; Krupka and Thibault, 1980; Marinucci and Bartha, 1979; Nassef,
1983; Pfaender and Bartholomew, 1982; Zitrides, 1978). Specific cyclic and
aromatic hydrocarbons which have been shown to be susceptible to microbial
attack are identified by compound in Appendix A. Also identified are the
microbial communities or specific microbes that are capable of degrading these
compounds.
2.2.4.3 Polycyclic Aromatic Hydrocarbons '| . •
Microbial degradation of polycyclic aromatic hydrocarbons (PAH) compounds have
been identified (Atlas, 1981; Cohen and Gabriele, 1982; Herbes, 1981;
Kobayashi and Rittman, 1982; Sherrill and Sayler, 1980). However, uniform
degradative pathways, comparable to those for the aliphatic and aromatic
compounds, have not yet been determined (Atlas, 1981). Among the naturally
occurring systems which degrade these compounds are: the fungi Polyporus
versicola and Poria monticola which have the capacity to degrade lignite coal
(Cohen and Gabriele, 1982); the microbial transformation of anthracene and
benz(a)anthracene by stream water and sediment bacteria (Herbes, 1981); and
the biodegradation of phenanthrene in fresh water environments i(Sherrill and
Sayler, 1980). These organisms are usually found downstream of surface water
pollution sites (Furukawa and Chakrabarty, 1982; Shiaris et al., 1980). Rates
of PAH biodegradation by naturally occurring microbial populations are
relatively slow when compared to degradation rates for aliphatic and aromatic
compounds. The degradation rates of PAH have been shown to be directly
related to historic environmental pollution of the sampling site, the length
of biodegradation assessment, temperature and the molecular size of the
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substrate (Sherrill and Sayler, 1980). Transformation rates in microbial
communities shift slowly in response to changes in PAH concentration, but have
been shown to remain elevated for more than a year after the removal of the
PAH source (Herbes, 1981).
Microbial degradation of two and three ring PAH in the environment has been
demonstrated (Aranha and Brown, 1981; Brilon et al., 1981b; Cerniglia et al.,
1980; Doggett, 1983; Furukawa and Chakrabarty, 1982; Kiyohara et al., 1982;
Knap and Williams, 1982; Kobayashi and Rittman, 1982; Reichartdt et al.,
1981). Where identified, PAH concentrations were at or below 100 ppm. Rates
of degradation in the environment were highly variable. Native stream
bacteria were shown to degrade anthracene (presumably at trace concentrations)
over a period of 64 days (Furukawa and Chakrabarty, 1982; Kobayashi and
Rittman, 1982). However, a plasmid assisted Alcaligenes was capable of
degrading this compound in one to three days (Kiyohara et al., 1982). Native
soil bacteria were shown to degrade 100 ppm of naphthalene in 48 hours (Aranha
and Brown, 1981). The substituted and unsubstituted forms of napthalene are
also degraded by Pseudomonas aerobically (Brilon et al., 1981a) and by
phototrophic bacteria anaerobically (Cerniglia et al., 1980). Biphenyl was
found to be degraded aerobically by plasmid assisted strains of Acinetobacter
and Arthrobacter (Furukawa and Chakrabarty, 1982), by Alcaligenes aerobically
in one to three days (Knap and Williams, 1982) and by Beijerinkia (Kobayashi
and Rittman, 1982).
2.2.4.4 Halogenated Organic Compounds
Halogenation is often implied as the reason for the presistence of an organic
compound in the environment. Some of the characteristics that promote
environmental persistence include: the location of the halogen atom oil the
organic compound; the halide involved; and the extent of halogenation
(Kobayashi and Rittman, 1982). Anaerobic reductive dehalogenation (removal of
a halogen atom by oxidation-reduction), either biological or abiological, has
been identified as the critical factor in the biodegradation or chemical
transformation of halogenated organics (Bouwer and McCarty, 1983a; Bouwer and
McCarty, 1983b; Edgehill and Finn, 1983; Guenzi and Beard, 1967; hill and
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McCarty, 1967; Kallman and Andrews, 1963; Kobayashi and Rittman, 1982;
Marinucci and Bartha, 1979; Reichartdt et al., 1981; Schreiber et al., 1980).
Reductive dechlorination is reported to be significant only when the
environmental oxidation-reduction potential (E,) is at or below 0.35V, with
the exact requirements dependent upon the compound involved (Kobayashi and
Rittman, 1982). Kobayashi and Rittman (1982) report that compounds degraded
via anaerobic reductive dechlorination include many pesticides as well as one
and two carbon halogenated aliphatic compounds. However, it is important to
note that polychlorinated biphenyls and halogenated benzenes have been found
to be degraded only under aerobic conditions (Kobayashi and Rittman, 1982).
A great deal of concern has been expressed concerning the persistence of
organic pesticides in the environment, particularly the more persistent
chlorinated pesticides. Hill and McCarty (1967) report that although these
compounds are resistent to aerobic decomposition they degrade more quickly
under biologically active anaerobic conditions. This degradation may be a
complete mineralization or a partial degradation to other organic end products
(Guenzi and Beard, 1967; Hill and McCarty, 1967; Kallman and Andrews, 1963).
One and two carbon halogenated aliphatic organic compounds at trace
concentrations were found to be subject to dehalogenation and degradation
under anaerobic but not, aerobic conditions (Bouwer and McCarty, 1983b; Bouwer
et al., 1981; McCarty et al., 1981). Several one and two carbon halogenated
aliphatic organic compounds present at low concentrations (less than 100 ug/1)
were degraded under methanogenic conditions in a continuous flow fixed film
biological reactor. Greater than 90 percent biodegradation was observed after
two days under continuous flow methanogenic conditions (Bouwer and McCarty,
I983a).
A number of halogenated aromatic and aliphatic compounds have been reported to
be dehalogenated in sewage (DiGeronimo et al, 1979; Jacobson and Alexander,
1981) and soil (Edgehill and Finn, 1983; Marinucci and Bartha, 1979).
Jacobsen and Alexander (1981) have reported the dechlorination of 4-chloro-3,
5-dinitro~ benzoic acid as a result of microbial growth both in the light (in
the absence of added nutrients) and in the dark (in the presence of acetate).
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Axenic (pure) bacterial cultures of Chlamydomonas and sewage microfauna
release chlorine from the compound and the latter produces alpha
hydroxymuconic semialdehyde as an endproduct. This material reportedly serves
as a substrate for further metabolism by a strain of Streptomyces (Jacobson
.and Alexander, 1981). The degradation of 1,2,3- and 1,2,4-trichlorobenzene
has been reported in soils with CO^ evolution (Marinucci and Bartha, 1979).
Pentachlorophenol degrading bacteria of the genus Arthrobacter, capable of the
complete mineralization of the compound, have recently been isolated from
soil, water and sewage (Stanlake and Finn, 1982). Direct inoculation of
Arthrobacter cells into pentachlorophenol contaminated soils reduced the
half-life of the pesticide from two weeks to less than one day, using 106
Arthrobacter cells per gram of dry soil at 30°C (Edgehill and Finn, 1983).
Microorganisms in sewage have been reported to degrade S^-dichlorobenzoates,
meta-, para- and orthobenzoates (Di Geronimo et al., 1979), and a
3-chlorobenzoate grown strain of Pseudomonas sp. B13 readily degrades
monofluorobenzoates (Schreiber et al., 1980).
Recent advances in microbial genetics have shown that improved degradation of
halogenated hydrocarbons can be achieved with constructed strains (Schwien and
Schmidt, 1982). In this study a Pseudomonas strain B13 able to degrade
3-chlorobenzoate and 4-chlorophenol, could transfer the ability to degrade
chlorocatechols to an Alcaligenes strain A2 recipient capable of growing on
benzoate and phenol. The transconjugant was able to use all three isomeric
chlorophenols, a property not possessed by either parent.
Chlorinated and polychlorinated biphenyls have been shown to be degraded by a
variety of plasmid assisted bacteria (Doggett, 1983; Furukawa and Chakrabarty,
1982; Kobayashi and Rittman, 1982; Reichartdt et al., 1981) with the rate of
degradation being inversely proportional to the level of chlorination.
Organisms capable of degrading chlorinated biphenyls include Acinetobacter
(Furukawa and Chakrabarty, 1982), Arthrobacter (Reichartdt et al., 1981),
Pseudomonas, Flavobacter, Archromobacter, Chromobacter and Nocardia (Kobayashi
and Rittman, 1982). Compounds in this category which have been shown to be
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degraded microbiologically are identified along with the agent of degradation
in Appendix A.
Degradation of chlorinated biphenyls have been observed in a mixed marine
mlcrobial community with estimated turnover rates of one year at concentra-
tions of 0.1 ug/liter or less, and higher turnover times probable at higher
concentrations (Reichartdt et al., 1981). However, an Arthrobacter strain M5
contaminant of an Acinetobacter strain P6 culture grown on biphenyl and
chlorinated biphenyls showed properties similar to the P6 strain (Furukawa and
Chakrabarty, 1982) as a result of a presumed plasmid transfer. The
Acinetobacter P6 strain can degrade 33 pure isomers of chlorinated biphenyl
including di-, tri- and tetrachlorobiphenyls. A combined culture of the
chlorinated biphenyl degrading P6 and M5 strains and genetically constructed
mono- or dichlorobenzoate utilizing pseudomonads (harboring the TOL, pAC25,
plCFl, pAC21 and pAC30 plasmids which regulate the degradation of these
aromatic compounds) allowed greater than 98 percent utilization of mono- and
dichloro-biphenyls, with the liberation of equivalent amounts of chloride
ions. Once dechlorinated, these compounds are degraded by mechanisms as
previously described. Appendix A describes those systems capable of degrading
chlorinated hydrocarbons.
2.2.4.5 Herbicides And Pesticides
Some herbicides and pesticides undergo fairly rapid decomposition in anaerobic
ecosystems (Guenzi and Beard, 1967; Hill and McCarty, 1967; Kallman and
Andrews, 1963; Lewis and Holm, 1981; Reddy and Sethunathan, 1983) and this
ability can be enhanced through genetic modifications (Chatterjee et al.,
1982; Kellogg et al., 1981; Kilbane et al., 1982; Serdar et al., 1982).
Microorganisms can also degrade these materials through cometabolism (i.e.,
not using these organics as a primary nutrient, but degrading them ,as an
ancillary action of normal metabolic activity) (Fogel et al., 1982; Jacobson
et al., 1980; Patil et al., 1972). Chlorodimeform has been shown to be
hydrolyzed by Chlorella and Oscillatoria to toluedide, which was, deformylated
to yield toluedine followed by fission of the aromatic nucleus (Benezet and
Knowles, 1981). Microorganisms were able to accomplish the complete or
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partial degradation of lindane, heptachlor, endrin, aldrin, heptachlor
epoxide, DDT, ODD and dieldrin in anaerobic digester sludge (Hill and McCarty,
1967). DDT is converted to ODD by yeast (Guenzi and Beard, 1967; Kallman and
Andrews, 1963). Low concentrations of methyl parathion are degraded under
aerobic conditions by aufwuchs (attached to a substrate rather than
free-floating) bacteria (Lewis and Holm, 1981). Endosulfan can be degraded by
16 fungi, 15 bacteria and 3 actinomycetes (Martens, 1976). Parathion is
mineralized by bacteria in the rice rhizosphere under anaerobic (flooded) and
aerobic .(non-flooded) conditions (Reddy and Sethunathan, 1983), and various
organophosphate insecticides have been cleaved under aerobic conditions and
mesophillic temperatures (Rosenberg and Alexander, 1979). In addition, the
cometabolism of trifuralin, profluralin, fluchloralin and nitrofen (Jacobson
et al., 1980) methoxychlor (Fogel et al., 1982), DDT, dieldrin, aldrin and
endrin (Patil et al., 1972) in various environments has been reported. The
degradation of these compounds by microbes is identified in Appendix A. When
identified in the literature, general environmental conditions and byproducts
have also been compiled.
Recent studies have shown that Pseudomonas cepacia AC1100 was' capable of using
2,4,5,-trichlorophenoxyacetic acid (2,4,5-T or Agent Orange) as a sole source
of carbon at concentrations of 1 mg per gram of soil (Chatterjee et al., 1982)
and 1 mg per milliliter (Kilbane et al., 1982), within one week. Optimum
degradation rates occurred at 30°C and 25 percent moisture content (Chatterjee
et al., 1982). Another organism, Pseudomonas diminuta, was found to have
enhanced capabilities to hydrolyze parathion because of plasmid pCSI (Serdar
et al., 1982). The degradation of a number of chlorinated hydrocarbons such
as 3-chloro or 4-chlorobenzene has recently been reported (Kellogg et al.,
1981). Kellogg et al. (1981) have demonstrated that plasmid pAC25 which
encodes the complete degradation of 3-chlorobenzoate does not allow host cells
to use 4-chlorobenzoate. However, the introduction of the TOL plasmid, which
specifies for xylene and toluene degradation, provides the microbe with a
broad substrate-specific benzoate oxygenase which allows the host cell to
degrade 4-chlorobenzoate and extends this cell's metabolic range to other
chlorobenzoates as well. These plasmids appear to evolve by recruitment of a
variety of genes from other plasmids arid interact among themselves to extend
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the substrate range of host cells to a wide variety of xenobiotic compounds.
Kellogg et al. (1981) report having developed by plasmid assisted molecular
breeding a culture of microorganisms harboring a variety of plasmids (such as
CAM, TOL, SAL, pAC21 and pAC25) which were capable of degrading
2,4,5-trichloro-phenoxyacetic acid at concentrations of 1.5 to 2 mg/ml. These
and similar laboratory derived microbial systems are being developed for
commercial use (Doggett, 1983; Kopecky, 1983; Zitrides, 1978). These include
the nine BI-CHEM systems (Kopecky, 1983), the PHENOBAC systems (Zitrides,
1978) and the strains of pseudomonads being developed by Doggett (Doggett,
1983).
2.2.4.6 Phthalate Esters
Esters of phthalic acid are industrially important chemicals used mainly in
the manufacturing of plastics, pesticides, and cosmetics and are ubiquitous in
the environment (Aftring, 1981; Benckiser and Ottow, 1982). Aftring et al.
(1981) reported that mixed cultures of bacteria from aquatic sediments were
capable of degrading phthalic acid, isophthalic acid and terephthalic acid
under anaerobic conditions. Benckiser and Ottow'(1982) have reported on the
metabolism of di-n-butylphthalate by a denitrifying strain of Pseudomonas
pseudoalcaligene s B20 bl. They suggested that one butanol moiety mostly
served as the carbon source for growth and denitrification. Others have
identified the mineralization of di(2-ethylhexyl) phthalate in lake water at
trace concentrations (Rubin et al., 1982) and the biodegradation of this
phthalic acid ester in a marine environment (Subba-Rao et al., 1982). Wolf et
al, (1980) have identified second order microbial degradation rate constants
for four phthalate esters obtained from sediment microorganisms and correlated
them with second order alkaline hydrolysis rate constants. The plasticizer
diethyl phthalate was also reported degradable by aufwuchs bacteria (microbial
growth attached to submerged surfaces) (Lewis and Holm, 1981). Microbial
systems reported to degrade phthalate esters have been identified in Appendix
A. ' ' ••
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2.2.4.7 Nitrosamines
Nitrosamines have received a great deal of recent attention because of their
carcinogenicity, mutagenicity and teratogenicity, and their presence in foods,
drugs and pesticides. These compounds have also been found in soils and water
and the potential for formation in the environment has been reported (Kobayashi
and Tchan, 1978; Yordy and Alexander, 1980).
In polluted waters, the compound dimethylnitrosamine has been shown to occur
as a result of sludge decomposition. However, photosynthetic bacteria and
other microorganisms were found to anaerobically metabolize this compound
(Kobayashi and Tchan, 1978). The carcinogen n-nitrosodiethanolamine (NDEIA)
was shown to degrade slowly at low concentrations (1 ug/ml) in samples of
sewage and lake water under anaerobic conditions (Yordy and Alexander, 1980).
The products formed appeared to be modified dimers of NDEIA and were slowly
mineralized in sewage. The bacterial degradation of nitrosamines is identified
in Appendix A. When provided in the literature, general environmental
conditions and byproducts are identified.
2.3
Application to Waste Deposits
Commercial operations already exist (Aquifer Remediation Systems, 1985;
Doggett, 1983; Flathman and Caplan, 1985; Jhaveri and Mazzacca, 1983; Kopecky,
1983; Kretschek and Krupka, 1984; Yaniga, 1982; Zitrides, 1978; USEPA, 1984)
which either have microbial strains in stock capable of degrading organic
wastes in situ or in portable biological reactors, and have the facilities to
adapt these organisms, or native microbes, to specific waste reclamation
problems. Among the companies contacted, four were willing to provide
information on their products, treatment processes, and site applications and
are identified by reference in this document (Jhaveri and Mazzacca, 1983;
Doggett, 1983; Kopecky, 1983; Zitrides, 1978). The remaining organizations
contacted considered their treatment processes or products proprietary
information or simply had insufficient information on product application to
be useful and therefore were not referenced in this document.
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In renovating hazardous waste sites, site operators may choose to develop
native populations to degrade 'wastes or may wish to use a commercial operation
to treat a site. The information presented below was developed from case
histories on treating surface soils (Doggett, 1983; Kaplan and Kaplan, 1982;
Kilbane et al., 1982; Kopecky, 1983; Zitrides, 1978), deep soils and
groundwaters (API, 1982; Jhaveri and Mazzacca, 1983; Kaplan and Kaplan, 1982;
Zitrides, 1978), lagoons or surface impoundments (ZitrideSj, 1978) and
industrial waste treatment plants (Zitrides, 1978).
I,
2.3.1 Site Assessment
Before a waste site can be reclaimed, the extent and degree of contamination
must be assessed as described in Chapter 1 of this report and in Repa and Kufs
(1985). This includes chemical analysis to identify and quantify hazardous
materials. The waste pile and soils surrounding the site should also be
tested for porosity, pH, nitrogen, phosphorous and trace minerals, to
establish the nutritional content of the soils and materials to be treated.
The proper microorganisms or groups of microbes must be selected to treat the
waste. Commercial firms use their past experience, laboratory screening,
onsite test plots, or any combination of these procedures to identify the
proper agents (either native populations or constructed strains) for waste
site renovation. If native microorganisms are selected, the laboratory
cultivates the microbes in the presence of low waste concentrations.
The initial waste concentration used is determined by performing waste
toxicity studies on the native populations. In order to breed organisms
capable of degrading specific wastes or waste groups, it may be necessary to
initially isolate and test individual species from the native population for
their ability to degrade identified waste groups (Kellogg et al., 1981) or
simply develop a waste degrading system using the entire native population
(Kaplan and Kaplan, 1982). The microbes or native populations are then
innoculated into laboratory scale systems that model the environment of the
contaminated site with respect to soil moisture, pH, temperature and pE
(dissolved oxygen content). In many instances a chemostat, fermentor or other
dynamic modeling system (microcosm) can be used for this purpose (Doggett,
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1983; Flathman and Caplan, 1985; Jhaveri and Mazzacca, 1983; Kaplan and
Kaplan, 1982; Kellogg et al., 1981; Kopecky, 1983; Zitrides, 1978). Native
microbial populations or microorganisms selected from these populations for
their ability to degrade specific waste groups can then be fed increasing
concentrations of the waste groups involved until a population develops that
is capable of degrading the hazardous organic components at on-site
concentrations under ambient environmental conditions. Laboratory studies of
this nature can take up to a year or longer to complete (Kellogg et al., 1981;
Jhaveri and Mazzacca, 1983).
Commercially available systems and adapted native microflora will use wastes
as sole carbon sources. However, these organisms also require sources of
nitrogen, phosphorus and trace elements which may not be present at the waste
treatment site in sufficient concentration to support optimum growth.
Generally, the desired ratio of carbon:nitrogen:phosphorus is 100:15:3. To
achieve this ratio, an analysis of the site soil matrix is required. The
contaminated site is modeled in the laboratory and augumented with commercially
available fertilizer sources (ammonium nitrate, sodium phosphate, etc) until
the desired ratio is obtained (Jhaveri and Mazzacca, 1983). Two of the firms
contacted for this study (Kopecky, 1983; Zitrides, 1978) have proprietary
formulations available for use as part of their treatment package, but must
still determine concentrations required for optimum waste degradation by
modeling the system in the laboratory. This process usually takes four to six
weeks (Kopecky, 1983; Zitrides, 1978).
Site temperatures, waste type or concentration, or other environmental factors
may render the waste insoluble. Emulsifying agents (surfactants) may be
required to increase the microbial availability of low solubility waste
constituents. Optimum treatment occurs when wastes are solubilized at a rate
that will allow maximum microbial catabolic (degradation) rates under the
environmental conditions imposed. If wastes are solubilized too slowly, then
maximum microbial growth rates will not be achieved due to insufficient
substrate concentration. If wastes are solubilized too rapidly, then
microbial growth may be inhibited by excess substrate in the environment.
Therefore it is important to determine, in the laboratory or in test plots at
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the site, the optimum concentrations of microbes, emulsifier a;nd fertilizer
required to support maximum biological activity. Two waste treatment
companies (Kopecky, 1983; Zitrides, 1978) have identified emulsifiers that are
available as part of their treatment packages.
Optimal microbial activity occurs in partially or fully saturated soil
conditions (-0.1 to 1.0 bars soil water vapor pressure, USEPA,, 1984). The
degradation rates of organic compounds may thus be enhanced by addition of
water (via irrigation, flooding, injection, etc. — see Section 1.2) or
drainage of saturated soils (via drainage ditches or wellpoints — see Section
1.3).
Depending on the waste types and microbial degradation pathways to be used,
aerobic or anaerobic conditions may be required (see Section 2.2). Oxygenation
can be accomplished by surface filling (USEPA, 1984), injection of air (Jhaveri
and Mazzacca, 1983), ozone (Nagel, 1982), or hydrogen peroxide (Wetzel et al.,
1985; Aquifer Remediation Systems, 1985). Anaerobic conditions can be
generated by flooding (without oxygen injection) and addition of excessive
amounts of easily biodegradable organic matter (to utilize available oxygen)
(USEPA, 1984). In addition, the surface may be covered with a synthetic
membrane liner, compacted or temporarily sealed to reduce the influx of oxygen.
The soil or groundwater pH may also require alteration, since optimal
microbial growth is in the pH range of 6-8 (Fannin et al., 1981),, Soil pH is
also an important factor in determining the effects of pesticides on soil
microbes (USEPA, 1984). Crushed limestone, lime products, or soda ash can be
used to increase the pH while acid-producing materials (aluminum or ferrous
sulfate) or sulfur will lower the soil pH (USEPA, 1984).
It has been shown in extensive laboratory testing that supplemental carbon and
energy sources (easily-biodegradable organic matter) can stimulate the
biodegradation of recalcitrant organic compounds through cometabolism (Fogel
et al., 1982; Jacobsen et al., 1980; Kaplan and Kaplan, 1982; Patil et al.,
1982; USEPA, 1984 and references therein). This process has been used to
promote the biodegradation of recalcitrant chlorophenol compounds at the
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Picillo Farm Site, Rhode Island (see Section 2.3.5; also Flathman et;al., 1983
and Flathman and Caplan, 1985). An interesting aspect of soil amendment-with
organic matter is the use of analog enrichment to promote cometabolism. In
this approach, a non-hazardous chemical analog of the hazardous compound is
added to the waste deposit to stimulate the native microbes' degradative
pathways for that type of compound. The structurally-similar hazardous
compound is often cometabolized (Alexander, 1981; USEPA, 1984 and references
therein). For example, addition of biphenyl to test soils stimulated the
biodegradation of polychlorinated biphenyls (PCBs) (Brunner et al., 1985).
The potential use of non-specific organic amendments or specific chemical
analogs to promote cometabolism would require evaluation during laboratory and
field pilot studies.
Once the proper microbial strains, site saturation, pH and oxygen
requirements, fertilizer formulations and emulsifier concentrations have been
identified, and the degradation rates and application rates are known, the
time course and economics of treatment can be identified for each site. Using
information gained from laboratory studies, scaled-up pilot studies may be
required to model waste treatment systems under field conditions to confirm
technical and economic feasibility of biological waste treatment (Zitrides,
1978). This study would be most appropriate if continuous long term treatment
is required. Having completed the above steps, sufficient quantities of
biological agents can be cultured and freeze-dried for transport, storage and
use at the site.
Based on the information developed above, the procedures to properly implement
an in situ biodegradation system (Zitrides, 1978) are summarized below:
o Collect data on waste sample analysis, soil composition and
indigenous microbial populations;
o Obtain monitor well data if applicable;
o Collect any other site data (soil type, moisture, pH, pE, temperature,
nutrients, etc.), necessary to complete a bench-scale study;
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o Analyze data and choose or breed proper microorganisms;
i
o Mutate and culture those microbes to perform desired functions;
o Inoculate waste samples with selected microbes under accepted
scientific procedures, and observe biodegradation rates;
o Determine optimum soil moisture, pH, oxygen, fertilizer and
eraulsifier requirements;
o Establish technical and economic feasibility of biological approach
to waste treatment;
o If technically and economically feasible, perform pilot study if
required;
o Construct treatment system and begin in situ waste treatment.
2.3.2 Case Histories of In Situ Treatment of Surface Waste Deposits or
Spills
2.3.2.1 Cleanup of Crude Oil Spill at an Oil Storage Site
A 1.6 hectare (four acre) spill area with crude oil penetration to a depth of
0.5 meters (1.5 feet) was restored to a condition where oil could not be
detected in the soil (Zitrides, 1978; Kretschek and Krupka,; 1984). The
cleanup process began by flooding the spill area to float unabsorbed crude
oil. Vacuum trucks removed floated oils, leaving oil residues absorbed in
soil. The site was tilled to create an aerated matrix of soil and crude oil.
Approximately 180 kg (400 pounds) of nutrient slurry (POLYBAC N) and 18 kg (40
pounds) of nonionic dispersant (POLYBAC E) were sprayed over the site to
precondition it for optimal microbial growth. A total of 23 kg (50 pounds) of
HYDROBAC bacteria (a commercially available, adapted, mutant bacterial
culture) were reconstituted with 1900 liters (500 gallons) of clean water and
sprayed over the contaminated area and the soils tilled again,, Nutrients,
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emulsifier and bacteria were reapplied after six weeks; the site was tilled as
required for aeration and to assure the complete mixing of microbes,
fertilizers and emulsifiers with the waste materials. Soil moisture was
maintained through application of water to keep the soil moist but not
flooded. Following two months of treatment, the site supported vegetation and
the appearance of the area was approaching normal. The treatment reduced oil
concentrations in the soil by 66% during the first five weeks (Kretschek and
Krupka, 1984). Treatment continued until crude oil residues could not be
observed in soils.
2.3.2.2 Applications of Adapted Bacterial Cultures to Surface Waste
Deposit Sites
The process of selection and application of DETOXSOL bacterial cultures for
specific contamination problems involves laboratory screening, testing at
onsite pilot plots and review of past experience with similar problems
(Kopecky, 1983). In general, the selected bacterial culture is sprayed evenly
o
over the site at the rate of one kilogram of bacteria per 25 m (one pound
2
of bacteria per 120 ft ). The bacteria are applied at weeks 1, 2, 4, and 6
of the treatment period, and every two weeks thereafter as required. Soils
are assayed for removal of organic contaminants before each application. The
contaminated site is usually watered daily to keep soils damp but not to the
point of flooding. The ground is tilled weekly when soil aeration or mixing
is required.
Soils are monitored for ammonia-nitrogen and orthophosphates. If these
minerals are found to be less than 5 ppm, then four kg of 8-8-8 (8% each of C,
2 2
N and P) fertilizer are added per 25 m (four pounds per 120 ft ) with the
bacterial application. Soils may be covered with polyethylene to stabilize
temperature and moisture. This method was described as effective to depths of
eight to twelve inches depending on soil porosity (Kopecky, 1983). However
with extended time and by using injection wells, sites can be detoxified . to
depths of several feet or more. Four case histories (Kopecky, 1983) in which
this system was successful in the treatment of styrene, atrizine, petroleum
distillate and trichlorophenate are shown in Table 2-1.
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TABLE 2-1
EXAMPLES OF BIOLOGICAL RENOVATION AT CONTAMINATED SURFACE SITES1
Waste Waste Treatment
Site Concentration Site Biological Tine
Contanioant (ppm) Characteristics Agent (Days)
Styrene 25
Atrazlne Saturated
Soil
Sludge containing 300
vo Tricholorophenate
oo
Petroleum 12,000
Distillate
Acrylonitrile 1,000
Formaldehyde 1,400
Ortho-Chloro- 15,000
phenol
Railroad tankcar spill, fll-CHEH SOS-B 21
area soils contaminated
to a depth of 8
Inches
50 Acre field BI-CHEM-PBO-6
Sludge spread on soil Bl-CHEM-GEC-1 28
to a depth of 6 inches
Spill covering 4 BI-CHEM-SUS-8 21
acres at an oil tank
farm
Soil and groundwater PHENOBAC 90
Contamination
Soil and groundwater PHENOBAC 22
contamination
Soil and groundwater MUTANT 274
Contamination BACTERIA
Residual
Concentration Reference
(ppm)
less than 1 1
1
less than 1 1
less than 1 1
less than 1 2
less than 1 2
less than 1 2
1 Application of - biological agents and site treatment for these examples are similar to those procedures
described in the text to treat hazardous organic wastes.
References: 1) Kopecky, 1983
2) Zitrides, 1978
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2.3.2.3 Biodegradation of Formaldehyde in Surface Soils
The main valve on a railroad tank car containing a 50% formaldehyde solution
was inadvertently opened and about 80,000 liters of the solution spilled over
the railroad ballast, into an adjacent ditch, through an orchard irrigation
system and into a river (Kretschek and Krupka, 1984). The ponded formaldehyde
was removed by vacuuming, and biological treatment was selected as the least
disruptive and most cost-effective approach for cleaning the contaminated soil
and railway ballast, which contained 700 to 1400 ppm formaldehyde. HYDROBAC
bacteria, an adapted mutant culture, was tested in the laboratory (in media
supplemented with soil extract and nutrients) and shown to be capable of
degrading formaldehyde (Kretschek and Krupka, 1984).
A 75,000 1 (20,000 gallon) bioreactor was filled with fresh water, nutrients,
surfactant and bacterial innoculum (HYDROBAC) and aerated (Kretschek and
Krupka, 1984). The solution from the tank was sprayed over the railroad track
ballast at a rate of 190-380 1/min. Leachate was collected in a drainage
ditch. The solution was continuously recycled and fresh water, nutrients,
surfactants and microorganisms were added daily. During the first week of
treatment, the formaldehyde concentration in the leachate was reduced from 750
ppm to 250 ppm (Kretschek and Krupka, 1984). Following three weeks of
treatment, residual formaldehyde in the leachate was less than one ppm.
2.3.3 Case Histories of In Situ Treatment of Subsurface Waste Deposits
or Spills
Subsurface waste deposit renovation poses problems relating to oxygen supply,
temperature, permeability and accessibility (API, 1982; Kopecky, 1983;
Zitrides, 1978) not encountered with surface disposal sites. Waste treatment
involves pumping selected microbes, including emulsifiers, fertilizers and an
oxygen source into wells penetrating the waste deposit and into peripheral or
downgradient wells (API, 1982; Doggett, 1983; Jhaveri and Mazzacca, 1983;
Kopecky, 1983; Zitrides, 1978) as required. Thus not only the waste pile is
treated but any groundwater plumes that may , be migrating from the site may be
renovated as well.
99
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Liquids recovered from the waste deposit are monitored to determine waste
degradation rates and may be used in the formulation of bacteria-emulsifier
and oxygen source-fertilizer preparations. Additional liquid treatment may be
required and may be cost-effective at the site surface prior to its
reinjection back into the waste site (Jhaveri and Mazzaccaj, 1983). If
recovery liquid is insufficient for this purpose, it may be supplemented with
fresh water. Sites low in moisture content can be moistened by the injection
of fresh water along with the treatment preparations.
The practicality of subsurface waste site renovation ultimately depends on
soil and waste pile permeability and site temperature. The treatment of waste
sites in high clay content soils, wastes containing large concentrations of
highly insoluble waste, or a combination of these factors may make biological
renovation of waste sites impractical. Waste site temperatures are controlled
by in situ soil temperatures and biological activity. Any environmental or
biological factors which may cause a site to be too cold or too hot will
adversely affect biological waste treatment. Temperatures at waste disposal
sites should be high enough to support microbial growth. Low temperature has
been reported as a limiting factor for microbial growth and this is reflected
in the 8 to 24 month renovation time required for subsurface waste treatment
using biological agents.
In addition to in situ treatment with biological agents, a water-emulsifier
mixture can be pumped into the waste deposit and the waste-bearing mixture
pumped to the surface and treated in a biological reactor {Jhaveri and
Mazzacca, 1983; Kopecky, 1983; Switzenbaum and Jewell, 1980; Zitrides, 1978).
This procedure allows for more accurate temperature and environmental control
than conventional in situ treatment. Alternatively a trench or pond may be
used as a biotreater (Zitrides, 1978) depending on environmental conditions,
economics and geological considerations. These biotreaters may be used as
either suspended microbial reactors or attached film expanded-bed reactors
(Doggett, 1983; Jhaveri and Mazzacca, 1983; Kopecky, 1983; Switzenbaum and
Jewell, 1980; Zitrides, 1978). The attached film process has been shown to
have twice the efficiency of the suspended population system under aerobic
conditions (Switzenbaum and Jewell, 1980).
100
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Effluent from biological reactors can be polished by passage through carbon
filters or adsorptive resins (such as XAD resins) if further treatment is
required (Doggett, 1983). Due to the capital investment for equipment, this
procedure is most economical for long term treatment of heavily contaminated
areas. However, smaller portable biological reactors are also available for
short term treatment of contaminated sites (Zitrides, 1978). Examples of
application of biological treatment to subsurface wastes are described below.
2.3.3.1 Renovation of Groundwater and Deep Soils Contaminated by
Gasoline
A bench scale study on the removal of leaded gasoline from subsurface soil
strata and groundwater recommended a combined biological/physical treatment as
the optimum approach (API, 1982). The contaminated area would be injected
with nutrients and a hydrogen peroxide solution at levels above and below the
water table in order to continuously bathe the gasoline contaminated region
with oxygenated, nutrient-filled water. Microbial action would degrade and
emulsify the gasoline, aiding in its mobilization. Emulsified gasoline
byproducts could then be pumped out and renovated at the surface by physical
means such as activated carbon filters.
2.3.3.2 Bioreclamation of a Subsurface Organic Solvent Spill
An underground storage tank at a generic pharmaceutical company (Biocraft,
Waldwick, NJ) leaked a mixture of methylene chloride, acetone, n-butyl alcohol
and dimethyl-aniline into subsurface soils and groundwater, with surface
intrusion to nearby storm sewers and contamination of a local brook (Jhaveri
and Mazzacca, 1983) (Figure 2-1). The total volume of leakage was not
accurately known but was estimated at 113,000 liters (30,000 gallons).
A biological reactor system consisting of a downgradient dewatering trench and
dewatering well system, two mobile biological activating tanks and two mobile
settling tanks, and two upgradient reinjection trenches was installed.
Contaminated groundwater was pumped into the bioreactors where biodegradation
rates were significantly increased by supplying air and nutrients. Sludge was
settled from the treated water in the settling tanks and reintroduced to the
101
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FIGURE 2-1
BIOCRAFT SITE PLAN
MAIN BUILDING
STORM SEWER.
LEAK
'A! SUBSURFACE TANK FARM
DEEP WELL
PLUME (> 10 mg/l COD)
(SOURCE: JHAVERI AND MAZZACCA, 1983)
102
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activating tanks. Renovated waters were discharged to the reinjection
trenches. Figure 2-2 illustrates the basic process flow diagram of this
system. Groundwaters were treated at the rate of 52,000 to 76,000 liters
(14,000 to 20,000 gallons) per day with a median contaminant reduction of 60
percent per pass. The site operators estimate that about 40 percent of the
biodegradation of wastes occurs in the deposit itself as a result of
reinjection of bioactive (microbes and nutrient-supplemented) water. This
treatment process is described in greater detail in Section 6.5 of this report.
An essentially similar system (with the addition of an initial air stripping
step) for remediation of soil and groundwater contamination by dichlorobenzene
and methylene chloride is described by Quince and Gardner (1982).
2.3.3.3 Leaking Underground Gasoline Storage Tank
Ten domestic water supply wells in Montgomery County, PA were contaminated by
low level, long-term loss of an undetermined amount of gasoline from a below
ground storage tank at a nearby service station (Yaniga, 1982). Soil and
groundwater in the area were contaminated, but no free product was found.
Monitoring wells and domestic well samples showed that a plume extended
several hundred feet from the site, with dissolved hydrocarbon concentrations
of up to 15 ppm.
The initial bioreclamation system consisted of a central pumping well to
capture the contaminant plume and an injection gallery located at the original
spill source (the tank pit). Recovered groundwater was passed through an air
stripping tower to remove volatile organics and oxygenate the water.
Nutrients were added batchwise to the treated groundwater and injected through
the gallery. Additional oxygen was added to the site through six (6) air
sparger wells located on the periphery of the plume. In the first 20 months
of operation, maximum hydrocarbon concentrations in groundwater samples were
reduced to 2.5 ppm (Aquifer Remediation Systems, 1985).
"Enhanced Bioreclamation" was used for the second phase of remediation. This
consisted of the addition of nutrients (Restore 352 Microbial Nutrient:
103
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10GPM
PUMP
15 MIL
PLASTIC-*
SHEET
WASHED
STONE
FIGURE 2-2
BIOCRAFT BIODEGRADATION TREATMENT SYSTEM
BASIC PROCESS FLOW DIAGRAM
GRAVEL
SLOTTED
PIPE
COLLECTION
TRENCH
-300'
PROCESS
INFLUENT
BIOACTIVE
WATER
r—AIR (4 PSIG)
I I I I I I I I I
MANHOLE
EXCAVATED
SOIL
COVER
I I I I I
I I I I I
I I I I
I I I I
till
I I I I
I I I I I I I I I
I I I I I I I I I
I I I I I I I I I
I I I I I I I | i
I I I I I I I I I
. , . i . 44 4 4
9 EQUALLY
SPACED AERATION
WELLS
G
R
O
U
N F
D L
O
ww
A
T
E
R
AERATION SETTLING
TANKS TANKS
RECHARGED
PIPE
WASHED
STONE
15 MIL
PLASTIC
SHEETING 10'
'SAND
BASE
RECHARGE
TRENCH
-------
ammonium chloride, sodium phosphate and trace elements) and an oxygen source
(hydrogen peroxide) through four injection wells (Aquifer Remediation Systems,
1985).
In response to nutrient addition, there was a ten fold increase in total
bacteria and a 200 fold increase in hydrocarbon degraders. Over a period of
2-1/2 months, the hydrocarbon concentration in groundwater was reduced to
about 250 ppb. Activated carbon adsorption was used in the final phase to
"polish" the groundwater to acceptable residual concentrations (Aquifer
Remediation Systems, 1985).
2.3.3.4 Leaking Underground Storage Tanks
Vapors discovered in a laboratory building at a midwestern industrial facility
were traced to leaking tanks in a below-ground tank vault for storing fuels
and solvents. Free product was found to be confined to the vault area.
Groundwater contamination was confined primarily to the vault, with some
dissolved hydrocarbons being detected in the clay strata immediately adjacent
to the tanks. Soils throughout the vault were saturated with aromatic and
aliphatic hydrocarbons. Total contamination was calculated to be about 2500
liters of free product, and 1100-3400 liters of hydrocarbons adsorbed to soils
(Raymond et al., 1976).
Following free product removal, bioreclamation was used to treat the
contaminated soil and groundwater. Laboratory investigations verified that
the site contained acclimated native bacteria capable of gasoline degradation.
Thirty percent solutions of ammonium sulfate, disodium phosphate and monosodium
phosphate were introduced using injection wells to provide nutrients. An
average of ten aeration systems pumping at 28.3 1/min (2.5 cfm) were employed
to provide oxygen. Over the next twelve months, eighty-seven tons of
inorganic nutrients were introduced into the area (Raymond et al., 1976).
The introduction of nutrients lead to an average one hundred fold increase in
the number of gasoline-utilizing bacteria in wells within the spill area.
When nutrient, addition was stopped after about one year, the water at the
105
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producing wells contained between 0-2.5 ppm of gasoline. Within six months
this level dropped further, to a nondetectable level (Raymond et al., 1976).
During the project, thirty two bacterial cultures capable of degrading
gasoline were isolated. Most of these isolates were unable to degrade many of
the individual components of gasoline. This suggests that significant
cometabolism occurs in the subsurface environment. The cultures were
identified as primarily Norcardia, Pseudomonas and Acinetobacter (Raymond et
al., 1976).
2.3.3.5 Bioremediation of Groundwater Contaminated with Fuel Oil
and Solvents
Groundwater contamination by fuel oil, benzene, xylene, toluene;, naphthalene
and styrene was discovered on an industrial site near Frankenthal, West
Germany. After recovery of free fuel oil, it was estimated 20 to 30 metric
tons of adsorbed and dissolved hydrocarbons remained in the ground (Stief,
1984). Combined hydraulic flushing and induced biodegradation was used to
treat this residual contamination.
The local water authority required that nutrients injected into the aquifer to
accelerate biodegradation and the flushed contaminants be kept within a defined
area so that the surrounding aquifer was not contaminated. Two separate
recirculation lines were installed, one for the flushing water (5 I/sec) and
the second for clean injection water (20 to 30 I/sec). The recirculated
flushing water, contaminated with hydrocarbons and biodegradation by-products,
was stripped and filtered before re-infiltration. Biodegradation was enhanced
by controlling the dosage of the nutrient nitrate and by increasing the water
temperature 10°C (Stief, 1984).
Biodegradation of aromatic hydrocarbons was simpler than degradation of
aliphatic hydrocarbons, and benzene biodegradation was better than that of
xylene and toulene. After three months it was found that aromatics had been
degraded in the whole area, and aliphatics were reduced to about one-third of
their initial concentration (Stief, 1984).
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2.3.3.6 Biodegradation of Ethylene Glycol in Groundwater
Approximtely 15,000 liters (4000 gallons) of 25% ethylene glycol solution
leaked from a storage lagoon at the Naval Air Engineering Center, Lakehurst,
NJ, contaminating the soil around the lagoon and creating a downgradient plume
(Flathman et al., 1984; Flathman and Caplan, 1985). A feasibility study
showed that the environment was not toxic to the native microfauna, which were
already adapted to and biodegrading the ethylene glycol in situ, although pH
and nutrient adjustment would be necessary to optimize bacterial degradation
rates (Flathman et al, 1984; Flathman and Caplan, 1985).
The biotreatment system included:
1) a series of injection well points (1.5 m spacing, 5 m deep) to
inoculate the soil and groundwater with adapted indigenous microbes
and nutrients (inorganic nitrogen and phosphorus) and adjust the pH;
2) five recovery wells to withdraw contaminated groundwater from beneath
the lagoon and from the plume; and
3) a surface aerator/bioreactor (activated sludge system) to further
treat the recovered groundwater.
Surface application was also used to flush ethylene glycol from the
unsaturated zone in the soil. Biodegradation of the ethylene glycol took
place both in situ and in the reactor (Flathman et al., 1984).
During the initial treatment period (26 days), groundwater concentrations of
ethylene glycol were reduced by 85-93% (Flathman et al., 1984). the subsequent
maintenance program focused on removal of the remaining pockets of
contamination by continued surface application of lime (to raise the pH) and
nutrients which are washed into the soil by natural precipitation (Flathman et
al., 1984).
107
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Two interesting observations can be made regarding this remedial process:
1) Natural (adapted) microfauna were used to accomplish the
biodegradation; augmentation with commercial strains was not required.
2) Ethylene glycol held in the unsaturated zone by capillary action was
aggressively flushed during the treatment process and biodegratied,
along with that in the groundwater.
2.3.4 Liquid Surface Waste Deposits
Lagoons, ponds and industrial waste treatment plants are amenable to
renovation using biological treatment. In essence, the entire water body
becomes a biological reactor. Optimum concentrations of bacteria, emulsifiers
and fertilizers are introduced and maintained in these systems until
renovation is complete. These surface waters can be monitored daily for
biological oxygen demand (BOD) reductions, which indicates the progress in
degradation of the catabolizable organics present. Renovation times for ponds
or lagoons containing high BOD levels range from 3 to 12 months depending on
the level of contamination (Zitrides, 1978). Unlike ponds, industrial
treatment plants are usually designed as flow-through chemostats in which high
concentrations of organic waste mixtures can be treated on a continual basis.
Treatment of waste streams in excess of 500,000 mg/1 BOD have been reported
(Zitrides, 1978).
A lagoon containing 500,000 ppm waste oil and grease (floating, dispersed and
3
deposited as sludge) in 15,000 m (four million gallons) of liquid was
treated by biological degradation (Zitrides, 1978). This !system was
inoculated with 68 kg (150 pounds) of PETROBAC, 68 kg (150 pounds) of
PHENOBAC, micronutrients (POLYBAC N) and emulsifiers (POLYBAC E). Ongoing
treatment consisted of regular addition of bacteria, nutrients and
emulsifiers. Freezing temperatures forced the shutdown of pumps and
compressors that served to aerate the system during the winter. Treatment was
resumed in a second phase using the procedures described above.
108
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High biological activity was observed within four weeks of initial treatment.
The bacteria were able to degrade 99 percent of the waste oil within seven
months of initial startup. Lagoon wastewaters, after renovation, were
o
discharged to local sewers at the rate of 75 m (20,000 gallons) per day
with no adverse effect on the municipal trickling filter system.
2.3.5 Renovation of Waste Disposal Sites
Only one practical example was found in the literature on the use of
biological agents for in situ renovation of sites at which chemical wastes
were intentionally disposed (Flathman et al., 1983). The use of native
(Mancinelli et al., 1981) or naturally adapted (Kellogg et al., 1981)
microorganisms for waste disposal site renovation has also been suggested in
the literature, and a number of significant advances have been made in the
last five years on in situ biological treatment methods for hazardous waste.
The most promising in situ approach is the work of Kellogg et al. (1981) and
others (Rosenberg and Alexander, 1979; Schwien and Schmidt, 1982; Serdar et
al., 1982) using constructed strains of microbes. Proprietary systems
(Aquifer Remediation Systems, 1985; Doggett, 1983; Jhaveri and Mazzacca, 1983;
Kopecky, 1983; Zitrides, 1978) are assumed to be derived by similar but not
necessarily identical mechanisms of selection and adaption as those previously
described (Kellogg et al., 1981; Perry, 1979; Schwein and Schmidt, 1982;
Serdar et al., 1982).
2.3.5.1 Biodegradation of Phenolic Compounds in Contaminated Soils
An explosion and fire led to the discovery in 1977 of approximately 10,000
buried drums of hazardous wastes at Picillo Farm, Coventry, Rhode Island. The
initial remedial measure at this Superfund site was the excavation and removal
of these drums; this left approximately 1300 m3 (46,000 ft3) of
phenol-contaminated soils, containing approximately 770 kg (1700 Ibs) of
phenolic compounds (Flathman et al., 1983; Flathman and Caplan, 1985).
109
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A feasibility study indicated that the soil contained native microbes capable
of degrading the phenolics (Flathman et al., 1983; Flathman and Caplan,
1985). The biotreatment system designed for the site consisted of a 0.28
hectare (0.69 acre) secure cell (land farm) draining to a 190,000 liter
(50,000 gallons) bioreactor which collected and treated leachate for
recirculation through a perforated pipe delivery system. The area was tilled
to a depth of 46 cm (1.5 ft) to aerate the soil, and moisture content was
maintained by irrigation through the perforated pipe. A mixture of commercial
bacterial strains (160 kg or 350 pounds) was seeded into the soil to augment
the native microbial population and pH and nutrient content adjusted as
required (Flathman et al., 1983).
Dxiring the first two weeks of treatment the average total recoverable phenols
concentration dropped at a rate of 150 ppm per week (Flathman et al., 1983;
Flathman and Caplan, 1985). Over the next four weeks, however, the
concentration dropped at a rate of only 4 ppm/week as the more readily
biodegradable phenolics were destroyed, leaving more refractory phenolic
compounds. At week 6 a cosubstrate (supplementary energy source) was added to
increase microbial population and activity. This led to an increase in the
rate of phenol destruction, to about 25 ppm/week (Flathman et al., 1983;
Flathman and Caplan, 1985). By day 304 of the treatment process,,, the average
total recoverable phenol concentration was 61 ppm, more than an order of
magnitude, lower than the initial concentration and within the 100 ppm goal of
the project (Flathman et al., 1983; Flathman and Caplan, 1985). :
2.4 Summary
The use of biological agents for the treatment of hazardous organic wastes is
a relatively new concept and is creating a biological technology for the large
scale treatment of such materials (Aquifer Remediation Systems, 1985; Doggett,
1983; Flathman et al., 1983; Jhaveri and Mazzacca, 1983; Kellogg, 1981;
Kopecky, 1983; Zitrides, 1978). As in all new applications, more information
concerning the use of appropriate microorganisms and the pathways that they
use to degrade specific compounds will be needed before the full extent of
their usefulness can be known. This will require major advances in the
110
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understanding of the genetic structure of many microbes and the creation of
additional strains that can function in existing waste treatment systems or at
disposal sites.
A variety of microbiological methodologies have been developed to treat sites
contaminated by organic materials. Site characteristics dictate the
appropriate treatment technology applicable to site renovation. These
technologies have been described above, with case histories identified in
Table 2-1. These technologies are summarized in Table 2-2 and are briefly
reviewed below. Organic waste sources that can be metabolized by
microorganisms are identified in Table 2-3.
Reclamation of surface waste spills or piles may involve the use of native
bacteria if contaminant compounds are nonhalogenated acyclic or simple
unsubstituted aromatics, or low concentrations of halogenated compounds. More
complex and/or halogenated compounds may require the use of adapted, mutated,
plasmid assisted or constructed bacteria plus fertilizer and emulsifiers to
renovate surface soils. Optimum conditions are aerobic, moist environments
with a pH between 5 to 7. Anaerobic microenvironments may be required for
reductive dehalogenation. Average renovation times are one to three months.
Treatment of deep soils, subsurface waste deposits and groundwater involves
the stimulation of native microbes or the injection of adapted or genetically
constructed bacteria with fertilizers, emulsifiers and an oxygen source
directly into and around the contaminated zone. Based on the studies and case
histories identified above and the commercial systems presently in use
(Aquifer Remediation Systems, 1985; Doggett, 1983; Jhaveri and Mazzacca, 1983;
Kopecky, 1983; Zitrides, 1978), biological systems to treat various organic
contaminants present in hazardous waste deposits may now exist, and the
methodology for breeding specific cultures that can degrade persistent
compounds has been developed (Doggett, 1983; Kellogg, 1981; Kobayashi and
Rittman, 1982; Kopecky, 1983; Rosenberg and Alexander, 1979; Schwein and
Schmidt, 1981; Serdar et al., 1982; Zitrides, 1978). Applications to specific
waste sites will involve , the ability to control temperature, pH, dissolved
111
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TABLE 2-2
SUMMARY OK MICROBIOLOGICAL TREATMENT TECHNOLOGIES1
Treatment Method
ro
Adapted Plasmid
Contamination Zone Native Native Mutant Associated Constructed Bioreactors Aerobic/
Waste Disposal Sites
Surface Waste +
+
+
+
+
+? 4
+ +
'? + +/+
Piles or Deposits
Subsurface
Waste Deposits
Groundwater
Surface Liquid Waste
Deposits
+
+
+
+
+
+
Information concerning the degradation of specific waste groups or organic species and the
microorganisms degrading those materials are cited by reference in Appendix A.
"+"s Technology is available
"-": Technology is not available
"?": Technology may be developed in laboratory
-------
Respiration
TABLE 2-3
SUMMARY OF ORGANIC GROUPS SUBJECT TO MICROBIOLOGICAL METABOLISM1
MODES OF MICROIAL METABOLISM
RING Fission
SUBSTRATE
COMPOUNDS
Fermen- Oxida- Co-oxi- Oxida- Reduc- Dehalo- Esteri- Ester- Dehydro- Deamina-
Aerobic Anaerobic tation tion dation tive tive genation fication ases genation tion
Photome- Uegrad-
taboiisui atlon
Straight Chain Alkanes
Branched Alkanes
Saturated Alkyl
Halides
Unsaturated Alkyl
Halides
Esters, Glycols, Epoxides
Alcohols
Aldehydes, Ketones
Carboxylic Acids
Amides
Esters
Nitriles
Amines
Phthalate Esters
Nitrosamines
Thiols
Cyclic Alkanes
Unhalogenated Aromatics
Halogenated Aromatics
-------
TABLE 2-3 (Continued)
SUMMARY OF ORUANIti GROUPS SUBJECT TO MICROBIOLOGICAL METABOLISM1-
Respiration
MODES OF MICROIAL METABOLISM
Ring Fission
SUBSTRATE
COMPOUNDS
Feroen- Oxida- Co-oxt- Oxida- Reduc- Dehalo- Euteri- Eater- Dehydro- Deamina-
Aeroble Anaerobic tation dation dation tive tive genatiou fication uses genatiou tlon
Photone- Degrad-
tabollen ation
Simple Aromatic
Nltro Compounds
Aromatic Nitro Compounds
With Other Functional
Groups
Phenols
Halogenated Side Chain
Aromatics
Fused Ring Hydroxy
Compounds
Nitrophenols
Halogenated Phenols
Phenols - Dihydrides,
Polyhydrides
Two & Three Ring Fused
Polycyclic Hydrocarbons
Biphenyls
Chlorinated Biphenyls
Polychlorinated Biphenyls
Four Ring Fused
Polycyclic Hydrocarbons
Five Ring Fused
Polyeyclic Hydrocarbons
-------
TABLE 2-3 (Continued)
SUMMARY OF ORGANIC GROUPS SUBJECT TO MICROBIOLOGICAL METABOLISM1
SUBSTRATE
COMPOUNDS
MODES OF MICROIAL METABOLISM
Respiration Ring Fission
Fermen- Oxida- Co-oxi- Oxida- Reduc- Dehalo- Esteri- Ester- Dehydro- Deamina-
datlon tive tive genation fication ases genation tion
Aerobic Anaerobic tatlon tion
Photome- Degrad-
tabolism ation
Fused Polycyclic
Hydrocarbons
Organophosphates
Pesticides and
Herbicides
I This table is a condensed version of Appendix A. Please refer to the Appendix for specific organics and the biological agents participating in
the metabolism of these compounds.
-------
oxygen, moisture, nutrients, solubility of waste materials, and microbial
predation. Current technology may be applied to the in situ treatment of
wastes in a manner similar to that of Jhaveri and Mazzacca (1983) and Flathman
et al. (1983). Additional advances may include the breeding of microorganisms
with laboratory evolved plasmids capable of degrading a variety of xenobiotic
compounds (Kellogg et al., 1981). Mobilized waste could be pumped to the
surface from perimeter wells (API, 1982) for treatment in bioreactors (Jhaveri
and Mazzacca, 1983) prior to final renovation by activated carbon or ionic
filters (API, 1982; Aquifer Remediation Systems, 1985: Doggett, 1983). The
range of delivery/recovery systems applicable to various waste deposit
settings is discussed in Section 1.
Process applications may require several years. In cases where biological
treatment cannot produce complete treatment, its use in conjunction with
chemical and physical treatments may be preferable to using any one technology
alone.
Renovating liquid waste deposit sites primarily involves the use of adapted,
mutant, plasmid assisted or genetically constructed bacteria in conjunction
with fertilizer application and aeration. Systems may be microaerophillic to
anaerobic upon diffusion into soils. Average renovation times are three
months to a year. •
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References*
1.
2.
3.
4.
5.
6.
7.
8.
9.
10.
11.
12,
13.
Aftring, R. P., B. E. Chalker and B. F. Taylor. 1981. Degradation of
Phthalic Acids by Denitrifying, Mixed Cultures of Bacteria. Appl.
Environ. Microbiol. 4.1:1177-1183.
Alexander, M., 1981. Biodegradation of Chemicals of Environmental
Concern. Science. 211:132-138.
American Petroleum Institute, Committee on Environmental Affairs. 1982.
Enhancing the Microbial Degradation of Underground Gasoline by Increasing
Available Oxygen, Final Report. Submitted to A.P.I, by Texas Research
Institute, Inc., 5902 W. Bee Caves Road, Austin, Texas.
Aquifer Remediation Systems, 1985. Press Release Kit. FMC Corp.,
Princeton, NJ.
Aranha, H. G. and L. R. Brown. 1981. Effects of Nitrogen Sources on End
Products of Naphthalene Degradation. Appl. Environ. Microbiol. 42:74-78.
Atlas, R. M. 1981. Microbial Degradation of Petroleum Hydrocarbons: An
Environmental Perspective. Microbiol. Rev. 45:180-209.
Atlas, R. M. and R. Bartha. 1973. Inhibition by Fatty Acids of the
Biodegradations of Petroleum. Antonie van Lieuwenhoek J. Microbiol.
Serol .39_: 257-271
Benckiser, G. and J. C. G. Ottow. 1982. Metabolism of the Plasticizer
di-n—Butylphthalate by Pseudomonas pseudoalcaligene s Under Anaero- bic
Conditions, With Nitrate as the Only Electron Acceptor. Appl. Environ.
Microbiol. 44:576-578.
Benezet, H. J. and C. 0. Knowles.
Algae. Chemosphere. 10:909-917.
1981. Degradation of Chlordimeform by
Bitton, G. and C. P. Gerba. 1984. Groundwater Pollution Microbiology. J
Wiley and Sons, New York. 377 pp.
Boethling, R. S. and M. Alexander. 1979a. Microbial Degradation of
Organic Compounds at Trace Levels. Environ. Sci. Technol. 13:989-991.
Boethling, R. S. and M. Alexander. 1979b. Effect of Concentration of
Organic Chemicals on Their Biodegradation of Natural Microbial
Communities. Appl. Environ. Microbiol. 37:1211-1216.
Bouwer, E. J. and P. L. McCarty. 1983a. Transformations of 1- and 2-
Carbon Halogenated Aliphatic Organic Compounds Under Methanogenic
Conditions. Appl. Environ. Microbial. 45:1286-1294.
* References are numbered to correspond with references in Appendix A.
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SECTION 3
SURFACTANT - ASSISTED FLUSHING
3.1 Introduction
Flushing or mobilization of wastes can serve two purposes: to promote the
recovery of wastes from the subsurface for treatment at the surface, or to
solubilize adsorbed compounds in order to enhance the rate of other in situ
treatment techniques (such as biodegradation or hydrolysis). Flushing or
mobilization using water alone may be sufficient for relatively soluble
compounds such as phenols; however, the use of chemicals such as surfactants
will be required for significant solubilization of insoluble (hydrophobic)
compounds. In addition, acid solutions can be used to mobilize certain
organics (amines, ethers, anilines) and basic solutions can mobilize some
phenols, chelating and complexing agents (USEPA, 1982).
Surfactants (surface active agents) are a class of natural and synthetic
chemicals whose abilities to promote the wetting, solubilization, and
eraulsification of various types of organic chemicals have found widespread
application. These properties make surfactants of possible use in the in situ
treatment of certain organic fractions in waste deposits. Used in conjunction
with various groundwater flooding and dewatering techniques, surfactants may
offer a means of improving the removal efficiency of these organics over the
results likely to be obtained with water alone.
An evaluation has been made of the feasibility of using surfactants for in
situ waste treatment processes. Since very little information exists on the
use of surfactants at waste sites, this evaluation has focused on a review of
the available literature on the application of surfactants to subsoil systems
and a consideration of fundamental chemical characteristics of the principal
surfactant classes with respect to their applicability to in situ organic
waste treatment.
126
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3.2 Background and Theory
Surfactants are a general class of chemicals whose amphipathetic molecular
•structures generally consist of a hydrophobic group which has little affinity
for the solvent phase (water) and a hydrophilic group which is readily soluble
in the solvent phase (Shaw, 1976). The terms lyophobic and lyophilic are
applied to systems where the carrier solvent is not water. This
characteristic of surfactants results in their tendency to concentrate
preferentially at phase interfaces (liquid-liquid, liquid-solid, liquid-gas)
and is responsible for their unique abilities to alter certain properties of
aqueous solutions. Surfactants might be used to enhance the effectiveness of
in situ treatment technologies by improving both the detergency of aqueous
solutions applied to waste deposits and the efficiency with which organics may
be transported by aqueous solutions from the subsurface waste deposit to the
surface.
Surfactants can increase the detergency ("cleaning power") of aqueous
solutions through a number of processes. These include the following:
o Preferential Wetting - Surfactants can improve the ability of an
aqueous solution to wet a solid surface (such as soil particles) by
decreasing the interfacial tension between the aqueous phase and the
solid phase (Rosen, 1978). By preferentially wetting the solid
surface, an aqueous solution can partially or completely displace an
adsorbed organic fraction. This reduction in the "strength" with
which an organic fraction adheres to soil particles may enhance the
effectiveness of contaminant recovery during groundwater pumping and
dewatering operations.
o Solubilization - The addition of surfactants can enhance the ability
of aqueous solutions to solubilize organic compounds.
Solubilization results from the interaction of the amphipathetic
surfactant molecules with molecules of the organic fraction. In
practice, significant Solubilization of organic material generally
127
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requires relatively high surfactant concentrations (above the
"critical micelle concentration") which lead to the formation of
surfactant micelles in the solution (Mukerjee, 1979). Micelles are
discrete clusters of surfactant molecules within an aqueous phase in
which the surfactant hydrophobic groups are directed toward the
interior of the micelle and hydrophilic groups toward the surrounding
solvent (water). Micelles can effectively incorporate or
"solubilize" susceptible organic compounds either within their
interiors (hydrophobic regions) or at their external peripheries
(hydrophilic regions).
o Emulsification - Surfactants can enhance the detergency of an aqueous
solution by promoting the dispersion of an insoluble organic phase
within the aqueous phase (emulsification). Emulsion formation
generally requires some minimal source of mechanical energy input.
As such, emulsification processes suffer the disadvantage of often
being readily reversible. Thus in a waste deposit spontaneous
separation of emulsified phases may occur prior to removal of the
emulsion.
3.3 Surfactant Chemical Characteristics
Surfactants are generally classified on the basis of the chemical
characteristics of the hydrophilic groups. The principal surfactant classes
(Rosen, 1978) are described below:
Anionic - The surface active portion of the surfactant molecule bears
a negative charge, for example RCJELSO-Na (a sodium
alkylbenzene sulfonate). Anionic surfactants find widespread use as
detergents and wetting agents, and are the largest surfactant class
in terms of usage and importance. Most groups of anionic surfactants
display limited to good water solubility.
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o Cationic - The surface active portion of the surfactant molecule
bears a positive charge, for example, RNHgCl (salt of a long
chain amine). This is a relatively small surfactant class, many of
whose members find somewhat specialized uses requiring surface
adsorption and surface coating.
Nonionic - The surface active portion of the surfactant molecule
does not bear apparent ionic charge, for examplej RCOOCH^CHOHCELOH
(monoglyceride of a long chain fatty acid). Nonionic surfactants are
the second most important class in terms of use. They do not display
charge effects, are generally soluble in water, and many nonionics
are soluble in organic solvents.
Amphoteric - Both positive and negative charges may be present
in surface active portion , of the molecule, for example,
R NH2CH2COO~ (a long chain amino acid). This is a small
surfactant class used in situations where specialized charge
properties are required.
The hydrophobic portion of surfactant molecules is typically comprised of a
long chain hydrocarbon residue. Common surfactant hydrophobic groups include:
o Branched chain, long alkyl groups (Cg - C20),
o Long-chain (Cg ~ C-ic) alkybenzene residues,
o Alkylnaphthalene residues (C^ and greater-length alkyl groups),
o Rosin derivatives,
o High-molecular weight ethylene oxide and propylene oxide polymers,
o Long chain perfluoroalkyl groups, and,
o Polysiloxane groups.
Table 3-1 provides general information on the four major surfactant types and
their principal classes. - , •
129
-------
TABLE 3-1
SURFACTAHT CHARACTERISTICS
CO
O
SURFACTANT TYPE
AND CLASSES
ANIONIC
1) Carboxllc Acid Salts
2) Sulfuric Acid Ester Salts
3) Phosphoric & Polyphosphoric
Acid Esters
4) Perfluorinated Anionlcs
5) Sulfonic Acid Salts.
SELECTED PROPERTIES
AND
Good Detergency
(1. 2, 3, 4, 5)
Good Wetting Agents
(1, 2, 3, 4, 5)
Strong Surface Ten-
sion Reducers (4,5)
Good Oil In Water
Emulsifiers - (5)
SOLUBILITY
Generally Water Soluble
(1, 2, 3, 4, 5)
Soluble in Polar Organlcs
(5)
REACTIVITY
Electrolyte Tolerant
(2, 3, 4, 5)
Electrolyte Sensitive
(1)
Resistant to Biodegradation
(4,5)
High Chemical Stability (4)
Resistant to Acid and Alkaline
Hydrolysis (3, 4, 5)
CATIONIC
1) Long Chain Amines
2) Dianines & Polyamines
3) Quaternary Ammonium Salts
Emulsifying Agents
(1, 3, 4)
Corrosion Inhibitor
(1)
Low or Varying Water Solubility Acid Stable (1, 3)
(1, 2, 4)
Water Soluble (3)
Surface Adsorption to Silicaeous
Materials (2)
4) toiyoxyechyienated Long Chain
Amines
-------
TABLE J»l (Cont'd)
SURFACTANT CHARACTERISTICS
SURFACTANT TYPE
AMD CLASSES
NONIONIC .- . . '
1) Polyoxyethylenated Alkyl phenols
Alkylphenol Ethoxylates
2) Polyoxyethylenated Straight
Chain Alcohols & Alcohol
Ethoxylates
3) Polyoxyethylenated Poly-
oxypropylene Glycols
4) Polyoxyethylenated Mercaptans
5) Long-Chain Carboxylic Acid
: Esters >• .
SELECTED PROPERTIES
AND USES(1)
Emulsifying Agents
(1, 5)
Detergents
(1, 2, 4, 6)
Wetting Agents
U, 7)
Dispersents (3)
Foam Control (3)
SOLUBILITY
REACTIVITY
Generally Water Soluble
Good Chemical Stability
(1, 6)
Water Insoluble Formulations Resistant to Biodegradation
(1, 6, 7) (1)
Relatively Non-Toxic (all)
Subject to Acid and Alkaline
Hydrolysis (.i, b, 7)
6) Alkylolamine "Condensates",
Alkanolamides
7) Tertiary Acetylenic Glycols
AMPHOTERICS . ,
1) pH Sensitive
2) pH Insensitive
Solublizing Agents
Wetting Agents
Varied (pH dependent)
Non-Toxic
Electrolyte Tolerant
Adsorption to Negatively Charged
Surtaces
11) Numbers refer to applicable classes within a given surfactant' type
-------
One of the most important properties of surfactant solutions with respect to
waste treatment lies in their ability to reduce organic/water interfacial
tensions (thereby potentially enhancing wetting, emulsification and transport
of organics). Unfortunately, interfacial tension data have been determined
for relatively few of the specific aqueous/organic systems of potential
interest with respect to waste treatment.
A simple approach to evaluating the potential use of surfactantus in organic
waste recovery involves consideration of the aqueous solubility of the organic
phase. The aqueous solubilities of selected organic compounds of interest are
listed in Table 3-2. In addition, selected octanol/water partition
coefficient values (K ), which are commonly used as relative measures of
ow
the tendency of organic compounds to adsorb to soil particles (Wasik et al.,
1981; Karickhoff et al., 1979) are also listed. These data demonstrate the
wide ranges in water solubilities (and soil adsorptivity) possessed by
potential organic constituents of waste deposits. In general, surfactants
would be most effective in promoting the mobilization of organic compounds of
relatively low water solubility and high log KQW values. Conversely,
surfactants may be of more limited value for mobilizing relatively water
soluble substances.
General prerequisites for effective lowering of the interfacial tension
between aqueous and organic phases by a surfactant include: :
o Spontaneous surfactant adsorption at the interface, >
o Molecular interaction between the hydrophobic portion of the
surfactant and the organic phase, and
I
o Strong molecular interaction between the hydrophilic portion of the
surfactant molecule and the aqueous phase.
132
-------
TABLE 3 2
PROPERTIES OF SELECTED ORGANIC COMPOUNDS WHICH INDICATE THE POTENTIAL EFFECTIVENESS OF SURFACTANTS1
CO
CO
CONTAMINANT CLASS
Normal Hydrocarbons
Unsaturated
Hydrocarbons
Halogenated
Hydrocarbons
Ethers
Aldehydes and
Ketones
COMPOUND
n-Pentane
n-Hexane
n-Heptane
n-Octane
1-Hesene
1-Heptene
1-Octene
1-Nonene
1-Pentyne
1-Chlorobutane
1-Chloroheptane
Trlchloroethylene
Trichloroethylene
Carbon Tetrachloride
Dllsopropyl Ether
Ethyl Ether
2-Butanone
3-Pentanone
2-Heptanone
2-Octanone
Heptaldehyde
Methyl n-Butyl Ketone
Methyl Ethyl Ketone
WATER SOLUBILITY (M)
5.65 x 10~4
1.43 x l(f4
3.57 x llf4
9.66 x 10~4
8.38 x 10~4
1.85 x 10~4
3.65 x 10~5
8.85 x 10~6
1.54 x 10"2
9.43 x 10~3
1.01 x 10~4
1.04 x 10~2
3.2 x 10~2*
5.2 x 10~4*
8.8 x 10~2*
9.3 x 10"1*
1.89
0.53
3.57 x 10~2
8.85 x 10"3
-
1.7*
4.9*
OCTANOL WATER PARTITION INTERFACIAL TENSION (Y)2
COEFFICIENT (log Kow) (dynes/cm)
3.
4.
4.
5.
3.
3.
4.
5.
2.
2.
4.
2.
-
0.
0.
1.
2.
-
-
-
62
11 50.0
66
18 50.8
47
99
88
35
12
55
15
53
37.4
45.0
17.9
10.7
69
99
98
76 -
13.7
9.7
3.0
-------
TABLE 3-2 (Cont'd)
CONTAMINANT CLASS
COMPOUND
WATER SOLUBILITY (H)
OCTANOL WATER PARTITION
COEFFICIENT (log K)
JNTERFAC1AL TENSION (Y)
(dynee/cn)
OJ
Esters
Alcohols , , : , •; •
Carboxylic Acids
Aromatic
Hydrocarbons
Methyl-nonanoate
Methyl Decanoate
Ethyl Acetate
n-Butyl Acetate
, 1-Butanol
1-Pentanol
1-Hexanol
1-Heptanol
Octanol
Heptylic Acid
Toluene
Ethyl benzene
o-Xylene
n-Butyl benzene
Benzene
m-Cresol
1.33 x 10
-5
2.05 x 10
0.726
5.77 x 10~2
0.854
0.133
4.14 x 10~2
1.13 x 10~2
2.0 x 10~3*
1.8 x 10~2*
6.28 x 10~3
1.76 x 10~3
2.08 x 10~3
1.03 x 10~4
2.2 x 10~2
2.59 x 10"2*
4.32
4.41
0.68
1.82
0.78
1.53
2.03
2.57
2.65
3.13
3.13
4.28
2.13
1.96
6.8
14.5
8.5
7.7
36.1
33.9
Phenols
Halogenated •
Aromatics
Phenol
2.6-Dimethylphenol
Halogenated Phenols m-Cresol
Bromobenzene
o-Dlchlorobenzene
0.81
7.90 x 10
,-2
2.59 x 10
-3
2.62 x 10~3
6.7 x 10"4*
1.45
2.31
1.96
2.98
3.38
36.5
-------
TABLE 3-2 (Cont'd)
to
CJi
CONTAMINANT CLASS COMPOUND
Fused Polycyclic
Hydrocarbons
Aalnes
Napthalene
1-Methylnaphthalene
Aniline
WATER SOLUBILITY (M)
2.39 x 10
2.23 x 10'
-*
OCTANOL WATER PARTITION
COEFFICIENT (log K)
3.35
3.6 x 10
rz«
1 Source - experimental data presented by Wasik et al. (1981) and Grain
2 In water at 20-25°C
* Values approxinated from data of Verscheueren (1977) IT-20°C)
1NTEKFAC1AL TENSION (Y)
(dynes/en)
2
5.8
-------
Little comprehensive, quantitative data is currently available on the relative
effectiveness and efficiency with which different surfactants reduce the
interfacial tensions of different aqueous-organic systems. In general, within
a given surfactant class the efficiency with which a surfactant reduces the
interfacial tension of a water/organic system has been reported (Rosen, 1978)
to:
o increase with increasing carbon number for straight chain surfactant
hydrophobic groups
o increase with the addition of phenyl groups, and
o increase less with branched chains than with straight chains of the
same carbon number.
The efficiency of polyoxyethylenated nonionic surfactants generally decreases
slowly with increasing oxyethylene content of the surfactant molecule.
3.4 Surfactant Application to Subsurface Deposits: Existing :
Information on Surfactant Behavior
A review of the literature did not reveal much information on the use of
svirfactants for the in situ treatment of organic waste deposits. The most
comprehensive evaluation of the potential use of surfactants for the
subsurface recovery of organic compounds has been in conjunction with tertiary
oil recovery technologies. In addition, several laboratory studies have
evaluated the feasibility of enhancing the recovery of spilled petroleum
products in groundwater systems by using surfactants.
3.4.1 Tertiary Oil Recovery '
The feasibility of utilizing surfactants to enhance the relatively poor
recovery efficiencies obtained in tertiary oil recovery by water flooding have
been studied extensively (Shah, 1977; Morgan et al., 1979). Research to date
136
-------
has focused on chemical characterization of those surfactants capable of
generating the "ultra low" interfacial tensions (less than 0.1 dynes/cm) which
calculations indicate are required for significant increases in oil recovery
efficiencies under the pressurized flooding conditions attainable in well
fields.
A variety of studies (Doe et al., 1977; Cayias et al., 1977; Wilson and
Brandner, 1977) have demonstrated that certain sulfonates and petroleum
sulfonate mixtures are particularly effective in reducing the interfacial
-2 -4
tensions of aqueous/oil systems to very low values (10 to 10
dynes/cm). These studies and others (Cash et al., 1977; Barakat et al., 1983;
Morgan et al. , 1979) have shown that the extent of interfacial tension
reduction in these sulfonate systems (and by analogy, possibly other
surfactant classes) is affected by a variety of physical/chemical factors
including the composition of the oil phase, the structure and concentration of
the surfactant, the solution electrolyte concentration, temperature, pH and
the molecular weight, structure and concentration of any surfactant
solubilizing additives (organic alcohols). In general it has been found that
significant interfacial tension reduction is observed at only a specific
surfactant concentration (or within a very narrow range of concentrations).
Data indicate that maximum interfacial tension reduction is observed only when
the surfactant chemical characteristics (equivalent weight and structure) are
closely correlated to those of the oil phase, and only for surfactant
concentrations at or in excess of the critical micelle concentrations.
Decreases in surfactant concentrations below critical micelle concentration
values lead to abrupt increases in interfacial tension. In addition,
interfacial tension has been shown to be highly sensitive to electrolyte
concentrations, with both insufficient and excessive electrolyte
concentrations decreasing surfactant effectiveness.
137
-------
3.4.2 Petroleum Spills
The feasibility of using surfactants to recover spilled petroleum products has
been studied by the Texas Research Institute (1979). In these studies the
ability of a series of commercial surfactants to enhance the displacement and
recovery of gasoline was evaluated in laboratory simulations of subsurface
spills. Significant reductions in interfacial tension at the gasoline/water
interface were considered to be a prerequisite of potential surfactant
effectiveness as an agent to displace gasoline. Selected results of these
studies are summarized in Table 3-3. The results indicate that the magnitude
of the gasoline/water interfacial tension reduction was greater for anionic
and nonionic surfactants than for fluorocarbons. However, within each class
significant variations in interfacial tension were observed depending upon the
specific surfactant employed.
Several surfactants which demonstrated significant interfacial tension
reductions were tested for their ability to enhance gasoline displacement from
laboratory sand systems after initial water flushing. Only one of the
surfactants tested (Richonate YLA, an alkylaryl sulfonate) measurably
increased gasoline recovery. Significantly, this was not the surfactant which
had displayed the greatest reduction in interfacial tension. In addition,
poor recovery flow rates were observed for this surfactant. The experiments
suggested that this was caused by the formation of a viscous emulsion of
surfactant solution and gasoline. Recoveries were improved (up to 40%) by the
use of a mixture of anionic Richonate YLA and nonionic Hyonic PE-90.
Subsequent studies (Texas Research Institute, 1982) in large scale model
aquifer systems have confirmed that surfactant solutions can enhance gasoline
recovery, but that recovery efficiencies are influenced by the method of
surfactant application. '
Based on the results of the Texas Research Institute (1979, 1982) studies, a
laboratory study of the;solubilization of various common contaminants by water
washes and by a surfactant mixture was conducted by Ellis et al. (1984). The
contaminants tested included:
138
-------
TABLE 3-3
SUMMARY OF EXPERIMENTS ON SURFACTANT-ENHANCED GASOLINE RECOVERY
(1)
SURFACTANTS^2)
STRUCTURE
WATER
SOLUBILITY
INTERFAC1AL
TENSION (Y)
(dynes/cm)
ENHANCED
GASOLINE RECOVERY
Deionlzed Water
11.5
NONIONICS
Hyonic PE-90 (DS)
Hyonic PE-190 (DS)
£ Hyonic PE-120 (DS)
Poly-Tergent'B-500 (0)
Alrosol (0)
ANIONIC
Polethoxylate nonylphenol
Polyethoxylate nonyl phenol
Polyethoxylate nonyl phenol
Polyethoxylate nonyl phenol
Oleic fatty acid amide
Clear Solution
Clear Solution
Clear Solution
Soluble
Dispersible
0.12
0.72
1.2
1.3
1.9
None Observed
(3)
None Observed
None Observed
Dupanol G (D)
Aerosol-OT (C)
Alfonic 1412-S (CO)
Richonate YLA (R)
C-550 Slurry (CO)
Aerosol-MA (C)
Sarkosyl-NL (CG)
Fatty alcohol amlne sulfate
Na dioctyl sulfosuccinate
Linear
H
Dodecyl benzene sulfonate,
isopropylamine salt
Linear alkyl benzene
sulfonate, Na salt
Na dihexyl sulfosuccinate
Lauroyl sarcosinate,
Na salt
50%
15 g/1
Soluble
Soluble H.,0/ETHOH
Soluble alcohol
343 g/1 25°C
Soluble as Na salt
1.0
1.2
2.4
O.t.1
1.2
7.1
1.8
None Observed
Moderate
-------
SABLE 3-3 (Cont'd)
SURFACTANTS^2)
STRUCTURE
WATER
SOLUBILITY
INTERtACIAL
TENSION (Y)
ENHANCED
liaSOLlNK KECOVERY
(dynes/cm)
FLUOROCARBON
Zonyl FSN 0.05% (D)
Zonyl FSN 0.05% (D)
Lodyne S-102 0.1%(CG)
Lodyne S-102 0.05%(CG)
Lodyne S-lll 0.05%(CG)
Fluorocarbon
Fluorocarbon
Sodium fluorinated alkyl
sulfonate
Sodium fluorinated alkyl
sulfonate
Sodium fluorinated alkyl
sulfonate
Greater than 2%
Greater than 2%
Soluble
Soluble
Soluble
3.9
11.0
2.3
3.2
6.9
Notes
(1) Adapted from American Petroleum Institute, (1979). Surfactant concentration 0.1% unless otherwise noted.
(2) Letters in parenthesis refer to manufacturers:
DS - Diamond Shamrock
0 - Olin
D - DuPost .-
C - Cyanamid
CO - Conoco
CG - CIBA - GEIGY
R - Richardson
(3)
Dash (-) indicates not tested.
-------
o intermediate and high molecular weight aliphatics and polynuclear
aromatics (PAH) derived from crude oil,
o PCBs in chlorobenzenes (Askarel), and
o Di-j tri-, and pentachlorophenols.
The soil used was a fine-to-coarse loam (gravelly silty sand) with a
O /
permeability of 10 to 10 cm/sec (28 to 0.28 ft/day) but low organic
carbon content (0.1%). A series of shaker table extractions and 1 meter (3
foot) long soil column extractions were performed. The former gave the
maximum extraction efficiency under soil washing conditions with agitation,
while the latter tests showed the potential extraction efficiencies under
gravity flow without agitation. Initially, a mixture of 2% Richonate YLA and
2% Hyonic PE-90 was tested. However, this mixture tended to suspend
(disaggregate) silt and clay grains, which clogged the soil columns. A
mixture of 2% Hyonic PE-90 with 2% Adsee 799 (both nonionic surfactants) was
subsequently used.
Table 3—4 gives the results of these studies. Water washes were ineffective
in solublizing either the aliphatic/PAH or the PCB mixtures. However, after
three pore volumes of the surfactant had passed through the soil columns, only
11% of the aliphatic/PAH mixture and 14% of the PCB mixture remained in the
soils. After ten pore volumes of surfactant flushing 7 and 3 percent
respectively of the aliphatic/PAH and PCB mixtures remained in the soil.
Subsequent water rinses did not reduce these residual concentrations.
In contrast, the initial water washes removed over 99% of the phenol mixture
(Table 3-4), with the surfactant washes removing much of the residual 1%.
These results demonstrate the efficiency of surfactant solubilization of
hydrophobic compounds such as aliphatics, PAH and PCB, and the fact that
surfactant—assisted flushing is not necessary for hydrophilic compounds such
as phenols.
141
-------
I
TABLE 3-4
RESULTS OF SURFACTANT-FLUSHING OF CONTAMINANTS
FROM TEST SOIL
Percent of Contaminants Remaining in Soil
Contaminant Test
Mixture
A
A
B
B
C
Type
Shaker
Column
Shaker
Column
Column
Water Washes
1234-7 8-10
96 93 91 -
b
- 100 - 100 100
100 100 100 -
b
- 100 - 100 100
b
- 3 - 1 0.8
Surfactant Washes
1234-7 8-10
42 26 27 -
b
- 11 - 9 7
35 18 12 -
b
- 14 - 3 3
b
- 0.1 - 0.1 0.1
Water Rinses
1.2 3 4-7 8-10
25 10 10 -
b
9 - 7 7
5 9 7
b
- 3-42
b
- 0.1 - 0.1 0.1
A = high MW allphatlcs and polynuclear aromatics.
B " PCBs In chlorobenzenes
C " di-, trl-, and pentachlorophenols
a. number of pore volumes of eluant (water or surfactant solution)
b. pore values 1-3 combined for column tests
-------
3.5 Surfactant Application to Subsurface Deposits: Geochemical
and Environmental Factors
The application of surfactant solutions to organic waste deposits requires
consideration of not only the chemical characteristics of the surfactant and
the waste but also of the environmental and geochemical factors which may
affect surfactant use. The latter can impose a variety of constraints on the
potential effectiveness of surfactant applications by impairing surfactant
delivery to the waste deposit, altering the chemical activity of the
surfactant or generating an environmental chemical hazard resulting from the
surfactant itself or a side reaction product. The principal geochemical
constraints may arise through interactions between the surfactant and site
soils or groundwater.
3.5.1 Groundwater Chemistry
The chemical composition of site groundwater can alter or inhibit the
effectiveness of surfactants. Inhibition can result from a variety of
reactions which either remove the surfactant from solution (precipitation) or
reduce the effectiveness of the surfactant (neutralization, complexing).
Among those groundwater chemical conditions which may influence surfactant
effectivness are ionic strength, polyvalent ion concentrations, and pH
levels. Many surfactants are optimally effective only within limited ranges
of ionic strength and electrolyte composition. In particular, many
surfactants lose their effectiveness or precipitate at high divalent ion (Ca
and Mg) concentrations (i.e., in "hard" waters - this is why water softeners
or ion exchange resins are used to pretreat such waters before addition of
detergents in industrial and household applications).
3.5.2 Soil Chemistry
Surfactant effectiveness may also be inhibited by chemical adsorption to soil
particles, thereby reducing the aqueous surfactant concentration. The extent
of adsorption of a given surfactant will be a function of several factors,
143
-------
including surfactant structure, soil composition, particle size and surface
area, and groundwater chemical composition. In general, soils of small
particle size and high surface area per unit weight (e.g., high clay content)
are likely to provide conditions under which maximum surfactant adsorption may
occur.
An example of the combined influence of soil and groundwater chemical
interactions on surfactant adsorption to soils is depicted in Figure 3-1.
Figure 3-1A shows that the adsorption of the anionic surfactant 4-phenyl
dodecyl benzene sulfonate markedly increases as the aqueous solution pH is
decreased below the point of zero charge (PZC) of the kaolinite substrate
(approximately pH 5, below which the clay surface has a net positive charge
and will adsorb the negatively charged surfactant hydrophilic groups).
Kaolinite clay particles and most other silicate mineral surfaces possess PZCs
in the acidic pH range, and are therefore negatively charged under most
natural water pH conditions (Parks, 1967), where adsorption of anionic
surfactants would not be a problem. Increasing electrolyte concentrations
also tend to increase the amount of surfactant adsorption to the kaolinite
substrate (Figure 3-1B), possibly through a neutralization of the negative
charge on the clay particle surfaces. Figure 3-1C depicts the influence of
substrate composition on surfactant adsorption. These results demonstrate
that surfactant adsorption increases with increasing solid phase surface area,
with the greatest adsorption being to kaolinite clay. Overall, these results
suggest that for this surfactant, and probably other anionic surfactants,
minimum soil adsorption losses would occur under conditioms of alkaline
solution pH, and at low electrolyte concentrations in soils of low particle
surface area.
The principal implications of the preceeding and other available information
concerning geochemical interactions likely to be observed in specific
surfactant classes may be summarized as follows:
144
-------
FIGURE 3-1
THE EFFECTS OF SOLUTION pH,
ELECTROLYTE COMPOSITION AND SOIL
COMPOSITION ON SURFACTANT
ADSORPTION TO SOIL
SALINITY - 1% NaCI
SUBSTRATE - KAOLINITE
468
SOLUTION pH
A.- SOLUTION pH EFFECTS
32 -
16 -
B. - ELECTROLYTE
3 -
II
So
-16 -
SUBSTRATE- KAOLINITE
0.5% NaCI
0.05% NaCI
0.0% NaCI
400 800 1200 1600 2000 2400
EQUILIBRIUM CONCENTRATION
( fl MOLE/I)
SALINITY- !%NaCI
KAOLINITE 15.1 M2/gm
BEREA 1.05M2/gm
SILICA 0.73 M2/gm
C.- SOIL COMPOSITION EFFECTS
500 1000 1500 2000
EQUILIBRIUM CONCENTRATION
(A< MOLE/I)
(ADAPTED FROM WADE ET AL, 1980)
145
-------
3.5.2.1 Cationic Surfactants
The surfaces of typical soil particles (such as clays) are negatively charged
under typical soil pH conditions (pH 5-8). Therefore, most if not all
cationic surfactants are likely to be readily adsorbed to soil particles under
these conditions and are not likely to be effective for application to waste
deposits. 'At low pH conditions, however, the soil particles may have a net
positive charge (Figure 3-1A), favoring the use of cationic surfactants.
3.5.2.2 Anionic Surfactants
The sensitivity of anionic surfactants to solution electrolyte concentrations
varies widely depending upon the specific surfactant. Many anionic
surfactants, including certain members of the fluorocarbon, sulfonate and
sulfosuccinate classes, may be precipitated in groundwater with hardness
levels which exceed several hundred ppm. By virtue of their negative charge,
anionic surfactants are likely to be significantly less prone to adsorption
(and consequently more mobile in groundwaters) than are cationic surfactants.
The data on 4-phenyl dodecyl benzene sulfonate discussed above indicate that
soil adsorption of anionic surfactants is likely to increase with decreasing
solution pH, increasing solution electrolyte concentration, and increasing
soil particle specific surface areas.
3.5.2.3 Nonionic Surfactants
Many nonionic surfactants may potentially adsorb to soil particles through a
combination of nonionic interactions. Polyoxyethylenated nonionics are known
to adsorb to nonpolar substrates from aqueous solutions via dispersion forces
or hydrophobic bonding off the hydrophobic surfactant groups (Rosen, 1978).
Conversely, polyoxyethylenated nonionics have also been demonstrated to adsorb
on polar solid surfaces via hydrogen bonding between ether linkages of the
polyoxyethylene chain and polar surface groups (such as hydroxyl)!. Adsorption
processes of this type may account for the apparent ineffectiveness of
146
-------
nonionic surfactants in enhancing gasoline recovery, despite the fact that
these surfactants display excellent aqueous/gasoline interfacial tension
reduction properties.
3.5.2.4 Amphoteric Surfactants
The presence of both positive and negative charge sites (the amphoteric
characteristic) on these surfactants suggests that they are more likely to
adsorb to soil particles than otherwise similarly structured anionic
surfactants under similar solution chemical conditions. Based on their
structural characteristics, amphoteric surfactants would be expected to
undergo ionic adsorption to soil particles with maximum adsorption occuring
under pH conditions wherein the surfactants display cationic charge
properties.
3.5.2.5 Anionic - Nonionic Surfactant Mixtures
Limited available data (Wade et al., 1980) indicate that mixtures of anionic
and nonionic surfactants demonstrate complex adsorption behavior with the
extent of the adsorption to a soil dependent at least in part upon the mole
ratios of the surfactants in the mixture. Of interest was the observation
that at certain mole ratios, mixtures of an anionic surfactant (3-phenyl
undecyl benzene sulfonate) and certain nonionic surfactants (ethoxylated
nonylphenols) demonstrated less adsorption to a kaolinite substrate than did
either of the surfactants tested individually. These results suggest that
anionic/nonionic surfactant mixtures might be formulated to minimize
geochemical interactions in applications to solid waste deposits.
3.6
Environmental Effects
The introduction of surfactant solutions into surface and groundwater systems
requires consideration of possible .adverse environmental effects* The
surfactant characteristics of principal environmental concern are
biodegradability, toxicity to plants and animals, and human health hazards
both during and after application.
147
-------
3.6.1 Biodegradability
The application to soils and groundwaters of any surfactant which is strongly
resistant to biodegradation may result in the generation of new environmental
chemical problems at a waste site beyond those which already exist. However,
in the case of application to waste deposits, a converse problem also exists
— surfactants which are too rapidly biodegraded may not retain surface
activity for time periods sufficient to complete the in situ treatment
process. Available evidence indicates that most commercially available
surfactants are biodegradable, although degradation rates of surfactants under
the range of geochemical conditions of interest in waste treatment are
lacking. Sivik et al. (1982) reviewed the biodegration of selected major
surfactants including C12 homologs of the following classes:
o linear alkylbenzene sulfonates,
o alkyl sulfates,
o alpha olefin sulfonates,
o secondary alkane sulfonates,
o alcohol ethoxy sulfates,
o alkyl phenol ethoxylates, and
o alcohol ethoxylates.
Results of BOD, C0_ evolution, and simulated treatment process tests
Indicated that for all of the tested compounds significant degradation
(greater than 50%) was observed in less than 20 days. Die away tests
suggested 90-100% decreases in surfactant concentrations in less than 10 days
for the compounds tested (Sivik et al., 1982).
Available manufacturer's information included in Table 3-5 indicates generally
rapid degradation of certain sulfosuccinates, sulfonates, and alkyl sulfates
and somewhat slower rates for alcohol ethoxylates. Quantitative data for the
fluorocarbons were unavailable but biodegradation is likely to be somewhat
slower than for the other listed compounds. Data for nonionic surfactants
indicate alkyl chains to be more rapidly degraded than ethylene oxide chains,
and alkylphenol ethoxylates to be somewhat more slowly degraded.
148
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TABLE 3-5
ENVIRONMENTAL CHEMICAL PROPERTIES OF SELECTED COMMERCIAL SURFACTANTS
U)
SURFACTANT CLASS
Fluorocarbons
EXAMPLE
ELECTROLYTE
TOLERANCE
WATER SOLUBILITY (HARPNESS - PPM)
Sulfonates (Anionics)
Alcohol Ethozylates
(Nonionlc)
o Lodyne Series o Soluble
(C1BA-GEIGY)
o Zonyl Series o >2gms/100gms
Alkanol Series o Soluble
(DuPont)
o Merpol Series o Generally
(DuPont) > 30X
o 300
o Electrolyte
Tolerant?
PH OF AQUEOUS
SOLUTIONS^)
5.0 - 8.5 (IX)
7.5 - 10.0 (IX)
6.0-9.0 (IX)
BIODEGRADATION
o Slow?
o Slow?
o Biodegraded
o Biodegraded
(20-60X In 20 days
with acclimated
bacteria)
TOXIC
0 3-10 go/kg
o Acute Dermal LD50
(Rabbit) 3-10gm/kg
o 1-25 gn/kg
o Acute Oral
Tozlclty for Fish
1-6 ng/L
Sulfosucclnates OT & Aerosol o l-60gms/100ol o 500-2500
(Anlonlc) Series (Cyananid)
5.0 - 8.0
o 50-100X- 8 days
(CSMA-Shake
Culture Test)
o 1-10 ml/kg
Alkyl Sulfates
(Anlonlc)
Duponol Series
(DuPont)
o Soluble
o Electrolyte
Tolerant
7.5 - 11.0 (3X)
o Biodegraded
(days to weeks)
o. 2-20 gB/kg
o Acute Oral
Toxicity to Fish
5-20 mg/1
(1) Parenthesis Indicate concentration of surfactant.
(2) Tozlclty reported as acute oral for rats unless otherwise specified.
-------
In general, within a given surfactant class, biodegradation rates were found
(Sivik et al., 1982) to vary with:
o the length of alkyl chains,
o the positions of phenyl groups, and
o the extent of chain branching.
Under the anaerobic conditions which may exist in organic waste deposits and
associated groundwaters, degradation rates are likely to be considerably
slower than under aerobic conditions. Qualitative data suggest that at least
certain types of surfactants (sulfates and sulfonates) will eventually degrade
in anaerobic environments (Sivik et al, 1982).
3.6.2 Toxicity
A detailed evaluation of the toxicity of the many commercially available
surfactants is beyond the scope of this study. For comparative purposes,
manufacturer-supplied toxicity data are included in Table 3-5. Sivik et al.
(1982) noted that commonly reported LC5Q values (for 24-96 hour studies)
ranged from 1-50 mg/1 for fish and 1-300 mg/1 for invertebrates for:
o linear alkylben'zene sulfonates,
o alkyl sulfates,
o alpha olefin sulfonates,
o secondary alkane sulfonates, ,
o alcohol ethoxy sulfates,
o alkylphenol ethoxylates, and
o alcohol ethoxylates.
In general, increases in carbon chain length up to C , were observed to
significantly increase toxicity, with toxicity decreases observed for longer
chain lengths. Based on an evaluation of rat LD50 data for the health
hazards to humans posed by surfactants, Sivik et al. (1982) indicated that
surfactants in general possess a relatively low level of acute mammalian
150
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toxicity. Acute oral LD5Q values for rats were generally found to range
from 650 mg/kg to greater than 3000 mg/kg. Other data from the literature
suggest no chronic effect levels in the range of 0.1% - 1.4% in diet or 0.01%
in drinking water.
3.7 Summary
The results of this study suggest that while selective surfactant applications
might be utilized to effectively enhance recoveries of organics in certain
waste and soil conditions, a substantial amount of detailed laboratory and
pilot scale research will be required for the specific waste/soil/groundwater
conditions at each site. Specific qualitative conclusions of this study
include the following:
o Surfactant application to waste deposits warrants serious
consideration as a means of reducing aqueous/organic interfacial
tensions, making the organics more accessible to other means of
degradation, and possibly enhancing the ability of aqueous solutions
to flush insoluble organics from subsurface soils.
o Although comprehensive interfacial tension data are lacking for most
surfactant/water/organic systems, there appear to be a number of
relatively inexpensive and environmentally • safe, commercially
available classes of surfactants which should significantly reduce
interfacial tensions in many aqueous/organic systems of interest in
waste treatment. Surfactant classes which may be particularly
effective in this regard include anionic fluorocarbons, anionic
sulfonates, and nonionic alcohol ethoxylates.
o Available information suggests that a single chemical chracteristic
of a surfactant or surfactant/water/organic system (for instance
interfacial tension) can not effectively predict the overall
likelihood, of surfactant effectiveness ;in waste treatment.
151
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Experimental measurements of properties (particularly interfacial
tension) for relevant organic systems can, however, provide a method
of initially screening specific surfactants for potential inclusion
in more detailed studies (i.e., batch or column soil studies).
The feasibility of solubilizing organics in concentrated surfactant
solutions warrants further consideration but may be somewhat
constrained because of the high surfactant solution concentrations
and volumes required to effect significant solubilizatlon of large
organic deposits, and difficulties in surfactant recovery for
recycling.
The use of surfactants to promote emulsification is likely to be
constrained by the complexity of emulsification processes, including
the possible need for mechanical energy to generate an emulsion, the
potential for reversibility of oil emulsions in aqueous systems, and
the potential for phase separation and organic readsorption to soils.
The likelihood of chemical interactions with soil particles or
groundwater constituents presents a potentially serious constraint to
the use of any surfactant type in soil systems. Anionic and nonionic
surfactants should be least affected by soil adsorption reactions at
normal soil pH (when soil particle surfaces are negatively charged),
while cationic surfactants would be adsorbed most strongly. With
decreasing soil pH, the soil particles will eventually become
positively charged (Figure 3-1A), at which point the anionic
surfactants will be adsorbed most while cationic surfactants would be
least adsorbed. Adsorption to soils and the resultant loss in
effectiveness preclude the consideration of cationics for waste
treatment. The "custom synthesis" of anionic-nonionic surfactant
mixtures should be considered as a means of minimizing adsorption
effects.
152
-------
o Within given classes, surfactants possess sufficiently varied
chemical characteristics with respect to aqueous solubility,
electrolyte and solution pH tolerance such that these properties
should not pose insurmountable limitations to application in most
groundwaters of low to moderate hardness (1-500 ppm).
o Under aerobic conditions, most commercial surfactants are effectively
biodegraded in relatively short time frames (days to weeks), and
effectiveness for in situ treatment might actually be inhibited by
overly rapid degradation rates. Under anaerobic conditions
degradation rates may be much slower and of greater environmental
concern in removing residual surfactants, particularly with respect
to anionic fluorocarbons and nonionic ethoxylated phenolics, which
may degrade very slowly.
In order to better define the likelihood of success of surfactant applications
to organic waste deposits additional information is required. In view of the
chemical complexity of organic waste mixtures and the apparent limitations of
aqueous/organic interfacial tension measurements in predicting the
effectiveness of surfactants, the emphasis of further research should be on
laboratory scale studies, possibly including:
o initial screenings of the effectiveness of specific surfactants to
reduce the interfacial tensions of various pure and mixed organic
phases, followed by,
o tests of the efficiency and effectiveness of specific surfactants to
remove organics in soil columns or similar simulation systems.
Emphasis of such studies should be placed on the investigation of mixed
surfactant systems since it is less likely that single surfactant systems will
possess the combination of surface active characteristics required for maximum
surface activity while simultaneously possessing the optimal characteristics
to minimize processes such as soil adsorption.
153
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3.8 Conclusions
For waste deposits containing organic compounds of relatively high water
—2
solubility (greater than 5 x 10 M), flushing with aqueous solutions alone
(without surfactant addition) may prove to be an effective treatment process
and should be considered. For deposits containing significantly less soluble
organic compounds which 'possess moderately high octanol-water partition
coefficient values (log K greater than 2) flushing with aqueous solutions
alone may prove to be of limited effectiveness. For these deposits, the use
of surfactant solutions may enhance recovery efficiencies. However, prior to
the application of surfactant solutions to waste deposits, laboratory research
must be conducted to determine both the most appropriate surfactant (or
mixture) for a particular waste in terms of the desired surface chemical
properties and also the most effective surfactant in terms of minimizing
unwanted interactions with subsoils.
The limited data base on surfactant use in soil systems is largely confined to
considerations of surfactant application to petroleum and petroleum derived
compounds and mixtures, including various component aliphatic and aromatic
hydrocarbons. Therefore, it is for these types of waste deposits that
surfactant applications may hold the greatest near term potential.
Application to other types of organic contaminants is possible but would
require considerably more background research.
154
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References
Barakat, Y., L.N. Fortney, R.S. Schechter, W.H. Wade, and S.H. Yiv 1983.
Criteria for Structuring Surfactants to Maximize Solubilization of Oil and
Water. J. Colloid Interface Science, 92 (2): 561-574.
Cash, L. J.L. Cayias, G. Fournier, D. Macallister, T Schares, R.S. Schechter
and W.H. Wade. 1977. The Application of Low Interfacial Tension Scaling Rules
to Binary Hydrocarbon Mixtures. J. Colloid Interface Science, 59 (1): 39-44.
Cayias, J.L., R.S. Schechter, and W.H. Wade, 1977. The Utilization of
Petroleum Sulfonates for Producing Low Interfacial Tensions between
Hydrocarbons and Water, J. Colloid Interface Science, 59 (1): 31-37.
Doe, P.H., W.H. Wade, and R.S. Schechter. 1977. Alkyl Benzene Sulfonates for
Producing Low Interfacial Tensions Between Hydrocarbons and Water. J. Colloid
and Interface Science, 59 (3): 525-531.
Ellis, W.D., J.R. Payne, A.N. Tafuri and F.J. Frastone. 1984. The
Development of Chemical Countermeasures for Hazardous Waste Contaminated
Soil. EPA-600/D-84-039. Municipal Environmental Research Laboratory, US
Environmental Protection Agency, Cincinnati, OH.
Morgan, J.C., R.S. Schechter and W.H. Wade. 1979. Ultra-Low Interfacial
Tension and Its Implications in Tertiary Oil Recovery. In: Solution Chemistry
of Surfactants, Volume 2, K.L. Mittal (ed.), Plenum Press, New York, NY.
Mukerjee, P. 1979. Solubilization in Aqueous Micellar Systems. In: Solution
Chemistry of Surfactants, Volume 1, K.L. Mittal (ed.), Plenum Press, New York,
NY.
Park, G.A. 1967. Aqueous Surface Chemistry. In: Equilibrium Concepts in
Natural Water Systems, R.F. Gould (ed.), Advances in Chemistry Series No. 67,
ACS Washington, D.C.
Rosen, M.J. 1978. Surfactants and
Interscience. New York, NY.
Interfacial Phenomena.
Wiley
Shah, D.O., ed. 1977. Improved Oil Recovery by Surfactant and Polymer
Flooding. Academic Press, New York, NY.
Shaw, D.J. 1976. Introduction to Colloid and Surface Chemistry. Butterworths,
London.
Sivik, A., M. Gouer, J. Perwak, P. Thayer. 1982. Environmental and Human
Health Aspects of Commercially Important Surfactants. In: Solution Behavior
of Surfactants: Theoretical and Applied Aspects, Volume 1, K.L. Mittal, and
E.J. Fendler ed., Plenum Press, New York, NY.
Texas Research Institute, 1979. Final Report Underground Movement of: Gasoline
on Groundwater and Enhanced Recovery by Surfactants, prepared for American
Petroleum Institute, Washington, DC.
155
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Texas Research Institute. 1982. Test Results of Surfactant Enhanced Gasoline
Recovery in a Large-Scale Model Aquifer, prepared for American Petroleum
Institute, Washington, D.C.
USEPA. 1982. Handbook for remedial action at waste disposal sites.
EPA-625/6-82-006. Municipal Environmental Research Laboratory, US
Environmental Protection Agency, Cincinnatti, OH.
Wade, W., R.S. Schechter, M. Bourrel, M. Baviere, M. Fernandez, C,, Kourkounis,
H. Lim, A. Gracia, C. Nunn, and J. Scamehorn. 1980. Tertiary Oil Recovery
Processes - Annual Report. DOE/BC/20001-6, Prepared for U.S. Department of
Energy, Washington, DC.
Wasik, S.P., Y.B. Tewari, M.M. Miller, and D.E. Martire. 1981. Octanol-Water
Partition Coefficients and Aqueous Solubilities of Organic Compounds. NTIS
#PB82-141797, National Bureau of Standards Report to the Environmental
Protection Agency.
Wilson, P.M. and C.F. Brandner. 1977. Aqueous Surfactant Solutions which
Exhibit Ultra-Low Tensions at the Oil-Water Interface. J. Colloid Interface
Science, 60 (3): 473-479;
156
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SECTION 4
HYDROLYSIS
4.1 Introduction
Hydrolysis is a chemical reaction in which a compound reacts with water,
leading to cleavage of a bond in the compound. A common form of hydrolysis
can be expressed as a displacement reaction,
RX + H90 - ROH + HX,
(4-1)
where R represents an organic moiety and X the cleaved group in the hydrolysis
reaction. In aqueous systems under typical environmental conditions,
hydrolysis represents a major degradation mechanism for many organic
chemicals. However, data with which to evaluate the contribution of
hydrolysis to degradation of chemicals in waste deposits is limited.
Therefore this section will review the basic controlling mechanisms of
hydrolysis in the environment, present methods of estimating hydrolysis rates
in waste deposits, and evaluate means of accelerating hydrolysis rates as a
potential treatment method for waste deposits.
The primary data sources for this section are recent reviews which cover
hydrolysis under environmental conditions (Harris, 1982; Mabey and Mill, 1978;
Mill 1979; Radding et al., 1977; Versar, Inc, 1979). These reviews include
extensive compilations of hydrolysis data including, in many cases, hydrolysis
rate constants, conditions, half-lives and other data for a wide variety of
organic compounds. These compilations, however, are not complete, and
considerable additional data on hydrolysis are available in the recent
literature. A comprehensive compound-by-compound review of the literature is
beyond the scope of this work. However, the reader should be aware that a
compound-specific search of the literature may provide data on hydrolysis
rates for numerous organic compounds not included in this report or the
primary data sources listed above.
157
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4.2 Hydrolysis Mechanisms and Kinetics
In general, hydrolysis proceeds by attack of a nucleophile (e.g., water or
hydroxyl ion) on an electrophile (e.g., carbon or phosphorus), resulting in
displacement of a cleaved group. The reaction rate may either be independent
of nucleophile concentration (unimolecular reaction) or be a function of the
nucleophile concentration (bimolecular reaction).
Hydrolysis may occur through a variety of reaction pathways. In some cases,
various hydrolysis mechanisms may be competing in a molecule with multiple
functional groups. For example, Harris (1982) reports studies of malathion
hydrolysis in which both carboxylate ester cleavage and phospliorodithioate
ester cleavage can be significant. While it is recognized that various
hydrolysis pathways can result in different by-products, an attempt has not
been made in this report to identify all of the potential products of
hydrolysis for the compounds considered. Of course, this would be required
for a specific application to ensure that the products do not present a
greater contamination problem than the parent compound(s).
Hydrolysis rates discussed in this section are based upon the disappearance
rate of the parent compound only, without respect to mechanism or by-product
formation. The rate of hydrolysis reactions can be described by kinetic rate
expressions. In almost all cases, hydrolysis appears to occur as a
first-order or pseudo-first-order reaction in which the rate of disappearance
of the substrate, RX, is proportional to the concentration of substrate:
-d(RX)/dt = k(RX)
(4-2)
The persistence, in terms of half-life, for a given substrate can be expressed
as:
(In 2)/k - 0.693/k
(4-3)
As will be discussed in greater detail below, half-lives for hydrolysis of
organic chemicals may range from seconds (or less) to millions of years
158
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(Harris, 1982; Mabey and Mill, 1978). The overall hydrolysis rate for a
compound may be comprised of several separate reaction rates, namely those
appropriate for neutral hydrolysis (rate independent of pH), acid-catalyzed
hydrolysis (rate proportional to hydrogen ion concentration), and
base-catalyzed hydrolysis (rate proportional to hydroxyl ion concentration).
This will be described further below.
4.2.1 Hydrolyzable Organic Groups
Harris (1982) has tabulated organic functional groups which are susceptible to
hydrolysis, as well as those which are resistant to hydrolysis. Compounds
resistant to hydrolysis include unsubstituted hydrocarbons (aliphatic and
aromatic), halogenated aromatics, PCBs, phenols, aromatic amines, and many
other classes (Table 4-1). Organic functional groups susceptible to
hydrolysis include alkyl halides, carbamates, nitriles, phosphoric and
phosphonic acid esters, and several other functional groups (Table 4-2). It
should be noted that a given organic molecule may contain both hydrolyzable
and non-hydrolyzable functional groups since it may contain more than one
functional group.
The reviews of hydrolysis cited in this report have focused primarily on the
overall hydrolysis rate under neutral conditions (pH 7) at or near 25°C
(77 F), as an indicator of the persistence of various .chemicals under
typical environmental settings. This approach provides a reasonably
conservative estimate of half-life via hydrolysis, although in some cases, the
half-life at pH 7 may be orders of magnitude shorter than the half-life at the
minimum hydrolysis rate. For example, Zepp et al. (1973) have shown that the
minimum hydrolysis rate of 2,4-D occurs at pH 3 to 4, and the hydrolysis rate
at pH 7 is approximately three orders of magnitude above the minimum value.
From the perspective of stabilizing waste deposits via hydrolysis of organic
contaminants, the hydrolysis rate at typical environmental conditions (i.e.,
at pH 7) is of limited interest, since chemicals which hydrolyze rapidly under
these conditions would not be persistent in waste deposits. Therefore, it is
159
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TABLE 4-1
GROUPS OF ORGANIC COMPOUNDS THAT ARE
GENERALLY RESISTANT TO HYDROLYSIS3 (Harris, 1982)
Alkanes
Alkenes
Alkynes
Benzenes/biphenyIs
Polycyclic aromatic hydrocarbons
Heterocyclic polycyclic
aromatic hydrocarbons
Halogenated aromatics/PCBs
Dieldrin/aldrin and related
halogenated hydrocarbon pesticides
Aromatic nitro compounds
Aromatic amines
Alcohols
Phenols
Glycols
Ethers
Aldehydes
Ketones
Carboxylic acids
Sulfonic acids
a. Multifunctional organic compounds in these categories may be
hydrolytically reactive -if they contain a hydrolyzable
functional group in addition the functionality listed above.
160
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TABLE 4-2
GROUPS OF ORGANIC COMPOUNDS THAT ARE
POTENTIALLY TREATABLE BY HYDROLYSIS (Harris, 1982)
Alkyl halides
Amides
Amines
Carbamates
Carboxylic acid esters
Nitriles
Phosphoriic acid esters
Phosphoric acid esters
Sulfonic acid esters
Sulfuric acid esters
161
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the purpose of this section, to examine the factors which control the
hydrolysis rate of chemicals in order to determine the feasibility of
accelerating hydrolysis rates of persistent chemicals in waste deposits.
Factors controlling hydrolysis rates include pH, ionic strength, temperature,
solvent composition, and catalysts. These factors are discussed below.
4.2.2 Effect of pH on Hydrolysis Rates
The hydrolysis rate of a given compound may be the sum of the neutral, acid-
catalyzed, and base-catalyzed processes. The respective rate expressions for
these processes are as follows:
-d(RX)/dt s kN(RX)
-d(KX)/dt s kA(RX)(H+)
-d(RX)/dt s kB(RX)(OH-)
(neutral hydrolysis) (4-4)
(acid-catalyzed hydrolysis) (4-5)
(base-catalyzed hydrolysis) (4-6)
Where kN, k., and kg are the neutral, acid-catalyzed, and base-catalyzed
rate constants, respectively. The overall rate of hydrolysis of a. compound is
given by the sum of the rates of the contributing reactions:
-d(RX)/dt - kN(RX) + kA(RX)(H+) + kB(RX)(OH~) (4-7)
At a fixed pH, the sum of these reactions appears as a pseudo'-first order
reaction in RX, where
-d(RX)/dt - k^RX)
The overall hydrolysis rate constant, k, , is given by
(4-8)
(4-9)
Figure 4-1 illustrates the effect of pH on overall hydrolysis rate (Mabey and
Mill, 1978). It is important to note that for substances where significant
acid- or base-catalyzed reaction rates apply, the effect of a one-unit pH
change is a one order of magnitude change in the overall hydrolysis? rate.
162
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FIGURE 4-1
pH DEPENDENCE OF HYDROLYSIS RATE BY
ACID, NEUTRAL, AND BASE PROMOTED PROCESSES
cr>
co
HYDROLYSIS
RATE
ACID
HYDROLYSIS
RATE
OVERALL
HYDROLYSIS
RATE
BASE
HYDROLYSIS
RATE
NEUTRAL
HYDROLYSIS
RATE
ACIDIC
BASIC
-------
In order to discuss the effect of pH on half-lives for hydrolysis for specific
compounds, kinetic data (rate constants) and other estimates of hydrolysis
rates have been compiled from Mabey and Mill (1978), Harris (1982), Radding et
al. (1977) and Versar Inc. (1979). These data are presented in Tables 4-3
through 4-10, and an example (hydrolysis of ethyl acetate) is illustrated in
Figure 4-2.
Where possible, hydrolysis half-lives as a function of pH have been calculated
from these available data. In many cases, data required to calculate the
effect of pH are not available. It must also be noted that this compilation
of hydrolysis rates is not exhaustive. The primary literature undoubtedly
contains more data on the hydrolysis rates of specific chemicals of concern
not included in this report, both under laboratory conditions and in
environmental settings.
4.2.3 Effect of Temperature on Hydrolysis Rates
Several methods for estimating the effect of temperature on the hydrolysis
rate constant are commonly applied in the study of kinetics (Zepp et al.,
1975). One example is illustrated by reference to the Arrhenius relation:
Ae
-EA/RT
(4-10)
where A is a constant, E. is the Arrhenius activation energy, R the gas
constant, and T is absolute temperature.
The dependence of hydrolysis rates on temperature must be considered in
evaluating data, and may represent a significant source of error in
extrapolating laboratory hydrolysis data to environmental conditions.
Although the temperature dependence of hydrolysis rates is compound-specific,
a generalized estimate that a 10 C decrease in temperature produces a factor
of 2.5 decrease in hydrolysis rate is reasonable in the range of 0 to 50°C
for most organic compounds (Mabey and Mill, 1978).
164
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FIGURE 4-2
EFFECT OF pH ON HYDROLYSIS
OF ETHYL ACETATE
to 10000
1 000
100
10
1
<
Q
UJ
u.
_j
<
11
pH
165
-------
Estimates of hydrolysis rates in the environment, including rates tabulated in
this report, are usually based on for a temperature of 25°C. Waste deposit
temperatures, in general, can be expected to be less than 25°C (unless
containing an internal heat source, e.g., organic decomposition in a
landfill). Typical non-thermal groundwater temperatures in the United States
vary primarily with latitude, and generally range from about 5 to 27°C
(40°F to 80°F) (Pettyjohn et al., 1979; Repa and Kufs, 1985). Seasonal
temperature variations may be observed near the surface. If waste deposit
temperatures are near ground water temperatures, hydrolysis rates in waste
deposits in northern climates could be approximately a factor of 5 to 6 slower
than those reported for 25°C (77°F).
4.2.4 Effect of Solvent Composition on Hydrolysis Rates
Hydrolysis rates are affected by solvent composition, with rates in water
greater than rates for mixed water and organic solvents. For example, the
hydrolysis rate of t-butylchloride increases approximately four orders of
magnitude with a change in solvent composition from 90% ethanol/10% water to
100% water (Mabey and Mill, 1978). Hydrolysis rates reported for mixed
solvents should be considered conservative estimates of hydrolysis rates for
compounds dissolved in water. With water as the solvent for hydrolysis, rates
may be affected by ionic strength (a measure of the total concentration of
dissolved constituents), and increasing ionic strength can either accelerate
or retard hydrolysis (Mabey and Mill, 1978). Total ionic strengths of less
than 0.1 M (which is equivalent to a salinity of about 3000-6000 ppm,
depending on the major ions present) are unlikely to have a significant affect
on hydrolysis rates (less than 5 to 10%) according to Harris (1982).
4.2.5 Catalysis
Mabey and Mill (1978) report that alkaline earth and heavy metal ions can
catalyze hydrolysis, apparently by increasing the effective OH~ ion
concentration. If this postulate is correct, metal ion catalysis would appear
to favor base-catalyzed hydrolysis processes. Specifically, copper, manganese,
magnesium, and cobalt have been found to catalyze various reactions. However,
166
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Mabey and Mill (1978) indicate that metal ion catalysis is unlikely to affect
hydrolysis rates at typical metal ion concentrations in the environment,
although additional research would be necessary to determine the exact
contributions of catalysis in specific environments. Harris (1982) reports
that apparent hydrolysis rates in surface water in excess of predictions from
laboratory results have been attributed to catalysis. At this time, the
potential catalytic properties of metal ions, clay surfaces, etc., which may
be present in a waste deposit cannot be predicted with any certainty.
Catalysis and retardation of hydrolysis rates of organic molecules by
surfactants has also been reported (Fendler and Fendler, 1970; N L Wolfe,
USEPA, Athens, GA, Personal Communication). However, simple rules for
estimating this effect in environmental settings are not available.
4.3 Acceleration of Hydrolysis Rates in Waste Deposits
While a variety of factors may contribute to or affect the hydrolysis rate of
organic chemicals in the environment, only the effects of pH and temperature
can be characterized in a predictable manner. Temperature effects may be
significant (e.g., a 25°C temperature increase may result in an order of
magnitude increase in hydrolysis rates); however, it is unlikely that such
temperature increases can be achieved in any field setting for extended
periods of time without enormous energy expenditures. For this reason, the
only apparent feasible method of increasing hydrolysis rates of chemicals in a
waste deposit is by controlling the pH regime of the deposit. In particular,
hydrolysis appears to be a potentially attractive in situ treatment method for
a number of organic substances subject to base-catalyzed hydrolysis, for which
the hydrolysis rates can be increased dramatically in the range of pH 7 to 11,
as discussed below. Acid-catalyzed hydrolysis is less desirable, since acid
conditions can mobilize significant concentrations of naturally-occurring or
pollutant trace metals, creating other contamination problems.
The selection of pH 11 as an upper limit of the pH range for hydrolysis rate
calculations in this report is somewhat arbitrary, since higher pH values
could theoretically be achieved by addition of. strong bases to waste
167
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deposits. However, the solubilities of major soil constituents (silica,
alumina and aluminosilicates such as kaolinite) increase at high pH (above pH
9 to 10) and increase substantially above pH 11 (Stumm and Morgan, 1980). For
this reason, a pH value near 11 [(OH)~ = 1 x 10~ moles/liter] is probably
a reasonable upper limit of pH achievable under field conditions. The value
of increasing the pH in order to accelerate the hydrolysis rate must consider
the ratio of k,/k . With pH = 11 being the approximate upper limit of pH
achievable (i.e., OH = 10~3) kb must be 103 times k for the
A
base—catalyzed rate to equal the neutral rate at a pH of 11, 10 times k
at a pH of 10, etc (see equations 4-4 through 4-6). If the base catalyzed
contribution to the overall hydrolysis rate is less than that obtained at
neutral conditions alone, raising the pH is probably not advisable.
Hydrolysis in the waste deposit can be assumed to be primarily an aqueous
phase reaction. Degradation rates of a compound sorbed to solid phases or
present in organic phases may be significantly different from the rate
applicable to the aqueous phase. As such, hydrolysis rates reported in this
section are probably most valid for relatively soluble species, and in waste
deposits where sorption is limited. The potential effect of increasing
hydrolysis rates through base catalysis for various chemical classes
susceptible to hydrolysis is discussed below.
4.3.1 Alkyl Halides
Alkyl halides generally hydrolyze according to Equation 4-1:
RX
EOH + HX
(4-11)
Hydrolysis rates for alkyl halides are influenced by both neutral hydrolysis
and base-catalyzed processes. Table 4-3 presents the calculated hydrolysis
half-lives for a number of alkyl halides in the pH range of 5 to ill, based on
reported hydrolysis rate constants. With the exception of polyhalomethanes
(e.g., CHBr.Cl), base catalysis is not significant below pH values of 11 to
3
13 (i.e., k. is less than 10 times k ). Furthermore, hydrolysis
b n
half-lives for alkyl halides (with the exception of some methyl and benzyl
168
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TABLE 4-3
HYDROLYSIS OF ALKYL HALIDES
Class/Compound Data
Alkyl Halides
Methyl Fluoride
Methyl Chloride
Methyl Bromide
Methyl Iodide
Methylene Chloride
CH2CHCH2C1
C,H_CH.C1
652
CbH5CHCl2
C(,Hi,CCl3
CH2CHCH2Br
CH2CHCH2I
P~CH3COH4CH2C1
C6H5CH2Br
p"CH3C(jH4CH2Br
CH2BrCl
CHC13
CHBrCl2
CHBr2Cl
CH2Br2
CHBr3
CHIC12
CHFIC1
CC14, Ippm
CC14, 1000 ppm
1 ,1 ,1-Trichloroethane
Tetrachloroethene
1 , 2-dichloroe thane
1 , 2-dibromoethane
Ethylchloride
1 , 1-dichloroethane
Hexachlorocyclopentadiene)
Source
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
" 1
1
1
2
3
3
3
2
2
kA
7
2
4
7
3
1
1
1
6
1
4
4
1
2
-
-
-
-
-
-
-
-
-
-
-
-
-
-
-
—
kN
.44E-10
.37E-8
.09E-7
.28E-8
.2E-11
.16E-7
.28E-5
.56E-3
.3E-2
.67E-5
.01E-6
.5E-4
.45E4
.67E-3
-
-
-
-
-
-
-
-
-
-
-
-
-
-
-
—
t 1/2 at pH
kfl
5.82E-7
6.18E-6
1.41E-4
6.47E-5
2.13E-8
6.24E-5
_
-
-
-
-
—
-
-
-
6.9E-5
1.6E-3
8.0E-4
-
3.2E-4
8.0E-4
2.2E-1
second
order
second
order
-
-
-
-
~
5
30y
0.93y
20d
llOd
704y
69d
15h
O.lh
19s
12h
2.0d
0.43h
1.32h
4.3m
-
350,000y
13,700y
27,400y
-
6,600y
27,500y
-
-
-
-
-
-
-
14d
6
30y
0.93y
20d
llOd
704y
69d
15h
O.lh
19s
12h
2.0d
0.43h
1.32h
4.3m
-
35,000y
1370y
2740y
-
6860y
2750y
-
-
-
-
-
-
-
-
14d
7
30y
0.93y
20d
llOd
704y
69d
15h
O.lh
19s
12h
2.0d
0.43h
1.32h
4.3m
44y
3500y
137y
274y
183y
686y
275y
l.Oy
7000y .
7y
0.5y
0.7y
SO.OOOy
5,000y
40d
14d
8
30y
0.93y
20d
llOd
704y
69d
15h
O.lh
19s
12h
2.0d
0.43h
1.32h
4.3m
-
350y
13. 7y
27.4y
-
68. fey
27.5y
36.5d
-
-
-
-
-
-
-
14d
9
30y
0.93y
20d
llOd
704y
69d
15h
O.lh
19s
12h
2.Ud
0.43h
1.32h
4.3m
-
35y
1.37y
2.74y
-
6.86y
2.8y
3.7d
-
-
-
-
-
-
-
14d
10
30y
0.93y
20d
lOOd
704y
69d
15 h
O.lh
19s
12h
2.UQ
0.43h
1.32h
4.3m
-
3.5y
50d
lOOd
-
250d
0.28y
0.37d
-
-
-
-
-
-
-
—
11
15y
0.93y
20d
55d
350y
44. 9d
Ibh
O.lh
19s
12h
2.0d
0.43h
1.32h
4.3m
-
0.35y
5a
lOd
-
25d
lOd
0.037d
-
-
-
-
-
-
-
—
Note: All values reported for 25+ 5°C.
s = second
m = minute
h = hour
d = day
y = year
Data Sources: 1. Mabey and Mill, 1978
2. Versar, Inc, 1979
3. Radding et al., 1977
Rate constants in sec
"1
-------
halides) are generally in the range of years. As such, attempts; to increase
in situ hydrolysis rates of alkyl halides through control of pH are unlikely
to be effective, for the range of reasonably achievable pH values.
4.3.2 Halogenated Ethers, Epoxides, and Alcohols
Hydrolysis rates of halogenated epoxides and ethers are generally high even at
neutral pH, with half lives reported in the range of days or: less (Table
4-4). As such, these substances are unlikely to persist in the environment.
Acceleration of hydrolysis of these substances is unlikely to be necessary.
Hydrolysis rates were collected for only two halogenated alcohols, 2-chloro-
ethanol and l-chloro-2-propanol; their reported hydrolysis half-lives at pH 7
are 21 years and 2 years, respectively. Data on the effect of pH on
hydrolysis rates for these substances were not obtained.
In summary, the available data do not indicate the potential for enhancement
of hydrolysis rates by pH control for in situ treatment of halogenated ethers,
epoxides or alcohols in waste deposits.
4.3.3 Epoxides
Epoxides are hydrolyzed by acid-catalyzed and neutral processes.: Hydrolysis
rates for epoxides are generally high, with half-lives in the range of 15 days
at pH 7 (Table 4-5). While hydrolysis rates for epoxides can be increased at
low pH, this process is unlikely to contribute significantly to the treatment
of these substances in waste deposits because of their limited persistence at
neutral pH values.
4.3.4 Esters (Carboxylic Acid Esters)
Esters hydrolyze to form carboxylic acids and alcohols according to the
reaction:
R1C(0)OR2
R1C(0)OH
(4-12)
170
-------
TABLE 4-4
HYDROLYSIS OF HALOGKNATED Kl'HHRS,
EPOX1DES, ALCOHOLS
Compound Data Source kA kN kB
Chloromethylmethylether
bis(Chloromethyl)ether
2-(Chloroethanol
l-Chloro-2-propanol
2-Chloroethylvinylether
3-Chloro-,l,2 epoxy,
2-methyl propane
Alpha-epichlorohydrin
Epibromohydrin
Note: All values reported
s - second
d - day
y - year
1 - - . - •
1
1 - -
1 - -
2 - -
3 1.84E-3 SE-7
3 8.0E-4 9.8E-7
3 6.1E-4 5E-7
for 25°C. Rate constants In see .
t 1/2 at pH
56 789
0.007s
38s
21y
2 y
0.48 y -
16d 16d 16d 16d 16d
8.2d 8.2d 8.2d 8.2d 8.2d
16d 16d Ibd Ibd 16d
10 11
-
-
-
-
-
Ibd Ibd
8.2d 8.2d
16d lod
Date Sources: 1. Radding et al., 1977
2. Versar, Inc. 1979
3. Mabey and Mill, 1978
-------
TABLE 4-5
HYDROUfSIS 01? ETOXIDES
-vl
ro
Compound Data
1,2-Epoxy ethane
1,2 Epoxy propane
1,2 Epoxy-2-aethyl
propane
1,2 Epoxy-3-hydroxy
propane
1 , 2-Epoxy-2-methyl-3-
hydroxy propane
1 , 2-Epoxy-2-tne thyl-3-
chloropropane
Trans-2,3 epoxy butane
Cis-2,3-epoxy butane
1,2-epoxy-l-phenyle thane
trans-1 , 2-epoxy-l
phenylpropane
Source
1
1
1
1
1
1
1
1
2
2
"A
1E-2
4.6E-2 ,
7.3EO
2.5E-3
1.1E-2 <
1.84E-3
1.2E-1
2.4E-1
-
kN kB
6.7E-7
5.5E-7
1.1E-6
2.84E-7
4E-7
5E-7 .
5E-7
5E-7 - - -
-
12d
7d
O.ld
26d
13d
15d
4.7d
2.8d
_
t 1/2 at pH
12d 12d 12d
14. 6d 14.6d 14. 6d
.95d 4.4d 6.8d
- 28d 28d 28d
16d 16d 16d
16d 16d 16d
13d 15. 7d 16d
lid 15.3d Ibd
lid - '
4d -
12d
14. 6d
7.2d
28d
16d
16d
16d
Ibd
12d
14. 6d
7.3d
28d
16d
Ibd
16d
16d
lid
14. 6d
7.3d
28d
16d
16d
Ibd
Ibd
Note: All values reported at 25°C. Kate constants in sec~l.
d ~ day
Data Sources: 1. Mabey and Mill, 1978
2. Radding et al., 1977
-------
Acid-catalyzed, neutral, and base-catalyzed processes may contribute to
hydrolysis of esters. Base-catalyzed processes dominate hydrolysis for many
esters above pH values in the range of 5 to 7. Half-lives for hydrolysis of
numerous aliphatic and aromatic acid esters are tabulated as a function of pH
in Table 4-6. Reported half-lives at pH 7 for esters cover a wide range (from
less than one day to over one hundred years). For essentially all esters
listed in Table 4-6, hydrolysis half-lives can be reduced to tens of days or
less in the pH range of 8 to 10. For example, the hydrolysis half-life of
t-butyl acetate is reduced from 140 years to 5.5 days by increasing pH from 7
to 10. Thus, the acceleration of hydrolysis rates in waste deposits through
base-catalyzed hydrolysis represents a potentially feasible method for in situ
treatment of carboxylic acid esters.
4.3.5 Amides
Amides hydrolyze by acid- and base-catalyzed processes, forming carboxylic
acids and amines according to the reaction (Mabey and Mill, 1978):
+ H20 = RC(0)OH +
(4-13)
Table 4-7 lists hydrolysis half-lives as a function of pH for various amides.
These half-lives are generally long (years to thousands of years) at pH 7, but
can be reduced substantially by increasing the pH, with calculated half-lives
in the order of years or less at pH 11. Base-catalyzed' hydrolysis may
therefore provide a feasible degradation mechanism for amides in waste
deposits, especially for chlorinated amides.
4.3.6 Carbamates
Carbamates may degrade by acid-catalyzed, neutral, or base-catalyzed
hydrolysis processes, although data in Mabey and Mill (1978) and Ryckman
(1984) indicate that base-catalyzed processes predominate. Carbamates are
hydrolyzed to alcohols, amines, and CO™ according to the reaction:
ROC(0)NR1R2
ROH
(4-14)
173
-------
TABLE 4-6
HYDROLYSIS OF ESTERS
Coopound Data Source kA
Ethyl Acetate
Isopropyl Acetate
Butyl Acetate
Vinyl Acetate
Allyl Acetate
Benzyl Acetate
0-acetyl phenol
2,4-dlnltrophenyl acetate
C1CH2C(0)°CH3
C12CHC(0)OCH3
C12CHC(0)OC6H5
F2CHC(0)OC2H5
C13CC(0)OCH3
F3CC(0)OC2H5
F3CC(0)OCCCH3)3
CH3SCH2C(0)OC2H5
CH3S(0)CHC(0)C2H5
(CH3)2SCH2C(0)-
OC2H5
C2H5C(0)OC2H5
C3H7C(0)OC2H5
(CH3)2CHC(0)OC2H5
CH2CHC(0)OC2H5
trans-CH3CHCHC(0)-
OC2H5
CHCC(0)OC2H5
C6H5C(0)OCH3
C6H5C(0)OC2H5
C^^3(0)OCH(CH3)2
CjHjCCOOCH^CjH^
p-N02-C6H4C(0)OCH3
p-N02-C6H4C(0)OCH3
p-N02-C6H4C(0)-
OC2H5
1-C5H4NC(0)OC2H5
o-C6H4[C(0)OC2H5]2
o-C6H4[C(0)OCH2-
6 5 J
P-C6H4(CCO)OCH3]2
P-C6H4[C(0)OC2H512
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1.1E-4
6.0E-5
1.3E-4
1.4E-4
-
1.1E-4
7.8E-5
-
8.5E-5
2.3E-4
-
—
-
_
-
-
-
3.3E-5
1.8E-5
1.2E-6
6.3E-7
-
4.0E-7
-
-
-
4.3E-7
1.4E-7
-
-
-
-
"n
1.5E-10
-
-
1.1E-7
-
-
6.6E-8
1.1E-5
2.1E-7
1.5E-5
1.8E-3
5.7E-5
7.7E-4
3.2E-3
1.3E-3
-
-
-
-
_
-
-
-
-
-
-
-
-
-
-
-
-
-
"»
1.1E-1
2.6E-2
1.5E-3
1.0E1
7.3E-1
2.0E-1
1.4EO
9.4E1
1.4E2
2.8E3
1.3E4
4.5E3
-
~
9.2E-1
1.3E1
2.0E2
8.7E-2
3.8E-2
2.3E-2
7.8E-2
1.3E-2
4.68EO
1.9E-3
3.0E-2
6.2E-3
8.0E-3
7.4E-2
6.4E-1
2.4E-01
5.4E-1
l.OE-2
1.7E-2
2.5E-1
6.9E-2
S
16y
35y
1.6y
73d
30y
17y
119d
17h
23d
llh
6.4m
3.3h
15m
3.6m
8.9m
24y
1.7y
40d
52.3y
lOOy
960y
244y
1140y
4.7y
3720y
730y
3500y
2700y
282y
34y
92y
41y
2200y
1300y
880y
320y
6
16y
6By
78y
67d
3.0y
10.4y
lOOd
16 h
5.0d
4.5h
6.1m
1.9h
15m
3.6m
8.9m
2.4y
62d
96h
24.4y
55y
96y
35y
160y
170d
1160y
73y
350y
270y
30y
3.4y
9.2y
4.1y
220y
130y
88y
32y
t 1/2
7
2.0y
8.4y
140y
7.3d
HOd
l.ly
38d
9.4h
14h
38m
3.7m
23m
15m
3.6m
8.9a
87d
6.2d
9.6h
2.5y
5.8y
9.6y
3.5y
17y
17d
118y
7.3y
35y
27y
3.0y
0.34y
0.92y
0.41y
22y
13y
0.88y
3.2y
at pti
b
U.2y
308d
1.5y
O.Bd
lid
40d
5.5d
1.8h
1.4h
4.1B
47s
2.6m
15m
3.6>
8.9m
8.7d
0.62d
57m
91d
212d
350d
128d
1.7y
1.7d
11 ".By
0.73y
3.5y
2.7y
0.3y
12. 4d
34d
15d
2.2y
1.3y
32d
117d
9
7.3d
31d
015y
1.9h
l.ld
4.0d
0.57d
0.2h
B.3m
255
4.7s
16s
15m
3.6m
8.9m
0.87d
1.5h
5.7m
9. Id
21d
35d
13d
62d
0.17d
1.2y
27d
12Bd
99d
lid
1.2d
3.4d
1.5d
80.3d
47. 5d
3.2d
11. 7d
1U
0.73d
3. Id
5.5d
12m
2.6h
9.6h
Ih
1.2m
50s
2.5s
0.5s
1.6s
15m
3.6m
8.9m
2.1h
9m
35s
0.9d
2. Id
3.5d
1.3d
6.2d
24m
0.12y
2.7d
IZ.bd
9.9d
l.ld
2.6h
0.34d
3.6h
B.Od
4.8d
7.7h
1.2d
11
U.073d
0.31d
U.5d
1.2m
0.2bh
Ih
5»
1+lm
SB
0.2s
0.05s
0.2s
3.6m
8.9m
13m
3.5s
2.2h
5h
0.35d
0.13d
0.6M
2.4m
4.4d
0.27d
1.3d
l.Ud
2.6h
0.2bh
O.Bh
0.3bh
0.8d
0.48d
0.77h
2.6h
Note: All values reported for 25°C. Rate constants in sec"*-.
s = second
m - minute
h = hour
d = day
y = year
Data Source: 1. Mabey and Mill, 1978
-------
TABLE 4-7
HYDROLYSIS OF AMIDES
C71
Compound
Acetamlde
Valeramide
Isobutyamide
Cyclopentanecarboxamide
Hethoxy acetamlde
Chloroacetamide
Dichloroacetamide •
Trichloroacetamlde
firomoacetamide
N-me thy lace tamide
N-e thylace tamide
Data Source
1
1
1
1
1
1
1
1
1
1
1
kA kN
8.36E-6
5.43E-6
4.63E-6
2.34E-5
7.84E-6
1.1E-5
-
-
-
3.2E-7
9.36E-8
KB
4.71E-5
1.41E-5
2.40E-5
1.67E-5
3.95E-4
1.5E-1
3.0E-1
9.4E-1
1.03E-5
5.46E-6
3.10E-6
t 1/2 at pH
5
262y
404y
470y
93. 9y
280y
84. 5y
73y
23y
2xlO&y
6900y
23,000y
6
2490y
3950y
4500y
931y
1860y
14. 6y
7.3y
2.3y
2xl05y
58,600y
1.8xl05y
7
3950y
ll,300y
7700y
5500y
500y
1.46y
0.73y
0.23y
21,200y
38,000h
70,000y
b
465y
1560y
915y
1300y
55. 6y
0.15y
26. 6d
8.4d
2120y
4020y
7090y
9
46. 5y
156y
91. 5y
1.32y
5.6y
5.5d
2.7d
0.84d
212y
402y
709y
10
4.65y
15. 6y
9.2y
13. 2y
0.56y
0.55d
6.5h
2. Oh
21. 2y
40y
71y
11
0.47y
l.Oy
0.92y
132y
20d
1.3h
O.bh
12m
2.1y
4.0y
7.1y
Note: All values at 25°C.
m * minute
h = hour
d " day
y " year
Data Source: 1. Mabey and Mill, 1978
Rate constants In sec
-1.
-------
Table 4-8 lists calculated hydrolysis half-lives as a function of pH for a
number of carbamates. Hydrolysis half-lives for carbamates at pH 7 range from
minutes to thousands of years. Calculations indicate that half-lives can be
reduced to the order of years or less for most carbamates when pH is increased
to the range of 10 to 11. The data in Table 4-8 indicate that base-catalyzed
hydrolysis can contribute significantly to the degradation of a wide variety
of carbamates in waste deposits through control of pH (see also Section 4.4
and Ryckman, 1984).
4.3.7 Phosphoric and Phosphonic Acid Esters
Phosphoric and phosphonic acid esters are hydrolyzed primarily by base
catalyzed processes, resulting in P-0 bond cleavage as illustrated in the
reaction (Mabey and Mill, 1978):
+ ROH
(4-15)
Cleavage of C-0 bonds in these esters may also occur through acid catalyzed or
neutral processes (Mabey and Mill, 1978). Table 4-9 presents a compilation of
hydrolysis half-lives as a function of pH for a variety of phosphonic and
phosphoric acid esters, many of which are of environmental significance as
pesticides and chemical warfare agents.
Half-lives for hydrolysis at pH 7 are generally in the range of years to
thousands of years for many of these compounds, although half-lives on the
order of days apply to several. Since base-catalyzed hydrolysis is the
dominant mechanism for almost all of these compounds in the pH range of 5 to
11, increasing pH in a waste deposit can be expected to have a marked effect
on their degradation. With few exceptions, pH values in the range of 9 to 11
result in calculated hydrolysis half-lives of one year ox less for phosphoric
and phosphonic acid esters. For this reason, control of pH in a waste deposit
appears to be a feasible method of in situ treatment for these compounds.
176
-------
TABLE 4-8
HYDROLYSIS OF CARBAMATES
Compound
C2H50(CO)N(CH3)C6H5
C6H50(CO)N(H)C6H5
CgHsOCCOM CH3) CgHs
C6H5
m-ClC6H4<)C(0)N(H)-
C6H5
p-N02C6H40C(0)N(H)-
^6"5
p-N02C6H40C(0)N-
(CH3)C6H5
1-C10H90C(0)N(H)CH3
1-C1UH90C(0)N(CH3)2
(C2H3)2NCH2CH2OC-
(0)N(H.)C6H5
( U2H3) 2NCH2CH2OC-
(0)N(H)CbH3(CH3)3
CH3
(CH3)3NC6H40C(0)N-
(CH3)2
C1CH2CH2OC(0)N(H)-
C128HCH2OC(0)NHC6H5
CC13CH2OC(0)NHC6H5
CF3CH2OC(0)NHC6H5
Ethyl carbamate
C2H50C(0)NHCH3
C2H50C(0)N(CH3)2
Date Source kA
1
1
1
I
1 . -
1
1
1
1
1
1
1
1
I
1
1
1
1
2
2
2
kN kB
5.5E-5
5.0E-6
5.42E1
4.2E-5
2.5E1
1.8E3
2.7E5
8.0E-4
9.4EO
1.8E-11
2.bE-i
9.4E-7
6.7E-1
2.8E-4
1.6E-3
5.0E-2
3.2E-1
l.OE-1
-
-
t 1/2 at pH
5
4xl05y
4.4xl06yr
150d
5.2xl05y
320d
4. 3d
43m
27,5UOy
2.3y
12UOy
B.ixlU-'y
2.3xlU7y
33y
89,500y
14,000y
440y
69y
220y
_
_
6
40,000y
4.4xlOSy
15d
52,000y
32d
llh
4.3m
2750y
83d
12UOy
83,UOUy
2.3xl06y
3.3y
7850y
1400y
44y
6.9y
22y
_
7
4,000y
44,000y
1.5d
5200y
3.2d
l.lh
26sec
27iy
8. 3d
12UOy
84UUy
2.4x10^
12Ud
785y
140y
4.4y
252d
2.2y
ll.OOOy
38,000y
39,000y
8
400y
4,400y
3.6h
520y
7.7h
6.4m
2.6sec
27. 3y
2Uh
12UOy
830y
23,OUOy
12d
78. 5y
14y
160d
25d
80d
„
_
"
9
40y
440yr
21m
52y
46m
39sec
U.3sec
2.7y
2h
1200y
83y
23UUy
1.2d
7.9y
1.4y
16d
2. 3d
8d
_
"
10
4y
44y
2m
5.2y
4.6m
3.9sec
U.03sec
lUOd
1.2m
12UOy
8.iy
2JUy
2.9h
268d
30d
1.6d
6h
20h
_
"
11
146d
4.4y
13 sec
191d
28sec
0.4sec
0.003sec
lUd
Msec
120Uy
3lua
23y
1.7m
29d
3d
36m
2h
_
"
Mote: All values at 25° C. Rate constants in sec"^-.
m a minute
h *= hour
d • day
y = year
Data Sources: 1. Mabey and Mill, 1978
2. Redding et al., 1977
-------
TABLE 4-9
HYDROLYSIS UK PHOSPHORIC AND PHOSPHON1U ACID ESTERS
CO
Compound Data Source KA KN KB
CH3PCOXOCH3)2
CH3P(0)(OCH(CH3)2)2
CH3P(0)(OC2H5)?0-p-
C6H4N02)
C2H5P(0)(OCH(CH3)2)2
CgH5P(0)(OC2H5)2
(CI130)3PO
(C2H50)3PO
(C2H5S)3PO
(C6H50)3PO
(C2H50)P(OX-p-
C6H4N02)
(p-C6H4N02)3PO
(CH30)2P(S)p-
C6H4N02
CH3OP(S)SCHCH-
(C02C2H5)2
CC2H50)2P(SXp-
b 4 2
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1.36E-9
1.7E-9
6.4E-9
1.2E-7
3.2E-9
1.1E-9
-
-
-
-
-
-
-
-
_
-
-
-
-
1.8E-8
4E-9
1.4E-9
2.7E-11
3.3E-6
l.OE-3
1.1E-7
-
3E-9
2.5E-3
2.2E-4
3.2E-7
4.0E-2
3.7E-8
5E-4
1.3E-4
8.2E-6
1.2E-2
1.7E-2
5.3E-1
3.43E-1
5.95E-3
4.3E-0
2.2E-4
5
8700y
93,OUOy
3.4xl05y
530y
6.9xl05y
43,000y
1.2y
5.5y
Iby
550y
2d
12m
73d
5y
7y
b
880y
9980y
2.3xl06y
55y
6.2xl06y
4400y
1.2y
5.5y
14y
112y
2d
12m
73d
187d
7y
t 1/2
7
88y
990y
b.6x!05y
5.5y
5.5xL06y
440y
1.22y
5.5y
8.5y
13y
2days
Urn
72d
18d
7yr
at pll
8
«. By
lUOy
b9,OUOy
200d
5.9xl05y
44y
1.2y
5.5y
1.6y
1.3y
2d
llm
b9d
1.9d
7yr
9
321d
lUy
b9UUy
20d
59,UUOy
4.4y
l.ly
5.4y
6bd
47d
9h
9m
47d
4.ih
4yr
1U
32d
iy
69Uy
2d
59UOy
IbOd
2bOd
4. by
7d
3d
3h
3m
lid
27m
321d
11
3.2d
3bd
b9y
4.8h
590y
Ibd
54d
l.By
Ibh
llh
22m
2Usec
1.3d
3m
3bd
Note: All values reported for I=Z25°C. Kate constaats in sec"1.
m ™ minute
h *= hour
d = day
y ™ year
Data Source: 1. Mabey and Mill, 1978
-------
4.3.8 Alkylating Agents, Pesticides and other Compounds
A number of compounds of potential environmental concern, including numerous
pesticides, do not fit conveniently into a single chemical class. However,
hydrolysis rate constants are available for some of these substances, and are
compiled in Table 4-10. While generalities cannot drawn with respect to
structure-reactivity relationships, it is apparent that base-catalyzed
hydrolysis can contribute significantly to the degradation of numerous
pesticides in the pH range of 5 to 11, which can have hydrolysis half-lives on
the order of one year or less at pH 11 or less (Table 4-10). As such, control
of pH in a waste deposit may be capable of increasing the degradation rate of
these ,substances. Additional data upon which to evaluate the potential
hydrolysis rates of other pesticides may be available in the primary
literature, by search on a compound-specific basis.
4.4 Case History of Base-Catalyzed Hydrolysis
A warehouse fire at an agricultural warehouse in Hillsboro, IL led to
contamination of soil and surface waters by a combination of 21 different
pesticides, including carbamates, anilines, pyridines, organophosphates and
benzoic acids (Ryckman, 1984). Bench scale studies were performed to evaluate
potential treatment technologies, which included aeration, evaporation,
alkaline hydrolysis, solar oxidation/photolysis, carbon adsorption and
oxidation with hydrogen peroxide (Ryckman, 1984).
The contaminated surface waters were treated by aeration, solar oxidation,
evaporation and powdered activated carbon adsorption/clarification. Forty
thousand cubic feet of soils contaminated up to depths of 3 feet were
detoxified in situ. Soda ash and powdered activated carbon were disced and
plowed into the soil. Periodic application of soda ash maintained a ph of 9,
and a water mist served to activate the ash. Some of the pesticides were
degraded by soda ash alkaline hydrolysis. The activated carbon mitigated
odors and absorbed agricultural chemicals to prevent further migration. In
addition, the black carbon absorbed solar radiation, thereby elevating soil
179
-------
TABLE 4-10
HYDROLYSIS OF MISCELLANEOUS COMPOUNDS (INCLUDING PESTICIDES)
OO
O
Compound Data Source ^A
beta-Proplolactone
CH2CH2S(02)
Dimethyl sulfate
Bls(chloroaethyl) ether
Phenyldimethyltriazine
Benzoyl chloride
(CH;j);jNCU
CH3OC(0)
Hethoxychlor
Uaptan
Atrazine
Malathion
Parathion
Paraoxou
Diazlnon
Diazoxon
Chlopyrifos
Seyin
Sevin
Baygon
Pyrolam
Dimetilan
P-Nitrophenyl-N-methyl
carbamate
2,4-D,m-butoxyethylester
Methoxychlor
DDT
2,4-D,methylester
1
1
1
1
1
1
1
1
1
1
1
2
2
2
2
2
2
2
2
2
2
2
2
2
2
2
2
_
-
-
-
-
-
-
-
-
-
3.9E-5
4.8E-5
-
-
2.1E-2
b.4E-l
-
-
,
-
-
-
-
2.0E-5
-
-
kN
3.3E-3
2.15E-5
1.66E-4
2.8E-2
2.75E-5
4.2E-2
2.5E-3
5.64E-4
2.99E-8
1.87E-5
7.bE-5
7.7E-V
4.5E-8
4.1E-8
4.3E-8
2.8E-7
l.E-7
-
-
-
-
-
4E-5
2.0E-5
2.8E-8
1.9E-9
KB
-
~
1.48E-2
-
-
-
-
-
3.b4E-4
5.7E2
—
b.SEO
2.3E-2
1.3E-1
5.3E-3
7.bE-b
1E-1
7.7EO
3.4EO
4.6E-1
1.1E-2
5.7E-5
3.0E3
3.02E1
2.8E-4
9.9E-3
1.7E1
5
3.5m
8.9h
1.2h
25sec
7h
Ibsec
4m
20m
270d
lOh
2.5h
l.bh
178d
195d
32d
1.2d
80d
2.9h
6.5y
48y
2000y
3.9xl05y
4.5h
9.6h
28bd
12y
1.3yr
6
3.5m
8.9h
1.2h
25sec
7h
Ibsec
4m
20m
270d
8h
2.5h
128d
177d
190d
125d
9d
79d
104d
23bd
4.8y
200y
39,000y
2.8h
9.5h
28bd
lly
47d
t 1/2
7
3.5m
8.9h
1.2h
25sec
7h
Ibsec
4m
20m
270d
3h
2.bh
14d
170d
149d
176d
23d
73d
lOd
24d
174d
20y
3900y
34m
8.4h
286d
7.6y
4.7d
at pll
8
3.5m
8.9h
1.2h
25 sec
7h
Ibsec
4m
20m
2b7d
20m
2.5h
1.5d
118d
47d
Ib5d
28d
40d
Id
2.4d
17d
2y
390y
3.8m
4.4h
283d
1.9y
llhr
9
3.5m
8.9h
1.2h
25 sec
7h
16 sec
4m
20m
241d
2m
2.bh
3.bh
29d
bd
14d
29d
7d
2.5h
5.7h
1.7d
73d
39y
23sec
3bm
252d
79d
l.lhr
10
3.5m
8.9h
l.lh
25sec
7h
Ibsec
4m
2Um
IZld
12sec
2.5h
21n
3.4d
Ibh
14d
29d
19h
15m
34m
4.2h
7.3d
3.9y
2.3sec
4m
122d
8d
b.8min
11
3.5m
8.9h
l.lh
2bsec
7h
Ibsec
4m
20m
2Ud
Isec
2.bh
2m
tj.4h
l.bh
1.5d
29d
2h
1.5m
3.4m
25m
18h
141d
0.23sec
23sec
20d
19h
41sec
Note: All Values for 25+5°C. Rate constants in sec"1.
m = minute
h = hour
d - day
y = year
Data Sources: 1. Mabey and Mill, 1978
2. Harris, 1982
-------
temperatures and catalyzing pesticide destruction (Ryckman, 1984). Periodic
discing and soil aeration accelerated pesticide degradation by solar oxidation
and volatile evaporation.
4.5 Summary
The discussion above indicates the potential application of base-catalyzed
hydrolysis to accelerate degradation of a variety of organic compounds in
waste deposits. However, it must be recognized that this discussion is based
almost exclusively on data obtained in laboratory studies in controlled
systems. At the present time, only limited investigations of hydrolysis rates
in soil or sediment systems have been conducted (Wolfe, 1983), and there is
very little practical field experience for control of hydrolysis rates in
waste deposits. The data reported in this section should thus be used as a
guide in selection of field situations where control of hydrolysis may provide
a viable treatment alternative. Laboratory bench scale treatability studies
and field pilot tests using site specific soil and waste matrices should be
conducted prior to actual field implementation (e.g., Ryckman, 1984).
Although only one field experience with base-catalyzed hydrolysis has been
reported in the literature (Ryckman, 1984), it is likely that conditions
favorable to hydrolysis can be readily produced in many situations using
available equipment, since the primary reagent is water, which can be
introduced to a waste deposit using methods described in Chapter I (Delivery
Systems). The alkaline conditions required to accelerate hydrolysis rates
could be produced by addition of lime or soda ash to the soil surface followed
by surface application of water (e.g., spraying or ponding). For deeper
deposits, subsurface application of alkaline solutions could be utilized.
It is difficult, considering the scarcity of currently available data, to
assess the potential interference of soil or waste deposit matrices on the
hydrolysis process. Studies on a limited number of compounds in
sediment-water systems conducted by Wolfe (1983) indicate that base-catalyzed
processes are retarded for sorbed organic compounds, while neutral hydrolysis
is unaffected. Hydrolysis rates of compounds in the aqueous phase of
181
-------
sediment-water systems are unaffected. Where much of the waste material is
sorbed to solid phases, base-catalyzed hydrolysis rates may be limited
kinetically by desorption rates which may be slow, since effective hydrolysis
may occur primarily in the aqueous phase. Retardation of hydrolysis through
sorption to solid phases can be expected to be greatest for compounds with a
high octanol-water partition coefficient (this parameter represents a useful
indicator for potential distribution of a compound between the aqueous and
soil phases, see Section 3.3). Sorption of organic compounds in general is
expected to be greatest for soils or deposits with high clay and organic
content, and lowest for sands and gravels. However, the effect of sorption on
hydrolysis can probably be determined quantitatively only through
site-specific testing of a given waste material and solid matrix system.
Hydrolysis appears to present a relatively economical option for long term
treatment of waste deposits since infrequent applications of chemicals (e.g.,
water and bases) can be expected to produce relatively long term modification
of deposit conditions to favor degradation of those chemicals amenable to
alkaline hydrolysis. However, dilution by groundwater flow, adsorption of
atmospheric carbon dioxide, and other sources of acidity to the deposit are
likely to require periodic addition of bases to maintain the desired pH in the
deposit for long term treatment.
Based upon the data and calculations discussed in this chapter, the following
classes of compounds are considered candidates for additional testing of the
feasibility of base-catalyzed hydrolysis as an in situ degradation method:
o Esters,
o Amides,
o Carbamates, ,
o Phosphoric and Phosphoric Acid Esters, and
o Certain Alkylating Agents and Pesticides.
Potential application of base-catalyzed hydrolysis for various classes of
compounds is summarized in Table 4-11.
182
-------
TABLE 4-11
APPLICABILITY OF BASE CATALYZED HYDROLYSIS AS A
TREATMENT METHOD FOR ORGANIC COMPOUNDS
Class of Compounds
Application of Base-Catalyzed
Hydrolysis Indicated
Aliphatic Hydrocarbons
Alkyl Halides
Ethers
Halogenated Ethers and Epoxides
Alcohols
Glycols, Epoxides
Aldehydes, Ketones
Carboxylic Acids
Amides
Esters
Nitriles
Amines
Azo Compounds, Hydrazine Derivatives
Nitrosamines
Thiols
Sulfides, Disulfides
Sulfonic Acids, Sulfoxides
Benzene and Substituted Benzenes
Halogenated Aromatic Compounds
Aromatic Nitro Compounds
Phenols
Phosphoric and Phosphoric Acid Esters
Halogenated Phenolic Compounds
Nitrophenolic Compounds
Fused Polycyclic Hydrocarbons (PNAs)
Fused Non-Aromatic Polycyclic
Hydrocarbons
Heterocyclic Nitrogen Compounds
Heterocyclic Oxygen Compounds
Heterocyclic Sulfur Compounds
Organophosphorus Compounds
Carbamates
Pesticides
(1)
(2)
(3)
(4)
(4)
(5)
+
+
+ (6)
(1) Requires pH above 11. .;.. . ; . <
(2) Hydrolysis rates generally high at neutral pH.
(3) Glycols resistant to hydrolysis; Epoxides hydrolyze readily at
neutral pH.
(4) Groups are potentially hydrolyzable. Available rate data limited.
(5) Sulfonic Acid esters are hydrolyzable.
(6) Application of base-catalyzed hydrolysis is compound specific.
183
-------
References
Fendler, E. J. and J. H. Fendler. 1970. Micellar Catalysis in Organic
Reactions: Kinetic and Mechanistic Implications. Advances in Physical
Organic Chemistry, 8:271-406.
Harris, J. C. 1982. Rate of Hydrolysis. In: Handbook of Chemical Property
Estimation Methods (Chapter 7). Lyman, W. J., W. F. Reehl and 0. H.
Rosenblatt (eds). McGraw Hill, New York, NY.
Mabey, W., and T. Mill. 1978. Critical Review of Hydrolysis of Organic
Compounds in Water under Environmental Conditions. J Phys Chem Ref Data,
7(2): 383-415.
Mill, T. 1979. Structure Reactivity Correlations for Environmental
Reactions. EPA-560/11-79-012, U.S. Environmental Protection Agency,
Washington, B.C. •
Pettyjohn, W. A., J. R. J. Studlick, R. C. Bain, and J. H. Lehr. 1979. A
Ground-Water Quality Atlas of the United States. National Water Well
Association, Worthington, OH.
Radding, S. B., D. H. Liu, H. L. Johnson, and T. Mill. 1977. Review of the
Environmental Fate of Selected Chemicals. EPA 560/5-77-033, U S Environmental
Protection Agency, Washington, B.C.
Ryckman, M.O. 1984. Detoxification of Soils, Water and Burn Residues from a
Major Agricultural Chemical Warehouse fire. In: Proceedings of the 5th
National Conference on Management of Uncontrolled Hazardous Waste Sites.
HMCRI, Silver Spring, MD. pp 420-426.
Stumm, W. and J. J. Morgan. 1980. Aquatic Chemistry. Wiley-Interscience,
New York, NY. 583 pp.
Versar, Inc. 1979. Water Related Environmental Fate of 129 Priroity
Pollutants. Vols I and 11. EPA/440-4-029 a and b, US Environmental
Protection Agency, Washington, B.C.
Zepp, R. G., N. L. Wolfe, J. A. Gordon, and G. L. Baughman. 1975. Dynamics
of 2, 4-D Esters in Surface Waters. Hydrolysis, Photolysis and Vaporization.
Environ. Sci. Technol., 9(13): 1144-1150.
184
-------
SECTION 5
CHEMICAL OXIDATION
Chemical oxidation is a process in which the oxidation state of a substance is
increased, which is equivalent to the loss of electrons by the oxidized
moiety. Although oxidizing agents most often supply oxygen during the
oxidation process, other electron acceptors can be utilized. Examples of
chemical oxidation include the oxidation of formaldehyde by hydrogen peroxide:
2HCHO + H,
/
HCOOH + H,
= 2HCOOH +
2H20
(5-1)
(5-2)
or the oxidation of phenol by ozone:
= 6C0
3H20
(5-3)
This section discusses the application of various chemical oxidation processes
In treating organic compounds in water and waste treatment, and evaluates
their potential application in waste deposit stabilization. The chemical
oxidants evaluated in this report, which may be suitable for in situ
stabilization of organic wastes, are hydrogen peroxide, ozone, and
hypochlorites. The use of these oxidants for treatment of waste and
wastewater is well documented. However, very little published information or
data from manufacturers was found on the application of chemical oxidation for
in situ degradation of organic compounds in waste deposits. Therefore, the
evaluation of the in situ application potential must be regarded as generally
hypothetical and untested in field situations. Because a single oxidizing
agent can oxidize a wide variety of compounds, each at different rates and
producing different oxidation products, bench and pilot-scale studies will be
required to determine the in situ oxidation rates of the contaminants in
question and ensure that undesirable (i.e., toxic) by-products are not
generated.
185
-------
5.1 Hydrogen Peroxide
5.1.1 Properties of Hydrogen Peroxide ;
Hydrogen peroxide (H^O,) *s a weakly acidic, clear colorless liquid, fully
miscible with water. It is commercially available in aqueous solution over a
wide concentration range. Properties of pure hydrogen peroxide and aqueous
hydrogen peroxide at various concentrations are listed in many chemical
handbooks, including Kirk-Othmer (1979).
The major chemical reactions and uses of hydrogen peroxide are based on its
molecular structure which includes a covalent oxygen-oxygen bond. The
principal reaction is oxidation, although some applications involve
decomposition, molecular additions, substitutions and reductions.
reactions of hydrogen peroxide can be expressed as:
These
H2°2
2H2°2
H2°2
Y
H202 H- KX
H2°2
Z =
WO + H20
2H20 + 02
YH2°2
ROOH'+ HX
ZH + 0
(oxidation)
(decomposition)
(molecular addition)
(substitution)
(reduction)
(5-4)
(5-5)
(5-6)
(5-7)
(5-8)
Hydrogen peroxide may react directly or after it has first ionized or
dissociated into free radicals. In the presence of catalysts, particularly
ferrous and ferric ions, hydrogen peroxide is decomposed to hydroxyl and
perhydroxyl radicals. These are very powerful oxidants and are the basis of
the Fenton reaction (Dorfman and Adams, 1973) which is used to effect a
variety of oxidations. The following equations show the pathways of hydroxyl
radical formation:
Fe
2+
3+
°T
H.,0,
Fe + OH + OH (hydroxyl radical)
Fe
2+
H + H02 (perhydroxyl
" radical)
(5-10)
186
-------
Hydrogen peroxide is a moderate strength chemical oxidant compared to
chlorine; its advantage is that hydrogen peroxide does not produce unwanted
and potentially hazardous chlorinated reaction products. However, the
reaction of hydrogen peroxide with high concentrations of some organic and
inorganic wastes can be strongly exothermic (heat-producing). Wastes
containing amines, cyanides, formaldehyde, phenols, ferrous ion or hypochlorite
at much greater than 1000 ppia have shown rapid temperature increases and
possible splattering or explosion due to gas evolution.
5.1.2 Oxidation of Organics by Hydrogen Peroxide
Hydrogen peroxide is used in municipal wastewater treatment to control
hydrogen sulfide generation, promote BOD and COD reduction, and for bulking in
activated sludge plants. In industrial wastewater treatment hydrogen peroxide
is used to detoxify cyanide and organic pollutants including formaldehyde,
phenol, acetic acid, lignin sugars, surfactants, amines and/or glycol ethers
and sulfur derivatives. A wide variety of organic compounds can be oxidized
by hydrogen peroxide. These include aldehydes, amines and amides, phenols,
various nitrogen and sulfur compounds, aliphatic and aromatic hydrocarbons and
others. Table 5-1 lists various chemical classes' reactivity with hydrogen
peroxides and any special conditions required (if known).
5.1.3 Application Potential of Hydrogen Peroxide for In Situ Treatment
At the present time, there is no actual field experience upon which to
evaluate the potential efficiency of hydrogen peroxide in oxidizing chemical
contaminants in waste deposits. As such, laboratory and/or pilot plant
studies utilizing the actual waste deposit matrix to study the effectiveness
of treatment by hydrogen peroxide would be required prior to any actual usage.
The documented application of hydrogen peroxide in treating different classes
of chemical wastes are: aldehydes, phenol, mercaptans, amines, hydroquinones,
hydrogen sulfide, cyanide, sulfides and disulfides. In addition hydrogen
peroxide is known to react with a wide variety of other organic compounds.
187
-------
TABLE 5-1
ORCAHIC CHEMICAL CLASSES
ABILITY 10 REACT HUH HYDROCEH PEROXIDE
00
00
1. Aliphatic Hydrocarbons
2. Alkyl Halides
3. Ethers
4. Halogenated Ethers and Epoxides
5. Alcohols
6. Glycols, Epoxides
7. Aldehydes, Ketones
8. Carboxylic Acids
9. Amides
10. Esters
11. Nitriles
12. Amines
13. Azo Compounds, Hydrazine
Derivatives
14. Nitrosamines
15. Thiols
16. Sulfides, Disulfides
17. Sulfonic Acids, Sulfoxides
18. Benzene and substituted Benzene
19. Halogenated Aromatic Compounds
20. Aromatic Nitro Compounds
YES
x
NO UNKNOWN COMMENTS
x (Saturated alkaneo unreactive; unsaturated compounds fora epoxideo and poly
hydroxy compounds).
10
Requires Fe catalyst; forms acetic acid and CO.,.
May require Fe*2 catalyst and alkaline conditions (pH 9-11);
forms organic acids. Reaction time * minutes.
Forms amides.
Primary amines react to form hydroxylamines, Azo, Azoxy, nitroso and nitro-
compounds; secondary amines react to form di-N-substituted (R2NOH) hydroyxl
amines. Reaction time ™ minutes to hours.
May require catalyst.
May require catalyst; may require low pH, Fe+ catalyst or elevated tempera-
ture (80°C). dialkyl sulfides (KSK. yield sulfoxides; dialkyl disulfides
(RSSR) yield sulfonates.
Requires Fe catalyst; forms phenol.
Conversion of nitrobenzene to nitrophenyl hydroxylamine reported; requires
+2
acetic acid and FE catalyst.
-------
TABLE 5-1 (Cont'd)
HASTE CHEMICAL CLASSES
ABILITY TO REACT WITH HYDROGEN PEROXIDE
YES
NO
UNKNOWN
COMMENTS
OO
21. Phenols x
22. Halogenated Phenolic Compounds x
23. Nltrophenollc Compounds x
24. Fused Polycyclic Hydrocarbons x
25. Fused Non-Aramotic Polycyclic Hydrocarbon x
26. Heterocyclic Nitrogen Compounds x
27. Hetrocyclic Oxygen Compounds x
28. Hetrocyclic Sulfur Compounds
29. Organophosphorus Compounds
30. Hydroquinones x
31. Mercaptans x
32. Olefins x
+2
Requires Fe catalyst and acid conditions (pH"=3-5). Forms organic acids which
can be completely degraded to C02. Reaction time = minutes to hours.
Forms quinones; further oxidation results in ring cleavage.
+2
Requires Fe catalyst and acid conditions (pH 3-5); forms organic acids.
Reaction time = 30 minutes.
Requires alkaline conditions; may require Fe
+2
catalyst; forms sulfonic acids
30-60 minutes.
Form .epoxy derivatives; further oxidized to glycols or polyhyiroxy compounds.
(RS03H) and disulfides (RSSR). Reaction time
-------
The reactivity of hydrogen peroxide with 29 different classes of chemical
wastes, based on the literature search and information provided by the
manufacturers of hydrogen peroxide, is indicated in Table 5-1. For efficient
+2
oxidation of complex organics, a catalyst such as Fe is often required.
It must be recognized, however, that the data presented in Table 5-1 are merely
suggestive of potential applications, and that one or more of the following
factors may limit the application of hydrogen peroxide as an in situ treatment
agent for waste deposits:
1. Hydrogen peroxide may react with explosive force with organic compounds
and/or reducing agents. Therefore, careful determination of waste
deposit chemical characteristics must be performed, and dilute solutions
of peroxide maybe required from the safety standpoint.
2. The potential stability (half-life) of hydrogen peroxide in waste
deposits is unknown. Decomposition of the peroxide may occur during
transport to or through a deposit.
3. Effective treatment of many organic substances with hydrogen peroxide may
require special conditions (pH, temperature or the presence of catalysts)
which may be difficult or impossible to attain in a waste deposit.
4. Hydrogen peroxide may react with organic or inorganic substances present
in the waste deposit other than the target compounds, greatly limiting
treatment efficiency.
5. Treatment with hydrogen peroxide may in some cases result in production
of degradation products more toxic than parent compounds (e.g., epoxides
and nitrosamines).
6. Compounds strongly sorbed to .the solid matrix or insoluble in water may
be difficult to treat effectively.
190
-------
In the event that these potential problems can be overcome at a given site,
the use of hydrogen peroxide as an oxidant offers certain advantages, namely:
1. It is available commercially as a liquid in various concentrations.
2. It is miscible in water at all concentrations, simplifying mixing with
water on the surface and in the waste deposit.
3. The density and viscosity of dilute hydrogen peroxide solutions are close
to those of water, allowing use of standard designs for delivery and
recovery systems.
However, in the absence of field information indicating its effectiveness for
in situ treatment of waste deposits, hydrogen peroxide can only be considered
as a potential treatment reagent whose application awaits further laboratory
and field testing before use at existing waste sites can be contemplated.
Hydrogen peroxide may, however, be used as an oxygen source to promote aerobic
biodegradation (see Section 2 and Wetzel et al., 1985).
5.2 Ozonation
5.2.1 Properties of Ozone
Ozone (0_), a blue gas with a characteristic odor, is a strong oxidizing
agent capable of oxidizing a variety of organic and inorganic compounds.
Ozonation is a common method of waste treatment, but it requires certain
precautions because ozone is an extremely reactive gas. It cannot be shipped
or stored, and must be generated on site immediately prior to its
application. Ozone rapidly decomposes to oxygen in aqueous solutions
containing impurities (such as organics or particulate matter), although the
decomposition proceeds more slowly in pure water or in the gaseous phase.
Figure 5-1 illustrates the decomposition of ozone in different water types at
20°C (68°F). This figure shows that in double-distilled water only 10% of
the ozone is decomposed after 85 minutes (an extrapolated half-life of about 9
191
-------
FIGURE 5-1
DECOMPOSITION RATES OF OZONE
IN VARIOUS WATERS (20oC)
ro
z
o
8
cc
I-
z
ui
O
z
O
U
w
Z
O
N
O
DOUBLE-DISTILLED WATER
10
20
30 40 50
TIME (MINUTES)
60
70
80
SOURCE: HANDBOOK OF OZONE TECHNOLOGY AND
APPLICATIONS, R. RICE AND A. NETZER, 1982
-------
hours), but if organics are present in the water the decomposition rates
increase dramatically (half-lives of about 18 minutes in groundwater and less
than 10 minutes in some lake waters).
Physico-chemical characteristics of ozone can be found in many chemical
handbooks, including Kirk-Othmer (1979) or Masschelein (1982). Its solubility
in water is dependent on equilibrium constants as defined by Henry's Law.
Impurities in water can have a substantial influence on the solubility of
ozone, either increasing or decreasing it.
Table 5-2 summarizes the ability of many waste chemical classes to react with
ozone. Mallevialle (1982) has compiled an extensive review of individual
reaction by-products and precursors for ozonation of a wide variety of
compounds.
5.2.2 Oxidation of Organics by Ozone
Oxidation of organic compounds with ozone can occur along three different
pathways. These pathways are (Masschelein, 1982):
o
o
Direct oxidation of the organic compound by ozone,
Oxidation of the organic by hydroxyl free radicals formed from
decomposed ozone, or
Oxidation reaction induced by interaction between ozone and the
solute.
Each of these oxidation pathways will result in different types of end
products. Therefore, the specific oxidation mechanism of the organic
compounds in question should be known so that undesirable (toxic) compounds
are not produced.
Oxidation rates of solutions of organic' materials are rapid during the early
stages of ozonation, but then the rates slow considerably. This is explained
by both the concentrations of readily oxidizable organic materials becoming
193
-------
TABLE 5-2
ORGANIC CHEMICAL CLASSES
ABILITY TO REACT WITH OZONE
YES
1. Aliphatic Hydrocarbons
saturated
unsaturated X
2. Alkyl Halides
3. Ethers X
4. Halogenated Ethers and Epoxides
5. Alcohols X
6. Glycols, Epoxides
7. Aldehydes, Ketones X
8. Carboxylic Acids X
9. Amides
10. Esters X
11. Nitriles
12. Amines X
13. Azo Compounds, Hydrazine Derivatives
14. Nitrosamines
15. Thiols
16. Sulfides, Disulfides X
17. Sulfonic Acids, Sulfoxides X
18. Benzene and substituted Benzene X
19. Halogenated Aromatic Compounds X
20. Aromatic Nitro Compounds X
21. Phenols X
22. Halogenated Phenolic Compounds X
23. Nitrophenolic Compounds X
24. Fused Polycyclic Hydrocarbons X
25. Fused Non-Aromatic Polycyclic Hydrocarbons
26. Heterocyclic Nitrogen Compounds
27. Hetrocyclic Oxygen Compounds
28. Hetrocyclic Sulfur Compounds
29. Organophosphorus Compounds X
NO
UNICNOWN
X
X
X
X
X
X
194
-------
lower and the organic oxidation products of ozonation being more refractory to
oxidation. Many compounds which are oxidized slowly by ozone will react 100
to 1,000 times faster in the presence of ultraviolet radiation or ultrasonic
energy.
The reaction rate of ozone with organics is also affected by pH. At a high pH
the slower hydroxyl free radical reaction will dominate; thus pH can be used
to control the reaction rate (USEPA, 1984).
5.2.3 Applications of Ozonation
Ozone has been used in the United States and more extensively in Europe for
the treatment of drinking water supplies, municipal wastewater treatment,
industrial waste treatment and in a few isolated cases the treatment of
contaminated groundwater.
5.2.3.1 Drinking Water Treatment
Unlike chlorination, using ozone as an oxidizing agent for the treatment of
potable water does not lead to the formation of undesirable chlorinated
organic substances such as trihalomethanes (THM), which are believed to be
carcinogens. Ozonation is usually not used in conjunction with chlorination
since it has been found that ozonation prior to chlorination increases the
formation potential of compounds such as THM (Katz, 1980).
Ozonation of drinking water has been used successfully for the following
applications (Rice and Netzer, 1982):
Bacterial Disinfection
Viral Inactivation
Oxidation of Soluble Iron and/or Manganese ,
Decomplexing Organically Bound Manganese (Oxidation)
Color, Taste and Odor Removal (by Oxidation of Organics)
Algae Removal
Oxidation of Organics (Phenols, Detergents, Pesticides)
195
-------
Microflocculation (Oxidation) of Dissolved Organics
|
- Oxidation of Inorganics (Cyanides, Sulfides, Nitrites)
- Turbidity or Suspended Solids Removal (Oxidation)
Pretreatment for Further Biological Treatment (Oxygenation of
groundwater or oxidation of complex organics to simpler, more
biodegradable compounds).
I
5.2.3.2 Industrial and Municipal Wastewater Treatment
The major application of ozonation in municipal wastewater treatment has been
for disinfection following primary and/or secondary treatment. It has also
been successfully used for lowering levels of biochemical oxygen demand (BOD)
or chemical oxygen demand (COD); oxidation of ammonia; removal of color,
organics, or suspended solids; and odor control.
Ozonation has successfully been used in treatment of industrial wastewaters
for the following purposes (Rice and Browning, 1981):
— Oxidation of cyanide in electroplating wastewaters
- Decolorization of dye stuffs
- Removal of phenolic compounds
- Recovery and reuse of spent iron cyanide photoprocessing bleach waters
- Treatment of acid coal mine wastewaters
- Processing wastewaters of mixed ore
- Oxidation of organic waste streams. '
Several experiments have been carried out to determine the required doses and
effectiveness of ozonation on various organic compounds. The results have
been compiled in Katz (1982) and are summarized in Table 5-3. It must be
realized that these results can be evaluated only in the context of the
experimental design and conditions of the studies reported in Katz (1982).
Variables which would alter the experimental results include pH, contact time,
ozone dosage, temperature, method of contact, presence of a catalyst, or
% '
presence of other competing compounds.
196
-------
TABLE 5-3
OZONATION OF VARIOUS COMPOUNDS IN WATER
Compound(s)
Petroleum
Gasoline
Benzene
Diethylbenzene
2,2,4-dinitro-
phenol
DDT
Malathion
Methyl-
parathion
Trichloro-
methyl
parathion
Dinitro-
orthocresol
Initial Concn.
(mg/1)
10
50
200
125 - 100
50
3
0.5
10
10
10
0.5 - 1
10
Ozone Dose Final Concn.
(mg/1) (Bg/1)
4.5 0.2 - 0.3
1.29 1
5.1 0.1
20 5
150 - 10 5 - 12.5
100 0.35
17 0.05
14 0
13.8 0.25
3.5 2
9.8 1
2.6 0
4.5 0.5
9.5 0.1
8-10 0.07
3.5 - 4.5 0
5-6 0
Percent
Reduction
97-98
98
99.8
97.5
88-96
99.3
98.3
100
50
80
90
100
95
99
99.3
100
100
197
-------
TABLE 5-3 (Cont'd)
OZONATION OF VARIOUS COMPOUNDS IN WATER
Compound(B)
Hethanol
Ethanol
looamyl alcohol
Glycerine
Hydrazine
Carbon Bisulfide
Hydrogen Sulfide
Phenol
tj-Cresol
Hydroquinone
Salicyclic Acid
Gasoline
Benzene
Toluene
Xylene
Acetone
Initial Concn.
(mg/1)
2000
1000
1000
1000
100
100
10
100
100
100
100
1000
500
500
500
100
Final Conca.
(mg/1)
160
90
80
0
0
0
0
0
0
0
0
0
0
0
0
30
Percent Reduction
92
91
92
100
100
100
100
100
100
100
100
100
100
100
100
70
198
-------
5.2.3.3 Groundwater Pollution Abatement
There are very few cases in which ozonation has been used for groundwater
pollution abatement. The most likely use of ozone in this context would be
for treatment of contaminated groundwater which has been pumped to the
surface, or as an oxygen source for biodegradation. A case history of the use
of ozonation for in situ treatment of contaminated groundwater in Karlsruhe,
West Germany was reported by Nagel (1982) and Rice (1984).
A municipal well field had been contaminated by hydrocarbons from a nearby
railway yard to the north and cyanides from a chemical waste disposal site to
the south (Figure 5-2). The well water had high turbidity, one well had
elevated cyanide concentrations, and oxygen and nitrate were totally consumed
by the contaminants' COD. As a result, two of the four wells in the well
field had been closed and the remaining two were threatened. To protect the
aquifer against further spread of the contaminants and decline in water
quality, an ozone treatment process was developed.
Water from the well contaminated by hydrocarbons was pumped from all levels of
the aquifer to a depth of 25 meters (82 ft). Ozone was produced on site from
3 3
dry air at a concentration of 25 g/m (0.0016 Ib/ft ) and was introduced
to the contaminated groundwater by a static tube mixer, where the organics
reacted in a bubble column and the ozone decomposed to oxygen. The water was
then returned to five fully-penetrating infiltration wells at a rate of 100
3
m /h (26,420 gallons/hr). Figure 5-3 illustrates the water flow diagram for
the ozonation process.
The oxygenated recharge water was recycled via the municipal wells after 30 to
40 days retention in the aquifer. This contact time was sufficient to reduce
the dissolved organic carbon (by biodegradation) from 3.5 mg/1 to 1.5 mg/1 in
the most heavily polluted well within about 2 months after the start-up of
ozonation. As a result the dissolved oxygen content of the groundwater rose
to several mg/1 and biological activity within the aquifer increased. The
water quality was improved sufficiently to allow the well field to remain in
operation.
199
-------
FIGURE 5-2
CITY WATERWORKS
KARLSRUHE WEST GERMANY
450M
MARSHALLING
YARD
WATERWORKS
'DURLACHER WALD
INFILTRATED
WELLS
OZONIZATION
PLANT
FEEDING WELL
SERVICE WELLS
INFILTRATED WELLS
STORAGE OF
CHEMICALS
GROUNDWATER
ELEVATIONS
200
-------
FIGURE 5-3
BAS3C FLOW DIAGRAM
FOR OZONATION OF
GROUND WATER AT KARLSRUHE
INFILTRATION
WELLS
OZONIZED
WATER INJECTION
INFILTRATION
WELLS
201
-------
The application reported in this case history diverges somewhat from strictly
in situ treatment for the following reasons:
o The groundwater was collected and ozonized conventionally (above
ground) to reduce the organic content.
o The ozonized water improved the quality of the recharge water and the
groundwater both as a direct result of oxidation of organics by ozone
and because decomposition of ozone in the recharge water raised the
oxygen content of the groundwater, which in turn promoted native
biological activity.
It should be noted that by the time the recharge water left the ozonation
plant, it contained no residual ozone but had high concentrations of dissolved
oxygen. There was thus no in situ chemical oxidation of organics by the ozone.
5.2.4 Application Potential of Ozone for In Situ Treatment
As illustrated above, ozonization has many applications for water and
wastewater treatment, and in some instances for groundwater pollution
abatement. In no cases, however, have any of the reported applications
involved the direct injection of ozonized water into a waste deposit to
oxidize organic materials. Furthermore, the literature survey did not reveal
any studies to evaluate the effectiveness of subsurface (in situ) ozonation.
The only study of the effect of soil on oxidation of organics using ozone is
an experiment with pesticides (Katz, 1980). It was found that the oxidation
proceeds rapidly in clean water, but significantly slower when humic materials
or soil particles are present. It was suggested in this study that dissolved
organics may be adsorbed onto humic or soil materials and become more
resistant to oxidation. These results indicate that ozonation in soil may be
difficult, and laboratory and pilot-scale experimentation simulating the in
situ conditions will be necessary.
202
-------
It is particularly important to accurately characterize the waste to be
ozonized and to perform tests on that specific mixture of compounds for two
reasons. First, various organic compounds in aqueous solution may compete for
ozone in the oxidation process, since ozone's oxidizing action is
non-specific. This could result in acceptable removal of one compound, no
removal or very slow removal of another compound of equal concern, or removal
of non-toxic natural soil organics but no removal of the more refractory
organics of concern. In addition, it is important to design sufficient
oxidant into the process to acomplish the amount of oxidation desired.
Secondly, the mechanisms of oxidation of the organic compounds originally
present must be understood in order to evaluate what oxidation by-products
might be formed. Some compounds are oxidized first to intermediates which are
more toxic than the starter materials, before being further oxidized to
innocuous compounds. An example of this would be the ozonation of the
pesticides parathion and malathion, which produces paraoxon and maloxon,
respectively (USEPA, 1984). These intermediates are more toxic than the
starter materials, but continued ozonation degrades the oxon intermediates.
Other examples of toxic byproducts include the ozonation of dimethylhydrazine,
2-hydroxyethylhydrazine and benzidine, which produce mutagenic compounds of
varying stability.
Much is still unknown about the mechanism of formation and chemical
characteristics of intermediate products. A better understanding of the
chemistry of the materials to be oxidized would be necessary to determine the
treatability potential of a waste deposit, and to properly design the
ozonation system. The major problem with in-situ treatment using ozonation,
however, seems to be the rapid decomposition of the ozone in aqueous
solution. The half-life of ozone in natural waters is about 10-25 minutes,
which is insufficient for delivery or significant contact time when introduced
into the soil - the ozonized solution would probably decompose before it
reaches the waste deposit. For these reasons, the in situ chemical oxidation
of waste deposits using ozone does not appear to be promising.
203
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5.3 Hypochlorites
5.3.1 Properties of Hypochlorites
Hypochlorites as a chemical class are the reaction products of chlorine with
an alkali. They are used principally as a means of delivering chlorine
without the necessity of handling pure chlorine as a liquid or gas.
Hypochlorites are strong oxidizing agents (stronger than hydrogen peroxide)
and are almost always used in aqueous solution. The two most common forms of
hypochlorite produced commercially are calcium hypochlorite (Ca(OCl) ^) and
sodium hypochlorite (NaOCl). Sodium hypochlorite is usually produced on a
commercial scale in two strengths, 5.25 wt% (household liquid bleach) and
13.03 wt% (commercial strength bleach). Calcium hypochlorite is produced
commercially in a form containing about 70% available chlorine. Other
hypochlorites used occasionally include barium, lithium and alkyl forms; these
are not produced commercially due to poor stability and/or price
considerations.
The three basic mechanisms for the reaction of chlorine with an organic
compound are addition, substitution and oxidation. Addition and substitution
(chlorination) result in the production of chlorinated organic compounds such
as trihalomethanes (THM) which, in most cases, are undesirable. The oxidation
reaction is the principal waste effluent treatment mechanism but is effective
only for a limited number of organic compounds (USEPA, 1979).
In strong solutions at low pH, the chlorination reaction predominates. In
weakly acidic solutions the oxidation reaction is primary. The treatment of
phenols by hypochlorite provides an example of both reactions. Mono-, di- and
tri- substituted phenols are readily formed in solution. These chlorophenols
are subsequently degraded to aliphatic acids by excess hypochlorite in an
oxidation reaction (Eisenhauer, 1964). Table 5-4 summarizes the ability of
various waste chemical classes to react with hypochlorites.
204
-------
TABLE 5-4
ORGANIC CHEMICAL CLASSES
ABILITY TO REACT WITH HYPOCHLORITES
t\3
O
Ol
1. Aliphatic Hydrocarbons
'i. Alkyl Halides
3. Ethers
4. Halogenated Ethers and Eposides
5. Alcohols
6. Glycols, Epoxides
7. Aldehydes, Ketones
8. Carboxylic Acids
9. Amides
10. Esters
11. Nitriles
12. Amines
13. Azo Compounds, Hydrazine Derivatives
14. NItrosamines
15. Thiols
16. Sulfides, Bisulfides
17. Sulfonic Acids, Sulfoxides
18. Benzene, Substituted Benzene
19. Halogenated Aromatic Compounds
20. Aromatic Nitro Compounds
21. Phenols
VES
X
X
NO UNKNOWN
COMMENTS
Possible chlorination and formation of chloramines
Chlorinated product possible
Forms alkylhypochlorites, hazardous and explosive
Used in preparation of Epoxides and Glycols from Halohydrin
Reaction of acetaldehyde yielding Chloroform (CHC13)
Chlorinated byproducts possible
Forms chloramines, hydrolysis of C-N bond, possible NC13 formation
Will not react unless unsaturated bonds are available for chlorohydrin
formation
Will not react unless unsaturated bonds are available for chlorohydrin
formation
Forms chloramines
Forms chloramines
Forms chloramines
Sulfides oxidize to sulfoxides without forming sultones
Forms chlorinated aromatic
Forms chlorinated aromatic, possible oxidation
Forms chlorinated aromatic or chloramine
Forms chlorinated phenols, oxidized to aliphatic acid
-------
1'ABLK 5-4 (Cont'd)
ro
o
YES
22. Halogenated Phenolic Compounds X
23. Nitcophenolic Compounds X
24. Fused Polycyclic Hydrocarbons (PNA's) X
25. Fused Non-Aromatic Polycyclic Hydrocarbons X
26. Heterocyclic Nitrogen Compounds
27. Heterocyclic Oxygen Compounds
28. Heterocyclic Sulfur Compounds
29. Organophosphorous Compounds
WASTE CHEMICAL CLASSES AND THK1K
ABILITY TO REACT WITH HYPOCHLORITES
NO
UNKNOWN
COMMENTS
Oxidized to aliphatic acid
ChlorinaCion of aromatic ring
Chlorinated and oxidized products (e.g., phenols and quinollnes)
Chlorinated product formed
Chlorinated product formed
Chlorinated product formed
-------
5.3.2 Treatment Applications of Hypochlorites
The major uses of hypochlorite include disinfection of potable water supplies
and sewage effluents, control of algae and biofouling organisms, and bleaching
of textiles and pulp and paper products. Hypochlorite has had some usage in
industrial waste treatment, primarily as an oxidizing agent for cyanide and
ammonium sulfide/sulfite wastes. Other uses include taste and odor control
(e.g., by oxidation of reduced sulfur or chlorophenols), and removal of
reduced iron and manganese species in water (White, 1978).
Chlorine substitution and addition appear to be the most common reactions
between chlorine and organics in aqueous solution. Only in a few cases do
these reactions proceed beyond this stage, in which they may be considered
"oxidative degradations". Thus treatment of organic chemicals in wastewater
using hypochlorite appears to have limited potential because the intermediate
products are often at least as toxic as the original waste material. This is
also true for treatment of drinking water supplies, where THM are reportedly
formed from humic acids and other naturally-occurring organic materials
(Jolley et al., 1978; Stevens et al., 1978; Rock, 1980).
The production of numerous chlorinated byproducts in waters treated by
chlorination has been reported at low concentrations of chlorine addition,
typical of municipal water and wastewater drinking effluents. Carlson and
Caple (1978) reported the substitutive chlorination of phenol, anisole,
acetanilide, and toluene under acid conditions, and Snider and Albey (1980)
reported the chlorination of biphenyl to mono- and dichlorinated biphenyls,
although the rate of reaction was slow above pH 6.2. Bieber and Trehey (1983)
reported formation of dichloroacetonitriles through chlorination of natural
waters. Both chlorinated and oxidation byproducts (including phenols and
quinones) result from chlorination of polynuclear aromatic materials
(Liukonnen et al., 1983). Ghanbari et al. (1983) reported incorporation of
chlorine into fatty acids, fatty acid esters and triglycerides.
207
-------
Increased chlorine doses and contact times can be expected to increase
formation of chlorinated byproducts. Heavy chlorination (2000 - 4000 ppm of
hypochlorite) of municipal wastewater has been reported by Glaze et al. (1978)
to result in substantial increases in chlorinated byproducts. Under these
conditions, chlorinated byproducts were formed from non-activated substances
(e.g., benzene, toluene, benzyl alcohol) which are generally not observed as
byproducts of chlorination at lower doses.
Disinfection with chlorine is well established as a public water supply
treatment but the utility of hypochlorite as an oxidant for organic substances
in water and wastewater remains doubtful. The only organic wastes that have
been treated successfully by oxidative degradation with hypochlorites are
phenols and chlorinated phenols. The degradation mechanism leads to the
formation of aliphatic acids by cleavage of the aromatic ring. Evidence of
other successful organic waste stream treatments by oxidative degradation
remains extremely limited.
An important application of hypochlorite oxidation for inorganic waste is that
of cyanide waste stream treatment. Cyanide is first oxidized to the less
toxic cyanate and then to harmless bicarbonates and nitrogen. This process
is capable of achieving an efficiency of 99 percent.
5.3.3 Potential for In Situ Treatment of Waste Deposits Using
Hypochlorite
The principal uses of chlorination have been for biological treatment
(disinfection) of water, wastewater, sewerage and for cleaning swimming
pools. The potential use of hypochlorites (in aqueous solution) for in situ
treatment o± organic wastes is, at best, extremely limited because the chief
products of chlorination are usually undesirable chlorinated organics (Table
5-4). ihe greatest potential use of chlorination for organic waste treatment
resides with phenols and phenolic compounds, where documented oxidative
degradation to aliphatic acids has been achieved, or with cyanides (see
below). However, control of conditions in a waste deposit to achieve this
degradation would be difficult. This information indicates that, except for
208
-------
some specific situations, use of aqueous solutions of hypochlorites is not
generally adviseable for in-situ treatment of organic chemicals due to the
possible formation of chlorinated organics, as well as the lack of available
information on in-situ treatment using hypochlorites.
5.3.3.1 In Situ Oxidation of Acrylonitrile Using Sodium Hypochlorite
A freight train derailment in Ohio led to the spillage and burning of 31600
liters (8360 gallons) of acrylonitrile (CH2CHCN) (Harsh, 1978). Following
the initial cleanup it was decided to oxidize the remaining acrylonitrile in
in surface ponds and soils by first raising the pH of the contaminated area
above 10 using lime, and oxidizing the cyanide portion of the acrylonitrile
molecule using sodium hypochlorite (HTH). The reaction would proceed in three
stages (Harsh, 1978):
1) CN + HOC1 = CNC1 + OH
2). CNC1 + 20H~ = CNO~ + Cl +
2CO.
3) 2CNO + 30C1 + H20
3C1
20H
A total of 4360 kg (9600 Ibs) of lime was spread over the area first to raise
the pH. Then a water solution containing 410 kg (900 Ibs) of HTH was sprayed
over the area, and an additional 180 kg (400 Ibs) of HTH was applied to
acrylonitrile pools (Harsh, 1978). Workers were forced to wear gas masks
because of strong chlorine gas fumes. Subsequent monitoring indicated no
residual acrylonitrile (Harsh, 1978).
209
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References
Bieber, T. L. and M. L. Trehey. 1983. Dihaloacetonitriles in Chlorinated
Natural Waters. In: Water Chlorination: Evironmental Impacts and Health
Effects, R. L Jolly (Ed.), Ann Arbor Science, Ann Arbor, MI. 4(1): 85-96.
Bower, E. J., M. Reinhard, T. Everhart, and P. L. McCarty. 1980. Organic
Materials Formed Through Decolorization of Coffee Wastewater With Chlorine and
Chlorine Dioxide. In: Water Chlorination: Environmental Impacts and Health
Effects, R. L. Jolley (Ed.), Ann Arbor Science, Ann Arbor, MI. 3: 315-323.
Carlson, R. M. and R. Caple. 1978. Organochemical Implications of Water
Chlorination. In: Water Chlorination: Environmental Impact and Health
Effects, R. L. Jolley (Ed.). Ann Arbor Science, Ann Arbor, MI. 1:65-75.
Dorfinan, L. M., and G. E. Adams. 1973. Reactivity of the Hydroxyl Radical in
Aqueous Solution. US Department of Commerce, National Bureau of Standards,
Washington, D.C.
E I duPont de Nemours and Company, Inc.
Handbook. Wilmington, Delaware.
Hydrogen Peroxide, Waiste Treatment
Eisenhauer, H. R. 1964.
Control Fed. J. 36:1124.
Oxidation of Phenolic Wastes. Water Pollution
FMC Corporation, Industrial Chemical Group. Industrial Waste Treatment with
Hydrogen Peroxide. FMC Corp., Philadelphia, PA.
Ghanbari, H. A., W. B. Wheeler, and J. R. Kirk. 1983. Reactions of Chlorine
and Chlorine Dioxide with Free Fatty Acids, Fatty Acid Esters and
Triglycerides. In: Water Chlorination: Environmental Impacts and Health
Effects, R. L. Jolley (Ed.) Ann Arbor Science, Ann Arbor, MI. 4(1):167-177.
Glaze, W. H., J. E. Henderson, IV, and G. Smith. 1978. Analysis of New
Chlorinated Organic Compounds Formed by Chlorination of Municipal Wastewater.
In: Water Chlorination: Environmental Impacts and Health Effects. R L Jolley
(Ed.), Ann Arbor Science, Ann Arbor, MI. 1:139-159.
Harsh, K. M. 1978. In Situ Neutralization of an Acrylonitrile Spill. In:
Proceedings of the Conf. on Control of Hazardous Materials Spills. HMCRI,
Silver Spring, MD. pp 187-189.
Jolley, R. L., G. Jones, W. W. Pitt, and James E. Thompson. 1978.
Chlorination of Organics in Cooling Waters and Process Effluents. In: Water
Chlorination: Environmental Impact and Health Effects. R. L. Jolley (Ed.),
Ann Arbor Science, Ann Arbor, hi. 1:105-138.
Katz, J. 1980. Ozone and Chlorine Dioxide Technology for Disinfection of
Drinking Water. Pollution Technology Review No. 67, Noyes Data Corp., Park
Ridge, NJ.
210
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Oyler- "• T-
. »• A. cox, z. j. *„,
. L.
, Ann Arbor Science, Ann Arbor, Ml. 4(1): 151-0185
Masschelein, W. 1982. Ozonization Manual for Water and Wastewater
Treatment. John Wiley & Sons, New York, NY. Wastewater
. L. Jolley (Ed.), Ann Arbor Science, Ann Arbor, MI. 1:21-35.
u a
Treatment, W. Masschelein (Ed.), John Wiley & Sons, New York, NY
; Browning. 1981. Ozone Treatment of Industrial Wastewater.
Technology Review No. 84, Noyes Data Corp., Park Ridge, NJ.
Rook J J. 1980. Possible Pathways for the Formation of Chlorinated
£t- B vDUr-inS Chlorination °f Humicacids and Resorcinol ! In:
nation: Environmental Impacts and Health Effects. R. L Jollev
.), Ann Arbor Science, Ann Arbor, MI. 3:85-98. Jox±ey,
Schumb W c., C. N. Satterfield and R. L. Wentworth. Hydrogen Peroxide
Reinhold Publishing Co, New York, NY nyuiugen reroxiae.
Snider, E H and F. C. Albey. 1980. Kinetics of Biphenyl Chlorination in
Aqueous Systems m the Neutral and Alkaline pH Ranges. In: Water
Chlorination: Environmental Impacts and Health Effects, R. L? Jolley (Ed )
Ann Arbor Science, Ann Arbor, MI. 3:219-225. JOJ.j.ey t^d.J,
- Seeger,-and G. G. Robeck. In: Water
Effects' R- L-
in
Effects
211
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SECTION 6
APPLICATION AND DESIGN OF SYSTEMS TO ACCELERATE
STABILIZATION OF WASTE DEPOSITS
6.1 Introduction
?
Federal (CERCLA) and state hazardous waste programs to date have focused
primarily on the need to identify uncontrolled waste deposits, perform
remedial investigations (RIs) to assess the nature and extent of threats
presented by these deposits, and undertake immediate remedial measures (IRMs),
if necessary, to reduce significant public health threats. The next step in
waste deposit remediation, long term mitigation of the waste hazard, has until
recently consisted predominantly of containment or excavation and off-site
management rather than waste treatment or detoxification at the site.
Currently, however, attention is being directed to the performance of
feasibility studies (FS) on the potential for in situ stabilization of waste
deposits (USEPA, 1984a; USEPA, 1985). |
The purpose of this report is to present a systematic review of the potential
for in situ stabilization of organic waste deposits. Of utmost importance, it
must be remembered that:
o The process technologies for in situ treatment have not been
established with confidence. In fact, there is very little
documentation of field pilot and full-scale testing in this regard
with the exception of biodegradation (information gaps with regard to
in situ stabilization are discussed further in USEPA, 1984a).
o The delivery/recovery systems for implementing in situ treatment
methodologies can probably be adapted from other existing
applications (i.e., wastewater treatment, ground water collection and
aquifer management, construction site dewatering, subsurface
injection of waters and grouts, irrigation engineering). However,
212
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these applications have had limited objectives, have been utilized in
benign applications where the implications of threat to public
health, welfare and the environment have not been an issue, and have
generally not been tested for in situ waste treatment.
o The feasibility and cost-effectiveness of conceptual delivery/
treatment/recovery systems have in general not been established
(except for a few cases such as the Biocraft Site, described in
Section 6.5), especially as these relate to achieving different
degrees of remediation within required time frames.
This report must therefore be viewed only as a guidance document with respect
to potential technologies for in situ waste stabilization as they currently
exist, that is in their conceptual or developmental stage. The following
sections describe the steps required for the evaluation of biodegradation,
surfactant-assisted flushing, hydrolysis and oxidation applications for in
situ treatment of wastes. In addition, a methodology for selection of
delivery and recovery systems is presented. The processes of defining the
remedial objectives for a site and selection of possible alternatives
(including in situ stabilization), which is a prerequisite in the National
Contingency Plan (NCP) procedure (USEPA, 1985) before in situ stabilization
can be considered, and of comparing alternatives and selection of the remedial
action(s) to be implemented, are described in A.D. Little (1983), USEPA
(1984a), and Repa and Kufs (1985).
6.2 Remedial Investigation
Definition of the nature and extent of the wastes at the site, and the
geohydrologic and geochemical conditions of the site, is a necessary
prerequisite to the evaluation of the feasibility of any remedial approach (in
situ stabilization or any other actions). These data are usually collected
213
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during the remedial investigation (RI) stage of site evaluation, prior to the
feasibility study (FS) stage, during which the evaluation of in situ
stabilization technologies described in this report would take place.
The extent of contamination is determined from the area and depth occupied by
the wastes, and the chemical composition and concentrations of the various
waste components. The waste forms (free liquid, solid, contaminated soils)
should also be determined. This information may be available from the site
operator/owners or from regulatory agencies; otherwise a site investigation
will be required to obtain or confirm the information. Geohydrological
parameters (such as site stratigraphy and topography, soil types,
permeabilities, infiltration rate, and groundwater depth and flow direction),
as well as meteorological and local land use characteristics, should be
determined as required to fully characterize the site for the design of the
necessary delivery/recovery systems. The physical and chemical characteristics
of groundwater, which can affect the feasibility of the treatment technique,
should also be determined. These parameters should include pH, temperature,
and inorganic and organic chemical composition. In addition, a risk
assessment would be required to determine the potential routes of exposure and
risk to humans and the environment, and therefore the levels to which
remediation would be required. Assessment of waste, soil and groundwater
characteristics as well as local site conditions is further described in USEPA
(1984a) and Repa and Kufs (1985).
6.3 Feasibility Study
After the site conditions and contaminant characteristics have been defined
during the Remedial Investigation, a Feasibility Study, in which the various
in situ stabilization methods are evaluated, can be undertaken (USEPA,
1984a). The first step is to evaluate, using the information presented in
this report, whether any of the contaminants present may be susceptable to in
situ biodegradation, surfactant-assisted flushing, hydrolysis or chemical
oxidation. Table 6-1 summarizes the potential applications of these methods
to waste materials. If any of these methods appear to be promising, the
method(s) are further investigated as described in the following subsections.
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TABLE 6-1
POTENTIAL APPLICATIONS OF TREATMENT METHODS TO WASTE CONTAMINANTS
Treatment Technology
ro
i—"
en
Chemical
Class
Base-Catalyzed
Aliphatic Hydrocarbons
Alkyl Halldes
Ethers
Halogenated Ethers and Epoxides
Alcohols
Glycols/Epoxides
Aldehydes, Ketones
Carboxyllc Acids
Amides
Esters
Nitrlles
Atnine s
Azo Compounds, Hydrazine Derivatives
Nltrosamlnes
Thlols
.Sulfides, Dlsulfldes
Sulfonic Acids, Sulfoxldes
Benzene S Substituted Benzene
Halogenated Aromatic Compounds
Aromatic Nltro Compounds
Phenols • ." '
Halogenated Phenolic Compounds
Nltrophenollc Compounds
Fused Polycyclic Hydrocarbons
Fused Non-Aromatic Polycycllcs
Heterocycllc Nitrogen Compounds
Heterocycllc Oxygen Compounds :
Heterocycllc Sulfur Compounds
Organophosphorus Compounds
'. Carbamates
Pesticides
Blodegradation
Hydrolysls
(1)
Oxidation^ Hater Flushing^
Surfactant Flushing^
?
?
W Based upon calculated half-lives for hydrolysis of representative compounds at pH 9 to 11.
<2> Based upon literature for oxidation of chemicals in water and wasteuater by hydrogen peroxide.
SDecific°aDn?^MnS°1"?i1ity a;d °ctanol/water partition coefficient (Kow) of representative compounds.
apecinc application will depend on solubility and Kow for specific compounds.
+ = can be used
- = cannot be used
? = Further research necessary ' . -
-? •= Probably cannot be used
+? - Probably can be used
-------
There are certain potential problems or concerns which must be addressed when
considering any in-situ treatment system. These problems must be analyzed on
a site-specific and treatment-specific basis, and can only be discussed in
general terms here. Primary among these is the problem of waste
heterogeneity, both with respect to irregular contaminant distribution and
inhomogeneous waste/soil physical properties (hydraulic conductivity,
stratification, depth to water table). This may lead to incomplete saturation
of the waste deposit and poor or incomplete exposure of the contaminants to
the reactant solution. Another potential problem relates to excessive
dilution or attenuation of the reactant solution, which may occur if large
volumes of water are required to saturate the deposit, if groundwater flow
through the deposit is rapid, or if the delivery system does not apply the
reactant directly to the waste deposit (e.g., by using injection wells) so
that it is attenuated or diluted during passage through overlying or
upgradient soils. ,
All materials to be used in the delivery and recovery systems, particularly
pipes, pump internal parts, and liners, must either be tested or checked with
the manufacturer or available literature for compatibility with the treatment
solution(s) and waste components. This should take place during the bench
scale treatability or field pilot study steps. Potential materials that can
be used for delivery/recovery piping include polyvinyl chloride (PVC),
stainless steel, carbon steel, polyethylene, styrene rubber, ABS
(acrylonitrile-butadiene-styrene), concrete, fiberglass, bituminized fiber, or
clay. It is probable that one or more of these materials will meet the
chemical and physical requirements of the specific application.
Finally, it must be recognized that the waste may consist of a mixture of
compounds with varying treatability by these methods; thus more than one
in situ treatment method may be required. In addition, other remedial
actions (such as excavation and treatment at the surface) may be required to
treat the concentrated source material, while in situ methods are used to
treat the more extensive, but lower concentration, plume. Furthermore, a
particular technique (e.g., hydrolysis) may detoxify certain compounds, but
alter others into more toxic forms or produce toxic intermediate compounds or
216
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by-products. These possibilities must be carefully evaluated during the
feasibility study on a compound-by-compound basis.
6.3.1 Evaluation of Biodegradation for In Situ Stabilization of Waste
Deposits
A systematic approach is developed in this section, based on the discussion
presented in Section 2 and data in Appendix A, to evaluate the utilization of
biodegradation as an in situ method to renovate waste piles or deposits. As
shown in Figure 6-1, this approach consists of eight steps, each of which is
described below.
The preliminary steps are to understand the nature and extent of wastes at the
site, the site geohydrologic parameters and groundwater and soil chemistries.
These steps are performed during the RI (see Section 6.2).
6.3.1.1
Determine Nutritional and Biological Characteristics of the Wastes (Step I)
This step should be performed during the waste characterization step of the
RI. The objectives of this step are 1) to determine certain environmental
factors which affect the selection of proper microbes for in situ renovation;
2) to quantify the basic nutrients available at the site for supporting the
selected microbes (see Steps II and III; and 3) to identify the microbial
community at the waste site so that certain native microbes may be considered
for use in site remediation, and any predators of the selected microbes may be
identified. The physical and chemical parameters which should be measured
are pH, temperature, porosity, and moisture content of the soil, and redox
potential, phosphorus, nitrogen and trace metal concentrations in the
groundwater.
217
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FIGURE 6-1
EVALUATION OF BIODEGRADATION
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-------
In addition, if the wastes are in solid or particulate form (e.g., spent
resins), permeabilities of these wastes should also be determined via in situ
methods or in the laboratory as appropriate. These data are needed for
designing the delivery/recovery systems (see Section 6.4) for transporting
microbes, water, nutrients, and oxygen to the wastes.
6.3.1.2 Identify Potentially Applicable Microbes (Step II)
Certain microbes potentially capable of degrading the organic materials in the
waste deposit can be identified by review of Appendix A of this report. An
updated literature review should also be performed on biodegradation of these
organics because of the rapidly-evolving nature of this field. If no single
species of microbes can be identified, this evaluation should continue on to
Step V (i.e., skip Steps III and IV).
6.3.1.3 Assess Presence and Absence of Limiting Factors (Step III)
In this step, the optimum growth conditions of the potentially-applicable
microbes identified in Step II are compared with waste site conditions
determined in Step I. Through these comparisons, the physical and chemical
factors at the waste site capable of limiting the growth of the microbes
(e.g., oxygen availability - surface vs subsurface or aerobic vs anaerobic; pH
acidic vs alkaline; temperature; presence of toxins; nutrients) can be
identified. If these limiting factors cannot be corrected (e.g., adding
oxygen, buffer solution, fertilizers, etc.), the microbes identified would
have to be eliminated from further consideration.
6.3.1.4 Availability of Microbes (Step IV)
Commercial firms which culture specialized microbial strains for biological
treatment of specific organics, or which enhance and adapt native microbes to
more efficiently degrade the identified organics, should be contacted to
determine the commercial availability of the microbes identified in Step II,
or any new microbial strains capable of degrading the wastes. If such
219
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microbes are available, this evaluation should move on to Step VI; otherwise
the research described in the following step would be needed to develop
microbes capable of biodegrading the identified organics.
6.3.1.5 Conducting Research to Develop New Microbial Strains (Step V)
Microbes with the required characteristics may be developed by use of native
microflora (identified in Steps I and II), adapted microorganisms, or
genetically modified microbes (i.e., through specific gene mutation or
genetically assisted molecular breeding). These microbes should be developed
for optimal growth at the specific physical and chemical conditions at the
waste site as well as the waste concentrations. To develop new microbial
strains it will be necessary to employ trained microbiologists or to utilize
the services of a firm or university having the specialized expertise required
to develop new strains. Such testing may extend well beyond the expertise
usually associated with sanitary engineering applications to municipal sewage
treatment systems. !
The microbes should be tested to determine whether they can effectively
degrade the identified organics. The potential for inhibition of microbial
growth by the wastes or the site environment should also be evaluated. If
microbes which are effective under in situ conditions can be developed, this
evaluation should continue to the next step (Figure 6-1). Otherwise, this
evaluation should be terminated, and alternative in-situ treatment methods
should be investigated.
The potential for adverse environmental impact of the selected microbes
(either enhanced native microbes, adapted microbes, or new genetically
engineered strains) should be analyzed. Such impacts may include, but not be
limited to, groundwater contamination and subsequent human health effects, and
changes in the local microbial community and soil conditions i so that the
reclaimed land may not be capable of sustaining the original vegetation. The
most suitable microbes for in situ biodegradation of organics are those which
produce no health impact to human beings, and are very specific for the
220
-------
identified wastes (that is, they would not degrade other organics in the soil
or compete for nutrients with other microorganisms, and they will expire when
the wastes are no longer available).
6.3.1.6 Laboratory (Bench Scale) Simulation Tests (Step VI)
Those microbes which are acceptable based on the risk analysis and assessment
of in situ limiting factors would advance to laboratory simulation tests. The
purpose of these tests is to determine the maximum biodegradation rate which
can be achieved under simulated in situ conditions. In this simulation,
proper modification of site conditions to improve the biodegradation rate
should also be considered. These modifications may include:
o Addition of buffers to adjust pH,
o Addition of fertilizers to provide adequate nutrients,
o Addition of emulsifiers to solubilize the wastes,
o Addition of water to adjust the moisture content, and
o Addition of oxygen to support the aerobic microbes.
Based on the results of these tests, the following parameters should be
determined for the maximum biodegradation rate:
o
o
o
o
o
o
o
Microbial concentration,
Substrate concentration,
Buffer concentration and dosing rate (if needed),
Nutrient concentration and dosing rate (if needed),
Emulsifier concentration and dosing rate (if needed),
Water dosing rate (if needed), and
Oxygen concentration and dosing rate (if needed).
These data are required for conducting on-site pilot tests. In addition, any
end products and side reaction products in the soil or in the recovered
solution should be identified to check whether the waste organics are in fact
221
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biologically degraded, and whether any other toxic intermediates or
by-products are generated, which would require additional treatment. Any
gases generated should also be quantified, especially for the case of
subsurface waste deposits, which may require venting.
I
The simulation tests for surface waste piles may be performed as beaker or
simple microcosm tests. For the case of subsurface waste deposits, the
simulation tests can be performed as simple column tests or as more
complicated microcosm systems to better represent the waste site conditions.
If more than one microbial strain has promising characteristics, the
simulation tests should be performed first on the strain which produces higher
degradation rates and can grow better under the waste site conditions.
Whether the other strains should be tested will depend on the schedule and
budget of a given in situ renovation project.
6.3.1.7 Onsite Pilot Test (Step VII)
In this step, a representative plot at the waste site (either surface pile or
subsurface deposit) should be selected for conducting a pilot test to confirm
the results obtained from the laboratory simulation tests (mainly, the maximum
biodegradation rate and its associated physical and chemical requirements).
In addition, the spray systems and tilling operation (for the case of surface
piles) or delivery/recovery systems (for the case of subsurface deposits)
designed for this pilot test would provide design and operational guidance for
the full-scale treatment facility. A further description of pilot-scale
testing, including sampling, analysis and monitoring methodologies, is
provided in USEPA (1984a) and references therein. The delivery/recovery
system which may be used for this application are described in Section 1 of
this report.
222
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6.3.1.8 Facility Conceptual Design and Cost-Effectiveness Assessment
(Step VIII)
After the onsite pilot test has been run for a sufficiently long time to
demonstrate that the microbes are capable of in situ biodegradation of the
waste materials and that they will not produce any adverse environmental
impacts, the full-scale treatment system for the waste deposit can be
conceptually designed. Order-of-magnitude cost estimates should be performed
according to the conceptual design (i.e., costs for detailed design,
engineering and construction of the treatment system; operation and
maintenance costs; and costs for culture and preparation of the microbes for
storage, transport and in situ application). These costs should be compared
with those of the other remediation alternatives (if any) to determine whether
this method should be implemented (USEPA 1984a; USEPA, 1985; Repa and Kufs,
1985).
6.3.2 Evaluation of Flushing and Surfactants For Waste Deposit
Stabilization
Figure 6-2 illustrates a systematic approach to identifying Commercially
available surfactants (single or a mixture) which can effectively mobilize the
organic contaminants at a given waste disposal site for further in situ
treatment or for recovery and subsequent surface treatment. As shown in this
figure, this approach consists of five steps; each of these steps is described
below.
The preliminary steps are to characterize the waste disposal site during the
remedial investigation (see Section 6.2). With regard to surfactants, it is
particularly important to measure the total ionic strength, hardness and
concentration of polyvalent cations because these can reduce the effectiveness
of a surfactant.
223
-------
FIGURE 6-2
EVALUATION OF FLUSHING & SURFACTANTS
« m
WASTE
CHARACTERISTICS
_*.
t
ro
ro
GEOHYDROLOGICAL
CONDITIONS
•^
@ = FEASIBLE
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IS WATER
FLUSHING
ADEQUATE?
INO
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^^^--^"^
INO
in
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TREATABILITY
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V
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-------
6.3.2.1 Flushing With Water or Surfactants (Step I)
If water can dissolve the wastes, it should always be used because of its
safety, low cost and because it will not introduce a new chemical into the
waste site. Based on the organics identified during the RI, Table 3-2 of
Section 3 should be consulted to check the solubilities and octanol-water
partition coefficients (KQW) of these organics. If Table 3-2 does not
contain the required data, supplemental literature surveys or laboratory tests
should be conducted to determine the solubilities and K of the identified
ow
organic waste materials. If their solubilities are greater than 5 x 10 M
and log KQw values are less than 2, water can be used to flush and recover
the organics. Otherwise, surfactants should be considered.
6.3.2.2 Identify Commercially Available Surfactants (Step II)
This step will be performed only when water alone cannot dissolve the organic
wastes. Based on the physical, chemical and geohydrological conditions of the
waste deposit and the site (determined during the RI), certain commercially
available surfactants may be identified from Tables 3-1 and 3-5 of Section 3.
The selection of a surfactant will be dependent upon its degradation rate,
toxicity and effectiveness in the environment of the waste disposal site (see
Section 3 for detailed discussion).
6.3.2.3 Bench Scale Treatability Studies of Potential Surfactants
(Step III)
In this step, the potential surfactants selected from Step II are screened in
a series of laboratory tests so that the most effective surfactant (single or
a mixture) can be identified. An example of such studies is presented by
Ellis et al. (1984). The tests may include, but not be limited to, the
following:
o The interaction between the surfactants and the wastes - In these
tests, the effectiveness of the surfactants are determined by the
solubilities of the wastes in the surfactant solutions. The chemical
225
-------
characteristics of the surfactant solutions after contact with the
wastes should also be checked for toxicity (i.e., the wastes may be
made more available, and therefore more toxic, by emulsion/solubili-
zation).
The interaction between the surfactants and the soil - These tests
should determine whether the surfactants lose their effectiveness
because of adsorption to soil at the site.
The interaction between the surfactants and the groundwater - These
tests would evaluate the effects of groundwater chemistry on the
effectiveness of the surfactants (i.e., the potential for
precipitation, neutralization or complexing of the surfactant by
naturally-occurring constituents of the groundwater).
Biodegradation tests with native microbes - These tests would
determine whether the surfactants are biodegraded by the native
microbes at the waste site, and the degradation rates at various
surfactant concentrations.
The most effective surfactant can be selected by comparing these test
results. If the test results indicate that none of the tested surfactants are
likely to be effective, alternative in situ treatment methods may have to be
investigated.
Following the identification of potentially-useful surfactants, or the
determination that water flushing alone should be sufficient, the
effectiveness (i.e., organic waste removal/recovery rates) of either water or
the selected surfactant should be tested under laboratory simulation
conditions. These simulation tests can be simple column tests (Ellis et al.,
1984) or specifically-designed microcosm tests which can better represent the
waste site conditions. In these tests, the optimum range of water or
surfactant concentrations and flow rates for recovering the wastes should be
determined. In addition, the chemical characteristics of the recovered
surfactant solutions and their potential toxicity should also be analyzed.
226
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6.3.2.4 On-Site Pilot Test (Step IV)
In this step, a representative plot at the waste site should be selected for
running a pilot test to confirm the results obtained from laboratory screening
and simulation tests. In addition, the delivery/recovery system designed and
implemented for this pilot test would provide design and operation guidance
for the full scale treatment facility. The conducting of pilot-scale tests is
discussed further in USEPA (1984a) and Repa and Kufs (1985). The potential
delivery/recovery systems are discussed in Section 6.4.
6.3.2.5 Facility Design and Cost-Effectiveness Assessment (Step V)
After the on-site test has shown that the water or selected surfactant is
effective in recovering the wastes, and the surfactant (if used) is
environmentally safe, the in-situ treatment facility for the waste site can be
conceptually designed. Order-of-magnitude cost estimates should be made
according to the conceptual design. These costs and the effectiveness of the
system should be compared with those of other alternatives (if any) to
determine whether this method should be implemented (USEPA, 1984a; USEPA,
1985; Repa and Kufs, 1985).
6.3.3
Evaluation of Hydrolysis for Waste Deposit Stabilization
Hydrolysis represents a major degradation process for many organic chemicals
as reviewed in Section 4, and feasible in situ methods of accelerating
hydrolysis rates are available. Thus hydrolysis is a potential in situ
treatment method or mobilization method for waste deposits containing a
variety of organic chemicals. ... , . .......
The application of hydrolysis to waste deposits will require implementation of
a systematic evaluation approach as illustrated in Figure,6-3. This will be
undertaken following the initial site investigation and contamination
evaluation (RI, see Section 6.2). This approach includes:
227
-------
ro
ro
CO
FIGURE 6-3
EVALUATION OF HYDROLYSIS
•« Rl >•
GEOHYDROLOGICAL
CONDITIONS
WASTE
CHARACTERISTICS
| II III IV V
ARE YES IS IN SITE YES BENCH SCALE © F,ELD © n^™7/^
_ HYDROLYZABLE I^ HYDROLYSIS -*- TREATABILITY -*- plLO-TSTUDY «««MCMT
^ COMPOUNDS ACHIEVABLE STUDIES PILOT STUDY ASSESSMENT
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• IMPLEMENTATION
• FEASIBLE
EVALUATE
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-------
o Identifying the organic compounds in the waste susceptible to
hydrolysis ( Step I)
o Analyzing the effect of variables on hydrolysis rates and
completeness, and assessing the need for and method of accelerating
the hydrolysis rate (Step II)
o Bench scale treatability studies (Step III)
o Design and implementation of a field demonstration program (Step IV)
o Assessment of the field demonstration program and judging the
feasibility of the hydrolysis system ( Step V)
The procedures for assessing the potential application of hydrolysis are
discussed below.
6.3.3.1 Identify Organic Compounds Susceptible to Hydrolysis (Step I)
Based on their hydrolysis half lives and potential for acceleration of
hydrolysis, organic compounds can be divided into three general groups: a
hydrolysis-resistant group, a hydrolysis-susceptible group (i.e., with
catalysis or PH adjustment) and a hydrolyzable group. Tables 4-1, 4-2 and
4-11 summarize the types of chemical compounds that are generally resistant or
susceptible to hydrolysis and Tables 4-3 through 4-10 present the hydrolysis
half-lives for a variety of compounds. The application of hydrolysis is
suitable for the hydrolyzable group, and may be possible for the hydrolysis-
susceptible group. Organic compounds in the waste deposit are first identified
by group. If the waste deposit contains both hydrolyzable and non-hydrolyzable
compounds, more laboratory treatability testing and cost-effectiveness
analysis may be required to confirm the feasibility of hydrolysis and the need
for additional treatment of the non-hydrolyzable compounds.
229
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6.3.3.2 Effects of Site Conditions on Hydrolysis and Potential for
Acceleration of Hydrolysis Rates (Step II)
A hydrolyzable or hydrolysis-susceptible waste deposit may still be precluded
from in situ treatment by hydrolysis due to the effects of pH, temperature,
solvent composition, or difficulty of catalysis. The effect: of pH on
hydrolysis rate is pronounced in cases where acid- or base-catalyzed
hydrolysis is important, changing the overall hydrolysis rate by up to one
order of magnitude for one unit change in pH (see Figures 4-1 and 4-2).
Tables 4-3 through 4-10 summarize the effects of pH on a variety of organic
compounds.
Because the hydrolysis rate is a function of temperature, extrapolating
laboratory temperature hydrolysis data to environmental conditions may
represent a significant source of error. In general, a 10 C (18 F)
decrease in temperature produces a 2.5 times decrease in the hydrolysis rate.
Hydrolysis rates are also affected by solvent composition, i.e., ionic
strength. Increasing the ionic strength can either accelerate or retard
hydrolysis. Catalysis or retardation of hydrolysis by surfactants can also
significantly alter hydrolysis rates (N Wolfe, USEPA, Athens, GA: personal
communication). Each of these factors must be evaluated for the hydrolyzable
compounds present to determine whether their hydrolysis rates may be
unavoidably retarded by in situ conditions, or may be accelerated, by altering
these conditions.
Three types of hydrolysis processes, base-catalyzed, neutral and
acid-catalyzed hydrolysis, may contribute to or affect the overall hydrolysis
rates of organic chemicals in the environment. Base-catalysis of hydrolysis
appears to be the most promising approach. The chemical classes potentially
treated through acceleration of degradation by base-catalyzed hydrolysis are
described in detail in Section 4 and presented in Table 4-11. The primary
design concern for implementation of base-catalyzed hydrolysis in a waste
deposit will be the production and maintenance of high pH (pH 9 to 11)
conditions with saturation or high moisture content in the waste deposit.
230
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6.3.3.3 Bench Scale Treatability Studies and Field Pilot Study
(Steps III and IV)
Bench scale treatability studies and field pilot tests using site specific
soil/waste matrices should be conducted prior to full scale implementation of
hydrolysis treatment systems for waste deposits (USEPA, 1984a). The exact
requirements for base addition in hydrolysis acceleration and the effects of
pH alteration should be determined by laboratory tests. Acidic or highly
buffered deposits or soils will require greater additions of base than poorly
buffered, neutral or alkaline deposits. However, anionic and amphoteric
(capable of acting as either an acid or base) species may be mobilized, and
the sorption of organic species in the deposit may be affected by a
significant change in the pH of the waste deposit matrix.
6.3.3.4 Cost-Effectiveness Assessment and Conceptual Design (Step V)
The laboratory and field tests will indicate the susceptibility to hydrolysis
of the wastes, in situ hydrolysis rates and the potential for catalytic
acceleration. In addition, the potential effects of site geohydrologic
conditions (soil organics, groundwater chemistry) and side reaction products
(which may be toxic) can be evaluated. These variables must be evaluated to
determine the feasibility, environmental acceptability (risk assessment), and
effectiveness of in situ hydrolysis of the waste deposit. The facility is
conceptually designed and order-of-magnitude cost estimates are made so that a
cost and effectiveness analysis can be performed (USEPA, 1984a; USEPA, 1985;
Repa and Kufs, 1985).
It must be noted that at the present time, only limited investigations of
hydrolysis in soils have been conducted. There is no practical field
experience for the control of hydrolysis rates in waste deposits. However,
conditions favorable to base-catalyzed hydrolysis can be produced using
available equipment and reagents. That is, lime can be applied to the surface
and irrigated to produce the base catalysis; alternatively NaOH solution can
be used. Again, the selection of pH-controlling reagents will be based on the
results of bench scale treatability studies.
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6.3.4 Evaluation of Oxidation for Waste Deposit Stabilization
The potential application of three oxidants (ozone, hydrogen peroxide, and
hypochlorites) to waste deposits is evaluated in Section 5. While these
oxidants are reactive with a wide variety of organic compounds and have
demonstrated applications in wastewater treatment, significant potential
problems may preclude their use as in situ treatment agents for waste deposits.
Hydrogen peroxide is a weaker oxidizing agent than ozone, but its stability in
water is considerably greater. However, the decomposition of hydrogen
peroxide to oxygen may be catalyzed by iron or certain other metals; therefore
effective delivery of hydrogen peroxide throughout an entire waste deposit may
be difficult or impossible because of the relatively low transport velocities
achievable in waste deposits compared to accelerated in situ decomposition
rates. Prior to consideration of hydrogen peroxide as an in situ treatment
method, it will be necessary to investigate the stability (or rate of
decomposition) of hydrogen peroxide in a specific waste deposit matrix, as
well as its effectiveness in treating contaminants of concern. In the event
that hydrogen peroxide is not determined to be effective as a treatment agent,
it may find usage as a source of oxygen in a waste deposit to support aerobic
microbial degradation of the wastes (Wetzel et al., 1985).
If the effectiveness of hydrogen peroxide as an oxidizing agent for a waste
treatment can be demonstrated, its application to a waste deposit does not
appear to present significant problems with respect to equipment selection.
The approach described for the evaluation of hydrolysis (Section 6.3.3, except
for the determination of catalyst-accelerated hydrolysis) is directly
applicable to the evaluation of the feasibility of using hydrogen peroxide.
Hydrogen peroxide is available commercially in a variety of concentrations and
freely dissolves in water at all concentrations. At low concentrations
hydrogen peroxide solutions have densities and viscosities similar to water.
The potential hazard of violent reactions of certain organic materials with
hydrogen peroxide should, however, be recognized. Applications of dilute
solutions may be necessary to avoid possible explosive hazards. Since addition
232
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of very dilute hydrogen peroxide solutions to a waste deposit could result in
flushing of contaminants, recovery methods as well as delivery methods should
be included in system design.
Potential application of ozone to organic contaminants, in waste deposits is
discussed in Section 5.2. While ozone is an effective oxidizing agent for
many organic compounds in wastewater treatment applications, its relatively
low stability in aqueous systems, particularly in the presence of certain
chemical contaminants, may preclude its effective application to waste
deposits. As indicated in Section 5.2, the half-life of ozone in natural
waters is less than one-half hour. Considering that flow rates of water
through waste deposits are likely to be on the order of inches/hour or less,
it is unlikely that effective oxidant doses of ozone can be delivered outside
of the immediate vicinity of the point of application (i.e., within inches or
feet of an injection point). For this reason, design of a feasible
application system for in situ treatment by ozone is unlikely. However, ozone
may be used to provide an oxygen source for biodegradation (see Section 5.2,
and Nagel, 1982).
Demonstrations of the effectiveness of hypochlorite as an oxidizing agent for
organic materials are extremely limited (Section 5.3). In addition,
hypochlorite reacts with organic compounds as a chlorinating agent as well as
an oxidizing agent, and there is a significant chance that hypochlorite
additions to waste deposits may lead to production of undesirable chlorinated
by-products (e.g., chloroform) rather than oxidative degradation of the
wastes. Therefore, the usage of hypochlorite is not recommended.
6.4 Application and Design of Delivery/Recovery Systems for In-Situ Treatment
The successful application of chemical solutions to a hazardous waste deposit
for in situ treatment or mobilization of contaminants from the deposit
requires the selection of a technically sound and cost effective
delivery/recovery system. A systematic approach for site evaluation,
233
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treatment method selection and conceptual process design is required prior to
final facility design. Prior to the implementation of a delivery/recovery
system, a laboratory test program and a field demonstration program would be
required as needed to fill data voids.
The treatment techniques and the site parameters that govern the design and
performance of the delivery/recovery technologies were discussed in Section
1. Based on these discussions, a matrix of decision factors consisting of
relevant technical criteria was developed in order to guide in the conceptual
design of feasible delivery and recovery systems for a given set of site
conditions (Tables 1-5 and 1-6). The systematic approach for implementing a
delivery/recovery system as shown in Figure 6-4 consists of site evaluation, a
program of additional field and laboratory testing (as required),
identification of alternative methods for delivery and recovery of solutions,
conceptual design of the alternative remedial technologies, economic
evaluation, system selection and finally detailed design and implementation of
the delivery/recovery system. :'
6.4.1
Determining
(Step I)
the Requirements of a Delivery/Recovery System
Analysis of the chemicals present in and the characteristics of the waste
deposit would lead to the selection of the treatment methods to be used.
Biological agents, chemical hydrolysis, oxidation or flushing methods using
water or surfactants would be selected as described in Section 6.3 before the
specific delivery/recovery sytstem is identified. This determination would
define the processes required .to deliver the treatment agents and possibly
also dictate the recovery methods.
There are many alternative methods for combining the reagent with a delivery
system. For example in situ oxidation might require forced injection of
hydrogen peroxide directly into the waste because of its reactivity with
soil. Alternatively, flushing with water into the same deposit could rely
upon passive, gravity type delivery systems since there is no degradation of
the reagent (water) during its time of passage through the soil medium to the
234 :
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FIGURE 6-4
SYSTEMATIC APPROACH TO DELIVERY/RECOVERY
SYSTEM SELECTION
rss
oo
en
BIODEGRADATION
HYDROLYSIS
OXIDATION
FLUSHING/SURFACTANT
REQUIREMENTS
RECOVERY
APPLICATION
• SURFACE
• SUBSURFACE
APPLICATION
1— »•
1
EVALUATION
• DATA
ANALYSIS
• DATA
VOIDS
/
/
/
/
/
\
\
II
ADDITIONAL FIELD
INVESTIGATIONS
• WASTE DEPOSIT
CONFIGURATION
• SITE GEOHYDROLOGY
• IN SITU PERMEABILITY
•. FIELD INFILTRATION
• TOPOGRAPHY
* CLiMATOLOGICAL
DATA
IV
ADDITIONAL LABORATORY
TESTING
• WASTE PHYSICAL
CHARACTERISTICS
• SOIL CHARACTER-
ISTICS
1
ft'i
/
v/
- ,^ DELIVERY/
~" RECOVERY V
\
VIII
DETAILED
DESIGN &
OF DELIVERY/
RECOVERY SYSTEM
DELIVERY
METHOD
• HOMOGENEITY
• PERMEABILITY
• WASTE CON-
FIGURATION
METHOD
• PERMEABILITY
• DEPTH OF
TABLE
V
EVALUATE
METHODS
_ • CONCEPTUAL
DESIGN
• COSTESTIMAT
/
/
r
*•»•
i
\
\
\
ii
E
GRAVITY METHOD
• FLOODING
• PONDING
• DITCH
!• SURFACE SPRAY
GALLERY .
• INFILTRATION
BED
FORCED METHOD
• ELECTRO OSMO
SIS
GRAVITY METHOD
• OPEN DITCH
• BURIED DRAIN
• 1
FORCED METHOD
• WELLPOINT
• ELECTRO-OSMO
SIS
VI
PERFORM
FIELD ,
" DEMONSTRATION
PROGRAM
*>
g
-------
deposit. It is difficult to generalize as to how the treatment per se becomes
a factor in selecting the delivery or recovery system. In some cases the
treatment will be a key consideration; in others the site geohydrologic
conditions, the deposit location and the anticipated costs will influence the
decision to a greater degree.
i-
6.4.2 Site Evaluation (Steps II, III, and IV)
Characterization of the contaminants present in the waste deposit is necessary
to select the appropriate treatment methods, and determination of site
geohydrologic features is needed to establish the location, selection and
design of the delivery/recovery system. The field investigation data
generated during the RI should be reviewed to identify any data voids. The RI
should provide at least the following information:
- Extent and nature of the waste deposit
Site soil characteristics such as porosity and permeability
- Surface drainage characteristics
Groundwater table depth, groundwater flow direction and velocity
- Field permeability testing of the waste deposit and host materials
Surface infiltration rate determination
- Laboratory analysis of soil, waste deposit and groundwater samples
- Climatological data.
Additional field and laboratory investigations may be required for the
evaluation of certain delivery/recovery methods if it is determined that the
data provided in the RI are insufficient.
6.4.3 Selecting the Delivery and Recovery Methods (Steps V)
The selection of the most appropriate delivery/recovery methods and systems
would be based on the configuration of the waste deposit (areal extent and
vertical depth), hydrologic characteristics (surface and subsurface) of the
waste deposit, and surface and subsurface geohydrologic characteristics of the
materials surrounding the waste deposit.
236
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6.4.3.1 Delivery Methods
The matrix for selection of delivery methods is presented in Table 1-5. This
table illustrates that forced delivery methods are applicable for all
conditions. The choice of a gravity delivery method is more dependent on the
listed parameters. The design factors for the delivery methods and their
associated design criteria are discussed below.
Location of the Deposit in Relation to Existing Groundwater Table —
As may be seen from Table 1-5, if the waste deposit is located in the
unsaturated zone all of the gravity and forced delivery systems may be
applicable. Presence of the waste deposit in the saturated zone eliminates
virtually all of the gravity delivery methods (with the possible exception of
ponding). Forced delivery appears to be the most effective delivery method
for waste deposits located in the saturated zone.
Contamination Present at the Surface —
This consideration may eliminate certain gravity based delivery systems (such
as the use of ditches, infiltration galleries and infiltration beds) which
require excavation to construct and therefore cannot deliver solution to the
surface. In this case, gravity based delivery systems applied at the surface
(i.e., flooding, ponding and surface spraying) could be applicable. Injection
into a waste deposit via forced injection can treat waste below the surface
but it would need to be supplemented by a gravity method to assure complete
treatment of surficial as well as deeper waste materials.
Waste Deposit Covered by an Impermeable Layer —
This parameter has no bearing for the forced delivery methods, but it will
have a significant impact for gravity delivery methods. For example, flooding
and spraying cannot be utilized as delivery methods if the deposit is
separated from the surface by an impermeable layer of soil or is covered by an
impermeable synthetic material.
237
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Topography —
Topographic considerations will limit, in part, the extent of applicability of
gravity flow methods. For example, flooding or ponding delivery methods
cannot be utilized on a steep slope although trenches may be feasible.
However, topography will not affect the forced delivery methods.
Infiltration Rate —
The infiltration rate governs the application rate of reactant on the top
layer of the deposit or soil. It has no bearing in the selection of forced
methods. In surface gravity applications this will play a major role, and may
eliminate flooding, surface spraying, ponding or ditching as potential
delivery methods.
Hydraulic Conductivity of the Waste Deposit and Surrounding Soil —
The hydraulic conductivity of the soil and waste deposit will dictate the flow
characteristics within and around the deposit. If the hydraulic conductivity
of the deposit is high and is equal to or greater than that of the surrounding
soil, low net pressure and short time durations would be required for a
solution to pass through the deposit. In this case, gravity delivery systems
may be used. Low hydraulic conductivity of the waste deposit indicates that
the deposit will not be easily drainable and will require higher pressure and
longer times for a solution to move through the deposit. If the waste deposit
has a lower hydraulic conductivity than the surrounding soils, solutions
delivered by gravity methods would bypass the deposit. In either case, a
forced delivery system would be required.
Gravity delivery methods would be applicable if the waste deposit and
surrounding medium have hydraulic conductivities in the range of 1 x 10
cm/sec to 1 x 10~3 cm/sec (280 to 2.8 ft/day). Forced delivery methods
would be required for a waste deposit or soils with a hydraulic conductivity
238
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-3 -4
between 1 x 10 cm/sec and 1 x 10 cm/sec (2.8 to 0.28 ft/day). For a
hydraulic conductivity less than 1 x 10 cm/sec (0.28 ft/day), forced
injection assisted by electro-osmosis used as a recovery method may be the
only effective system.
Depth to Bottom of the Waste Deposit —
The cutoff for this parameter was chosen based on engineering judgement. If
the depth to the base of the deposit is too great it may take too long for a
solution to travel through the deposit under the force of gravity. Based on
this condition, a reasonable cut-off point for gravity delivery methods was
chosen at 5 meters (16 feet).
In addition to the parameters and conditions presented in Table 1-5, the
homogeneity of the waste deposit and surrounding soil media is important,
although very difficult to quantify. Waste deposits and soils with large
variations in hydraulic conductivity as a function of depth or lateral
location vastly complicate the delivery of treatment reagents. Gravity
methods are much more effective in relatively homogeneous deposit and soil
environments where the applied solution can be evenly distributed throughout
the deposit. In a heterogeneous environment, the waste deposit probably
cannot be effectively treated by gravity delivery methods. Only forced
delivery methods offer any promise in such cases.
In general, gravity delivery methods are effective when the waste deposit is
situated in an unsaturated zone at the surface or with a shallow, relatively
permeable overburden, and the depth to bottom of the deposit is limited to 5
__O
meters (16 feet) with hydraulic conductivity greater than 1 x 10 cm/sec
(2.8 ft/day). Forced delivery methods will be most effective for waste
deposits covered by thick overburdens of significant depth (more than 5
meters). A forced method utilizing electro-osmosis could be considered for
solution injection into a deposit with hydraulic conductivities lower than 1 x
10 cm/sec, although at present the applicability and effectiveness of
electro-osmosis has not been demonstrated. In general, forced methods should
239
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be highly effective for waste deposits with hydraulic conductivities in the
range of 1 x ICf1 cm/sec to 10~4 cm/sec (280 to 0.28 ft/day).
6.4.3.2 Recovery Methods
I
Table 1-6 indicates the applicability of various recovery methods for
different site characteristics. Two parameters, depth to the recovery zone
and hydraulic conductivity, are considered sufficiently important to include
in the matrix. Although additional parameters (such as effective porosity and
storativity) may play an important role in designing deep well or well point
systems, these two parameters are the most appropriate guide for the
preliminary selection of recovery methods. It should be noted that the
recovery of injected solutions will be from 'the saturated zone (water table
aquifer) and normally the recovery method(s) will be installed beyond the
boundary of the waste deposit. The depth to the recovery zone is chosen as a
prescriptive parameter because gravity methods are generally impractical
beyond a 5 meter (16 foot) depth. The hydraulic conductivity will dictate the
drainage characteristics and thereby control the selection of recovery
(dewatering) methods.
The following design criteria, condensed from the matrix of criteria for
selecting a recovery system (Table 1-6), could be used as a basis for
conceptual design:
1. Gravity recovery systems (open ditches and buried drains) or forced
recovery systems (well points or deep wells) are applicable for a
site having a hydraulic conductivity between 1 x 10 cm/sec and 1
x 10 cm/sec (280 to 2.8 ft/day). A vacuum well point system or
possibly a deep well system would be suitable for a site with a
—3
hydraulic conductivity in the range of 1 x 10 cm/sec to 1 x
» r
10 cm/sec (2.8 to 0.28 ft/day). The electro-osmosis* method may
be considered for low permeability conditions (below 1 x 10
cm/sec or 0.28 ft/day), but considerable experimentation and
laboratory simulation and testing would be a necessary precursor to
use of this method.
240
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2. A multi-stage well point or vacuum well point system would be
required for a depth between 5 meters and 12 meters (16 feet to 40
feet).
3.
Open ditches and buried drains should be limited to depths of less
than 5 meters (16 feet) and must be within the zone of saturation.
Deep wells would be practical for a depth of more than 5 to 12 meters
(16 to 40 feet).
In general, gravity recovery methods are suitable for a shallow recovery zone
(depth to water table from the surface should not be more than 5 meters). For
a deeper recovery zone, forced recovery methods must be employed.
6.4.4 Field Demonstration Program (Step VI)
A site specific field demonstration program for the selected feasible methods
is undertaken if necessary to evaluate the effectiveness of the methods and to
generate design information such as ditch spacing or well spacing which will
be required for proper delivery and recovery of the treatment agent.
6.4.5 Evaluating Alternative Methods (Step VII)
Based on the field demonstration program, alternative delivery/recovery
systems are developed and a conceptual design and associated cost evaluation
is performed. Based on the cost analysis and treatment system effectiveness,
final selection of a delivery/recovery system is made for subsequent
implementation (USEPA, 1984a; USEPA, 1985; Repa and Kufs, 1985).
6.4.6 Detailed Design and Implementation (Step VIII)
The final steps of engineering and design for installation of an in situ
treatment system would be the detailed design, specification preparation,
equipment procurement and installation of the following facilities necessary
to apply, distribute and collect treatment solutions:
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- Treatment agent storage, preparation and delivery facilities and
equipment,
Earthwork for site preparation,
Delivery and recovery system facilities and equipment, and
- Monitoring system facilities and equipment.
A number of recent publications give a comprehensive description of how in
situ treatment methodologies can be evaluated, tested and undertaken, and
reflect the general principles outlined in this report. These studies include
Jhaveri and Mazzacca (1983), USEPA (1984b), Ryckman (1984), Wetzel et al.
(1985), Flathman et al. (1983), Flathman et al. (1984), and Flathman and
Caplan (1985). Most of these studies, which present detailed case histories
of in situ treatment by biodegradation or hydrolysis (Ryckman, 1984), have
been described in previous sections of this report. ;
6.5 Case History of RI/FS and In Situ Treatment of Contaminated Soil and
Groundwater
6.5.1 Site Summary
In August 1975 contamination was observed in a small creek that discharges
into Allendale Brook in the Town of Waldwick, NJ. Biocraft Laboratories, a
small synthetic penicillin manufacturer is located on a 1.72 h.a (4.3 acre)
plot near the contaminated creek within an industrial park in Waldwick. It
was determined that leakage had ocurred in underground tanks used to store
waste solvent between 1972, when the plant commenced operation, and 1975, when
the contamination was discovered, (Jhaveri and Mazzacca, 1983; USEPA, I984c).
The waste solvents seeped into an adjacent storm sewer and thence drained into
the stream, where a fish kill in 1973 was attributed to the contamination.
The local shallow aquifer was contaminated and it was feared that a town
drinking water well was threatened by the plume. It is estimated (Jhaveri and
Mazzacca, 1983) that the following contaminants probably leaked into the
subsurface environment:
242
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methylene chloride
n-butyl alcohol
dimethyl aniline
acetone
6.5.2 Remedial Investigation
82310 kg
30305 kg
11925 kg
4840 kg
(181500 Ibs)
(66825 Ibs)
(26300 Ibs)
(10890 Ibs)
The Biocraft site was well characterized both geohydrologically and
chemically. Figure 2-1 (in Section 2) is a site plan for the Biocraft site
showing the approximate configuration of the contaminant plume prior to
remediation. Figure 6-5 is the configuration of the water table at the site
before implementation of remediation. Six groundwater monitoring wells were
installed at the site in January 1976, followed by 22 more wells in June 1976.
These were used both to monitor and selectively pump contaminated water. The
basic results of the monitoring and testing program were that:
1)
2)
The contaminant plume (as measured by COD greater than 100
mg/liter) roughly followed the plume outline show in in Figure
2-1 and was approximately 0.71 ha (1.75 acres) in area. It was
estimated that 9175 m3 (12000 yd3) of soil were contaminated.
The contamination had not penetrated a semi-consolidated
silt/fine sand layer located approximately 4 m (12 ft) below
grade and no contamination had entered the deep aquifer which is
the source of the town's water supply.
The tabulation below presents chemical data taken from the six sampling wells
at the site during Jan - June 1976.
Parameter Concentration Range
5.2 - 7.5
2 - 21000 mg/1
8 - 31QOO mg/1
2 - 9625 mg/1
5 - 6246 mg/1
pH
BOD
COD
TOG
243
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FIGURE 6-5
GROUNDWATER SURFACE CONTOURS
BIOCRAFTSITE
I GROURBWATER/V/
I FLOW /\
(SOURCE: JHAVERI AND MAZZACCA, 1983)
244
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An onsite test well sampled in 1981 (just before the biodegradation operation
began at the site) revealed groundwater concentrations of 85 mg/1 acetone, 55
mg/1 methylene chloride and 648 mg/1 COD (USEPA, 1984c).
The site is located in an area of unstratified and stratified glacial drift.
A layer of silt and gravel approximately 3 ft thick is found at the surface
and is underlain by glacial till to a thickness of 8 to 15 feet. Based on
slug tests at 5 on-site wells, the hydraulic conductivity of the glacial till
layer ranges from 1.7x10 3 to 9.4xl(T7cm/sec (0.02 to 36 gallons/day per
ft ). Approximately 12 m (40 feet) of semi-consolidated silt and fine sand
underlies the till layer. This layer, which lies at an average depth of about
4 m (12 ft) below grade, is considered to be an aquiclude.
For further detail the reader is referred to USEPA (1984c) and Jhaveri and
Malacca (1983) for a complete description of the extent of contamination at
the site, site characteristics (soil properties, aquifer conditions), the
monitoring program and the reactions and depositions of state and local
regulatory bodies and interested parties.
6.5.3 Feasibility Study
The evaluative process leading to the selection of biodegradation as compared
with other technologies (e.g., slurry walls, excavation) described in USEPA
(1984c). This section describes the methodology followed by Biocraft
Laboratories in technically developing the biodegradation system for their
site.
Biocraft and their subsidiary Groundwater Decontamination Systems (CDS), which
holds a patent on the system ultimately installed at the site, required
approximately 2-1/2 years to proceed through the research and development
stage of the biodegradation system, that is to proceed through Steps II
through VII as presented on Figure 6-1.
Biocraft and its consultants selected and developed the biodegradation
alternative in May 1979. The alternative included four elements:
245
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1) Collection of the contaminated plume in a downgradient buried trench,
2) Surface treatment of the collected groundwater in a bioreactor to
remove contamination and aerate the water,
3) Reintroduction of the treated water upgradient via infiltration
trenches in order to flush the soil and to supplement the subsurface
microbe population, and
4) Promotion of subsurface biological activity through the use of
aeration wells.
Establishment of biodegradation as an in situ treatment technology was
faciliated by:
1) The relatively homogeneous nature of the well-characterized
contaminant plume,
2) The presence of soil contamination within the saturated portion of
the surficial aquifer (i.e., above the aquiclude),
3) The permeability of the soils which lay within the feasible range for
carrying out in situ treatment, and
4)
The depth to the deposit (or plume) was less than 5 meters (16 ft).
These conditions indicate that gravity delivery and recovery systems would be
suitable for the site (see Sections 1.2 and 1.3 of this report).
When Biocraft started the investigation in July 1978 the biodegradability of
methylene chloride, the principal pollutant at the site, was not: clearly known
(Jhaveri and Mazzacca, 1983). To determine whether readily available microbes
would be capable of degrading methylene chloride (i.e., to perform Step II),
contaminated groundwater was inoculated with soil samples taken from the
246
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Biocraft site itself as well as uncontaminated (control) samples from the
homes of various employees. Completion of this step indicated that the onsite
soil sample held the most promising microbial population.
The research effort then turned to the identification of limiting factors,
i.e., Step III of Figure 6-1. A shaker flask study using .contaminated water
as the carbon source was undertaken to determine optimum growth conditions.
Jhaveri and Mazzacca (1983) provide details on nutrient media tested and USEPA
(1984c) provides a tabulation of the experimental results. The conclusions of
this step were:
1) Nitrogen and phosphorus addition increased cell growth,
2) Phosphorus addition (as dibasic phosphate) supplied buffering
capacity to the medium thus accounting for HC1 formation associated
with methylene chloride degradation, and
3) Anaerobic study results were not favorable.
Having established that the onsite microorganisms were effective in degrading
methylene chloride, Biocraft/GDS proceeded to bench scale treatability and
field pilot studies (i.e., Steps VI and VII). The basic elements of this
portion of the program consisted of:
1) Aeration and nutrient addition to an onsite well. This test
demonstrated the feasibility of subsurface aerobic activity: a 100
fold increase in cell count was observed after 7 days
2) Bench scale batch testing in fermentors. Various temperatures and
aeration rates were tested.
3) Bench scale continuous testing in fermentors. These tests
established the percent destruction of methylene chloride as a
function of the retention time, and
247
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4) Pilot plant studies conducted in two 208 liter (55 gallon) drum
reactors. Air sparging, temperature control systems, and nutrient
feed methods were tested. The process retention time, nutrient feed
rate and aeration rate were established.
The pilot plant showed that methylene chloride could be reduced by 99% of its
inlet concentration, butanol levels could be reduced by 96% and dimethyl
aniline concentrations could be reduced by 59%. More detailed data and
process difficulties are reported upon in USEPA (1984c) and Jhaveri and
Mazzacca (1983). Biocraft/GDS were sufficiently satisfied with results of
this series of studies to proceed to the development of a full scale reactor
which is described briefly below and in detail in both of the above references.
In summary, performance by Biocraft/GDS of Steps I-VIII as shown in Figure 6-1
established the feasibility of both an above ground reactor using onsite
microorganisms and of bringing about subsurface biological degradation. The
following subsection compares the site and waste characteristics at the
Biocraft site to the guidance presented in Section 1 of this report on the
selection of delivery and recovery systems and describes the entire treatment
system at the site.
6.5.4 Description of the Treatment System
The shallow groundwater table (0-3 meters below the surface) and depth to the
aquiclude of less than 5'meters (16 ft), combined with soil permeabilities in
the range of IxlO"3 cm/sec, indicated that a gravity delivery and recovery
system could be considered (see Section 1.5 of this report for a discussion of
parameters used in selection of delivery/recovery systems). As shown in
Figure 6-5, although a groundwater mound exists in the southeast corner of the
site, groundwater flow is generally to the northwest. Based on these
considerations, Biocraft and its consultants designed the recirculating in
situ and above ground (bioreactor) treatment system illustrated in Figure
2-2. The operation of this system would provide both plume containment and
248
-------
removal of the source. This system is relatively passive and unobtrusive: The
delivery and recovery systems are below ground and invisible to passersby, and
implementation of the system did not require disruption of the Biocraft
operation, which other actions such as excavation would clearly have done.
The delivery system consists of two "recharge trenches" (infiltration
galleries), one of which is illustrated in cross-section in Figure 2-2 The
trenches are approximately 30 m long, 1 m wide and 3 m deep, and were
excavated by backhoe (USEPA, 1984c). The trenches are lined on all sides but
the front (Figure 2-2) with a 15 mil plastic liner to direct the injection
water toward the waste deposit. The trenches are filled with 5 cm diameter
washed stone to the surface (Figure 2-2). A 5 cm slotted pipe placed along
the trench (1.5 m from the bottom) delivers recharge water at a rate of about
25,900 I/day (6850 gallons/day) per trench. Each trench also has two
monitoring wells (one at each end) which can also be used to flush the
trenches if sludge accumulates.
The recovery system, located approximately 90m (300 ft) northwest of the
infiltration trenches, consists of a buried trench and slotted pipe collection
system (Figure 2-2). The trench is about 24 » long, 1.2 m wide and 3 m deep,
and is filled with a layered, washed stone gravel pack (USEPA, 1984c)'
Groundwater is pumped at a rate of 38 1 (10 gallons) per minute from a slotted
central collection well which is also fed by the collection pipes.
Groundwater is also pumped from two bucket wells at the southern edge of the
site to intercept the southerly component of groundwater flow from the
groundwater mound (USEPA, 1984c; Jhaveri and Mazzacca, 1983).
The in situ aeration system (Figure 2-2) consists of nine aeration wells
spaced about 9 m away from each other and arranged in a rectangular matrix 9 m
wide and 30 m long (USEPA, 1984c). This arrangement was based on the
assumption of a 4.5 m radius of influence of each aeration point. Air is
continuously injected at a pressure of 28-62 kN/m2 (4-9 psi).
249
-------
The surface treatment system (bioreactor) consists of a dual system of two
aeration and two sludge settling tanks, each tank having a capacity of 20,000
1 (5400 gallons). The stainless steel, temperature-controlled tanks were
originally used for milk transport. Influent water from the collection trench
and two interceptor wells is pumped first to the aeration tanks, where most of
the biodegradation occurs (Jhaveri and Mazzacca, 1983). Air is added to each
tank through'a series of porous ceramic tube diff users at a rate of 0.8 m
per minute. Temperature is kept constant at 20°C (68°F) using a single
pass steam coil installed in the tanks. A nutrient solution is metered into
the aeration tanks as required. Effluent air from the aeration tanks is
passed through replaceable activated carbon adsorbers to remove any
volatilized organics.
The effluent stream from the aeration tanks is combined and pumped to the
sludge settling tanks in which some biomass solids are settled out and
recycled to the aeration tanks. The supernatant from the settling tanks is
pumped to the reinjection trenches. An important point is that much of the
biomass is allowed to pass with the supernatant into the recharge trenches in
order to continually inoculate the trench and subsurface with microorganisms.
Waste sludge production is approximately 42 1 (11 gallons) per month.
The system is presently operating at an average flow rate of 36 1 (9.5
gallons) per minute with a retention time in the aeration tank of 17.5 hours.
The system has the capability to handle a flow of up to 53 1 (14 gallons) per
minute or 76,000 1 (20,000 gallons) per day with a retention time of 12 hours
(Jhaveri and Mazzacca, 1983).
Biocraft personnel indicate that approximately ,60% of the total 'biodegradation
of the contaminants takes place in the surface reactors, and approximately 40%
takes place in situ (i.e., in the soil and groundwater) (Dr V Jhaveri,
personal communication, 1985). The average one-cycle removal efficiency for
the surface bioreactors is 88-98% for all contaminants except dimethyl aniline
(which is 64%) (Jhaveri and Mazzacca, 1983). The system began operation in
August, 1981. As of June 1985, Biocraft personnel report that the site is 95%
remediated, and operation is expected to be terminated in 1986.
250
-------
6.5.5 Cost Data for the Biocraft Site
The following information is presented to illustrate the cost of applying in
situ treatment using biodegradation. Although the Biocraft site data may not
be directly extrapolated to other sites of differing extent or differing
contaminant inventories, certain conclusions may be drawn from these data
which illustrate the decisions which have to be made, during the evaluation of
in situ treatment systems.
Table 6-2 presents data on the cost of remedial action at the Biocraft Site.
The data are taken from USEPA (1984c) and from personal communications with
personnel at the Biocraft Site. The total capital cost for the remediation
(sum of Items 1, 2 and 3 in Table 6-2) was $925,678 of which $446,280 (about
48%) was expended during the feasibility portion of the remediation. Further,
the feasibility study required about 2-1/2 years to complete. While the
percentage of the cost attributable to the feasibility study and its duration
would undoubtedly be reduced in future applications of the Biocraft/GDS system
at other sites, it illustrates the relative investment which may have to be
put into the feasibility study (i.e., Steps II through VII of Figure 6-1)
prior to startup of an in situ treatment application. Note that this figure
($446,280) does not include the RI portion ($73,948 or 8%) of the project nor
does it include the engineering design portions ($119,891 or 13%) of the
overall capital cost.
Independent of the particulars at the Biocraft Site, this discussion
illustrates that application of an in situ treatment system is likely to
involve a significant research and development component which will be site
specific, possibly requiring a lengthy development period and possibly
involving expenditures which will be a large fraction of the total capital
cost.
251
-------
TABLE 6-2
COSTS OF REMEDIAL ACTION AT THE BIOCRAFT SITE
(Cost Data From USEPA, 1984c)
Project Element
1. Activities Associated With
Remedial Investigation (Step I. Figure 6-1)
a. Monitoring Wells and Test
Borings Installation
b. Laboratory Testing (Independent
Laboratory plus 400 hrs Biocraft
Time @ $50/hr)
c. Consultant Charges (including
200 hrs of Biocraft time, unit
cost not specified)
Total RI Costs
2. Activities Associated With Feasibility
Studies (Steps II - VII. Figure 6-1)
a. Labor (including in-house labor)
b. Equipment
i) Pilot Plant (building, piping, pumps)
ii) Other
Total Equipment
c. Quality Control Lab
Total Feasibility Study Costs
3, Implementation Costs
a. Biostimulation Plant Design and
Construction
i) Engineering Design
- Biocraft in-house (360 hrs @ $50/hr)
- Engineering, Drafting
Total Biostimulation Plant Design
Expenditure
6874
27704
39370
73948
$ 296280
$ 40000
$ 60000
$ 100000
$ 50000
$ 446280
18000
40400
58400
252
-------
4.
ii) Masonry and Construction
iii) Equipment
Total Biostimulation Plant Costs
b. Delivery/Recovery System Design and
Construction
i) Design
- Laboratory Testing
- Labor (consultants $24673;
Biocraft in-house $26400)
Total Delivery/Recovery Design
ii) Installation (All Contractor Costs)
- Air and monitoring well points
- Trenches, air well construction
and site work
- Supervising Geohydrologist
- Engineering
Total Delivery/Recovery Installation
Total Delivery/Recovery
Systems Costs
Operating Costs (3 per day)
a) Utilities
- Electricity (26.4 Kw, 24 hrs/day)
- Steam (72 Ibs/day @ 90 PSI)
Total Utilities
$ 73975
$ 88832
$ 221207
$ 10418
$ 51073
$ 61491
$ 12740
$ 805QO
$ 21513
j 8000
$ 122753
-------
b) Maintenance Costs
- Quality Control Laboratory (technician)
- Fermentation Laboratory (technician)
- Maintenance
- Supervision
Total Maintenance
c) Nutrient Salts
Total Daily Operating Costs
Cost per gallon: $225.50/13680 gal/day - $0.0165/gallon
$ 24.40
$ 97.10
$ 20.26
$ 17.14
$ 158.90
* 19.20
$ 225.50
254
-------
References
Ellis, W. D., J. R. Payne, A. N. Tafuri and F. J. Freetone. 1984. The
Development of Chemical Countermeasures for Hazardous Waste Contaminated Soil.
EPA-600/D-84-039. Municipal Environmental Research Laboratory, U.S.
Environmental Protection Agency, Cincinnati, OH.
Flathman, P. E., W. C. Studabaker, G. D. Githens and B. W. Muller. 1983.
Biological Spill Cleanup. In: Proceedings of the technical seminar on chemical
spills, October 25-27, Toronto, Ontario, Canada. Technical Services Branch,
Environmental Protection Services, Environment Canada, pp. 117-130.
Flathman, P. E., J. R. Quince and L. S. Bottomely. 1984. Biological Treatment
Ethylene Glycol-Contaminated Groundwater at Naval Air Engineering Center
Lakehurst, NJ. In: Proc. 4th Nat. Symp. on Aquifer Restoration and Groundwater
Monitoring. Nat. Water Well Assoc. , Worthington, OH, pp 111-119.
Flathman, P. E. and J. A Caplan. 1985. Biological Cleanup of Chemical Spills.
\oor°CeedlngS of Hazmacon 85- Assoc. of Bay Area Governments, Oakland, CA,
pp 323-346.
Jhaveri, V. and A. J. Mazzacca. 1983. Bio-reclamation of Ground and
Groundwater-Case History. 4th Nat. Conf. on Management of Uncontrolled Hazardous
Waste Sites. HMCRI, Silver Spring, MD, pp 242-247. Also, V. Jhaveri, A. J.
Mazzacca and J. K. Mahon, personal communication, Groundwater Decontamination
Systems, Inc., Waldwick, NJ.
Little, A. D., 1983. Handbook for Evaluating Remedial Action Technology Plans.
EPA-600/2-83-076. Municipal Environmental Research Laboratory. U S*
Environmental Protection Agency, Cincinnati, OH.
Repa, E. and C. Kufs, 1985. Leachate Plume Management. Draft Report for
Hazardous Waste Engineering Research Laboratory, U.S. Environmental Protection
Agency j Cincinnati, OH.
Ryckman, M. D. 1984. Detoxification of Soils, Water and Burn Residues from a
Major Agricultural Chemical Warehouse Fire. In: Proceedings of the 5th National
Conference on Management of Uncontrolled Hazardous Waste Sites. HMCRI Silver
Springs, MD. pp 420-426.
USEPA. 1984a. Review of In-Place Treatment Techniques for Contaminated Surface
Soils. EPA-540/2-84-003a. Municipal Environmental Research Laboratory, U.S.
Environmental Protection Agency, Cincinnati, OH.
EPA-
o,r™ studies 1-23: Remedial Responses at Hazardous Waste Sites
-84-002b. MERL, U.S. Environmental Protection Agency, Cincinnati, OH.
USEPA. 1985. National Oil and Hazardous Substances Pollution Contingency Plan
Federal Register, 50(29) :5862 Feb 12, 1985.
255
-------
Wetzel, R. S., S. M. Henry, P. A. Spooney, S. C. James and E. Heyse. 1985.
In Situ Treatment of Contaminated Groundwater ans Soils, Kelley Air Force
Base Texas. In: Land Disposal of Hazardous Waste: Proceedings of the llth
Ann. Research Symp. EPA/600/9-85/013, HWERL, U.S. Environmental Protection
Agency, Cincinnati, OH.
256
-------
INDEX
Achromobacter, 78
Acinetobacter, 78, 85, 88, 106
Activated Carbon, for organics adsorption, 52-53 101 105 179
Adsorption (sorption) '
of organics, 52
of surfactants by soils, 143-147
Aeration of soils, 94, 249
Aerobic bacteria, 78-79
Aerobic biodegradation of organics, 75, 82-85, 94
Alcohols
biodegradation of, 81, 113
halogenated, hydrolysis of, 170-171
oxidation of, 188, 194, 205
polyoxyethylnated, as surfactants, 131
solubility of, 134
Aldehydes
biodegradation of, 81, 113
oxidation of, 188, 194, 205
solubility of, 134
Aliphatic hydrocarbons
biodegradation of, 81-84, 106, 113
oxidation of, 188, 194, 205
solubility of, 134
surfactant-assisted flushing of, 138-142
Alkanes, biodegradation of, 80-83, 113
Alkyl Halides
hydrolysis of, 168-170
oxidation of, 188, 194, 205
Alkylating Agents, hydrolysis of, 179-180
Alternatives for delivery and recovery systems, analysis of, 63-70 233-242
Amides and Amines
as surfactants, 130
biodegradation of, 113
hydrolysis of, 173, 175
oxidation of, 188, 194, 205
Ammonium salts, quaternary, as surfactants, 130
Amphoteric surfactants, 129, 131, 147
Anaerobic bacteria, 78
Anaerobic biodegradation of organics, 75, 85-90, 94
Analog enrichment for cometabolism. See Cometabolasm, analog enrichment for
Anionic surfactants, 128, 130, 146-147
Application rate of reactant, calculation of required, 17-20 40-43
Aromatic hydrocarbons '
biodegradation of, 79, 81, 83-84, 105-106, 113-115
chlorinated. See Aromatic hydrocarbons, halogenated
halogenated, biodegradation of, 85-88, 113
halogenated, oxidation of, 189, 194, 205
polynuclear (polycyclic), biodegradation of, 84-85
257
-------
oxidation of, 189, 194, 206
solubility of, 134
surfactant-assisted flushing of, 138-142
Arthrobacter, 78, 82, 85, 87-88
Assessment
of site characteristics, 92, 236
of waste characteristics, 92
Bacillus, 82, 99
Bacteria, biodegradation by, 77-79
Bench-scale (laboratory) testing, 221-222, 225-226, 231
Benzene
biodegradation of, 106
oxidation of, 188, 194, 205
Bi-Chem, 98
Biocraft Site remediation, 101-104, 243-255
Biodegradability of organics, relative measures of, 81
Biological degradation (Biodegradation), 74-125
aerobic, 75, 82-85
anaerobic, 75, 85-90
analog enrichment for cometabolism. See Cometabolism, analog enrichment
for
case histories', 96-110, 243-255
development of microbial agents for, 76-80, 92-96
evaluation of feasibility of, 217-223
parameters influencing, 80, 92-95, 99-101
oxygen sources for, 94
BOD, 81, 108, 196
Brevibacterium, 78
Buried pipes (drains), gravity recovery using, 48-52, 101, 104
Calcium hypochlorite
Carbamates
hydrolysis of.
Carboxylic
, oxidation of organics by. See Hypochlorites.
173, 176, 177, 179
as surfactants, 130
biodegradation of, 81-83, 113
oxidation of, 188, 194, 205
solubility of, 134
Catabolism of organics, 75-76
Catalysts
for hydrolysis, 166-167
for oxidation by hydrogen peroxide, 186, 188
Cationic surfactants, 129, 130, 146 '
CERCLA, 212
Chemical oxidation
of organics. See Oxidation of organics,
Climate, influence on selection of delivery and recovery systems,; 69-70
COD, 81, 196
Cometabolism, analog enrichment for, 88-89, 94-95
Commercial microbial products, 77, 91-93, 219
Conceptual design, 223, 227, 231
258
-------
Contamination depth
and selection of delivery systems, 15, 29, 40, 65-67 237
and selection of recovery systems, 44, 57, 65-66, 68,' 240-241
Corynebacterium, 78, 82
Deep wells for forced recovery systems, 56-58
Degradation techniques. See In situ treatment technologies.
Dehalogenation, 85-88
Delivery technologies, 13-43
Depth to contamination. See Contamination depth.
Detergents. See Surfactants.
Detoxsol, 97
Ditches
gravity delivery using, 29-31
gravity recovery using, 47-48
Effective porosity, 7-8
Elecro-osmosis, 60-62
Emulsifiers, 128. See also surfactants.
in biodegradation, 93, 96, 99-100
Epoxides
halogenated, hydrolysis of, 170-171
hydrolysis of, 170, 172
oxidation of, 188, 194, 205
Esters,
phosphonic and polyphosphoric acid, as surfactants, 130
carboxylic acid, hydrolysis of, 170, 172-173
phosphonic and phosphoric acid, hydrolysis of, 176, 178
phthalate acid, biodegradation of 90, 113
solubility of, 134
Extraction (soil flushing) techniques. See Flushing.
Fatty acids. See Carboxylic acids.
Fermentation, degradation of organics by, 75
Fertilizer. See Nutrients.
Flavobacterium, 78
Flooding, gravity delivery by, 21-24
Flushing
evaluation of feasibility of, 223-227
soil, using surfactants, 126-156
soil, using water, 126
Forced delivery systems, 4, 35-39 See also Infiltration wells, Open-ended
pipes, Slotted pipes .
parameters affecting selection of, 4-12, 35, 38, 39-43 63-70
Forced recovery systems, 4, 53-57 See also Deep wells, Well points, Vacuum
well points •,.,-..,...,...•.. f.. , ,
parameters .affecting selection ,of., 4-12, 44, 57, 59, 63-70
259
-------
Genetic engineering of microbes for biodegradation of organics, 76, 87-88
Geohydrologic parameters in selection of delivery and recovery systems,
4-12, 236-241
CDS. See Biocraft site remediation.
Glycols
as surfactants, 131
biodegradation of, 113
oxidation of, 188, 194, 205
Gravity delivery systems, 3, 15-35. See also Ditches, Flooding, Infiltration
beds, Infiltration galleries, Irrigation, Ponding
parameters affecting selection of, 3, 4-12, 16-20, 39-43, 63-70, 237-240
Gravity recovery systems, 4, 45-53. See also Buried drains, Ditches.
parameters affecting selection of, 4-12, 43-44, 57, 59, 63-70, 240-241
Green sand, glauconitic, as an adsorbent, 52-53
Groundwater
characteristics affecting biodegradation of organics, 80, 92-96, 217,
characteristics affecting hydrolysis of organics,
characteristics affecting oxidation of organics,
characteristics affecting selection of surfactants, 143, 145-147
depth to, influence on selection of delivery and recovery systems, 65,
67-69, 237, 240-241
velocity, 7
Halogenated organics. See also Alkyl halides, Phenols, PCBs.
biodegradation of, 85-88
hydrolysis of, 168-171
oxidation of, 188, 194, 205
solubility of, 133-134
Herbicides, biodegradation of, 88-90, 113
oxidation of, 189, 194, 206
Heterocyclic nitrogen, oxygen and sulfur compounds, oxidation of, 189, 194, 206
Hydraulic conductivity of soils, 7-10
and grain size correlations, 9-10, 11
as a parameter in selection of delivery systems, 15, 35, 66, 67
as a parameter in selection of recovery systems, 44, 59-60, 66,
methods of measurement, 9
Hydrobac, 92, 99
Hydrogen peroxide
as a source of oxygen for biodegradation, 94, 101, 105
limitations on in situ use of, 190-191
oxidation of organics by, 187-191
properties of, 186-187
Hydrogen sulfide, from anaerobic biodegradation, 75
Hydrolysis of organics, 157-184
acceleration of rates of, 167-180, 230
definition of, 157-158
evaluation of feasibility of, 227-231
parameters affecting hydrolysis rates, 162-167
rates of, 158-159
Hydroquinones, oxidation of, 189
68
260
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Infiltration bed, 33-35
Infiltration gallery, 31-33, 101, 103, 104, 250
Infiltration, soil, 5-7, 15, 20, 21
as a parameter for selecting delivery systems, 15-16, 67
methods of measurement, 6-7
Information requirements
for evaluation of biodegradation, 217, 219-223
for evaluation of hydrolysis, 229-231
for evaluation of oxidation, 232-233
for evaluation of soil flushing, 223, 225-227
for selection of delivery and recovery systems, 16-17, 37-39, 43-47,
63-70, 234-241
Injection wells, 35-39, 105, 107, 199-202
In-place treatment technologies. See In situ treatment technologies.
In situ treatment technologies
biodegradation, 74-125
hydrolysis, 157-184
oxidation, 185-211
selection of, 214-233
soil flushing, 126-156
Interfacial tension of organics in water, 132-137, 139, 140
Ion exchange, as an immobilization technique, 52-53
Irrigation, gravity delivery by spray. See Spray irrigation.
Kow (octanol/water partition coefficient), 132-135
Karlsruhe, W. Germany, ozonation of groundwater at, 199-202
Ketones
biodegradation of, 113
oxidation of, 188, 194, 205
solubility of, 133
See Bench-scale testing.
Laboratory (bench-scale) testing.
Lime and limestone
as an adsorbent, 52
to adjust soil pH, 52, 107
Mercaptan
as surfactants, 131
oxidation of, 189
Methanogenic bacteria, 78
Micelles, 128
Micrococcus, 78, 84
National Contingency Plan (NCP), 213
Nitriles
biodegradation of, 83, 113
oxidation of, 188, 194, 205
Nitrosamines
biodegradation of, 91, 113
oxidation of, 188, 194, 205
Nitro-substituted aromatics. See Aromatics, nitro compounds.
261
-------
Nonionic surfactants, 129, 131, 146, 147 ,
Nutrients
and microbial activity, 80, 93-97, 103, 105-106, 217, 248
Octanol/water partition coefficient (Kow), 132-135
Olefins, oxidation of, 189
Open-ended pipes, use in injection wells, 35-36
Organics. See also specific organic groups.
aerobic biodegradation of. See Aerobic biodegradation
anaerobic biodegradation of. See Anaerobic biodegradation
flushing of. See Flushing
hydrolysis of. See Hydrolysis of organics.
oxidation of. See Oxidation-of organics.
summary matrix of treatment technologies for, 215
Organophosphates. See also Pesticides
biodegradation of, 115
hydrolysis of, 179
oxidation of, 189, 194, 206 !
Oxidation of organics, 185-211
evaluation of feasibility of, 232-233
using hydrogen peroxide, 187-191
using hypochlorites, 204-209
using ozone, 193-203
Oxidizing agents. See Oxidation of organics.
Oxygen sources for aerobic biodegradation, 94, 103-105
Ozone
as a source of oxygen for biodegradation, 94, 199-202
oxidation of organics by, 193-203
Partition coefficient, octanol/water, See Octanol/water partition coefficient.
PCBs
biodegradation of, 87-88, 95, 114
surfactant-assisted Flushing of, 138-142
Perfluorinated anionic compounds as surfactants, 130
Permeability, soil. See Hydraulic Conductivity.
Permeable Treatment beds, 52-53
Peroxide. See Hydrogen peroxide.
Pesticides
biodegradation of, 86, 88-90, 115
hydrolysis of, 176, 179-180
oxidation of, 189, 194, 203, 206
Petrobac, 108
pH
adjustment of, 94, 179
effects on biodegradation, 94
effects on hydrolysis, 162-164, 167-180
effects on oxidation using ozone, 195
effects on surfactant-assisted flushing, 143-146
Phenols and halogenated phenolic compounds. See also PCBs.
biodegradation of, 83-84, 109-110, 113
oxidation of, 189, 194, 205, 207-208
262
-------
solubility of, 134
surfactant-assisted flushing of, 141
Phosphonic and phosphoric acid esters. See Esters, phosphonic and phosphoric
acid.
Phototrophic bacteria, 79
Phthalate esters. See Esters, phthalate acid.
Physical containment, 2
Pilot-scale testing, 95, 222, 227, 231
Pipes for delivery and recovery systems, selection of
composition of, 33
Plasmids, 78, 85, 87, 89-90
Polybac, 96, 108
Polychlorinated biphenyls. See PCBs.
Polynuclear aromatic hydrocarbons. See Aromatic hydrocarbons, polynuclear.
Ponding, gravity delivery using, 24-26
Porosity, soil, 11-14
Pseudomonas, 78, 82-84, 87, 89-90, 99, 106
Recovery technologies, 43-63
Recovery wells. See Deep wells, Well points, Vacuum Well points.
Redox potential, requirements for anaerobic biodegradation, 78
Remedial investigation/feasibility studies (RI/FS), 213-242, 243-255
Salt content. See TDS
Selection of in situ treatment technologies. See In situ treatment
technologies
Site specific characteristics
in selection of delivery systems, 15-17, 63-70, 237-240
in selection of recovery systems, 43—44, 63—70, 240-241
Slotted pipes, use in injection wells, 35-36
Sodium hypochlorite, oxidation of organics by. See Hypochlorites.
Soil characteristics. See Soil properties
Soil contamination depth. See Contamination depth.
Soil flushing. See Flushing
Soil moisture, effects on biodegradation of, 94
Soil nutrients. See Nutrients.
Soil pH. See pH.
Soil properties. See also Hydraulic Conductivity, Infiltration, Porosity,
Specific retention, Specific yield
and selection of delivery systems, 16, 65-67, 69-70, 235-239
and selection of recovery systems, 43-44, 65—66, 68-70, 240-241
and selection of treatment technologies, 214, 217, 219, 221-222, 225-226
Specific retention, 12
Specific yield, 12
Spray irrigation, gravity delivery using, 27-28
Sprinkler irrigation. See Spray irrigation.
Subsurface aeration. See Aeration of soils.
Subsurface drains. See Buried pipes.
Sulfonic acids and sulfoxides
as surfactants, 130
oxidation of, 188, 194, 205
263
-------
Superfund. See CEBCLA.
Surface drains. See Ditches.
Surface flooding. See Flooding.
Surface irrigation. See Spray irrigation.
Surface ponding. See Ponding.
Surfactants
as flushing solutions, 126, 136-142
biodegradation of, 148-150 ,
classes of, 128-131
evaluation of feasibility for in situ use of, 223-227
parameters affecting the use of, 143-147
properties of, 127-136
solubilities of, 132-135, 139-140
toxicity of, 149-151
TDS (total dissolved solids)
effects on hydrolysis of, 166
effects on oxidation of, 143-145
Technology selection. See In situ treatment technologies
Thiols, oxidation of, 188, 194, 205
Tilling to increase infiltration rates, 15, 20, 21
Topography in selection of delivery and recovery systems, 15, 67, 68
Treatability studies. See Bench-Scale Studies.
UOD, 81
Vacuum well points for forced recovery systems, 54-56
Velocity of groundwater. See Groundwater velocity.
Vibrio, 78
Waste characteristics in technology selection, 217, 219, 221, 223, 225-226,
229-230, 232-233
Water table. See Site specific characteristics
Well points for forced recovery systems, 54-56
Zeolites, as adsorbents, 52
Zero point of change (ZPC), 144-145
264
-------
APPENDIX A
BIOLOGICAL DEGRADATION OF
ORGANIC MATERIALS
A-l
-------
APPhHDlX A
EVALUATION OP SYSTEHS TO ACCtLtKAIE STABILIZATION OF HASTE PILtS OK ULFUS11S
BIOLOGICAL DEGRADATION OF ORGANIC MATERIALS
Acyclic hydrocarbons "
alkanea
methane
N-alkanea
. (C2 to €44) ;
^ ethane
| 5
ro
propane
butane
N-dodecane (C-12)
N-tridecane (C-13)
H-hexadecane (C-16)
H-dodecane +
K-hexadecane
N-tridecane +
N-hexadecane
BIOLOGICAL
Oxidation
Oxidation-
oxldases and
dehydrogcnases
Cooxidations while
grown on a* thane
Cooxidations while
grown OD methane
Oxidation of propane
Cooxidations while
• grown on methane
Cooxidations while
grown on acetate
PRODUCTS
C02, H20
Monoteralnal oxidations to primary
alcohols; to aldehydes; and to
MODOcarboxyllc acids
Acetic acid, ethanol, acetaldehyde,
C02. cell naterlal and
extracellular constituents
Propionlc acid, propanol, acetone.
C02, cell material and
extracellular constituents
Butyric acid, 1-butanol 2-butaaone
1,12 dodecanidoic acid
1,13 tridecanidolc acid
1,16 hexadecanldolc acid
1,12 dodecanidoic acid +
1,16 hexadecanldoic-acld
1,13 tridecanidolc acid +
1,16 hexadecanidolc acid
BIOLOGICAL
AGEHT(S)
Hethylotropic bacteria
In aolls
Microorganisms
Pseudoaonas aethanlca
Pseudottopas methanica
Hycobacterluga vaccae
Paeudoaonas methanica
Candida cloacae „
ENV1ROHHENTAL
REQUIREMENTS
aerobic, landfills, natural gas
leaks
aerobic, growth on cethanc
aerobic, growth on methane
aerobic
aerobic, growth on mettiane
aerobic, growth on acetate
SUBSTRATE CONTACT
CONCfcMTRAUON UhB KEPERENCES*
40-501 i 10X V/V 68
6, 59, 75,
100 114
52 ethane, 45X methane, 79
SOX air
30X propane, '40X methane, 79
SOX air
79
79
5X V/V 79
5X V/V
3* V/V
5X V/V
5X V/V
N-octadecane (C-18)
Esters produced via
cooxldatlons of an
n-Alkane substrate to
a hoeologous osygenated •
compound without degradation
CH,(CH-).,-COOCH,-(CH~)lfi-CH +CH,
COOCH2-CCH2)i6-CR3 (111 Sixture)J
Hicrococcus cerlfecanB aerobic, growth on alkane
substrates
Saturated soils
79
3, b, 115
* References listed following Section 2 of this report
-------
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EHV
R
IAL
HHE
IEMEHTS
-------
SUISTMTt
Saturate** aUyl lulU»
bream chloroM thane
•TMQll ( cMoro** tltitt*
broaofora
chlorinated hydrocarbon!
carbon tetrachlorlde
chloroacetic acid
chloroacetate
chlorobutyrlc acid
chloroethane
chlorofora
2-chloroproplonlc acid
dibronochlorone thane
dlbroooMthaa*
dlbroao* thane
dlchlo roe thane
1,2-dichloroethane
dlchlofop-nethyl
nethyl chloride
rs-gerve
propachlor
trf chloroethane
1 i 1 ,1-trichloroe thane
BIOLOGICAL
ACTIVITY niODOCTS
anaerobic raiplratloa, COj CM*, H*( Br~ and Cl~
rcductlre dtchlorinatlo*
Partial the*, tram for-
mation (|r) reductive
4*chloriaxti»a
anaerobic dechlorl na-
tion partial che».
tram formation (Br)
reductive dachlorlna- C02, CH4, {{**, Cl~
tlon, anaerobic and
aerobic reaplratlon
reductive dechlorln*- C02 CH^, K* and Cl~
tlon,_anatroblc
raaplratlon cooxidatlon
Dahslogenatlon,
oxidation
Dehalogenatlon, C02, Cl~
oxidation Beta
Dehalogenation
oxidation
reductive dechlorlna-
tlon
reductive d cch lor 1 na-
tion, anaerobic reaplra-
tlon cooxidatlon
Dehalogenatlona,
oxidation
reductive dechlorlna-
tion anaerobic rea-
plratlon, partial
chea. transformation
(Br) cooxidatlon
Dehalofenation
Partial Chen, trans-
formation (Br),
anaerobic respiration
cooxidatlon
Seductive dechlorina-
tion, anaerobic rea- C02, CH^, H+ and Cl~
plration cooxidatlon
Reductive dechlorina-
tlon decarboxylation
Reductive dechlorlna-
tlon
Dehalogenation
Dehalogenation
Reductive dechlorina-
tlon, anaerobic res-
piration cooxldstion
IIOUCIGU.
Soil bactcritui, «*tb*no-
(«Ric •iKcd culture with
aev<{« alodje
Bacteria, 8I-CHEH CEC-1,
BI-CHEM FOB- 6 and
BI-CHCM rEP-7
Kethanogenic Mixed cul-
ture with acetate aa
aubatrate
Sewage bacteria
PBeudononas ep.
Sewage bacteria
Soil bacteria
Kethanogenlc alxed cul-
ture with acetate aa
aubstrate
Sewage bacteria
Hethanogenic nixed cul-
ture with acetate a«
aubetrate, soil bacteria,
atwaf* aludge
Soil bacteria
Soil bacteria, nethano-
genic aixed culture with
acetate as substrate
Bacteria
Soil bacteria
Soil bacteria
Soil bacteria, mlcrobial
culture
Marine bacteria, sewage
sludge
Methanogenic mixed cul-
ture with acetate as
substrate
SftCUI. INITIAL
CKYIBOWttXTAt. StlfSTXATK
UQUIUXTXTS COKCgfTitATIQH
•rueroblc, contlnuoui flow 37 u$/l
fixed fll« reactor
fixed flla reactor
anaerobic, aerobic
54 ux/1
anaerobic, continuous flow
fixed flla reactora,
denltrtfication
47 ug/al
aerobic, 29'C
anaerobic
50 ug/nl-60 ug/1
anaerobic, continuous flow
fixed flla reactor
63 ug/1
anaerobic, continuous flow
fixed flla reactor.
deoitri flea tlon
Anaerobic
Anaerobic, continuous flow ?B u g/1
fixed filn reactor,
denitrlfication
Anaerobic, continuous flow
fixed fila reactor
Anaerobic
Aerobic, anaerobic
Anaerobic, continuous flow
fixed film reactor,
denltrif lea tion 53 u a/1
CONTACT
T1H1
2 day*
2 davi
16 duya
iOI iu 2 days
10 days
50 days
35X, 2 days
2 days
2 days
8 weeks
8 weeks
irmtxccs
57
13 15
1«, 57
2, U, 1
2U, 45,
51, 11
1J, 14
45
12
45
57
13, 15,
44
13, 14,
57
13, 14
57
13
2
57
2
2
57
13, 15
13. 15.
71
15
57
-------
en
E 11 B STRATH
1,1 ,2,2— tetrachloroe thane
trlchloronethane
tetrachloronethane
trlchlorofluorooe thane
•ethylene chloride
Bnuturated Alkyl Halldea
1,1-dlchloroethylene
1 ,2-d Ichloroethylene
1,2-tranad Ichloroethylene
1,3-dtchloropropylene
Heptachlor
perchloroethylene
trlchloroethylene
tetrachloroe thy lene
vinylldlene chloride
Ethers, Glyr.ols, Epoxldes
Isopropyl ether
bia-(2-chlorolsopropyl)
ether
1, 2-epoxyoctane
glycidol
ethylene glycol
glycidyl nitrate
glyccrol
BIOLOGICAL
ACTIVITY PRODUCTS
Reductive dechlorlna-
tlon, anaerobic res-
piration cooxidaeion
Reductive dechlorina- C02> CH^ K* and Cl*~
plration
Reductive dechlorina-
tion, anaerobic res-
piration cooxidaeion
Reductive
dahalogenatlon
Dechlorlnatlon,
oxidation
Dechlorlnatlon,
oxidation
Dechlorinatlon,
oxidation
Dechlorinatlon,
oxidation
Dechlorination,
oxidation
Epoxidatlon
Reductive dechlorina- C02, CH$, H* and Cl~
tion, anaerobic res-
piration, oxidation
Reductive dechlorina- C02, CH4, H+ and Cl~
tion, anaerobic res-
piration, oxidation
Hydroxlation of free
aethyl group, beta
oxidation
Oxidation C02, H20j C02, CH4, and H+
Oxidation C02, H20, glycollc acid, pyruvate
to C02 and H20
Oxidation Glycerol to C02, H20
Oxidation Pyruvate to C02, H20, CH4
BIOLOGICAL
ACENT(S)
substrate
HethanOgenic mixed cul-
Bubstrate
Methanogenic mixed cul-
subtrate
Sewage sludge
Soil bacteria
Sewage sludge
BI-CHEH CEC-1
Sewage sludge
BI-CHEH G2C-1
Sewage sludge
Sewage sludge
Hicrobial culture, soil
bacteria
PHENOBAC
Methanogenic culture
Methanogenic culture
Soil bacteria
Sewage sludge
Pseudomonads op.
Activated sludge micro-
organisms, digester
microorganisms
Salt requiring bacterium
T-52, Acetobacter sp. ,
Gluconobacter sp. ,
Acenltobacter ep.
Activated sludge micro-
organisms, digester
microorganisms
Activated sludge micro-
organisms, digester
microorganisms
SPECIAL INITIAL
ENVIRONMENTAL SUBSTRATE
REQUIREMENTS CONCEN1KAT10N
fixed flln reactor,
denltrification
Aerobic, anaerobic
reactor, denltrification
Anaerobic, continuous flow
Anaerobic
anoxlc conditions
Aerobic (?)
Aerobic (?)
Aerobic (?)
Aerobic (7)
aeration basin
Anaerobic
Anaerobic
anoxlc conditions
Aerobic
Aerobic
Aerobic, 30"C; Anaerobic, 3?*C 100 pp«
Aerobic, 30°C 10-5 g/1 '
Aerobic, 30'C; Anaerobic, 37*C 100 ppa
Aerobic, 30°C; Anaerobic, 37°C 100 ppn
CONTACT
tlHE REFERENCES
1J, 15, 57
1J, 57
57
57
57, 59
57, 59
57
57
2
114
8 weeks 13, 15, 57
8 weeks 13, 15, 57
57
114
57
26
-
50
36 hours ^0
20
20
50
20, 50
polyethylene glycol Polyethylene glycol
dehydrogenase
Bacterial culture
Aerobic
-------
SWISTMTt
Alcohol!
•cthinol
arcanol
laobutyl alcohol
lao-nonyl alcohol
2-ethyI hexanol
decyl alcohol
H-propyl alcohol
tart butyl alcohol
octyl alcohol
laopropyl alcohol
polyvlnyl alcohol
Aldtthydca and Ke tones
foraaldehyde
acetone
methyl ethyl Itetone
BIOLOGICAL
ACTIVITY
o,,..,,.
Oxidation
Oxidation
Oxidation
Oxidation
Oxidation
Oxidation
Oxidation
Oxidation
Oxidation
Oxidation
Oxidation
Oxidation
"~ ~~ sreclATC ~ imiAt ~ — ~ —
C02, H20 rHEHOUC, loll ind Aerobic
Mthanol oxidation
bacteria (Hethyloeoccua
and HethyloBonaa)
PIIEHOBAC
PIIEHOBAC
PHENOBAC
PHEHOEAC
PHENOBAC
PIIEHOBAC
PHENOBAC
PIIEHOBAC
PHENOBAC
PVA oxldaoe degradation by Pauedononaa putlda VHISA Bacterial symbiotic relation-
tlon, aerobic, Bloreactor
PHENOEAC Aerobic, digester, land 1 day
farnlng
PHENOBAC Aerobic, digestion, land 1 day
0V, 75
71
75
71
75
75
75
75
75
75
95
75, 114
75
75
acetaldehyde Oxidation
acrolcin
methyl-Iaobutyl-ketone
Carboxyllc Acids
acetic acid Oxidation
,
planning
Pyruvate to C(>2
C02 and H20
9ft
7i
75
Aerobic, anaerobic
-------
SUBSTRATE
N-C44
Short & Jong chain acids
C14~CI8 carboyllc aclda
dfcurboxyltr acids
branched dlcarboxyllc
acids
pentanolc acid
hexanolr acid
octanlc acid
pfopanedlolc acid
butanedolc acid
2-raethyIbutanldloIc acid
pentanedlolc acid
hexanedJolc acid
hepianedlolc acid
octanedtolc acid
nonanedlolc acid
decanedlolc acid
undecanedlolc acid
tridecanedlolc acid
dodecanedlolc acid
benzole acid
An Ides
acetanlde
Esters
methyl acetate
ethyl acetate
arayl acetate
dlethyl adlpate
dibasic eaten
trltMthylaalM
dlethylanine
dlethanalanine
dlnethylanlne
Phthalate Esters
m-phthallc acid
o-phthallc actd
p-phthalic acid
dlbutyl phthalate
nonobutyl phthalate
terephthalic acid
dl-n-butylphthalate
dI-(2-ethylhexyl)
phthalate
n-dlbutyl phthalate
di-n-octyl phthalate
dlethyl phthalate
dlbutyl phthalate
BIOLOGICAL
ACTIVITY PRODUCTS
Degradation of carboxyllc C02 and 1^0
aclda proceeds by beta
oxidation with the C02 and HjO
formation of acetyl- C02 and h^O
CoA, and a fatty acid
shortered by a 2 carbon C02 and HjO
unit. Acetyl-Co A is
netabolized to C02 and
cycle
Catahollem can require
Onega (dlterainal) oxi-
dation for dlcarboxyllc
acids and alpha oxida-
tion or beta alfcyl group
removal for branched
chains
Hydrolase Acetate to C02, antsonla to protein
Degradation
Blodegradation C02
Blodegradation C02
Blodegradation C02
Esterases
Oxidation
Esterasea, oxidation
Est erases, oxidation,
capable of netabolzlng
other phthalate esters
as well
Oxidation
BIOLOGICAL
AGEHT(S)
Bacteria,
Pseudomonads sp.
Brevebacterlum sp.
Flavobacterlum sn.
Hycobacterlua vaccae
Pseudononas sp.
Soil bacteria as
innoculum
Pseudomonas aeruginosa
BI-CHEH GEC-1
Pseudomonas sp.
Peeudomonas sp.
Pseudononas sp.
Mixed culture bacteria
Mixed culture bacteria
Nocardla sp.
Mixed culture
Pseudoroonas sp.
Ps eud oa 1 caTige n es sp.
ptithallc acid ester
degrading bacteria
Microorganisms
positive bacteria
SPECIAL INITIAL
ENVIRONMENTAL SUBSTRATE
Aerobic, anaerobic
Aerobic
Aerobic, anaerobic
Degradation of organic aclda
aerobic, 24 *C
Degradation of organic acids In
24 c 16.2 ppo aa C
1.8 ppa as C
5.2 ppn as C
7.4 ppa a C
9.9 ppa a C
9.9 ppn a C
9.5 ppn a C
7.5 ppn a C
2.6 ppn a C
8.8 ppa BB C
20 ppa as C
Aerobic
Axenlc culture Incubated
aerobic, at 29 "C, in the dark 210 ug/1
conditions;
Aerobic, by Acenetobacter sp.
Anaerobic denitrifying
conditions
Anaerobic denitrifying
conditions
CONTACT
6, 99
7
54
77
77
8b
86
86
86
86
8b
86
95X 86
reduction in 86
21 days 86
86
8Q-90X 86
reduction in 86
9 daya 86
86
86
86
86
86
77
114
114
114
114
114
59
11
50X In 4 days 11
11
1
1
1
1
1
1
8, 86
107
111
107
108
108
-------
110UCKAL
SWSTMTK ACWlTr
Wtylrhthalatt
butylktnyl phthalat*
dtfaofeutyi pHtKalate
dt~iiai.6nylptithlate
phthalate eitera
«.»«,«.
dUethylnltroaaalne
n-nltroBodlethanolaBlne
(NDEIA)
Thfole
ethanethlol
benzene thlol
Cyclic Alkanca
cycloalkaneB
cyclohexane
cyclohexane
^
I n-hexane
00
cycloparafflns
cyclopropane
cyclopentane
methyl cyclopentane
2-cyclopentene
cyclohexane
cycloseptane
cyclooctane
n-butyl cyclohexanol
Kltrllea, cyanidaa
acrylonltril*
acetonitrlle
acrylonltrlle
adlpoiiltrlle
phenylnltrlle
phthaloni trlle
hydrocyanic acid
sodiuta cyanide
djchlorophenyl l&ocyanate
Second order ntcroblal
Mineralized
Biodegradatlon
Mineralization
Oxidation
Cooxldatlons,
comenaalla.
Cooxldatlono,
Cooxldationa.
connensaliaa
with other soil
alcroorganisaa
utilize products as
sources of carbon and
energy
Cooxidation
Degradation
Degradation
Degradation
Degradation
Degradation
Degradation
Degradation
fSOMXTS
ml*
NontuMrlgenlc product
Hot identified but Buggeated to be
non-carcenogenlc dine re of NDEIA
C02
C02, H20
Cyclohexanone, ne totalized to
C02 + H20
Cyclohexanol used by second organisn
Propaldehyde
Cyclopentone
Methyl cyclopentone
Cyclopentanone
Cyclohexanone
Cycloslptanone
Cyclooctanone
Cyclohexaneacetlc acid
IIOU£ICAL EWIIOWKKTAL Mlt&XATE
ACTKT(S) irOUIBEHKHTS LOMCUfTIUTttlM
RhodopaeudOBonaB Anaerobic. .r»f«r» yvii iTr.
capaulata. ruMea organlaaa llluaenated \.\ v^/ml
.
•«terl« aevage pH 7.8 leisonal effect
Bacteria
Mixed Dlcroblal
population
Pseudomonas oleovorans Aerobic, fermentation
Two Pseudoeonss ap.
Mycobacterlum vaccse and
other soiln icroorganisms
Hocardla sp. Growth on n-alkanes
Mixed culture of yeast Aerobic
mold, protozoa bacteria;
activated sludge
Mutant nlcroorganlsn 20'C 500 og/l
Mutant microorganism 20°C 500 og/1
Mutant alcroorganlsn blotreator 100 ppm-1000 ppm
Mutant alcroorganism 20*C 500 og/1
Mutant microorganism 20BC 500 og/1
Mutant nlcroorganlsn 20°C 250 og/1
Mutant nlcroorganlsm 20'C 250 rg/1
Mutant microorganism 20°C 500 ng/1
BI-CHEM CS-1-9
IXAHCI
TDU: «tranias
1M
ion
iUo
114
IU
57.
20 d»yl 113
IS
79
26
7S
79
79
79
79
79
79
79
79
79
43
57
100X in 3 hrs 61,
100Z in 3 hrs 61,
1 ppn, 3 aonths 61,
1002 in 10 hrs 61,
1001 in 16 hrs 61,
100Z In 1 hrs 61,
1001 In 3 hrs 61,
100Z In 8 hrs 61,
59
58
75, 114
75, 114
75, 114
75, 114
75, 114
75, 114
75, 114
75, 114
-------
BIOLOGICAL
SUBSTRATE ACTIVITY PRODUCTS
anallne Degradation
diethylanaUne
.o-cliloroanal Ine Degradation
p-chloroanaline Degradation
4-chloroanallne Dechlorlnation
trichloroarullne Degradation
Aroiutlc hydrocarbon!
•roMtlca
alkyl benzene (ulfonate Cooxldation in the leopropanol & catechol both used as
presence of glucose growth substrates
followed by complete
netabollon of end
products
way aono & dloxygenaae
Heca ( xcradlol) C02
pathwa - hydration or
hydrog nation followed
by non xidaClve ring
fi*8iO
SPECIAL _ INITIAL
BIOLOGICAL ENVIRONMENTAL SUBSTRATE
ACENT(S) REQUIREMENTS CONCENTRATION
Mutant Aerobacter Bp. aOO m/1
Mutant Aerobacter op. 500 nfc/l
Mutant Aerobacter sp. iQO ng/1
Sewage nlcroflora
Mutant Aerobacter Bp. ^OQ «o/l
essential
Bacteria Anaerobic
Peeudomonaa put Ida, Aerobic
Sewage sludge Aerobic
Stabilization pond Aerobic (?)
nlcrobea
CONTACT
TIME REfr'EREMC
100X in 10 lirs 61, 75,
114
10UX in 20 hrb 61, 73,
1001 in IB hra 61, 75,
44
100X in 30 hr. 61, 75,
6
100X in 20 days 43
31
57
i7
57
114
114
114
114
-------
SOimATt
k*a*Mte
benzole acid
dlvlnyl benzene
tiydroxy benzoate
p-hydroxy benzoate
p-hydroxy benaoate
2,4 dlhydroxy benzoate
dlhydroxy banco* te
3 Bechyl benzoat*
benzyl f ornate
c 1 B-C 1 s-Mucona te
n- butyl benzene
ethylbenzene
n-propyl benzene
3-nethyiauconl£ acid
p-lsopropyl toluene
pyrolldone
toluene
o-xylene
p-xylene
BIOLOGICAL
AcriviTr
PtotDMtabotlia -
reduction fey a naval
pathway l»sdla| to
ring fltilc-Q
Kindtlau pathway
Mineralization
3-oioadipate beaioate
patbvay -
Oxidation
Hlcroblal
degradation of oil
abate retort water
Aerobic netabollea
4 , S-oxygenaae pathvay
Bets-ketoadlpstc
pathway
Photoaa a 1 Dilation
3-oxoadlpate pathway
dehydroxybenxoate
dahydrogamaae
Hattdelate pathvay
Oxidation
Muconate-cycloisoBerase
Cooxldation
Cooxldation
Cooxidation
Hucona te-cycLoisoaerase
Cooxldation
Cooxidation
Cooxldatton
ttOCCCTS
""""
fiucclnatt and aectyt-C(/\
by citric acid cycl« to COj * vater
co2
Cana carboxyl-alpha-hydroxy
Buconlc sealaldehyde and subsequent
break dawn involving pyruvate
netabolize to COj and BjO in
Benzoate which is degraded by the
be ta-ke toad 1 pate pathway to
acetyi-CoA and succlnate
(metabolites) of the citric acid cycle
Phenylacetic acid
Phenylacetic acid
Cinnanic acid
p-isopropyl benzoate
Glutanic acid
o-tolulc acid
p-toluic acid,2,3-dehydroxy-
I1DLCC1CAL
A£OT(S)
toodopaeudomus
palttatrta
PieudoMaaa rut Ma
Paeudomonaa sp. B13
Tranaconjugant
Alcaligenea sp, atraln
A 7-2 trans if e r of
ha locate chol-de grading
capacity froa
PaeudoBonaa ap. B13
Mixed population of aoil
bacteria
BI-CHEH GEC-1
ghodopseudoaona a ap.
Paeudoaonas puttda
RhodopeeudOBonas
paluatria
Pseudononas Bp. B13
PaeudoBonas ap., sewage
PaeudoBonas put id a
BhodopseudOBonas
palustria
Pseudononas sp.
Ho card la sp.
Nocardia sp.
Hocardla sp.
P&eudosonaa ap. B13
Hocardta sp.
Bacillus sp.
~ Bacillus sp.
Pseudooonas putlda
Hocardia op.
Hocardla ap.
CKVlKOfiWXTAL SUISTHATt
Aonroblc, vfcoto«yntbatic 20 u aoUa
coadttleoa
Aerobic
Aerobic w>000 ^^
Aerobic
Aerobic, low aenaitlvity to
phenols, 28 *C on shaker culture
Aerobic, phosphate auppleoent 21 ppa as carbon
Aerobic
Aerobic
Anaerobic, growth oo benzoate 20 u molea
in light
Aerobic
Aerobic, plasmld transfer led
to the ability to uae this
conpound as a sole energy and
carbon source
Aerobic
Aerobic, growth on p-hydroxy
benzoate
Aerobic
Growth OR N-alkanes
Growth on N-alkanes
Growth on N-alkanes
Aerobic
Growth on N-alkenes
Anaerobic
Aerobic
COKTACT
TIKE KOtXQCCZS
*1 BiMU* 29
77
99Z In V day* b9
90
91
901 ID 9 days 86
59
29
77
220 mi nut en 29
90
57
77
29
90
43
43
43
90
43
43
57
43
43
p-tolulc acid
tri-p-cresyl phosphate Oxidation
Activated return sludge Aerobic, 2l"C
1 u g/Bl
70-80Z, 24 hrs 62
-------
SUBSTRATE
Halogens ted Aromatlcs
alpha chtorotoluene
chlorobenzene
3-chl urobenzoate
p-j B-; o-chlorobenzoate
d 1 chlorobenzene
3 4-; 3 5-
dlchlorobenzoate
m-dlchlorobenzene
o-d Ichlorobenzene
p-d Ichlorobenzene :
3,4-dichlorobenzoate
2-;>;4-fluorobenzoic
acid
hexach 1 orob«n«n*
nonochlorob«iu*n«
aonochlorobenzoate
trlchlorobenzene
1,2,3- and 1,2,4-
tr Ichlorobenzene
1,3,5- tri chlorobenzene
1,2,4-trlchlorobentene
1,2, 3- tri chlorobenzene
2 ,3 ,6-trlchlorobenzoate
BIOLOGICAL
ACTIVITY
Dechlorlnatlon
Respiration
Dechlorlnatlon
Dechlorinatlon, growth
Dechlorlnation, growth
Use as sole carbon
sources
Degradation, ring
disruption
Degradation, ring
disruption
Degradation, ring
disruption
Use- as sole :carbon
source
3 chlorobenzoate
grown cells readily
cooetabollzed
Honofluorobcnzoates
Degradation, ring
disruption
Conetabollsv
Degradation, ring
disruption
Degradation, ring
disruption
Blodegradatlon
Blodegradation
Degradation, ring
disruption
Blodegradatlon
Dechlorlnatlon,
BIOLOGICAL
PRODUCTS AGENT(S)
Sewage microflora
Sewage microflora
Pseudomonas sp. B13
Alealigehes ap. A7-2
Pseudotnonas op.
Mutant Peeudomonaa SP.
Mutant Pseudomonaa sp.
Kutant PseudomofiaB sp.
Mlcrobia.l population
in sewage "
Benzoate pathways used 2 fluoro-cis- Pseudomonas sp. B13
product of 2- and 3-f luorobenzoate
Complete aeUbolis* of 4-fluorobenzoate >
occurs
Sewage sludge nutant
Paeudomonad sp.
Nocardia ap. t
aycobacteriua
Sevage sludge
2,6-j 2,3-dechlorobenzene, 2,4- & Soil microbes
2,5 di chlorobenzene; (X>2; slow
Mutant Pseudomonaa sp.
Kutant PseudononaB sp.
C02 soil bacteria
Hewport River,
K. Carolina aicrobial
community
Mutant Pseudononas sp.
C02 Soil bacteria
3,5-dlchlorocatechol Brevlbacteriuo sp.
SPECIAL INITIAL
ENV1ROKHENTAL SUBSTRATE
REQUIREMENTS COMCENTRATIOH
Aerobic, 27 °C to 6°C
Plasmid transfer led to ability
to metabolize these compounds
simultaneously; -sole energy and
carbon source
Aerobic, 30°C 200 •«/!
Aerobic, 30°C 200 ag/1
Aerobic, 30*C 200 Bg/1
25eC, dissolved 02 at 100 ug/nl
6-8 ug/&l
Aerobic
Aerobic, 30"C 200 w/1
Aerobic
Aerobic
Aerobic -
Aerobic, 30 "C 200 ng/1
Aerobic, 30"C 200 ag/1
Aerobic, 20°C, 66Z moisture, 500 ppH/g
fertilizer
Fresh, brackish and narlne Dry weight soil
water; 30 *C
Aerobic, 30"C 200 ng/1
A, 20°C, 66Z moisture
fertilizers
CONTACT
TIME
17 ng/l-l/hr-1,
(upstrean)
(estuarlne)
0.03 ng/l'1/
hr"* (Bwriue)
100X 29 hre
1002 26 hra
100X 25 hrs
9 dayu
100X, 15 hra
100X, 52 hrs
100X, 48 hre
9.1 ns/l'Vhr"1
100X, 40 hrs
REFERENCES
45
44
45
27| 57
61, 75, 108
114
61, 75, 108
114
61, 75, 108
114
27
SO
57
61, 75, 108,
57
57
57
61, 75, 108,
114
61, 75, 108,
114
69
BO
bl, 7i, 97,
114
69
43
Inducable enzyne
Induction
-------
StflCTKATT
Staple AroMtlc Nltroftn
compounds
4-amfnobeaKefie
tieMhydro-l ,!,$*•
trlnltro-
1.3,5-triatIne
nitrobenzene
3- and 4-
nltrobenzolc acid
1, 2- and 1,3-
3,5-dtnttroheniolc acid
2,4-dlnitratoluene
^
|
. ! 2,4 ,6— trinl trotoluene
ro
BIOUKICA1.
ACTIVITY
Rfns elevate, possible
•tairalUstlon
Blodeiradatlon.
successive reduction of
nltro groups to • point
where deatabllliatlon
and fragmentation of
the ring occur a.
Ron cyclic degradation
product! arise via
subsequent reduction
and rearrangement of
reaction products.
Ring clevage, possible
nlneraliiation
Aromatic Hltro compounds with
Other Functional Croups
benzylaaine
4-chlorobenzoni trite
4-chloro-3,5-
dlnltrobenzolc acid
4-chloro- 2,5-
dlnitrobenzoic acid
3-and 4-nltrotoluenes
2,6-dinltrotoluene
4-toluidine
Phenols
camphor
cine rone
Mineralization
Mineralization
Mineralization
Dechlorinatlon
Dehalogenation,
come ta bo 1 ism
Dehalogenation
Degradation
Konooxygenases
oxidation
hydroxylatlon
rxooucTS
V.
Hexshydro-I-nltroso-a.S-dlnltro-
1,3,5-trlazlnc, hexaliydro-1,3-
dlnl troso-S-nttro-1 .3 ,5-t rlnltro«o-
1.3,5-trlazIne,liydrarlnc,J,l-
dlmethyl-hydrazIne.l.Z.-dUethylhyd-
raclne, formaldehyde, and methmol
Anallne
Aninobenzolc acid
Aalnonitrobeneolc acid
•
2-anlno-4 ,6-dinl trotoluene;
4 anlno-2l6-dinitrotoluene;
2 , D-dlHmlno-4-nI trotoluene;
Alpha-hydroxymuconic
aemlaldehyde mineralized by
Streptomycea «p.
Toluldine
Aainoni trotoluene
3,4,A; trimethyl-5-carboxymethyl
delta 2-cyclopentone metabolized to
isabutyrate, converted to Isobutyl-
CoA which Is netaboltzed via vallne
catabolista
Cinerolone
llfttCCICAL
ACEKI(S)
S***&
Activated sludge
Stabilization pond
nlcrobea
Sewage
Sewage
Sewage
Stabilization pond
nlcrobea
nicroorganlsBs
Lake water, sewage
Sewage olcroflora
Chlanydofflonas sp. Al and
A2, sewage aicroflora
Sewage raicroflora
Sewase
Sewage
Sewage
Paeudoaonaa putids
Aspergillus nlger
SIZC1AL JK1T1AL
£«VIIOI«fXT*L SUIiTHAIt
UflUIIDEKIS COKCOmuTlOM
Atrollc
Anicroblc, deottrltlcitloo 500-100 ut/«l
Aerobic (1)
Aerobic
Anaerobic or aerobic
Anaerobic or aerobic
Aerobic (?)
1
55*C, 601 nolsture
1 u g/Bl
250 ng/nl or le»«
284 ps/«l
In the light In the absence of
nutrients and In the dark with
acetate
Anaerobic
Aerobic, anaerobic
Aerobic
- • " -
Aerobic, inducable enzyme
systems
CCIKIACI
TIM MTttDittS
(0
4 diys 72
57
40
40
40
40
57
91 da con ostl 51
500 ng/nl^/hr"1 B9
590 pg/nl'Vur"1 104
60 hours 105
45
45
45
40
40
40
77
43
-------
BIOLOGICAL
SUBSTRATE ACTIVITY PRODUCTS
cinnamate HintraUiad C02 and CH$
Hloaralixtd COi
dlphenyl hydrazine
p,p-dlchlorodlphenyl- Co oxidation p-chlorophenylacetate
•ethane
creaol Used as carbon and energy source
p-cresol Hinaraliutlon C02
creosotes
methyl creaotenate Mineralization
ethoxylated phenolo Hlneralizatlon
phenolics Biodegradatlon
Degradation ring
disruption
r— * Organic acida fermented to C02 & CHi
CO
Phenol degradation
affected by Iron con-
innoculua aize and pH
Hlneralizatlon
Mineralization
Mineralization
BIOLOGICAL
ACENT(S)
Sewage sludge
stream bacteria
Sewage sludge
Hydrogenononas ep.
Pseudomonas sp.
Aureobasldium pullulans
Adapted site bacteria
Hlcroorganlsos
BI-CHEH CEC-1
BI-CHEM TEX-4
Microorganisms
BI-CHEM COG-2
Mutant Pseudononas op.
nethanogenlc ecosy sterna.
Oil refinery settling
pond bacteria
Pseudomonas sp..
Vibrio ep.. Spirillum sp.
Bacillus sp. ,
Nocardla ap.
Chlanydoaonas
ulvarensis
SPECIAL INITIAL
ENVIRONMENTAL " SUBSTRATE
REQU IREMENTS CONCENTRATION
Anaerobic
Aerobic
Aerobic
Aerobic
Anaerobic, sediment and water SO ig/1
20 °C
Pits, ponds, lagoons soils and
waste water
30 °C
Anaerobic
100-400 as/liter
Aerobic, shaker culture 28 °C.
Maximum degradation rate is
100 mg/ liter with a pH optimum
between 7 and 8. Continuous
light.
Aerobic
Aerobic
Aerobic , light required
CONTACT
TIME REFERENCES
31
20X ID 500 bts 63
57
57
43
70 hr> M
23
59
59
28
59
61, 75, 108,
110, 11*.
31
34 to 100 bourn 45
57
57
57
-------
•lOUtCICAL
SVBSTUTt ACTIV1TT MCKXXTS
Humiliation
Himrallutlon
Hleerallxatloa
Dcsre4atfoa, rlcj COj, water, cell protopleta
disruption
Oxidation
Mineralization
Oxidation, ortho Beta ketoadlpate
pathway, phenol
hydfoxylaae, catechol
l,2-oxy£enaB0
Mineralization . COj
phenyl phenol
reaorclnol
Aroaattca ulth Halogenated
Side Chain.
l,l-dIphenyl-2,2,2-
trlehlorothane Coaetabollaa 2-ph5nyl-3,3,3, trlehloroproplonle
acid
p,p-dlchlorodlphenyl-
nethane Coaetabollaa p-chlorophenylacetate
2,4,5-trlchlorophenoxy
acetic acid Coaetabollaa 3,S-dlchloroentechol
atyrene
II0UCICAL
ACZXKSt
Ptorldlua fu»Mlarn».
SeeaedUaua baielllenaea
Eunlena traellua
Cerynebacterlaa J9.
Hjtant PaeudtMM>oaa tp.
Paeudoaonaa ap. atraln
813, AlcaTTaenea ap.
atraln A 7-2
Lake water bacteria
aewage bacteria
Trlchoaporon
cutanlua POB14
Freahwater and aewage
bacteria
BI-CHEM CEC-1
BI-CHEH CEC-1
BrevlbacterlUD ap.
Hydrogenoaonaa ap.
Hydroeenoaonaa ap.
BI-CUEM CEC-1
— mrao: • IKITI/U.
UVUO»C]ITAL suisnAit oataua
izgJittxzxTS coNUxnuriui TIW.
Aerablc
Aerobic, light required
Aerobic
Aerobic, 30*C Inorganic jco «/i J4 hrt
fertilliera (H and P)
culture
Aerobic, 29'C, no lllualnatlon
Trace 1-100 ug/al
Aerobic, 30*C growth In a
Karublahl jar feraentatlon 250 «8/l 100X 14 hra
Aerobic, pU 7.0 ,29*C
'i'rece
Aerobic
Aerobic
Ktmrxos
SI
a
17
61. 75, 108,
111 114
91
89
96
105
5V
59
43
43
43
59
BI-CHEH SUS-8. Detoxsol Aerobic. Contanlnated soil 25
depth Co 8 Inchea
Halophenola
pentachlorophenol
Oxidation uae aa sole C02
aource of organic
carbon and energy
Bacterltm KC-3
Aerobic, contlnuoua flow 200 gm/1
enrichaent culture pH 7.5, 25*C
aonochlorophenol
4-chlorophenol
Metabolized aa aole
carbon aource
Sals energy "nd csrbon
source
Oxidation, sole carbon
and energy source
1,2-dioxygenatlon of
chlorocatechola
Blodegradatlon
C02, cell ness
Soil microbes Anaerobic
BI-CHEM PEP-7
Arthrobacter ep. atrain Aerobic, chenoctac
NC '
Hocardla Bp., Aerobic
Hycobacterlun Bp.
PoeudomonaB sp. B13
Pseudomonas sp..
sewage
Pseudononas, ap.
strain B13,
ortho chlorophenol
A7-2 mutant bacteria
Mutant bacteria
Aerobic, use of plaenid
transfer to enhance netabollsn
Aerobic, pond leachates of
phenols and o-chlorophenols
Aerobic, contaminated soil and
pond spray injection leachate
systeo. biotreator pond
250 pp.
15,000 ppa
1 pp., 21 daya 59
73t ID 24 lira 24
57
59
97Z reduction 101
90
57
91
75Z In 36 days 110
1 ppm, 9 months 114
-------
I
01
SUBSTRATE
2 , 5-d ihromophenol
n-bromophenol
pen t a chl ore phenol
2,3,5-trichlorophenol
2,5-dlchlorophenol
p-chlorophenol
HI trophenola
nltrophenol
p-nl trophenol
fenl trotlilon
(0,0-dlnethyl-0|3-
ae thyl-4-n I trophenyl J
phosphorothionate)
Phenol B-D I hydride a,
Polyhydrldcs
t-butyl catechol
catechol
chlorocstechol
4-chlorocatechol ... .
3,5-dJchlorocatechol
3,4-chlorocatechoJ.
3-methylcatechol
4-chlororesorclnol
protocatechuate
Polycyclic Hydrocarbons
Two and Three Ring Fused
Polycyclic Hydrocarbons
anthracene
Aroclor 1221
BIOLOGICAL
ACTIVITY
Degradation and ring
disruption
Nltroreductlon
Mineralization
Conetabollan,
can be used as sole
carbon source
Oxidation
.,„-., ""•-
HetabollsD
HetabolliB
Degradation
Com*taboli.m, met*
cltavag* oxyganasa
Cometabollan, Beta
cleavage oxygenate
1,2 deoxygenaae
1,2 deoxygenaae
CoBetabollai
Comet a boll HB
Oxidized
protocatechuate; 4,5-
oxygenaae pathway
Hicrobial
trans fornat Ion
Degradation
Sole carbon source,
degradation, comae n-
surate growth
PRODUCTS
tlon
p-amlnophenol
co2
Beta ketoadipate, netabollzed further
to aucclnate and acetyl-Co A then
to C02 and H20
2-hydroxy-4-chloro-Buconlc-
••Bialdehyde
2 hydroxv-3,5-dechloro-»uconic-
«e*l aldehyde
2-hydroxy-3 nethyl-Buconlc
senialdehyde
Supports growth
C02, nonextractable bound C (90%)
B=L
Mutant Pseudomonaa ap.
Mutant Pseudononas ap.
Rumen microorganisms
.
marine bacterial conrauni-
tles (2- weeks adaption)
take Ontario sediments,
soil & activated sludge
Aclnetobacter sp..
Arononas sp.. Bacillus ep.
PBeudOnonaa sp..
Flavobacterluai sp. . etc.
BI-CHEM CEC-1
PaeudoBonas putlda
Pp&eudononas sp. B13
Alcalignee ap. A7-2
Paeudoaonaa ap. B13
AchroBobacter sp.
Achroaobacter ep.
Alcaligenes sp. A7-2
Alcallgenea sp. A7-2
Achroaobacter ap.
Sewage
'RhodopaeudomonaH
palua tree
Stream sediments down-
streaa of a coal coking
uastewater discharge site
Alcaligenes facalis
Plasmld harboring mixed
cultures
SPECIAL INITIAL
ENVIRONMENTAL SUBSTRATE CONTACT
Aerobic, 20°C 500 ^/i loox> o houra 61> ?i( 1(J8j
Aerobic, 30«C ,00 ^^ ^ ^ ^^
200 ng/1 100X, 29 hours
ifO ng/i 26X, 120 houm
20U L^/l 100X, 38 hours
2UU mg/1 1UUS, 4D hours
200 Bg/1 100X, 32 hours
Anaerobic
57
20-100 ppb 99
Anaerobic, cyclone Q^ fi?
fernentor
Aeroblc 0.5-1 day 67
59
Aerobic •*-,
Aerobic (JQ
Aerobic g.
Aerobic ^
43
43
91
91
/3
44
Aerobic 29
Aerobic, 20'C ^ dayfl ^ „
Aerobic, 30°C , , ,„ „ ,-
j. j aays jj
Chemostat 33
-------
SUISTFATT
blfbcoyl
4-chloroblphenyl
•ono and
dlchloroblphenyls
nonochloroblphenyla
mono and
d 1 ch 1 o robl ph enyl •
4-chloroblphenyl
4,4-dlchloroblphenyl
dlchloroblphinyl
fluoranthene
naphthalene
IIOLOCIUL
ACTIVITY
Dtg»4atto«. -U
Mthviy
dtgradaclon
Blodegradation
Degradation
Total degradation
Degradation
Blodegradation
High level
dehalogenaae
Degradation
Oxidation
laonerlzation
oxidation
Hlcrobial
transformation
Sediment absorbed
blodegradation
rtocucrs
Chlorobttuolc >eld
Cla-2(3-dlt}ydro*yblphtayl
4-hydroxyblphenyl beniolc acid
C02
co2l H2o, cr
Hajor end product C02
4-chIoro-4 hydroxy blphenyl
4 ,4-dichloro-3-hydroxyblpnenyl
Chlorinated benzole acid
C02, unidentified yellou conpound
Gentisate
1-naphthol, 4-hydroxy-l-tetralone,
cis-naphthalena dihydrodlol.
clB-1.2-dihydroxy-l,2-dihydro
naphthalene; 4-hydroxy-l-tetralone;
and 1-naphthol
COo, cell bound C
1-naphthol; cls-1, 2-dlhydroxyl-
1 , 2-dihydronaphthalene;
alpha-naphthol; beta-naphthol;
trans-l,2-dehydroxy-l,2-
dehydronapthaleoe; A-hydroxyl
SftctAL In IT HAL
•10LCCICAL CJWlKOiOCHTAL. 5VKTJW7E
ACtHT(S) UCOUDff»rrS COKCEMTRATICW
Aelaetohdctcr >?. F6 At robl e. »lit«ld
Artbrobacter ip. KS lavelvtMnt, 25*-30'C
AlcallBlnci facealfc Aerobic. 30*C
Vij«rlnckU 18/36 Aerobic
Qaelllatorla ip., Aerobic
Port Valdez. AK 10'C, no aeration 4.7-4.4 u nole/
aeauater-S • depth liter
Acinetobacter ip, P6 Aerobic* plaiald Involvement
25»-30'C
Arthrobacter ap. with
plaanld p AC 2 7- or PAC31-
harborlng Paeudononaa
putlda
Port Valder, AK, 10'C. no aeration 1.5-4.5 u mole/
Beauater 5 • depth liter
Pseudononas sp., Vibrio Aerobic
ap- i Splrellun sp.,
Flavobacter op.
Achronobacter ap..
Chrooobacter, ap..
BacllliB Bp.t Hocardla sp.
Fungi
Sewage sludge Aerobic
Naphthalene utilizing Aerobic, pH 7.0, ahaker culture 100 ng/l
nlcroorganisBB fron soil
Pseudononas sp. A3 A, 30"C, biostat fernentoro
Cyanobacterla and Photoautotrophlc conditions
nicroalgae
Osclllatorla sp. Photoauto trophic conditions
strain JCM
Stream sedinents down- Aerobic, 20*C
stream from a coal
coking uaatewater dis-
charge site sedlnent
Bacteria
BI-CHEH FOG-3
Agnenellun sp., Aerobic
Osclllatorla sp.,
Anabaena sp.,
Cunnlnghaaella
elegans
CONTACT
TIWE uymxos
38
1-3 d.yi 55
J7
57
9.3-&.U n Mle/ 85
liter/day
961 usage 'M
981 ID 1 week 3d
1.2-4.1 D mole/ 85
liter/day
57
57
57
57
94%, 48 hra 5
17
21
22
310 days 41
5fa
5y
57
-------
SUBSTRATE
CH3, OCH3, Cl or N02
p a enes
C02H, Ctf2C02H or S03H 1
substituted naphthalenes
napththalenesulfonate
2-naphthaleneaulfonate
phenanthrene
polycycllc arooetic
hydrocarbons
polyrhlorinated
blphenylB
Four Ring Fused Polycycllc
Hydrocarbons
d 1 benzan th racene
chryecne
polynuclear aromatics
BIOLOGICAL
ACTIVITY
Ability to breakdown
this compound common
Degradation
Degradation
Degradation
Degradation
Degradation
Biodegradation
Biodcgradttion
Degradation
Blodegradatlon
Blodegradatlon
Blodegradation appears
to extent of chlorina-
Degradation
Blodegradatlon
PRODUCTS
Hydroxylated In 7,8 positions
and 4-substituted sallcylatea are
accumulated
Hydroxylation In 1,2 position,
l,2-dehydroxy-l,2-dehydro
naphthalene-2-carboxylic acids are
formed
Unidentified end products which are
metabolized by other bacteria
Cl8-,3,4-dehydroxyr3,4-dehydro
phenanthracene
Very slow insignificant breakdown
BIOLOGICAL
AGENT(S)
HIcrocoleus sp. ,
Nostoc sp..
Coccochlorls sp..
Aphanocapsa sp. t
Chlorella sp. ,
Dunalilla ap.,
Chlamydamonas sp-.
Amphora sp.
Pseudomonas sp.,
Mj:allgeneji sp. ,
Corgslbacterlum sp.
Aeromonas sp. ,
Norcardla sp.
Pseudomonas sp. A3
Pseudomonas sp, C22
Pseudomonaa sp, A3
Pseudomonas sp. C22
Pseudomonas sp. A3r
Paeudbcionas flp. C22.
naphthalene degrading
sewage bacteria
Al call genes faecalis,
Beljerlnckla Bwt.
Pseudononae SPM64
Marine sediments
Beijerinckla ap.
Indigenous reservoir
microblal population
Microorganisos
Tndlgenlous reservoir
ralcrobial population
Microorganisms
Pseudononas sp.
Vibrio sp.. Spirillum sp,
Plavobacterlum so.
Stream sedlnents in
vicinity of coal
coking wastewater
discharge site
Cunninghamella elegans
'Activated sludge
Marine sediments
BI-CHEM FOG- 3
bacteria
SPECIAL INITIAL
ENVIRONMENTAL SUBSTRATE
REQUIREMENTS CONCENTRATION
Aerobic
Aerobic, 30°C, blostat
Aerobic, 30°C, biostat
Aerobic, 28°C
Aerobic, 30CC
Aerobic
25°C in dark 100 u g/100 aole
Pits, ponds, lagoons, soils
and waste water streams
25 °C, dark
Pits, ponds, lagoons, soils and
wastewater streams
Aerobic
••
Aerobic, 25"C
Aerobic
Aerobic
15 °C, dark
CONTACT
57
17
56
57
8UX, 4 ueeku 93, 94
28
56
28
34 days 41
57
56
4 to 8 weeks 94
-------
1
t— *
00
suism-rr
Mvt Rlag Tute* Pol cyclic
HydTocaraoaa
bcnio(a)pyraB»
Fused Pnlycycllc
Hydrocarboaa
coal
Organophoaphatea
aapon
azodrln
dasanlt
dcazlnon
•alathln
orthene
parathlon
trithion
dlnethoate
dylox
•ethyl parathlon
vapona
parathlon
•ethyl parathlon
Pesticides and Herbicides
trlchlorophenoxyacetlc
acid (agent orange)
•ethoxychlor
BIOLOGICAL
ACTIVITY
Degradation,
Hgnln digeatlon
Died aa sole phos-
phoroua aourcea-nona
of the organo-phoa-
phatea aerved a* a
carbon source
Hitroreduction
Ring cleavage
Mineralization
Hydfolyeea
Tranafonutloaa
Degradation
5-T degradation pathway
Dechlorinatlon ,
oxidation
De chlorination
raooocTS
Tran»-7,6-dlhy<]rorr-7,8-dlhydro-
twwoMrymt
Black liquid produce
Dimethyl phosphate
dlethyl phosphorothlonate
dlethyl phoaphorothlonata
dlHthylphoaphorodithloate
dlethylphoaphorotMoate
diethylphoaphorothloate
Aalno-pa rath ion
C02
Dlethylthlophosphoric acid and
p-nltro phenol
Cl~, chlorophenols and related
coapounda
Cl~, C02
l,l-dichloro-2,2-bls (p-methoxy-
BIDLOCICAL
ACOT(S)
Cuanlatha**lla el t tana
PaeudoBooaa a P.
Bajjerlnckla ap.
Polyporoua veralcola.
Porla, MntJcola
Bacteria Isolated fro*
aoll and sewage
Paeudononofl pertlda,
Paeudoeonaa 28
Bacillus aubtllus.
.Rhyzobrua ap. , Chlorella
pyrgnprdsa,
soil bacteria
Rice rhlzosphere
Pseudononaa dlninuta
Aufwuchs •Icroorganlaaw
Poeudoraonaa capaclo
AC 1100
Klcroorganisn from waste
duvp altea
Paeudononas cepacla
AC1 100
Soil bacteria
Bacillus Bubatills
Mocardla ap. ,
Streptoaycea sp.
Aetobacter aerogenea
— msffl, isnnr
UVIJtOWfCKTAL SUBSTIATi;
Aerobic
Aerobic
Aerobic
Lignite coal.
28'C, 801 relative huatdity
Aerobic, 29 *C shaker culturea .
Anaerobic
Flooded and non flooded
condittona
Aquatic microbial growth "
attached to aubaerged aurfacaa
or auapandad in atraaaars or
•ats.
Growth in soil, 30'C, 15-50Z i ^/g, BOu
DO is tu re
Plasalds pAC 25, TOL, CAM, SAL BOO u g/ad
pAC21; chaatostat envlronsent
Aerobic, 30*C ahaker culture 1 •g/al
Anaerobic, denitrifying 1,000 pp«
conditions
A=i"«ijie, shsker cultsrs 25*C G. 5-5.0 u §/!!£»
Aerobic
Aerobic, anaerobic
CCWUCT
TIM gffPOtCB
57
25
b7
87
87
87
87
87
87
87
87
87
87
87
57
13 d.ya 84
92
t4
95Z, 1 week 23, 52, b3
701, 7 days 51
971, 6 daya 52
90Z, 3 «ratha 37
30 =1=-:=: 45
57
21
phenyl) ethylene; 1,1-dichloro-
2,2-bla (p-»ethoxyphenyl)-ethene
Beta-hydroxypropionaldehyde
Site water nicrobea
-------
BIOLOGICAL
SUBSTRATE ACTIVITY
chlorodl.efom Biodegradatlon.
non enzymatic
degradation
DDT DDD
Degradation
" PRODUCTS
4-chloro-o-f ornotoluidllene ,
4-chloro-o-toluidene, 5-chloro
anthranlllc acid, n-formyl-
5-chloroanthraniIlc acid
DDD, DDE, kelthane, DBP and DBH
Reductive dechlorination DDD
Reductive dechlorination DDD
diuron Mineralization
•----
dleldrin Anaerobic degradation
«*drln D*iradatloa
Epoxidatloo
DDE
DDH, DDKS, TDE, DDE
Photodleldrin,
toxic epoxide >oiety reduced to
deflen
Chlordlne, chlordene epoxide
Photodleldrin, aldren dlol
DUldrln
BIOLOGICAL
AGEHTfS) '
Chlorella. Osclllatorla
Anaerobic digestion,
sludge
Soil nlcroorganlsns
CoBoercial yeast
Klebaiella pneumonia.
Escherichia Coll
Pseudfifflanas clostridunu
Paeudonenaa vulgarla
Oceanic conditions
Mixed culture of fungi
find bacteria
Anaerobic digester aludge
Anaeystls nidulana.
Agmeneloua
quardlplicatuna.
Pseudouonas ap.
Ruaen flutd
Actiaoaycetes
Ocean sedlaents
Anaerobic digester aludge
Site water microbes,
•ewage sludge
SPECIAL ' ' INITIAL
ENVIRONMENTAL SUBSTRATE
Aerobic, 25*C, presence of light 960 u g/«l
Anaerobic, 35 *C
Anaerobic
Anaerobic
Anaerobic, more than 20
bacteria species are reported
2 DDT
Aerobl c
Aerobic, anaerobic synbiotic
relationship
Anaerobic, 35"C 150 u g/^
Anaerobic
Anaerobic
Anaerobic, aerobic
Anaerobic, 35 "C x
CONTACT
14 days 9, 57
42
39
49
57
57
78
57
Slow 42
57
57
57
78
lOOt, 40 d*yn 42
57
Degradation
Aldrin diol, photodieldrin
Sea water, bottoa aedl- 26°C-27"C and pH 7.6-7.7
oente fron ocean and
astuarinaa
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BIOLOGICAL
SUISTMTt ACTIVITY
ft* tachto ro«l t ro tttnttM
phorate sulfate Sulfoxtde reduction
toxaphene
atrazlna Detoxification
Miscellaneous Hydrocarbons
che«lcal waste
crude pet role ua products Blodegradatlon
detergents (nonlonic,
anlonlc, catlonlc)
ethoxylated phenols
^> foams
fO halogenoted hydrocarbons Blodegradatlon
o
endosulfan
Degradation
endrtn Anaerobic degradation
heptachlor Anaerobic degradation
heptachlor epoxlde Anaerobic degradation
kepone Cometabollsn
iiotxciaL
Afpergellux «tg*r.
Fllsarlua solsjil.
SH^.
yletprcte,
F«ntcilliu» »p,,
Trtchodsrma
veridie
Soil bacteria
Corynibacterlua pyroReoet
BI-CHEH fOB-6
PHENQ1AC,
a Mixture of aerobic
organlaam
Non-toxic byproducts Hicroorganlim
BI-CHEH TEX-4
BI-CHEM T2X-4
BI-OffiH TEX-4
Non-toxic byproducts Klcroorganiaoa
BI-CHEH POB-6
BI-CHEM PEP-7
BI-CHEH FOG-3
Endosulfan, endodiol endohydroether Fungi, aoll bacteria,
sctlnoaycea
Endogulfate (fungi), Sixteen fungi,
endodiol (bacteria) and C02, fifteen bacteria,
endohydroxy ether, and 2 unidentified three actinonycea
products epeclea
Anaerobic digester
sludge
chlorine atona HicroeoccuB sp,, yeast,
soil orgaaieas
Ketoendrin (24, 4X) Fish pond algae
Anaerobic digester sludge
Activated digester sludge
bovine ruaen fluid
Treatment lagoon sludge
SKrC'UL ' ' IhlTUL
ITQUIIMXTS COHCOOUJICM
AtroVlc, only ^urlBf actlva
icowtb
Anaerobic
Anaerobic
Detoxvol saturated soils
Aeration tank, PH 5.5-6.0. 100-bOO p%/l BOD,
3-7i •«/!
phenoli
Contanlnated Bolls and aquatic
envlronaenta
Contaminated aoila, aquatic
eye teats
Aerobic, 22* to 27'C 1 us/ ml
Anaerobic, 35°C ^y^ ug/nl
Anaerobic, 35"C ^JQ yg/ai
Anaerobic, 35'C 150 ug/Ml
Aerobic
Anaerobic
Anaerobic
(XMTACT
Tint urooxos
57
57
57
59
Plant umder con- 114
trol in 15 dtya.
clou, 12 his.
2", ""
59
59
59
29
59
59
59
Fungi 90X, 57
6 ueeks
Bacteria 70
70-WK, 10 days;
actlnoayates
30X, 10 day.
100X 30 days 42
57
78
1001, 20 .Inutt. 42
501, 60 days 42
57
57
-------
SUBSTRATE
1 Indane
industrial surfactant*
oil
organic solvents
pesticides 6 herbicides
petroleum distillate
,__ phenollcs
>
ro
polycycllc aromatic
hydrocarbons
polychlorlnated blphenyls
polynuclear aroma tics
refined petroleum
products
sludge* (paper industry fl
and vegetable processing)
"sulfur compound*
waste oil
BIOLOGICAL
ACTIVITY
Anaerobic degradation
Blodegradatlon
•Biodegradatlon
Bl ode gradation
Biodegradstion
Blodegradatlon
Blodegradatlon
Blodegradatlon
Biodegradatlon
•h
Degradation
PRODUCTS
Fentachlorocyclohexane
gamtna-3,4,5,67-tetrachloro-l-
cyclohexane, alpha BHC
Non-toxic byproducts
Non-toxic byproducts
Non-toxic byproducts
Non— toxic byproducts
Non-toxic byproducts
Non-toxic byproducts
Non-toxic byproducts
BIOLOGICAL
AGENT(S)
Anaerobic digester slud
Chlorella vulgaria.
Chlamydaraonas
relnhardtll
Chlosteridlura sp. .
Pseudomonas sp.
Soil bacteria
-^
Sewage sludge
Microorganisms
BI-CHEM FOG-3
PHENOBAC
MicroorganisDS
Microorganisms
BI-CHEM SUS-8,
detoxsol
BI-CHEM COG-2
Microorganisms
Microorganisms
BI-CHEM POB-6
BI-CHEM FOG-6
Microorganisms
BI-CHEM PAC-5
BI-CHEM SUS-8
PHENOBAC, :
PETROBAC
SPECIAL INITIAL
ENVIRONMENTAL SUBSTRATE
REQU 1REMENTS CONCENTRATION
ge Anaerobic, 35"C JJQ ug/ai
Anaerobic (?)
Aerobic
Anaerobic
Anaerobic
Anaerobic
Contaminated soils, aquatic
systems
Aeration basin 0.5 MGD waste- Oil/grease
liquid storage facility
Contaminated soils, aquatic
systems
Contaainated soils, aquatic
sy at etna
4 acrea contaminated soil ^ 200 op*
systems
4 million gallon 800 pp«
lagoon POLYBAC N (nacronu-
CONTACT
TIME REFERENCES
100X 25 4ay* 42
57
57
57
57
57
28
59
SIX 24 hrs. 114
95X 72 hrs
reduced to 7 ng/1
28
28
1 ppn, 21 daya 59
26
59
28
28
59
59
26, 6b
59
26
992, 9 Booths 114
trlents nitrogen and
phosphorous)
-------
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.S. GOVERNMENT PRINTING OFFICE: 19^6_6t6~116f 40648
-------
-------
Jnited States
Environmental Protection
Agency
Information
Cincinnati OH 45268
BULK RATE
POSTAGE & FEES PAID
EPA
PERMIT No. G-35
Official Business
Penalty for Private Use, $300
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