Hazardous
Materials
In Situ
Stabilization

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f/r"
7 I
hazardous materials
In Situ stabilization
U.S. Environmental Protection Agency
Region III Information Resource
Canlcr (3PM52)
841 Chestnut Street
Philadelphia, PA 19107

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Printed in the United States of America
L.C. Card No. 87-81735
Copyright © 1987 Hazardous Materials Control
Research Institute
9300 Columbia Boulevard
Silver Spring, Maryland 20910

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PREFACE
In situ systems to accelerate the stabilization of waste deposits is another potential solution to the
management of hazardous wastes in Superfund sites. It may be used as the method of cleanup or in
concept with other site cleanup concepts. The fundamental concept for in situ stabilization is chemical
treatment of the waste within the soil medium. 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 reactions.
As with other Superfund site remediation technology, the application of these systems to uncontrolled
waste sites will require a site-by-site, customized approach that considers the uniqueness of the combina-
tion of the subsurface geohydrology, waste inventory and site history at each site. There is also extensive
need for complementary laboratory simulations and testing prior to implementation of this concept.

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ACKNOWLEDGEMENT
This publication was prepared for EPA under Contract No. 68-03-3113. HMCRI. The 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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CONTENTS
1.0	DELIVERY AND RECOVERY SYSTEMS FOR WASTE DEPOSIT STABILIZATION		
2.0	BIODEGRADATION 	19
3.0	SURFACTANT-ASSISTED FLUSHING	34
4.0	HYDROLYSIS 		
5.0	CHEMICAL OXIDATION 	52
6.0 APPLICATION AND DESIGN OF SYSTEMS TO ACCELERATE STABILIZATION	60
OF WASTE DEPOSITS
APPENDIX A 	73

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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).
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:
•	Waste deposits are in a solid state and no immiscible flow is
expected.
•	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.
•	Recovery of the reaction products and spent reactant will be
exclusively from a water table aquifer.
•	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).
1.1.2 Waste Deposit Settings and Methods of Delivery
and Recovery
Given the above assumptions, four basic methods (and com-
binations thereof) have been investigated and conceptually applied
as possible delivery and recovery systems. These are:
•	Gravity delivery by surface or subsurface means
•	Forced delivery
•	Gravity recovery
•	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:
•	Shallow depth to waste deposit from the surface (less than 5
meters or 16 feet)
•	Highly permeable cover material between land surface and the
waste deposit (greater than 10~3 cm/sec or 2.8 ft/day)
•	Highly permeable waste deposit and surrounding soil media
(greater than 10~3 cm/sec or 2.8 ft/day)
•	Waste deposit located above the groundwater table
•	High surface infiltration rate (greater than 10 cm or 4 inches
per day)
•	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.
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1.1.2.2	FORCED DELIVERY
Forced delivery methods can be used in all the conditions noted
above, 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 shallow waste deposits (less than 10 meters or 33 feet deep, and
less than 5 meters or 16 feet below the water table) with hydraulic
conductivities of 10~3 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 sur-
rounding soil media which can have an effect on the selection of
delivery and recovery systems are:
•	Surface soil infiltration rate
•	Hydraulic conductivity of the waste deposit, cover material and
surrounding soil media
•	Interrelated parameters of porosity, specific yield, specific reten-
tion 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 con-
dition is approached. Usually the steady-state infiltration rate is
used as the design criterion for the hydraulic loading rate of sur-
face 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 infiltra-
tion rate is approximated by:
where: I0
I,
A,
n
t
A,t"
infiltration rate at time, t (length/time)
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
(1-1)
The equation for expressing the cumulative intake of fluid at time,
t, is:
where: Y = A,t(n + *V(n + 1)	(1-2)
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 flooding involves using an
area ranging from 1 m2 (10 ft2) 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 in-
filtrometers 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
15-35 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 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 calcu-
lation, theoretical background, limits of applicability and poten-
tial 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:
Q = KIA, or	(1-3)
v = KI = Q/A,	(1-4)
where: Q = flow rate (length7/time)
K = hydraulic conductivity (length/time)
I = hydraulic gradient (dimensionless)
A = cross-sectional area (length2)
v = specific discharge ("Darcy velocity")
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: n;ff = 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, neff 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 (Kh) is not equal to the K for vertical flow
(Kv). The hydraulic conductivity K can also be expressed by
(Luthin, 1957):
K = C(d50)2pg/u
(1-6)
where:
C = proportionality constant for the medium, based on
2

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geometric characteristics (dimensionless)
d50 = mean grain diameter of the medium (length)
p = density of the fluid (mass/ length3)
u = kinetic viscosity of the fluid (mass/length x time)
g = acceleration of gravity (length/time2)
In this equation Cd50)2 (also defined as k, the specific or intrin-
sic 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 par-
ticles, 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.
Obtaining a representative value of the hydraulic conductivity
at a particular site is essential to establish the rate at which a treat-
ment solution can move through the waste deposit and surrounding
soil media. Not only is it important to obtain hydraulic conduc-
tivity 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 1 x 103 cm/sec
(2.8 ft/day) would normally be feasible; however, if the waste
material is surrounded by a host material with a hydraulic con-
ductivity of 1 x 103 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, imple-
mentation and interpretation of these methods are described in
detail in several sources, including USEPA (1977); USEPA (1980a),
Repa 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 conduc-
tivity 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
(d10) means that 10 percent (by weight) of the soil particles are
finer than the specified diameter. Table 1-1 shows that, for exam-
TABLE 1-1
HYDRAULIC CONDUCTIVITY, POROSITY AND DRAINAGE CHARACTERISTICS OF MATERIALS
10
28.000
K (HYDRAULIC CONDUCTIVITY)
1
2800
10r'
280
,0
28
10"3
2.8
10"*
0.28
10*°
0.03
10"«
0.003
10*'
0.0003
10"*
0.00003
CM/SEC
FT/DAY
-+-
101
-+-
Iff'
10"
w
.06
.02
-4-
10'5
.01
	(-
w
10"'
.006
—I—
.002
—I—
.001
-4-
10'
CM/SEC
MM
0.08 0.04 0.024	0.008 0.004 0.0024	0.0008 0.0004 0.0002	0.00008 0.00004
EFFECTIVE GRAIN DIAMETER. d10
GRAVEL I CS. SANO I MtP. IANO | FINE SAND | CS. SILT 1 MtO. WLT I FINC SILT |	ctAY
INCHES
CHAW GRAVELS
VERY FINE SANDS
CLf AN SANDS
COARSE FINE
	SANOGRAVEL MIXTURES. TILL
SILTS. ORGANIC * INORGANIC
VARVEO CLAYS. ETC.
HORIZONTAL K
	SAND SILT-CLAY MIXTURES. Till
I
HAICCS or
TYRCAl
RANGE OF POROSITY VALUES
AVERAGE K VALUES	FOR VARIOUS MATERIALS
Type
cm/sec
ft/day

Type
Porosity
very fine sand
5 x 10~3
14

clay
45 - 55
fine sand
2 * 10"2
57

sand
35 - 40
concrete sand
2 x 10"2
57

gravel
30 - 40
fine Co medium sand
5 x 10"2
140

sand and gravel
20 - 35
medium sand
H
1
©
280

municipal waste
0
1
o
f)
medium to coarse sand
1.5 x 10"1
425



clean gravel
10
2.8 x
10A


concrete gravel
25
7.1 x
10A

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pie, 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"1 and 1 x 10~3 cm/sec (280 and 2.8 ft/day).
Media with d10 between 0.1 to 0.005 mm (0.004 to 0.0002 inches)
will have a hydraulic conductivity between 1 x 10"3 and approxi-
mately 1 x 10"6 cm/sec (2.8 to 0.003 ft/day).
The field bore-hole method or single well pump test is a com-
monly used method for measuring the in situ hydraulic conduc-
tivity 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 hori-
zontal component of the hydraulic conductivity. The hydraulic con-
ductivity 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 equip-
ment 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 proper-
ties for modeling groundwater flow. 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 impor-
tant 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.
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):
Sy	=	ioo cy w/y,)	d-7)
where: Sy	=	specific yield (dimensionless)
Vw	=	volume of liquid removed by gravity (length3)
Vs	=	volume of soil (length3)
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(V/Vs)	1.8
where. R = specific retention (dimensionless)
Vr = volume of water retained by the soil against the
pull of gravity (length3)
Vs = volume of soil (length3)
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 (V/Vs) = 100 (VrV)/Vs	(1-9)
where: n = total porosity (dimensionless)
Vs = total volume of soil or rock (length3)
V	= volume of water required to saturate the sample
(length3)
V	= aggregate volume of the solid particles that make
up the sample (length3)
The porosity of a sample is best measured in the laboratory. Gibb
et al. (1985) describe experiments to measure the effective porosity
(n fr) 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), USEAP (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 con-
taminants in the waste deposits or flush contaminants from the
deposits to the groundwater table. Flushed contaminants can sub-
sequently be collected and treated above ground. The available
delivery methods are grouped into two generic categories: grav ity
and forced. Gravity methods apply the flushing or reactant solution
directly over the waste deposit (if the waste deposit is at the sur-
face) 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. Ground-
water flow analysis is described in most groundwater texts,
MAXIMUM 10% GRAIN SIZE
NOTE.SOIL SAMPLE FROM SOUTHERN COASTAL BASIN. CA.
CLASSIFICATION SHOWN IS THE LABORATORY CLASSIFICATION OF
THE SAMPLE.
SOURCE; USEPA, 1977
FIGURE 1-1
POROSITY, SPECIFIC RETENTION AND SPECIFIC YIELD
VARIATIONS WITH GRAIN SIZE
4

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including Freeze and Cherry (1979) and Cedergren (1977), and
mathematical modeling is reviewed in Wang and Anderson (1982).
The following sections present discussions of the design con-
sideration of the following delivery technologies:
•	Gravity
-	Flooding
-	Ponding
-	Spray Irrigation
-	Ditches
-	Subsurface Spreading (infiltration galleries and beds)
•	Forced
-	Injection Wells
1.2.1 Gravity Deliver} Methods
Gravity delivery methods are applicable at or near the ground
surface and can be classified into two groups: surface and sub-
surface spreading. Surface spreading involves the application of
the flushing or reactant solution over the waste deposit or over-
lying 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 infil-
tration 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.
As the depth to the top of the waste deposit increases, the effec-
tiveness 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 ground-
water 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 selec-
tion 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, the following conditions must
be known about the site:
•	Surface topography and area of the site
•	Sustained infiltration rate (I)
•	Configuration of the waste deposit
-	areal dimensions (LxW)
-	thickness of deposit (d)
-	depth to deposit (d0) from the surface
•	Aquifer thickness (above impermeable layer) before solution
application Hd)
•	Hydraulic conductivity (K) of the waste deposit and sur-
rounding soil media (based on minimum value)
•	Depth to water table (S0) from the surface
•	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 recover (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):
q, = K (Hc - Hd2) /L +2X)	(1-10)
where:
Hc = Total saturated thickness above aquiclude required
to meet the assumption that the waste deposit is
fully saturated (length)
Hd = Elevation of the recovery system above the imper
meable layer (i.e., the original water table eleva-
tion before solution application) (length)
X = Distance from the edge of the deposit to the
recovery system (length)
L = Length of the deposit parallel to the groundwater
flow (length)
K = Average hydraulic conductivity of the waste
deposit and surrounding soil (length/time)
q, _ Required application rate per unit area of the
deposit to satisfy the saturation criterion (length/
time).
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 treat-
ment 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, 1,, can be estimated by
solving equation (1-10).
Another equation expresses the relationship between the appli-
cation rate (q,), time (t) required for saturation of the waste
|—1^4.—j
FIGURE 1-2
GENERAL LAYOUT OF A GRAVITY DELIVERY
AND RECOVERY SYSTEM
5

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deposit and surrounding soil media, and the location of the recovery
system (X) as follows:
t = n(Hc + Hd>X/2qW	
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 determined by the areal
extent of the waste deposit. The total amount of solution (Q0) that
can infiltrate into the deposit would be:
Qo = AIs	(1-12)
where: Q0 = maximum application rate of the deposit
(lengthVtime)
A = area of application (length2)
Is = sustained infiltration rate (length/time)
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 Qo 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 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.
The parameter qo (= Q0/A, the application rate per unit area
of the delivery system) represents the surface infiltration or sub-
surface delivery rate (using Equation 1-3) achievable by a gravity
delivery method. If q, is set equal to qo 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, qo must be equal to or greater than qr If q0 is less
than q,, the next step would be to determine whether q0 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.
1.2.1.4 SURFACE APPLICATIONS
Flooding In effecting solution application by flooding, the solu-
tion 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 con-
ductivity are high (greater than 10"3 cm/sec or 2.8 ft/day). Pre-
ferred 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 sandy soils. This would be
the equivalent to approximately 5200 m3/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 is at 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 con-
struction 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 appli-
cation method ineffective. The natural infiltration rate can be
enhanced by tilling or furrowing the surface soil (USEPA, 1977;
USEPA, 1984b), but deeper impermeable layers may not be reached
by this method and other delivery systems will therefore be re-
quired. 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.
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, con-
structing 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 be 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 con-
struction 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.
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 rain-
fall 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 equip-
6

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TABLE 1-2
DESIGN LIMITATION FOR
INFILTRATION-PERCOLATION SYSTEMS
tor	Range of Feasible Values
TO TREATMENT-
Liquid loading rate
Annual application rate
Land required for
1-ngd (3785 a^/day)flov
Application
techniques
Source: Pound aod Crltes, 1973
10 co 300 ca/vk, or 2 x 10 ^ to
6 i 10 * ca/sec (4 to 120 In/wk)
5 to 155 a/yr (17 to 500 fc/yr)
1 to 25 hectare* (2 to 62 acres)
plus buffer tones
Usually surface
Rapidly permeable soils, such as
sands, loany sands and saody loans
PREVAILING GROUND
SLOPE GENERALLY
LESS THAN 3 PERCENT
DISTRIBUTION
OITCHES
DITCH OUTLET
SURFACE
COLLECTION
DITCH
SUPPLY DITCH
HAZARDOUS WASTE
DISPOSAL BOUNDARY
FIGURE 1-3
TYPICAL PLAN OF DITCH AND FLOODING TYPE
GRAVITY DELIVERY SYSTEM
TO TREATMENT
CLAY OR BEOROCK
FIGURE 1-4
GRAVITY DELIVERY SYSTEM USING PONDING
ORiCisal
GROUNDWATER
table
CLAY OR BEDROCK
FIGURE 1-5
GRAVITY DELIVERY SYSTEM USING PONDING
IN A THIN IMPERVIOUS STRATUM
TO
TREATMENT
W
minkkin)
i r
"y
-RECOVERY WELL
RECOVER* WELL-
/////////////
CLAY OR BEOROCK
FIGURE 1-6
GRAVITY DELIVERY SYSTEM USING SURFACE SPRAYING
ment 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 con-
ditions 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 con-
tained 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.
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 in-
filtration 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 cir-
cumstances 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.
7

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TO TREATMENT
TO TREATMENT
CLAY OR BEOROCK
FIGURE 1-7
GRAVITY DELIVERY SYSTEM USING DITCHES

////////////////////////////////)///////"
CL*V 0« l(D«OC«
FIGURE 1-8
SUBSURFACE GRAVITY DELIVERY SYSTEM
USING INFILTRATION GALLERIES
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 sus-
ceptible 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 pro-
vide for storage and easy delivery of the solution.
These methods are suitable where the waste deposit is subsur-
face 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 and wastes between 1 x 10"2
cm/sec and 1 x 1(H 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 con-
tact directly with more permeable strata. If the soils are of a silty
nature, with hydraulic conductivities between 1 x 10"1 cm/sec and
1 x 10~5 cm/sec (0.28 and 0.028 ft/day) (Table 1-1), these techni-
ques may still be used but the application rate and therefore the
treatment time will be much slower. As with ditches, the appli-
cation 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.
The solution can be introduced into the gallery by injection in
different locations along the length of the gallery or through per-
forated 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—Infiltation 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 in-
filtration 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.
-GROUNO LEVEL
CLAV ON AEOROCK
FIGURE 1-9
SUBSURFACE GRAVITY DELIVERY SYSTEM
USING AN INFILTRATION BED
8

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These systems would have the same limiting factors and effec-
tiveness 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 1 x 10~*
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 ex-
pendable and movable point for driving into the 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.
OPEN ENO	SLOTTED
FIGURE 1-10
INJECTION PIPES FOR FORCED DELIVERY
A forced delivery system, unlike the gravity systems, is concep-
tually 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 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 "in-
jectability" of the site soil in question. In general, a soil is not con-
sidered "injectable" if over 10 percent of the sample passes the


f
INJECTION WELLS
•~RECOVERY WELLS

O

o\


O
Q
° J
•
•
DIRECTION
OF
O
\

-WASTE DEPOSIT
GROUNDWATER
FLOW

o


PLAN VIEW
FIGURE 1-11
FORCED DELIVERY USING INJECTION WELLS
#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 35 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 deter-
mines 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
cause the fluid to travel toward the surface rather than seeping
through the formation. To avoid this, the injection pressure should
be kept below 1.5 x 10"4 N/m2 per meter (1 psi per foot) of over-
burden above the injection level (Winderkorn 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.
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):
where:	r = 0.62 (Qt/n)1 3	(1-13)
r = radial distance of solution penetration (length)
Q = rate of solution application (lengthVtime)

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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 in-
jection 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 con-
ductivity in the horizontal direction (i.e., K. = Kh). 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., K%
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 gravity delivery method
would depend upon: surface topography, infiltration rate, con-
figuration 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 10~4 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 con-
ditions. 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 paramaters and waste deposit
characteristics as well as the location and rate of recovery system
operation. An example is presented below to demonstrate the pro-
cedures to estimate the required/allowed application rate of treat-
ment solution.
In this example, it is assumed that a field geohydrological survey
was conducted and the following data were generated:
•	The waste deposits lie immediately above the water table.
•	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
m .
•	^ or'8inal water table above an impermeable layer
(Ha) is 3 m.
•	r ? thickness of the waste deposit is 1.5 m; thus the thickness
of induced saturation (H ) would need to be
1.5 m + Hd = 4.5 m.
•	At the site, the surrounding soil is sandy with a hydraulic con-
ductivity of 1 x 10 4 cm/sec and the soil porosity (n) is 45%.
•	Laboratory geotechnical and waste tests have determined that
the hydraulic conductivity 0f the waste deposit is approximately
x 10 cm/sec, and the reaction time (See Section 2-5) re-
quires a maximum of 10 minutes.
The first step in establishing a delivery system concept is to deter-
mine the allowable application rate of solution based on the
infiltration rate of the soil. Equation (1-12) is applied as follows:
O = AI	(1*14>
= 1200 x 3.5 X 10"6 = 4.2 x lO 'mVsec = 0.25 mVmin,
Q0 = and
n /a = I =2.1 x 10~4m3 per square meter
q0 = ^o'n . »
per minute.
Thus the maximum rate at which solution can be introduced
into the soil is 0.25 mVminute, or 2.1 x 10"4 m3 per square meter
^The second step in the process is to determine the required appli-
,,;nn rate of solution based on the hydraulic conductivity and
rproverv system location (X). This application rate represents the
TZl renuired to maintain the in situ treatment system at steady
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
o,i The hvdraulic conductivity of surrounding medium and waste
depo^(K) is 0 06 m/minute. Equation (1-10) is then applied as
follows:
q = K(Hc2 - HdJ) /L(L + 2X)
= 0.06 (4.52- 32) /30 (30 + 2X)
q, = 0.68/ (900 + 60X)
The third step in the process of establishing the conceptual design
is to set q = q0 10 define the minimum distance the recovery
system should be located from the edge of waste deposit (Xmjn).
Setting q equal to 2.1 x 10"4 m/min (from above) and solving
for X, one obtains Xmi = 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 = ' H^X
1	2qW	1 '
_ 0-45 (4.5 + 3) x 40 = L69
1 '	80q	q
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 on t = 6 days, the initial application
rate will be 0.27 mVmin. Thereafter, operated at 0.27 mVmin (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.
H = 4.5 m
Hj = 3 m
K - 1 x 10~4cm/sec = 6 x 10~5m/minute
L = 30 m
(Waste Deposit Area (A) = L x W = 30 x 40 m- = 1200 m-
n = 0.3 (30°7o)
= 7 x 10_5cm/sec
Determining q0 and q, yields:
q = 7 x 10-' cm/sec = 4.2 x 10"5 m3 /square meter
per minute
Qi = 6 x 10~5 (4.52 - 32) /30 (30 + 2X)
= 6.8 x 10"4 / (900 + 60X)
Letting qn = q, 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 J governs the application
process Thus, a solution for this situation can be achieved only by
10

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assuming the location of a recovery system, calculating the satura-
tion (q,) value, checking that it is less than q0, then checking this
against q and t estimated for the site.
As an example, assume a reasonable distance of 25 m for the loca-
tion of the recovery system downgradient of the waste deposit. Then,
q, = 3.8 x 10~6 m/minute
t = 0.30 (4.5 + 3) 25/80q
= 0.70/q
Letting q = q,, t = 1.9 x 105 min = 128 days.
The practical application flow rate (Equation 1-12) will be:
Q0 = q,A = 3.8 x 10~6 x 1200 =
4.6 x 10~3 mJ /min (1.2 gpm)
Delivering an approximate flow of 4 liters per minute over an area
of 1200 m2 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 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 con-
straint 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 ap-
plication and feasibility is hydraulic conductivity. As hydraulic con-
ductivity is reduced, the rate of spent solution recovery is decreas-
ed, thus increasing the period necessary for recovery. In a related
manner, forced recovery systems will require greater energy utiliza-
tion 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 (DJ distribution of a soil 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 Administra-
tion (1976). Repa and Kufs (1985) provide a good practical hand-
book 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 cnductivity and hydraulic gradient), in accordance with
Darcy's Law (Equation 1-3). Whereas hydraulic conductivity is a
function of 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 (H2 - H2) / 2q L - l/2	(1-15)
where:
K = Hydraulic Conductivity (length/time)
q: = Solution application rate/unit area of the deposit
(length/time)
L = Length of the deposit parallel to the groundwater flow,
(length)
Hc = Total saturated thickness required, i.e., distance from the
top of the waste deposit to the impermeable layer (length)
Hd = Height of the recovery system above impermeable layer
(length)
Therefore, X can be determined by measuring K and setting Hd,
measuring the depth of the deposit, and using the q, value deter-
mined for the delivery system.
SOLUTION APPLICATION RATE Iql
1)	RECOVERY DISTANCE IX) ¦ (Hc2 . Hd2)K/2q • 1/2
2)	MAXIMUM GRAVITY APPLICATION RATE (q0) • LI
(I -NATURAL INFILTRATION RATE)
SET Q ¦ 0oTO DEFINE MINIMUM X
IF X IS TOO LARGE FOR THE SITE, THEN:
•	INCREASE (INCREASE HYDROSTATIC HEAD; TILL SURFACE)
•	USE FORCEO OELIVERY SYSTEM
FIGURE 1-12
METHOD FOR CALCULATING LOCATION
OF A BURIED PIPE RECOVERY 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 col-
lection 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
11

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the side slopes. Ditches and trenches work best in permeable media
such as sands, where the hydraulic conductivity is greater than 1
x 10"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 approxi-
mately 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 (verticakhorizontal), 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.
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 (Hd) 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 recover ditches is described in Federal Highway Ad-
ministration (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.
intercepted and contaminated groundwater if the surrounding
materials have a moderately high permeability. Detailed informa-
tion on drain construction is provided in Federal Highway Ad-
ministration (1976), Luthin (1957), U S Dept of Agriculture (1972),
and Repa and Kufs (1985). A buried drain collection system is il-
lustrated 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 ex-
cavation 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 app-
ly 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.
The disturbed area caused by porous drains is relatively small com-
pared 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 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).
CftOUMO LEVEL
OtLIVEftY SYSTEM
clay or uorock
ORIGINAL GROUNDWATER
LEVEL
—WASHED
STONE
«-*EAfO«ATED
PIPE
FIGURE 1-13
GRAVITY RECOVERY USING A DITCH
1.3.1.2 Buried Drains
Buried drainage pipes containing either slots, performations, or
open joints are another type of gravity collection method similar
to the mfutrauon 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 backfill-
ed 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 fiowthrough of the
DIRECTION OF FLOW
////////////////////////
CLAY OR BEDROCK
FIGURE 1-14
GRAVITY RECOVERY USING BURIED PERFORATED PIPES
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 con-
12

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tamination is primarily in the aqueous phase (e.g., a spill), or in
combination with other treatment methods which remediate the con-
taminant 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 exposure 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 bee used for in situ
treatment of contaminants, although bench- and pilot-scale tests have
been performed to determine treatment effectiveness (Park, 1985;
Repa and Kufs, 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
hydrolysis of some organic contaminants in the groundwater (see
Section 5.0). Activated carbon is commonly used as a treatment
method of adsorption of hydrophobic (non-polar) organic con-
taminants in water. However, significant problems such as plugging
of the bed, short lifetime or saturation of the carbon, and desorp-
tion might occur with the use of activated carbon in permeable treat-
ment beds (Repa and Kufs, 1985). Glauconitic green sands are us-
ed 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 pro-
mising 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 severely 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 or 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 ground-
water 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.
1.3.2.1 WELLPOINT SYSTEMS
Wellpoint 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 condi-
tions. 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
deposit and surrounding soil media ranges between 1 x 10"1 and
1 x 10~3 cm/sec (280 to 2.8 ft/day). For deep groundwater condi-
tions, 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 ele-
vation) 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 well-
points are attached to 3.8 or 5 cm (1 Vi or 2 inch) riser pipes con-
nected 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 com-
bined vacuum and centrifugal pump which is connected to the header
and pumps water from the wellpoints. Generally, a stage of well-
points 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 conven-
tional 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 producing larger flows to the well-
points (Fruco and Associates, 1966). Vacuum wellpoint systems are
used in deposits with hydraulic conductivities from 1 x 10 ~3 to as
low as 1 x 10"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
WELLPOINTS (VACUUM
OR
CONVENTIONAL)
—HEADER
PLAN VIEW
DELIVERY SYSTEM
III III /—COMMON HEADER
	 TTTtIt i—	rr1	-mTBfirutUT
	Y
/ WASTE
DIRECTION OF / DEPOSIT
INDUCED FLOW *	f		
\

—SCREEN
CLAY OR BEDROCK ''
	
CROSS SECTION
FIGURE 1-15
WELLPOINT SYSTEM FOR FORCED 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.

OEEPWELL
WASTE I r\
OE POSIT / w
O
PLAN VIEW
1.3.2.2 DEEP WELL SYSTEMS
Deep well systems are particularly suited for recovering ground-
water 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 ob-
tained with a wellpoint recovery system. Mixed and axial flow pumps
powered by electricitiy, 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).
Deep wells for dewatering are similar 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. Con-
struction details are described in Repa and Kufs (1985) and order-
U.S. STANDARD SIEVE
OPENINGS IN INCHES
DELIVERY SYSTEM
I I I I I
- TO TREATMENT
t7 INITIAL GROUNDWATER TABLE	
		_ CONE OF
DIRECTION OF INDUCED FLOW



CLAY OR BEDROCK
CROSS SECTION
FIGURE 1-16
DEEP WELL RECOVERY SYSTEM
U.S. STANDARD SIEVE NUMBERS
100
z
o
5
>
CC
uI
Z

<
8
u
cc
100
0.001 MM
INCHES
GRAIN SIZE

gravel
SAND

COARSE
FINE
COARSE
MEDIUM
FINE
SILT OR CLAY
SOURCE: US DEPARTMENT OF THE NAVY, 1982
FIGURE 1-17
LIMITS OF RECOVERY METHODS APPLICABLE TO DIFFERENT SOILS
14

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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 well-
point 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 points and deep
wells, are limited to media with an effective grain size (d10) between
0.1 to 1 mm (0.004 to 0.04 inches) which generally have a hydraulic
conductivity between 5 x 10"1 to 10~3 cm/sec (1420 to 2.8 ft/day)
(Table 1-1). For media with effective grain size (d10) between 0.1
to 0.01 mm (0.004 to 0.0004 inches) which generally have a hydraulic
conductivity between 10~3 to 10~5 cm/sec (2.8 to 0.03 ft/day)
(Table 1-1), recovery of water may be possible using well points
equipped with vacuum pumps. Media with hydraulic conductivity
less than 10~5 cm/sec (0.03 ft/day) i.e., d10 less than 0.01 mm
(0.0004 inches) can only be dewatered by means of other enhance-
ment 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 "5 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 con-
ductivity 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 cathods (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, 1975) 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 preceding section) shows the grain size
distribution for which electro-osmosis should be considered. The
corresponding hydraulic conductivity for these materials ranges
from lxlO-5 to less than lxlO-7 cm/sec (0.03 to 0.0003 ft/day).
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 perform-
ed to determine the electro-osmotic transmission coefficient, kc
(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, Q , may be estimated from
the equation (Fruco and Associates, 196o).
Qe = keiea,z	(1-15)
where:
ke = coefficient of electro-osmotic permeability
(length2 /volts x time)
ie = 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):
I, = (Ac + B) /t	(1-16)
where:
I, = 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
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 effec-
tive means of dewatering or permeating certain fine-grained soils.
In an electro-osmotic dewatering system, the depth of the elec-
trodes 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 con-
figuration 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 elec-
trodes 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 ground-
water 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"7 cm/sec (0.0003 ft/day) prior to initiating electro-osmosis can
be made to exhibit hydraulic conductivities in the range of 10'5
to 10"6 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 conduc-
tivities 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.
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
15

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TABLE 1-3
DELIVERY AND RECOVERY SYSTEMS
Delivery /Recovery
Techniques	Alternatives
Deliver} Systems

Gravity Methods
Flooding, Ponding, Ditch, Surface

Spraying, Infiltration Gallery,

Infiltration Bed
Forced Methods
Injection Pipe (open end or slotted)
Recovery Methods

Gravity Methods
Ditch, Buried Drain
Forced Methods
Wellpoint with Vacuum

Wellpoint without Vacuum

Deep Well
Electro-Osmosis

1.5.2 Application of Various Systems
The selection of delivery or recovery methods depends primarily
on the geohydraulic characteristics of the disposal site. Two
matrices (one for delivery systems and one for recovery systems)
have been developed in order to guide the identification of feasible
delivery and recovery methods for a 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
TABLE 1-4
RELATIVE IMPORTANCE OF GEOTECHNICAL PARAMETERS
IN GRAVITY AND FORCED SYSTEMS
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.
Gravity	Forced
Methods	Methods
1.	Hydraulic Conductivity (K)	I	I
2.	Infiltration Rate (ls)	I	NI
3.	Application Rate (q,)	I	LI
4.	Configuration of Water Table and	1	LI1
Waste Deposit Location (Hc, Hd)
5.	Time to Reach Saturated Condition (t)	1	LI
6.	Homogeneity	I	LI
7.	Relation of Hydraulic Conductivity	I	LI
between Waste Deposit and Surrounding
Medium
8.	Relationship between Infiltration	I	NI
Rate and Hydraulic Conductivity
NOTE: I	= Important
LI	= Less Important
NI	= Not Important
1	= wellpoints are limited by depth to the watertable
TABLE 1-5
MATRIX FOR DELIVERY METHODS
0»tlvrrv
ti«nla
I
I Location n( th» depoalc	M j
I r.t.tl.. t.	ir.,,,,.,,.
I I'ouA4«at«r tikU	|
I
I
Thlcktittft
I
of	I
ovtrlyltif I Tapoftraphjr
Japtra*«kl« I (fUpt)
la
!!!_
I
Infiltration
Ut«
c«/hr
(iwtiwi/hf)
NpdrawlU Coadwctlvlty
«*/••< (ft/Aay)
I	Dapt* *• Ittlca
|	at IN*
I	Waat* D*petit
{	Hatara KD
lUnaatu-lPartiallyI
I <1.5.1 >l.Sa I
|.i-.i I .0.-.i |<.o»	„->-,o-4l io-'-io" I 11
UCAV1TT
1 - 1
1
! T r
1 1 1

1
1

j it
1
1

1 1
1
1

1 1
1 t

1 I
ViWVJ/ iwta/
1
1 1
1 I
I • ~ ImwII ng
1 1
1 I 1
IE I MA III
X
1
x 1
MA
1
1 HA
1
1 X

1 1
INA 1
1
¦ 1
X
1 1
1 NA 1
u
1 1
1 NA I »
1
IA | X
!u r*
J. Poadlni
1 I
u In* 111
I
x I

1 NA
1 X
1 X
INA 1
x 1
X
1 U 1

lull
IA 1 I
! " 1**
J- Swrlaet
"praying
1 X
1
HA | MA ill
1 1 f
X
X 1
1
HA
1 NA
1
1 1
1
1 X
1
1 1
lu 1
1 1
1
X I
1
X
1 1
1 NA I
1 1
u
1 1
1 NA 1 »
1 1
IA 1 1
1
iu r
*. 0itch«»
1
1 X
1 1 1
IE 1 HA 1 MA I
X
1
X t
1
1
I MA
1
1 X
1
1 X
1 1
1 X
1
X I
X
1 1
1 X 1
X
1 1
1 U |
1
IA 1 *
i i
1 U 1 «*
S. t"'1 It rat Ian
t:«ll«r1»a
1 X
1
IX I w 1 M 1
1 1 1
X
1
1
'
I X
I
1 X
1
1 X
1 1
1 1 1
I
X I
X
1 I
1 x I

1 1
1 IX I
1
IA 1 «
1 X 1 MA
| I
*• IMtitrattn*
1 X
1
> 1 1
1 "A I HA I
1 1 1

1
1
X
1
1 X
1
1 x
1
1 X
1 1
lHA 1
1
X 1
1
1 1
1 X 1

1 1
1 U 1
1
IA I X
1 1
1 X 1 NA

1
! 11

I

1

1
1
1

1

1 1
1
1 1
1 . 1 n l-rj tnn
Hp*.
1 <
1
1
" « ' 111'
1 j"'!
1—	1	1 1


1
1
t X
1
1
1 X
\
1
1 1
1
1
1 x
1
1
X I
1
X
1
1
1 X
1
X
1 1
1 1
1 x {
1 1
»"> ! «
1
1 1
1 X 1 I
1 1
1 * Applicable
'•t " Lr«a
Applicabl
U)
(2)
•	«« CMblnxl ,t..nr Con,.I
•	Applicable with tUutro-oatwaia.
16

-------
TABLE 1-6
MATRIX FOR RECOVERY METHODS
Reco very
Methods	Depth to Groundwater	Hydraulic Conductivity




> 10* 1 - 10-'
10"'- 10-4
JO*4- 10" *




cm/sec
em/sec
cm/sec

0-5 m
512 m
> 12 m
(> 280-2.8
(2.8-0.28
(0.28-0.00?

<0-16 ft)
(16-40 fl)
(>40 ft)
ft/day)
ft/dav)
fl'dav)
Gravity:
Open Ditche*
X
NA
NA
X
LE
NA
and Trenches






Porous Drains
X
NA
NA
X
LE
NA
Forces:






Wellpoini
X
X
NA
X
LE
NA
Deep Well
NA
X
X
X
LE
NA
Vacuum Well
X
X
NA
NA
X
LE
Point






Electro-
X
X
X
NA
NA
X
osmosis
X « Applicable
LE ¦ Less Effective
NA = Not Applicable
conductivities between 1 x 10"1 cm/sec and 1 x 10"3 cm/sec (280
to 2.8 ft/day). Forced delivery methods would be applicable for
situations where hydraulic conductivities are between 1 x 10"3
cm/sec and 1 x 10"4 cm/sec (2.8 to 0.28 ft/day). At a site where
the hydraulic conductivity is less than 1 x 10"4 cm/sec (0.28
ft/day), enhancement techniques such as electro-osmosis of
hydro fracturing would be required.
In terms of recovery systems, a site where the hydraulic con-
ductivity is between 1 x 10"1 cm/sec and 1 x 10"3 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 conductivity between 1 x
10~3 cm/sec and 1 x 10"5 cm/sec (2.8 to 0.03 ft/day). For a waste
deposit having a relatively low hydraulic conductivity (below
1 x 10"5 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 well-
points 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).
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 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 freezing 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 accom-
modate 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 reactant solution directly in 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 effec-
tively reached by gravity delivery methods.
The hydraulic gradient formed by the natural conditions can-
not 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.
REFERENCES *
1.	American Society of Civil Engineers (ASCE), 1972. Ground-
water Management Manual. ASCE Manual 40, ASCE, New
York, NY.
2.	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.
3.	Black, C. A. (ed). 1965. Methods of Soil Analysis, Part I:
Physical Properties, Agronomy 9, Amer. Soc. of Agron.,
Madison, WI.
4.	Bouwer, H., 1964. Measuring Horizontal and Vertical Con-
ductivities of Soil with the Double Tube Method. Soil Sci. Soc.
Amer. Proc. 28:19-23.
5.	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.
6.	Bouwer, H. and R. C. Rice, 1967. Modified Tube-Diameters
for the Double Tube Apparatus. Soil Sci. Soc. Amer Proc.
31:437-439.
7.	Cedergren, H. R., 1977. Seepage, Drainage and Flow Nets (2nd
edition). J. Wiley and Sons, New York, NY.
8.	Federal Highway Administration, 1976. Grouting in Soils.
FHWA-RD-76-27. FHA Office of Research and Development,
Washington, DC.
9.	Freeze, R. A. and J. A. Cherry, 1979. Groundwater. Prentice
Hall, Englewood Cliffs, NJ 604 pp.
10.	Fruco and Associates, 1966. Dewatering and Groundwater
Control for Deep Excavation. U.S. Army Engineering Water-
ways Experimental Station, Vicksburg, NJ.
17

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11.	Fry, A. W. and A. S. Grey, 1971. Sprinkler Irrigation Hand-
book. Rain Bird Sprinkler Mfg Corp., Glendora, CA.
12.	Gibb, J. P., M. J. Barcelona, J. D. Ritchey and M. H.
LeFaivre, 1975. Effective Porosity of Geologic Materials. In:
Land Treatment of Hazardous Wastes: Proc of the 11th
Annual Res. Symp. EPA/600/9/85-013. HWERL, US En-
vironmental Protection Agency, Cincinnati, OH. pp. 190-197.
13.	Huisman, L., 1972. Groundwater Recovery. Winchester Press,
New York, NY.
14.	Huisman, L. and T. N. Olsthorn, 1983. Artificial Groundwater
Recharge. Pitman Advanced Publishing, New York, NY.
15.	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.
16.	Little, A. D., 1983. Handbook for Evaluating Remedial Action
Technology Plans. EPA-600/2-83-076. MERL, U.S. U.S.
Environmental Protection Agency, Cincinati, OH.
17.	Lohman, S. W., 1979. Groundwater Hydraulics. U.S.
Geological Survey Professional Paper 708. U.S. Geological
Survey, Alexandria, VA.
18.	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.
19.	Luthin, J. N. (ed.), 1957. Drainage of Agricultural Lands.
American Society of Agronomy, Madison, WI.
20.	Martin Associates, A. W., 1978. Guidance Manual for
Minimizing Pollution from Waste Disposal Sites.
EPA-600/2-78-142. MERL, U.S. U.S. Environmental Pro-
tection Agency, Cincinnati, OH.
21.	Meinzer, O. E., 1923. Outline of Groundwater Hydrology.
U.S. Geological Survey Water Supply 494 (reprinted 1968).
U.S. Geological Survey, Alexandria, VA.
22.	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. O. Riggs, eds), ASTM STP 746. American Society for
Testing and Materials, Philadelphia, PA.
23.	Park, J. E., 1985. Permeable Materials for the Removal of
Pollutants from Hazardous Waste Leachates. Proceedings of
the 11th Anual Research Symposium on Land Disposal of
Hazardous Wastes, HWERL, USEPA, Cincinnati, OH.
p. 19-26.
24.	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.
25.	Repa, E. and C. Kufs, 1985. Leachate Plume Management.
Draft Report for HWERL, U.S. Environmental Protection
Agency, Cincinnati, OH.
26.	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. U.S. Environmental Pro-
tection Agency, Cincinnati, OH.
27.	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 Conser-
vation Service, Washington, DC.
28.	U.S. Department of the Navy, 1982. Soil Mechanics. NAV-
FAC SM-71. Naval Facilities Engineering Command, Alexan-
dria, VA.
29.	USEPA, 1973. Wastewater Treatment and Reuse by Land
Application. USEPA Office of Research and Development,
Washington, DC.
30.	USEPA, 1976. Erosion and Sediment Control—Surface
Mining in the Eastern U.S. EPA-625/3-76-006. USEPA,
Washington, DC.
31.	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.
32.	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.
33.	USEPA, 1980b. Design Manual for Onsite Wastewater Treat-
ment and Disposal Systems. EPA-625/1-80-012, USEPA
Office of Research and Development, Washington, DC.
34.	USEPA 1980c. Lining of Waste Impoundment and Disposal
Facilities. Manual SW-870, USEPA Office of Water and Waste
Management, Washington, DC.
35.	USEPA 1982. Remedial Action at Waste Disposal Sites.
EPA-625/3-76-006. USEPA, Washington, DC.
36.	USEPA, 1984a. Slurry Trench Construction for Pollution
Migration Control. EPA-540/2-84-001 MERL, U.S.
Environmental Protection Agency, Cincinnati, OH.
37.	USEPA, 1984b. Review of In-Place Treatment Techniques for
Contaminated Surface Soils. EPA-540/2-84-003a. MERL,
U.S. Environemental Protection Agency, Cincinnati, OH.
38.	USEPA, 1984c. Summary Report: Remedial Response at
Hazardous Waste Sites. EPA-540/2-84-002a. MERL, U.S.
Environmental Protection Agency, Cincinnati, OH.
39.	USEPA, 1984d. Case Studies 1-23: Remedial Responses at
Hazardous Waste Sites. EPA-540/2-84-002b. MERL, U.S. En-
vironmental Protection Agency, Cincinnati, OH.
40.	Urguhart, L. C. (ed.), 1968. Civil Engineering Handbook.
McGraw Hill Co., New York, NY.
41.	Wang, H. F. and M. P. Anderson, 1982. Introduction to
Groundwater Modeling. W. H. Freeman Co., San Francisco,
CA.
42.	Winterkorn, H. F. and H-Y. Fang, 1975. Foundation
Engineering Handbook. Van Nostrand Reinhold Co., New
York, NY.
18

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SECTION 2
BIODEGRADATION
2.1 INTRODUCTION
Recent developments in applied microbiology (Bitton and Berga,
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 in-
volves 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 com-
pounds include industrial surfactants, organic solvents, crude and
refined petroleum products, pesticides and herbicides, poly-
chlorinated biphenyls, polycyclic aromatic hydrocarbons and other
classes of organic compounds (Bitton and Berga, 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 ground-
water is recirculated from downgradient recovery wells, 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 treat-
ment 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 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 hydro-
carbons as the sole source of carbon and energy. These microbial
communities react to 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 (C02) 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 sus-
ceptible to further biodegradation. Microorganisms catabolize
hydrocarbons by anaerobic respiration in the absence of molecular
oxygen using inorganic substrates as terminal electron acceptors.
In anaerobic respiration, C02 is reduced to methane (CHJ,
sulfate (SCJ2) to sulfide (S~2), and nitrate (NCp to molecular
nitrogen (N2) or ammonium ion (NHJ. 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.
In most cases, naturally occurring microbial activity can decom-
pose organic materials of both natural and synthetic origin to
harmless or stable forms or both by aerobically mineralizing them
to C02 and water, or anaerobically decomposing them to CO,,
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 con-
ducive 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.
19

-------
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 occurring 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, 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, ground-
waters, 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/oxygena-
tion 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 and 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 demonstrated the abilities associated
with the organism identified by the strain number. In other in-
stances, 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 calls. These
episom.es 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 having 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 degra-
dation (Kogayashi 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 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 us-
ing oxygen as the terminal electron acceptor or anaerobically by
nitrate respiration.
Oligotrophic bacteria are defined as microbes that live under con-
ditions of low productivity (carbon flux of less than one mg/l/day).
The Caulobacters (Poindexter, 1981) are the best known group of
obligate oligotrophies 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-geen 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).
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, car-
bon 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 com-
pound, the number and physiological state of the organisms as a
function of available nutrients, the seasonal state of microbial
20

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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 association with the use of biological agents in organic waste
treatment. Components of this risk include the probabilities of
release, survival and growth of non-indigenous microbes, and the
consequent occurrence of some undesirable environment 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 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, substitu-
tions, 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 (BOD2.) to ultimate
oxygen demand (UOD), indicate what proportion of compounds
can be degraded biologically (estimated by BOD) compared to the
biorefractory portion which would require chemical decomposi-
tion (estimate by COD or UOD). BOD5/COD ratios and
BOD2|/UOD ratios (also called refractory index, or RI) 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, naphthalene) ap-
pear to have relatively high degradability using these relationships,
while halogenated phenols, aliphatics and aromatics appear to be
less readily biodegraded (Rapa 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 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, essentially because of their simple form and
availability in the environment. Native populations of
microorganisms have been shown to degrade aliphatic hydrocar-
bons ranging in concentration from 1 ppm or less up to approx-
imately 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 con-
centrations 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 hydro-
carbons including gasoline, fuel oil, and other petroleum products.
They found that waters containing less than 10 ppm hydrocarbons
generally had populations of less than 103 organisms per ml while
waters with hydrocarbon concentrations in excess of 10 ppm
generally supported populations on the order of 106 organisms per
ml. Species were identified as belonging mostly to the genera
Pseudomonas and Arthrobacter.
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). Represen-
tative 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, concen-
trations 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; Orn-
ston, 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.
Degradation rates for cyclic aliphatics and aromatic hydrocar-
bons by native microbes are slower than for acyclic compounds.
Native microorganisms will completely or partially degrade these
21

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compounds at concentrations below 100 mg/1 over a period of 10
to 90 days (Horvath, 1972; Rogers et al., 1981; Rubin etal., 1982).
Little data are available concerning mutant bacterial degradation
rates on unsubstituted aromatic hydrocarbons. However,
genetically altered strains of Pseudomonas, 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	POLYCLICIC AROMATIC HYDROCARBONS
Microbial degradation of polycyclic aromatic hydrocarbons
(PAH) compounds have been identified (Atlas, 1981; Cohen and
Cabriele, 1982; Herbes, 1981; Kobayashi and Rittman, 1982;
Sherrill and Sayler, 1980). However, uniform degradative
pathways, comparable to those for the aliphatic and aromatic com-
pounds, 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
(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 biodegra-
dation 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 substrate (Sherrill and
Sayler, 1980). Transformation rates in microbial communities shift
slowly in responst 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 environ-
ment 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; Reichardt 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 com-
pound 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 naphthalene are also degraded by Pseudomonas aero-
bically (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 on 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 McCarty, 1967; Kallman
and Andrews, 1963; Kobayashi and Rittman, 1982; Marinucci and
Bartha, 1979; Reichartdt et al., 1981; Schreiber et al., 1980). Reduc-
tive dechlorination is reported to be significant only when the
environmental oxidation-reduction potential (EH) 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 impor-
tant to note that polychlorinated biphenyls and halogenated
benzenes have been found to be degraded only under aerobic con-
ditions (Kobayashi and Rittman, 1982).
A great deal of concern has been expressed concerning the per-
sistence 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 dehalogena-
tion 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
/xg/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, 3 983a).
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).
Axenic (pure) bacterial cultures of Chlamydomonas and sewage
microfauna release chlorine from the compound and the latter pro-
duces 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 C02 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 pentach-
lorophenol 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
3,4-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 improv-
ed degradation of halogenated hydrocarbons can be achieved with
constructed strains (Schwien and Schmidt, 1972). 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 isometric 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,
22

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1982; Reichardt 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 (Reichardt et
al., 1981), Pseudomonas, Flavobacter, Archromobacter,
Chromobacter and Nocardia (Kobayashi and Rittman, 1982). Com-
pounds in this category which have been shown to be degraded
microbiologically are identified along with the agent of degradation
in Appendix A.
Degradation of chlorinated biphenyls have been observed in a
mixed marine microbial community with estimated turnover rates
of one year at concentrations of 0.1 ^g/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 pseudomonada (harboring the TOD, pAC25, pKFl,
pAC21 and pAC30 plasmids which regulate the degradation of
these aromatic compounds) allowed greater than 98 percent utilizing
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 decomposi-
tion 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 come-
tabolism (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 tuluedide, 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 partial degradation of lindane, heptachlor, endrin,
aldrin, heptachlor epoxide, DDT, DDD and dieldrin in anaerobic
digester sludge (Hill and McCarty, 1967). DDT is converted to
DDD 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 in-
secticides 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 AC 1100
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
plasmic 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 the cell's metabolic range to other chlorobenzoates
as well. These plasmids appear to evolve by recruitment of a variety
of genes from other plasmids and interact among themselves to
extend the substrate range of host cells to a wide variety of
xenobiotic compounds. Kellogg et al. (1981) report having
developed by plasmid assisted nolecular 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 Bl-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 pseudoalcatigenes 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.
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 invironment 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 (NDE1A) was
shown to degrade slowly at low concentrations (1 ^g/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 Deposite
Commercial operations already exist (Aquifer Remediation
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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 pro-
cesses, 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 con-
sidered their treatment processes or products proprietary informa-
tion or simply had insufficient information on product applica-
tion to be useful and therefore were not referenced in this
document.
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 ground-
waters (API, 1982; Jhaveri and Mazzacca, 1983; Kaplan and
Kaplan, 1982; Zitrides, 1978), lagoons or surface impoundments
(Zitrides, 1978) and industrial waste treatment plants (Zitrides,
1978).
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 Rapa 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 nutri-
tional content of the soils and materials to be treated.
The proper microorganisms of groups of microbes must be
selected to treat the waste. Commercial firms use their past
experience, laboratory screening, onsite test plots, or any combina-
tion 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 concentration. 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 inocu-
lated 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, 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 concentration 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 con-
taminated 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 treat-
ment 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
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 the site, the
optimum concentrations of microbes, emulsifier and fertilizer
required to support maximum biological activity. Two waste treat-
ment 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 supple-
mental 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 pro-
mote the biodegradation of recalcitrant chlorophenol compounds
at the 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, addi-
tion 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 evalua-
tion during laboratory and field pilot studies.
Once the proper microbial strains, site saturation, pH and oxygen
requirements, fertilizer formulations and emulsifier concentrations
24

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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, suf-
ficient 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:
•	Collect data on waste sample analysis, soil composition and
indigenous microbial populations;
•	Obtain monitor well data if applicable;
•	Collect any other site data (soil type, moisture, pH, pE,
temperature, nutrients, etc.), necessary to complete a bench-scale
study;
•	Analyze data and choose or breed proper microorganisms;
•	Mutate and culture those microbes to perform desired functions;
•	Inoculate waste samples with selected microbes under accepted
scientific procedures, and observe biodegradation rates;
•	Determine optimum soil moisture, pH, oxygen, fertilizer and
emulsifier requirements;
•	Establish technical and economic feasibility of biological
approach to waste treatment;
•	If technically and economically feasible, perform pilot study if
required;
•	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 con-
taminated area and the soils tilled again. Nutrients, emulsifier and
bacteria were reapplied after six weeks; the site was tilled as required
for aeration and to assure the complete mixing of microbes, fertili-
zers 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 con-
tinued 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 reviews of
past experience with similar problems (Kopecky, 1983). In general,
the selected bacterial culture is sprayed evenly over the site at the
rate of one kilogram of bacteria per 25 m2 (one pound of bacteria
per 120 ft2). 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, N and P) fertilizer are added per 25 m:
(four pounds per 120 ft2) 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.
2.3.2.3 BIODEGRADATION OF FORMALDEHYDE IN
SURFACE SOILS
The main valve on a railroad tank car containing a 50°io formal-
dehyde 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 bacteria 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.
, Liquids recovered from the waste deposit are monitored to deter-
mine 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 Mazzacca, 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 con-
taining large concentrations of highly insoluble waste, or a com-
bination 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
25

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TABLE 2-1
EXAMPLES OF BIOLOGICAL RENOVATION AT CONTAMINATED SURFACE SITES1
Waste
Site
Containment
Concentration
fppmj
Waste
Site
Characteristics
Biological
Agent
Treatment
Time
(Days)
Residual
Concentration
(ppm)
Reference
Styrene
Atrazine
Sludge containing
T richolorophenate
Petroleum
25
Saturated
Soil
300
12,000
Distillate

Acrylonitrile
1,000
Formaldehyde
1,400
Ortho-Chloro-
15,000
phenol

Railroad tankcar spill,
area soils contamin-
ated to a depth of 8
inches
50 Acre field
Sludge spread on soil
to a depth of 6 inches
Spill covering 4
acres at an oil tank
farm
Soil and groundwater
contamination
Soil and groundwater
contamination
Soil and groundwater
contamination
Bl-CHEM-SUS-8
BI-CHEM-PBO-6
BI-CHEM-GEC-1
BI-CHEM-SUS-8
PHENOBAC
PHENOBAC
MUTANT
BACTERIA
21
less than 1
28
21
90
22
274
less than 1
less than 1
less than 1
less than 1
less than 1
'Application of biological agents and site treatment
References: 1) Kopecky, 1983
2) Zitrides, 1978
for these examples are similar to those procedures described in the text to treat hazardous orgamc wastes.
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 temperature. Alternatively a trench or
pond may be used as a biotreater (Zitrides, 1978) depending on
environmental conditions, economics and geological considera-
tions. 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).
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 sub-
surface 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
-MAIN BUILDING
J SIORM_SEW|R._J
industrial park
IETd
SUBSURFACE TANK FARM]
deepwell
PLUME ( > 10 m«/t CODI
(SOURCE: JHAVERI AND MAZZACCA. 1983)
FIGURE 2-1
BIOCRAFT SITE PLAN
26

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chloride, acetone, n-butyl alcohol and dimethyl-aniline into sub-
surface 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 two mobile settling tanks, and two up-
gradient reinjection trenches was installed. Contaminated ground-
water 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 reintro-
duced to the 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 bio-
active (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 contami-
nation by dichlorobenzene and methylene chloride is described by
Quince and Gardner (19821-
• COUAILV
VACCO *CRATIO«
HEUt
FIGURE 2-2
BIOCRAFT BIODEGRADATION TREATMENT SVSTEM-BASIC
PROCESS FLOW DIAGRAM
2.3.3.3 LEAKING UNDERGRAOUND GASOLINE
STORAGE TANK
Ten domestic water supply wells in Montgomery County, PA
were contaminated by low level, long-terms 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 con-
centrations 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 ground-
water 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 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: 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 tenfold increase
in total bacteria and a 200 fold increase in hydrocarbon degraders.
Over a period of 2Vi months, the hydrocarbon concentration in
groundwater was reduced to about 250 ppb. Activated carbon
adsorption was used in the final phase to "polish" the ground-
water to acceptable residual concentrations (Aquifer Remediation
Systems, 1985).
2.3.3.4	LEAKING UNDERGROUND STORAGE TANKS
Vapors discovered in a laboratory building at a midwestem
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 hydro-
carbons 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 led 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 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 culturees capable of
degrading gasoline were isolated. Most of these isolates were unable
to degrade many of the individual componentss 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 GROUNDWATED
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 were used to treat
this residual contamination.
The local water authority required that nutrients injected into
the aquifer to acclerate biodegradation and the flushed con-
taminants 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 1/sec) and the second for
clean injection water (20 to 30 1/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 toluene. 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).
27

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2.3 3 6 BIODEGRADATION OF ETHYLENE GLYCOL IN
GROUNDWATER
Approximately 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 surfac 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 con-
centrations of ethylene glycol were reduced by 85-93 Vo (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).
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
biodegraded, 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 concen-
trations 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 deposited as sludge) in 15,000 m3 (four million
gallons) of liquid was treated by biological degradation (Zitrides,
1978). This system was inoculated with 68 kg (150 pounds) or
PETROBAC, 68 kg (150 pounds) of PHENOBAC, micronutrients
(POLYBAC N) and emulsifiers (POLYBAC E). Ongoing treat-
ment 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.
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 discharged to local sewers at the rate of 75
m3 (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 renova-
tion 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 Kellog 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 approxi-
mately 10,000 buried drums of hazardous wastes at Picillo Farm,
Coventry, Rhode Island. The initial remedial measure at this Super-
fund site was the excavation and removal of these drums; this left
approximately 1300 mJ (46,000 ft3) of phenol-contaminated soils,
containing approximately 770 kg (1700 lbs) of phenolic compounds
(Flathman et al., 1983; Flathman and Caplan, 1985).
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 irri-
gation 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).
During 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 treat-
ment 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
Remediaton 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 con-
cerning 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 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.
28

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TABLE 2-2
SUMMARY OF MICROBIOLOGICAL TREATMENT TECHNOLOGIES'
Treatment Method
Adapted	Plasmid
Contamination Zone	Native	Native	Mutant	Associated	Constructed	Bioreactors	Aerobic
Bacteria	Bacteria	Bacteria	Bacteria	Breeding	Anaerobic
Waste Disposal Sites
Surface Waste
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.
" + Technology is available
" —Technology is not available
Technology may be developed in laboratory
±	±	+?	+?	±	+. -
± +	+	+	+	-	+?
- +	+	+	+	±	+'-
+	+	+	+	+	+ / -
± +	+	+	+	+	+/ +
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 desribed 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, plasmic
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 micro-
environments may be required for reductive dehalogenation.
Average renovation times are one to three months.
Treatment of deep soils, subsurface waste deposits and ground-
water 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 identi-
fied 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 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 are numbered to correspond with references in
Appendix A.
29

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TABLE 2-3
SUMMARY OF ORGANIC GROUPS SUBJECT TO MICROBIOLOGICAL METABOLISM1
Substrate
Compounds
Respiration
Fermen-
Aerobic Anaerobic ration
Modes of Microbial Metabolism
RING Fission
Oxida- Co-oxi- Oxida- Reduc- Dehalo- Esteri- Ester- Dehydro-
tion dation rive live genation ftcation ases genation
Deamina-
rion
Phoiome-
tabolism
Degra-
dation
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
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	+	f	+	+ +
Halogenated Aromatics	+	+	+ + + +
Simple Aromauc	+	+	+ +
Nitro 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	+
Polycyclic Hydrocarbons	+
Fused Polycyclic	+
Hydrocarbons
Organophosphates	+	+	+	+
Pesticides and	+	+	+	+
Herbicides				
1 This table is a condensed version of Appendix A. Please refer to the Appendix for specific organics and the biological agents participating in the
of these compounds.
+
+
+
+
+
+
+
+
+
+
+
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E. Wildung. 1981. Microbiological Degradation of Organic
Components in Oil Shale Retort Water: Organic Acids. Appl.
Environ. Microbial. 42:830-837.
87.	Rosenberg, A. andM. Alexander. 1979. Microbial Cleavage
of Various Organophosphorous Insecticides. Appl. Environ.
Microbiol. 37:886-891.
88.	Rosenzweig, W. G. and G. Stotzky. 1980. Influence of
Environmental Factors on Antagonism of Fungi by Bacteria
in Soil: Nutrient Levels. Appl. Environ. Microbiol.
39354-360.
89.	Rubin, H. E., R. V. Subba-Rao and M. Alexander. 1982.
Rates of Mineralization of Trace Concentrations of Aromatic
Compounds in Lake Water and Sewage Samples. Appl.
Environ. Microbiol. 431133-1138.
90.	Schreiber, A., M. Hellwig, E. Dorn, W. Reineke and H. J.
Knackmuss. 1980. Critical Reactions in Fluorobenzoic Acid
Degradation by Pseudomonas Sp. B13. Appl. Environ.
Microbiol. 39:58-67.
91.	Schwien, U. and E. Schmidt. 1982. Improved Degradation
of Monochlorophenols by a Constructed Strain. Appl.
Environ. Microbiol. 4433-39.
92.	Serdar, C. M., D. T. Gibson, D. M. Munnecke and J. H.
Lancaster. 1982. Plasmid Involvement in Parathion
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Microbiol. 44:246-249.
93.	Sherrill, T. W. and G. S. Sayler. 1980. Phenanthrene
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Microbiol. 39:172-178.
94.	Shiaris, N. P., T. W. Sherrill and G. S. Sayler. 1980.
Tenax-GC Extraction Technique for Residual Poly chlorinated
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Biodegradation Assays. Appl. Environ. Microbiol.
39:165-171.
95.	Shimao, M., Y. Taniguchi, S. Shikata, N. Kato and C.
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by a Symbiotic Mixed Culture. Appl. Environ. Microbiol.
44:28-32.
96.	Shoda, M. and S. Udaka. 1980. Preferential Utilization of
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Appl. Environ. Microbiol. 39:1129-1133.
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33

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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
biodegradable 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, solubili-
zation, and emulsification of various types of organic chemicals
have found widespread application. These properties make surfac-
tants 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 sur-
factants 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 considera-
tion of fundamental chemical characteristics of the principal
surfactant classes with respect to their applicability to in situ organic
waste treatment.
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 ef-
fectiveness 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 solu-
tions 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:
• 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 absorbed 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.
•	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
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 incor-
porate or "solubilize" susceptible organic compounds either
within their interior (hydrophobic regions) or at their external
peripheries (hydrophilic regions).
•	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 RC6H4SO,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.
•	Cationic—The surface active portion of the surfactant molecule
bears a positive charge, for example, RNH3C1 ~ (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 mole-
cule does not bear apparent ionic charge, for example,
RCOOCHjCHOHCH2OH (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.
34

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•	Amphoteric—Both positive and negative charges may be present
in surface active portion and the molecule, for example,
R~NH,CH,COO~ (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 surfac-
tant hydrophobic groups include:
•	Branched chain, long alkyl groups (Cg-C^),
•	Long-chain (C8-C15) alkybenzene residues,
•	Alkylnaphthalene residues (C3 and greater-length alkyl groups),
•	Rosin derivatives,
•	High-molecular weight ethylene oxide and propylene oxide
polymers,
•	Long chain perfluoroalkyl groups, and,
•	Polysiloxane groups.
Table 3-1 provides general information on the four major
surfactant types and their principal classes.
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 surfactants
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 (Kow), which
are commonly used as relative measures of 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 Kow values. Conversely, surfactants may be of more limited
TABLE 3-1
SURFACTANT CHARACTERISTICS
Surfactant Type
and Classes
Selected Properties
and Uses1
Solubility
Reactivity
Anionic
1)	Carboxlic Acid Salts
2)	Sulfuric Acid Ester Salts
3)	Phosphoric & Polyphosphoric
Acid Esters
4)	Perfluorinated Anionics
5)	Sulfonic Acid Salts
Cationic
1)	Long Chain Amines
2)	Diamines & Polyamines
3)	Quaternary Ammonium Salts
4)	Polyoxyethylenated Long Chain
Amines
Nonionic
1)	Polyoxyethylenated Alkylphenols
Alkylphenol Ethoxylates
2)	Polyoxyethylenated Straight
Chain Alcohols & Alcohol
Ethoxylates
3)	Polyoxyethylenated Poly-
oxypropylene Glycols
4)	Polyoxyethylenated Mercaptans
5)	Long-Chain Carboxylic Acid
Esters
6)	Alkylolamine "Condensates",
Alkanolamides
7)	Tertiary Acetylenic Glycols
Amphoterics
1)	pH Sensitive
2)	pH Insensitive
Good Detergency
(1, 2, 3, 4, 5)
Good Wetting Agents
(1,2, 3, 4, 5)
Strong Surface Tension
Reducers (4, 5)
Good Oil in Water
Emulsifiers - (5)
Emulsifying Agents
(1, 3, 4)
Corrosion Inhibitor
(1)
Emulfifying Agents
(1, 5)
Detergents
(1, 2, 4, 6)
Wetting Agents
(2, 7)
Dispersents (3)
Foam Control (3)
Solublizing Agents
Wetting Agents
Generally Water Soluble
(1, 2, 3, 4, 5)
Soluble in Polar Organics
(5)
Low or Varying Water Solubility
(1, 2, 4)
Water Soluble (3)
Generally Water Soluble
Water Insoluble Formulations
(1, 6, 7)
Varied (pH dependent)
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)
Acid Stable (1, 3)
Surface Adsorption to Silicaeous
Materials (2)
Good Chemical Stability
(1, 6)
Resistant to Biodegradation
(1)
Relatively Non-Toxic (all)
Subject to Acid and Alkaline
Hydrolysis (5, 6, 7)
Non-Toxic
Electrolyte Tolerant
Adsorption to Negatively Charged
Surfaces
Numbers refer to applicable classes within a given surfactant type.
35

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TABLE 3-2
PROPERTIES OF SELECTED ORGANIC COMPOUNDS WHICH INDICATE THE POTENTIAL EFFECTIVENESS OF SURFACTANTS'
Oclanol Water Partition	Interfacial Tension (Yf
Contaminant Class	Compound	Water Solubility (M)	Coefficient (log Kow)	(dynes/cm)
Normal




Hydrocarbons
n-Pentane
5.65 x 10~4
1.43 x 10"4
3.62
—
n-Hexane
4.11
50.0

n-Heptane
3.57 x 10"4
4.66
—

n-Octane
9.66 x 10"4
5.18
50.8
Unsaturated
1-Hexene
8.38 x 10"4
3.47
—
Hydrocarbons
1-Heptene
1.85 x 10"4
3.99
—
1-Octene
3.65 x 10"5
4.88
—

1-Nonene
8.85 x 10"6
5.35
—

1-Pentyne
1.54 x 10~2
2.12
—
Halogenated
1-Chlorobutane
9.43 x 10~3
2.55
—
Hydrocarbons
1-Chloroheptane
1.01 x 10~4
4.15
—
T richloroethylene
1.04 x 10"2
2.53
—

T richloroethylene
3.2 x 10~2*
—
37.4

Carbon Tetrachloride
5.2 x 10~4*
—
45.0
Ethers
Diisopropyl Ether
8.8 x 10~2*
—
17.9

Ethyl Ether
9.3 x 10"u
—
10.7
Aldehydes and
2-Butanone
1.89
0.69
—
Keytones
3-Pentanone
0.53
0.99
—
2-Heptanone
3.57 x 10~2
1.98
—

2-Octanone
8.85 x 10"3
2.76
—

Heptaldehyde
—
—
13.7

Methyl n-Butyl Ketone
1.7*
—
9.7

Methyl Ethyl Ketone
4.9*
—
3.0
Esters
Methyl-nonanoate
1.33 x i
-------
• 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 surfactants 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 ground-
water 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 has focused on chemical characterization
of those surfactants capable of generating the "ultra low" inter-
facial 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 tensions of aqueous/oil systems to very low values
(10~2 to 10~4 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 cor-
related to those of the oil phase, and only for surfactant concen-
trations at or in excess of the critical micelle concentrations.
Decreases in surfactant concentrations below critical micelle con-
centration 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.
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:
•	Intermediate and high molecular weight aliphatics and
polynuclear aromatics (PAH) derived from crude oil,
•	PCBs in chlorobenzenes (Askarel), and
•	Di-, tri-, and pentachlorophenols.
The soil used was a fine-to-coarse loam (gravelly silty sand) with
a permeability of 10~2 to 10~4 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 solubilizing 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. Sub-
sequent 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.
3.5 SURFACTANT APPLICATION TO SUBSURFACE
DEPOSITS 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 contraints 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
37
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TABLE 3-3
SUMMARY OF EXPERIMENTS ON SURFACTANT-ENHANCED GASOLINE RECOVERY1"
Surfactants^1
Deionized Water
Nonionics
Hyonic PE-90 (DS)
Hyonic PE-190 (DS)
Hyonic PE-120 (DS)
Poly-Tergent B-500 (O)
Alrosol (O)
Anionic
Dupanol G (D)
Aerosol-OT (C)
Alfonic 1412-S (CO)
Richonate YLA (K)
C-550 Slurry (CO)
Aerosol-NIA (C)
Sarkosyl-NL (CG)
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-ll 1 0.05% (C'G)
Structure
Polethoxylate nonylphenol
Polyethoxylate nonyl phenol
Polyethoxylate nonyl phenol
Polyethoxylate nonyl phenol
Oleic fatty acid amide
Fatty alcohol amine sulfate
Na dioctyl sulfosuccinate
Linear C,2-C14 H,5_,9
(OCH,CH,)3OSO,-Na
Dodecyl benzene sulfonate,
isopropylamine salt
Linear alky] benzene
sulfonate, Na salt
Na dihexyl sulfosuccinate
Lauroyl sarcosinate,
Na salt
Fluorocarbon
Fluorocarbon
Sodium fluorinated alkyl
sulfonate
Sodium fluorinated slkyl
sulfonate
Sodium fluorinated alkyl
sulfonate
Water
Solubility
Clear Solution
Clear Solution
Clear Solution
Soluble
Dispersible
50%
15 g/1
Soluble
Soluble H,0/ETHOH
Soluble alcohol
343 g/1 25 =C
Soluble as Na salt
Greater than 2ro
Greater than 2ro
Soluble
Soluble
Soluble
Interfacial
Tension (Y)
(dynes cm)
11.5
0.12
0.72
1.2
1.3
1.9
1.0
1.2
2.4
0.61
1.2
7.1
1.8
3.9
11.0
2.3
3.2
6.9
Notes
"'Adapted from American Petroleum Institute, (1979). Surfactant concentration 0.1% unless otherwise noted.
(2)Letters	in parentheses refer to manufacturers:
DS - Diamond Shamrock C - Cyanamid
O - Olin	CO - Conoco
D - DuPont	CG - CIBA - GEIGY
R - Richardson
(3)Dash	(—) indicates not tested.
Enhanced
Gasoline
Recovery
None Observed
(?)
None Observed
None Observed
None Observed
Moderate
TABLE 3-4
RESULTS OF SURFACTANT-FLUSHING OF CONTAMINANTS FROM TEST SOIL
Percent of Contaminants Remaining in Soil
Contaminant Test
Mixture Type	Water Washes	Surfactant Washes	Water Rinses
1	2	3 4-7 8-10 1 2 3 4- 7 8-10 1 2 3 4-7 8-10
A
Shaker
96
93
91
—
—
42
26
27
—
—
25
10
10
—
—
A
Column
—
100b
-
100
100
—
llb
—
9
7
—
9b
—

7
B
Shaker
100
100
100
—
—
35
18
12
—
—
5
9
7
—
—
B
Column
	b
100
—
100
100
—
14b
—
3
3
—
3b
—
4

C
Column
—
3b
—
1
0.8
—
0.1b
—
0.1
0.1
—
0.1b
—
0.1
0.1
A = high MW aliphatic'- and polynuclear aromatics.
B =PCBs in chloroben/enes
C=di-, tri-, and pentachlorophenols
a.	number of pore volumes of eluant (water or surfactant solution)
b.	pore values 1-3 combined for column tests
38
-------
(neutralization, complexing). Among those groundwater chemical
conditions which may influence surfactant effectiveness 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, 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 conditions of alkaline solution pH, and at low electro-
lyte concentrations in soils of low particle surface area.
The principal implications of the preceding and other available
information concerning geochemical interactions likely to be
observed in specific surfactant classes may be summarized as
follows:
3.5.2.1	CATIONIC SURFACTANTS
The surfaces of typical soil particles (such as clays) are nega-
tively charged under typical soil pH conditions (pH 5-8). Therefore,
most if not all cationic surfactants are likely to be readily absorbed
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 postive charge (Figure
3-1A), favoring the use of cationic surfactants.
3.5.2.2	ANIONIC SURFACTANTS
The sensitivity of anionic surfactants to solution electrolyte con-
centrations 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 solu-
tion electrolyte concentration, and increasing soil particle specific
surface areas.
A.- SOLUTION pH CFFCCTS
a-iLfCTNOLVTE
CONCENTRATION €F FECTS
SUBSTRATC ¦ KAOLINITE
0 5*. N*Q
7
S?
13
COO 1200 1COO 1000 2400
EOUILIMIUM CONCENTRATION
< u MOLE/I)
C- SOIL COMPOSITION EFFECTS
(ADAPTED FROM WADE ET AL. W
0	MO	1000	ISOO	2000
EQUILIBRIUM CONCENTRATION
I v MOLC'H
FIGURE 3-1
THE EFFECTS OF SOLUTION pH, ELECTROLYTE COMPOSITION
AND SOIL COMPOSITION
ON SURFACTANT ADSORPTION TO SOIL
3.5.2.3	NONIONIC SURFACTANTS
Many nonionic surfactants may potentially adsorb to soil par-
ticles through a combination of nonionic interactions. Poly-
oxyethylenated 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 either linkages of the polyoxyethylene chain and
polar surface groups (such as hydroxyl). Adsorption processes of
this type may account for the apparent ineffectiveness of 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 con-
ditions. Based on their structural characteristics, amphoteric sur-
factants would be expected to undergo ionic adsorption to soil
particles with maximum adsorption occurring 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
39
-------
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 formu-
lated to minimize geochemical interactions in applications to solid
waste deposits.
3.6 ENVIRONMENTAL EFFECTS
The introduction of surfactant solutions into surface and ground-
water systems requires consideration of possible adverse environ-
mental effects. The surfactant characteristics of principal environ-
mental concern are biodegradability, toxicity to plants and animals,
and human health hazards both during and after application.
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 appli-
cation 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 commer-
cially 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 biodegradation of selected major surfactants including C12
homologs of the following classes:
•	linear alkylbenzene sulfonates,
•	alkyl sulfates,
•	secondary alkane sulfonates,
•	alcohol ethoxy sulfates,
•	alkyl phenol ethoxylates, and
•	alcohol ethoxylates.
Results of BOD, C02 evolution, and simulated treatment
process tests indicated that for all of the tested compounds sig-
nificant degradation (greater than 50%) was observed in less than
20 days. Die away tests suggested 90-100*% decreases in surfac-
tant 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.
In general, within a given surfactant class, biodegradation rates
were found (Sivik et al., 1982) to vary with:
•	the length of alkyl chains,
•	the positions of phenyl groups, and
•	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 com-
parative purposes, manufacturer-supplied toxicity data are included
in Table 3-5. Sivik et al. (1982) noted that commonly reported
LC^ values (for 24-96 hour studies) ranged from 1-50 mg/1 for
fish and 1-300 mg/1 for invertebrates for:
•	linear alkylbenzene sulfonates,
•	alkyl sulfates,
•	alpha olefin sulfonates,
TABLE 3-5
ENVIRONMENTAL CHEMICAL PROPERTIES OF SELECTED COMMERCIAL SURFACTANTS'
Surfactant Class
Example
Electrolyte
Tolerance	PH of Aqueous
Water Solubility (Hardness-PPM) Solutions1	Biodegradation
Toxic Levels1
Fluorocarbons
o Lodyne Series
(CIBA-GEIGY)
0
Soluble
o 300
5.0-8.5 (1%)
0
Slow?
o
0
3-10 gm/kg
Acute Dermal LD50
(Rabbit) 3-10 gm/kg

o Zonyl Series
0
> 2gms/ lOOgms
—

0
Slow?
0
1-25 gm/kg
Sulfonates (Anionica)
Alkanol Series
(DuPont)
0
Soluble
—
7.5-10.0 (1%)
0
Biodegraded


Alcohol Ethoxylates
o Merpol Series
0
Generally
o Electrolyte
6.0-9.0 (1%)
o
Biodegraded
o
Acute Oral
(Nonionic)
(DuPont)

>30%
Tolerant?


(20-60% in 20 days
with acclimated
bacteria)

Toxicity for Fish
1-6 mg/L
Sulfosuccinates
OT & Aerosol
0
l-60gms/100ml
o 500-2500
5.0-8.0
0
50-100%-8 days
0
1-10 ml/kg
(Anionic)
Series (Cyanamid)





(CSMA-Shake
Culture Test)


Alkyl Sulfates
Duponol Series
0
Soluble
o Electrolyte
7.5-11.0(3%)
0
Biodegraded
0
2-20 gm/kg
(Anionic)
(DuPont)


Tolerant


(days to weeks)
0
Acute Oral
Toxicity to Fish
5-20 mg/L
1	Parentheses indicate
2	Toxicity reported as
concentration of surfactant.
acute oral for rate unless otherwise specified.
40
-------
•	secondary alkane sulfonates,
•	alcohol ethoxy sulfates,
•	alkylphenol ethoxylates, and
•	alcohol ethoxylates.
In general, increases in carbon chain length up to C16 were
observed to significantly increase toxicity, with toxicity decreases
observed for longer chain lengths. Based on an evaluation of rat
LDjq 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 toxicity. Acute oral LD50
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 %-l .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 utlized 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:
•	Surfactant application to waste deposits warrants serious con-
sideration 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.
•	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, com-
mercially 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.
•	Available information suggests that a single chemical
characteristic 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.
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 surfac-
tant solutions warrants further consideration but may be
somewhat constrained because of the high surfactant solution
concentrations and volumes required to effect significant
solubilization 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 absorption
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 3-1 A), 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.
•	Within given classes, surfactants possess sufficiently varied
chemical characteristics with respect to aqueous solubility, elec-
trolyte 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).
•	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, par-
ticularly 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 inter-
facial tension measurements in predicting the effectiveness of
surfactants, the emphasis of further research should be on
laboratory scale studies, possibly including:
•	initial screenings of the effectiveness of specific surfactants to
reduce the interfacial tensions of various pure and mixed organic
phases, followed by,
•	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.
3.8 CONCLUSIONS
For waste deposits containing organic compounds of relatively
high water solubility (greater than 5 x 10~2M), 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 Kott 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 inter-
actions 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.
References
1.	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 Inter-
face Science, 92 (2): 561-574.
2.	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 Mix-
tures. J. Colloid Interface Science, 59 (1): 39-44.
Cayias, J.L., R.S. Schechter, and W.H. Wade. 1977. The Utili-
zation of Petroleum Sulfonates for Producing Low Interfacial
41
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Tensions between Hydrocarbons and Water, J. Colloid Inter-
face Science, 59 (1): 31-37.
3.	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.
4.	Ellis, W.D., J.R. Payne, A.N. Tafuri and F.J. Frastone. 1984.
The Development of Chemical Countermeasures for Hazard-
ous Waste Contaminated Soil. EPA-600/D-84-039. Municipal
Environmental Research Laboratory, US Environmental Pro-
tection Agency, Cincinnati, OH.
5.	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.
6.	Mukerjee, P. 1979. Solubilization in Aqueous Micellar
Systems. In: Solution Chemistry of Surfactants, Volume 1,
K.L. Mittal (ed.), Plenum Press, New York, NY.
7.	Park, G.A. 1967. Aqueous Surface Chemistry. In: Equilibrium
Concepts in Natural Water Systems, R.F. Gould (ed.), Ad-
vances in Chemistry Series No. 67, ACS Washington, D.C.
8.	Rosen, M.J. 1978. Surfactants and Interfacial Phenomena.
Wiley Interscience. New York, NY.
9.	Shah, D.O., ed. 1977. Improved Oil Recovery by Surfactant
and Polymer Flooding. Academic Press, New York, NY.
10.	Shaw, D.J. 1976. Introduction to Colloid and Surface
Chemistry. Butterworths, London.
11.	Sivik, A., M. Gouer, J. Perwak, P. Thayer, 1982. Environ-
mental and Human Health Aspects of Commercially Impor-
tant 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.
12.	Texas Research Institute. 1979. Final Report Underground
Movement of Gasoline on Groundwater and Enhanced
Recovery by Surfactants, prepared for American Petroleum
Institute, Washington, D.C.
13.	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.
14.	USEPA. 1982. Handbook for remedial action at waste disposal
sites. EPA-625/6-82-006. Municipal Environmental Research
Laboratory, US Environmental Protection Agency, Cincinnati,
OH.
15.	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, D.C.
16.	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.
17.	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.
42
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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 + H,0 = ROH + HX,	(4-1)
where R represents an organic moiety and X the cleaved group in
the hydrolysis reaction. In aqueous systems under typical environ-
mental 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 complications 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 com-
pounds not included in this report or the primary data sources listed
above.
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 concentra-
tion (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 car-
boxylate ester cleavage and phosphorodithioate 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 reac-
tions 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:
tl/: = (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 (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 unsub-
stituted hydrocarbons (aliphatic and aromatic), halogenated
aromatics, PCBs, phenols, aromatic amines, and many other classes
(Table 4-1). Organic functional groups susceptible to hydrolysis
include alky! 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 con-
tain 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. (1975) 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 environ-
mental 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 the purpose of
this section to examine the factors which control the hydrolysis
43
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TABLE 4-1
GROUPS OF ORGANIC COMPOUNDS THAT ARE
GENERALLY RESISTANT TO HYDROLYSIS3 (Harris, 1982)
Alkanes
Aromatic nitro compounds
Alkenes
Aromatic amines
Alkynes
Alcohols
Benzenes/biphenyls
Phenols
Polycyclic aromatic hydrocarbons
Glycols
Heterocyclic polycyclic
Ethers
aromatic hydrocarbons
Aldehydes
Halogenated aromatics/PCBs
Ketones
Dieldrin/aldrin and related
Carboxylic acids
halogenated hydrocarbon pesticides
Sulfonic acids
aMultifunctional organic compounds in these categories may be
hydrolytically reactive if they contain a hydrolyzable functional group in
addition the functionality listed above.
HYDROLYSIS
RATE
ACIO
HYDROLYSIS JX
RATE \
OVERALL
HYDROLYSIS ^
V RATE >
yyN BASE
>5^ HYDROLYSIS
^ RATE



NEUTRAL J



HYDROLYSIS



RATE
ACIDIC NEUTRAL	BASIC
pH 	»-
FIGURE 4-1
pH DEPENDENCE OF HYDROLYSIS RATE BY ACID, NEUTRAL.
AND BASE PROMOTED PROCESSES
TABLE 4-2
GROUPS OF ORGANIC COMPOUNDS THAT ARE
POTENTIALLY TREATABLE BY HYDROLYSIS (Harris, 1982)
Alkylhalides
Amides
Amines
Carbamates
Carboxylic acid esters
Nitriles
Phosphonic acid esters
Phosphoric acid esters
Sulfonic acid esters
Sulfuric acid esters
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 = kN(RX) (neutral hydrolysis)	(4-4)
-d(RX)/dt = kA(RX)(H+) (acid-catalyzed hydrolysis) (4-5)
-d(RX)/dt = kB(RX)(OH~)(base-catalyzed hydrolysis) (4-6)
Where kN, kA, and kB 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 the pseudo-
first order reaction in RX, where
- d(RX)/dt = kh(RX)	(4-8)
The overall hydrolysis rate constant, kh, is given by
kh = kN + kA(H) + kB(OH-)	(4-9)
CO
>-
<
O
<
I
pH
FIGURE 4-2
EFFECT OF pH ON HYDROLYSIS OF ETHYL ACETATE
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.
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 environ-
mental settings.
44
-------
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:
— EJKT
Kh = Ae A	(4-10)
where A is a constant, is the Arrhenius activation energy, R
the gas constant, and T is absolute temperature.
The dependence of hydrolysis rates on temperature must be con-
sidered in evaluating data, and may represent a significant source
of error in extrapolating laboratory hydrolysis data to environ-
mental 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).
Estimates of hydrolysis rates in the environment, including rates
tabulated in this report, are usually based on 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 varia-
tions may be observed near the surface. If waste deposit
temperatures are near groundwater 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 com-
pounds dissolved in water. With water as the solvent for hydrolysis,
rates may be affected by ionic strength (a measure of the total con-
centration 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 effec-
tive 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, Mabey and Mill
(1978) indicate that metal ion catalysis is unlikely to affect
hydrolysis rates at typical metal ion concentrations in the environ-
ment, although additional research would be necessary to deter-
mine 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 potential-
ly 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 mobile significant concentra-
tions 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 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"3 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 kb/kn. With pH =
11 being the approximate upper limit of pH achievable (i.e., OH
= 10 ~3) kb must be 103 times kn for the base-catalyzed rate to
equal the neutral rate at a pH of 11, 104 times kn at a pH of 10,
etc (see equations 4-4 through 4-6). If the base catalyzed contribu-
tion 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 + H20 = ROH + 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 11, based on reported hydrolysis rate con-
stants. With the exception of polyhalomethanes (e.g., CHBr,Cl),
base catalysis is not significant below pH values of 11 to 13 ti.e.,
kfe is less than 103 times kn). Furthermore, hydrolysis half-lives for
alkyl halides (with the exception of some methyl and benzyl 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
45
-------
TABLE 4-3
HYDROLYSIS OF ALKYL HALIDES
Cl« »«/Co«pourw3
*N
Alfcyl Helljei
Methyl Fluoride
Helhyl Chloride
Hethyl Bro«lde
fethyl Iodide
Methylene Chloride
CHjtMCMjCl
c.Wl
Cttn3CnClj
Cn^CnCHjBr
O^CttOi^l
t«N)CN2Br
p-UijC^H^CMjBr
ChjlrCl
ChCIj
CNBrClj
CHBrjCl
IT'1
CHIC12
CHFiCl
CCI4, lppe
CCI4, 1000 ppa
1.1 ,1-Trlchloroetheoe
let nchlorcethene
1,2-die hi or octane
1,2-dlbroeoet^ae
Ethylchlorld*
1, l-iJlChiofoe t heae
neiechlorocyclopeotedleoe)
7.44E-1Q
J.52E-7
30 jr
3Ojr
3 Siy
3Ujr

Jvy

2.37E-B
6.18E-6
U.9-Jy
Q.9Jy
U.»3y
0.93y
U.¥3y
u.tJy
'j.yjj
4.09E-7
1.4IE-4
20d
20d
20d
20d
2Ud
2ua
iwe
7.28E-8
6.47E-5
nh
l>h
lih
l>h
l>ri
Hi,
l.SM-J
-
O.lfi
U.lh
U.lh
V. 1ft
U.lh
m
v. 1 r.
*.36-2
-
lit
1*»
l*e
Ifu
1*»
lvc

1.67t-J
-
12ll
12h
l2h
Uti
I2n
l£Ti
nr.
4.U1C-0
-
2.1M
2.U4
i.va

2.0d
i.ua
i.*>£
4.5K-4
-
U.4Jft
U.4Jh
y.4jr»
u.eJft
u.4jn
w.

1.0*4
-
1.3211
l.Jlh
l.Sih
1. J2h
l.jih
1. j^r.
1. j;:.
if.67*-j
-
4.3«
4.3«
4.i«
4.3*
4.JS

4. m
-
-
-
-
44,
-
-
-
-
-

JXI,(KK)y
35,UOOy
JMH»y
JXJy

J.iy
y. j>.
-
l.fct-3
1J,?UUy
lil7Uy
U7y
13.7y
1.37y
>Ud
>0
-
e.ot-4
27,4UOy
274Uy
274,
27.4y
2. 74y
lUUd
luc
-
-
-
-
My
•
-
"
"
-
3.2S-4
t.tOUy
686U7

6«.by
0.86y
2>Ud
£>a
-
S.OC-4
27,>OOy
27MJy
27S,
27. >y
2.9y
U.26y
luc
-
2.2E-1
-
-
l.Oy
3B.*d
3.7d
0. 37a
u.vjr.
-
MCOfid
-
-
7000y
-
-
-
*

order







-
•econd
-
-
7y
-
-
"
*

order







-


-
0 .if
-
-
-
-


-
-
U.7y
-
-
*

-
-
-
-
JU.UJUy
-
-
"
*
-
-
-
-
i.OUUy
-
-
"


-
"
-
4IM
-
"


-
-
14a
14d
14 d
l<»d
14d
-
-
All vtiucc reported tor a®c. luce eoa*t«Bt* lo iec~*.
•	• eecond	~
•	* •loute
rt ¦ nour
d * a«y
y • ye«r
D«te Source*: 1. netey end Hill, 1*78
2.	Vereer, lnc, 1*79
3.	fUddlnc et el., 1977
substances is unlikely to be necessary.
Hydrolysis rates were collected for only two halogenated
alcohols, 2-chloro-ethanol and 1-chloro-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 treat-
ment 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:
R,C(0)0R2 + H20 = R,C(0)0H + R2OH	(4-12)
Acid-catalyzed, neutral, and base-catalyzed processes may con-
tribute 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):
RC(0)NR,R2 + H:0 = RC(0)0H + R,R2NH	(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
C02 according to the reaction:
ROC(0)NR,R2 + H,0 = ROH + HNR,R2 + CO, (4-14)
Table 4-8 lists calculated hydrolysis half-lives as a function of
pH for a number of carbamates. Hydrolysis half-lives for car-
46
-------
TABLE 44
HYDROLYSIS OF HALOGENATED ETHERS, EPOXIDES, ALCOHOLS
l HI at
CoapouDd Source	St?
CMoro*e thy l*c chyle tber	1	-	-	- 0.0U7*	-
bi*(CMoroMthyl}«th«r	1	-	-	3S«	- -
2-(Chioroeth*nol	1	-	-	21y	- -
1-Chloro-2-propanol	1-	-	-	-	2 y	• -
2-Chioroethylvlnyl«tb«r	2	-	-	0.41	jr	-
3-Chloro-,	1,2 epoiy,	3	1.S4E-9 il-7 -	ltd	ltd 164	164 164 IM	1»4
2-aethyl preptM
Alpha-eplehlorohydrlB	3	6.0E-4 9.®£-7 -	*.24	0.24 9.24	0.34 4.24 ft.2a	V.id
tplbro*ohy4rla	J	6.1S-4 SE-7 "	1M	154 164	164 164 164	1»4
hoc*: Ail vtlucc reported tor i5°C. Kit* couusti la mc l.
• " second
d • day
Jt - year
Date Source*: 1. Uddla| et ®1«» 1977
2.	Vtrwr, Ibc, 1979
3.	Kabey «arf Hill. 1976
TABLE 4-5
HYDROLYSIS OF EPOXIDES
t 1/^ 4t"pH"
Co«?o~n<2 Data
Source

kH

5
6
7
*
V
lu
11
1,2-Epo*y ethane
1
lt-2
6.7E-7
_
12d
12d
12d
12d
124
lid
lid
1,2 £po«y propane
1
4.6E-2
5.5E-7
-
74
14.6d
14.fed
14.6d
14. bd
14:64
1*. 6C
1 ,2 Epo*y-2-»ethyl











prop* ne
1
7.3E0
1.1E-6
-
O.ld
.9Sd
4.4d
6.8d
7.2d
7.3d
7.3d
1,2 E?o»y-3-hydro*y
1
2. 5E-3
2.84E-7

26d
28d
28d
2Bd
244
284
2»d
prvpane











I,2-Epo*y-2-•et^yl-3-
1
1.1E-2
4E-7
-
13d
ltd
16d
164
164
16d
16d
hyfiroxy propane











1,2-Epoly-2-•«thyl•3-
1
1.84E-3
tt-7
-
lid
16d
16d
164
164
l»d
16d
chlo^op^op*ne











Tt«pi«-2.3 epoiy bultoc
I
1.2E-1
51-7
-
4.7«
13d
15.7d
164
164
16d
led
CI r-2 ,3-e?o«y butane
1
2.4L-1
5£-7
-
2.»d
lid
15.3d
ltd
164
160
loo
I,^-epoiy-l-phenyl«than«
I
-
-

-
-
1J4
-
-
-
.
trans-* ,;-epoxy-l
1
*
-
-
*
-
ltd
-
-
-
-
prienyl?:op«n«
Note: 
a " day
Data Source*: 1.	Mabey and Hill, 1978
2.	tUddlng et el. , IS??
bamates 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 Phosphoric Acid Esters
Phosphoric and phosphonic acid esters are hydrolyzed primarily
by base catalyzed processes, resulting in P-O bond cleavage as
illustrated in the reaction (Mabey and Mill, 1978):
ROPR,Rj + H20 = HOPRjRj + ROH	(4-15)
Cleavage of C-O 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 com-
pounds 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 or 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.
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
47
-------
TABLE 4-6
HYDROLYSIS OF ESTERS
Coapou&a	Pat
tlft/l Acetate
l»o;rop?l Acetat*
Butyl Acetate
Vinyl Acetate
Allyl Acetate
ixnxyl Acetate
0-ac*tyl pheool
2 ,4-dloltrophenyk *c«tat«
C1C«2C(0)°CHj
CljCHCtOOCHj
CljCMCtONK^Hj
FjCHCOJOCjHj
CIjCCCO)OCHj
F jCC(0)OCjMj
F3CC<0>OctCHj)3
CH]SCHjC{0)OC2Hj
C«3S(0)CHC(0)C2H5
(CH312SCHjC(05-
OC-rf^
C2«5C(0)OC2MS
CjK7C(Q)OC2H5
(CH,)2CHC(0)0C2H5
CM2CrtCCO)OC2H5
trana-CHjCHCHCCO)-
°C2Ms
CHCC(COOC2Hs
C^H^CCOJoCHj
C^MjCCoJOCjM^
C^t £(o)OC>UQ< j) j
C^CluJOCHjC^
p-NUj-CjjMfcCv 0)0CH j
p-Wj-c^HACtuJoCMj
j»-.VJ2-CfcMC(0)-
oc2hs
1~CjM4HC(0 JCiCj**,
o-CftM4[C(u;oC2Hj)2
o-CbM4iC(0)XH2-
W*
p-Cbhi(CC0)DCH3|2
p-CtH^lCtOOC^n
1/2 at pH
I.It-*
b.Ot-i
1.3E-4
1.4E-4
1.1E-4
7.0E-5
8.5E-5
2.3E-4
l.its
1.2E-*
4.3E-7
4,06-7
6.6E-8
1.1E-5
2.IE-7
l.SE-5
1.81-J
S.7E-5
7.7E*4
3.2E-3
1.3E-3
1.1E-1
2.6E-2
l.SE-J
1.0EI
7.3E-1
2.0E-1
1.4E0
9. ill
1.4E2
2.8E3
1.JC4
4.5E3
9.2E-1
l.ttl
2.0E2
8.7E-2
3.8E-2
2.3E-2
7.6E-2
1.3E-2
4.6SEO
l.JE-3
3.01-2
6.2E-3
8.0E-3
7.4E"2
4.4E-1
2.4E-01
5.4E-1
1.0E-2
1.7E-2
2.5E-1
4.9E-2
5
1
b
19 y
7
2.Uy
«
0.2y
9
7. Jd
35y
«y
».9y
3UM
3ld
I.*y
7dy
14Uy
i.iy
U15y
7 Jd
«7d
7.Id
U.M
l.Vh
3Uy
i.Uy
1104
114
l.ld
17y
lu .4y
l.ly
4U4
4. yd
im
1 Wd
38d
5.5d
u.57d
17h
lfch
9.4h
1.8h
0.2h
214
5.04
14h
1.4h
It. 3a
llh
4.5b
Mm
4.1a
255
4.4a
6.1a
3.7*
47a
4.7a
3.3h
1.9h
23m
2.ka
19a
ISa
15s
15a
15a
15a
3.6«
l.ia
3.4a
3.4*
3.4a
8.9a
8.9a
t.9a
8.9a
8.9a
24y
2.4y
67d
8.74
0.874
1.7,
424
4.2d
0.624
1.5h
40d
96h
9.4h
57a
5.7a
52.3jr
24.4y
2.5y
914
9.14
100y
55y
5.»y
2124
214
960y
96y
9.6y
3504
354
24*y
35y
J.5y
1284
13d
ll*0y
16Uy
Uy
l.'y
624
4.7y
17Ud
174
1.74
0.174
J720y
UWy
ll#y
n.tff
1.2/
73Uy
73y
7.3y
U.73y
274
3MHjy
35Uy
3>y
i.5y
12*4
moy
2 7Uy
27y
2.7y
994
2«2y
3Uy
3.0y
U.Jy
114
34y
3.4y
U.Wy
12.44
1.24
92y
i.ty
U.92y
344
3.4d
41y
*.ly
0.41y
154
1.54
2200y
220y
22y
2.2y
80.34
130Cy
13Uy

l.iy
47. 3d
BttOy
B8y
0. My
324
3.2d
32Qy
32y
3.2y
ll7d
11.74
1U
3.14
>.JW
Lis
4.ah
*.6h
lh
1.2a
>0s
2.5a
0-5a
1.6 a
Ife
3.6a
8.9a
2.1h
9a
35a
0.94
2.14
3.34
1.34
«.2o
i4«
0.12y
1.	7i
W.ad
t.ta
l.ld
;.t»h
j.«h
B.Wd
4.00
7.7h
1.2(1
11
u.w7j«
u. Jifl
U.iu
l.*a
u.^or.
lft
>a
>»
U.4s
0.05«
0.2a
15a
3.6a
0.9b
13*
U
~.5a
2.2h
5h
U.3>4
0.13d
U.**«
i.4a
4.4d
U. *7d
i • ->4
1.Ud
i.4i»
~.isft
U.Oli
u. joh
v.«d
0.404
y.77t>
2.4h
hoct: All values repotted (or 2S*C. Plate constant* In	.
a * second
• ¦ alnute
h • hour
d • day
y • year
t)jid dourer: 1. Ha bey and Hill, 19?W
TABLE 4-7
HYDROLYSIS OF AMIDES
Coapo^fwi		li
Acctaaldc
Valeraslde
laobutyaaldc
Cyclopentanec»rbo*aald«
Methoiy icetaalde
CMofoacf taald*
UlcM or oicet aside
Tr Uhloroacet4kM«
Rruaoacelaaldv
t hyl*cetaside
S-e thylacelaaide
t 1/J
8.36E-6
5.43E-6
4.63E-6
2.34E-5
7.84E-6
1.IE-5
3.2E-7
9.36E-«
4.71E-5
262y
2490y
3950y
465y
46.5y
1.41E-5
4U4y
3950y
ll.JUUy
156Uy
I56y
2.40E-5
470y
4SOOy
77UUy
9l5y
91.5y
1.67E-5
93.9y
V3ly
5500y
I300y
1.32y
3.95E-A
280y
186Cly
500y
55.6y
5.6y
J.5E-1
84.5y
14.6y
1.46y
y.liy
>. id
3.0E-1
73y
7.5y
0.73y
26.64
2.74
9.4E-1
ih .
2.3y.
0.23y
N.4d
U.84d
1.03E-5
2110S
2*10*y
21»200y
212Uy
lily
5.44E-4
6*0Oy
58,400v
1.8*lU5y
3«.000h
4020y
402y
3.1UE-6
23,OOyy
7U.0O0y
7090y
7U9y
iu
*.t>y
15.fry
9.2y
13. 2y
U.iby
u.»d
6.in
2.Oh
21. If
4Uy
n*
11
U. 4 7y
1	.Wy
v.* if
li2y
204
1. Jt.
y.fch
Wa
2	.1 y
* .uy
J-it
Note: All value* at ^5°C. Kate conatanta In aec.
s • slouce
n • hovr
4 " «ay
y ¦ year
U4i» boy tee: I. HatNsjp a 0(4 Hill, lWtJ
generalities cannot be 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.
48
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).
-------
TABLE 4-8
HYDROLYSIS OF CARBAMATES
Date Sowrc*
t 1/3
Chj0C(0)N(
C2H50(C0)N(CH3;C6Hj
1
-
3.5E-3
4*lUsy
4U.UUUV
4.4sltf*y
4,0U0y
4UUy
4Uy

i4oa
1
"
3.0K-6
4.4mlU6yr
44,UU0y
4,400y
440yr
44y
4.4y
CfcrfjtXCOjNCHKtHs
1
~
5.42E1
lSUd
13d
1.3d
3.6h
21a
2a
1 i s«c
CbMs0CC0)ti(CH3)C^Ms
1
-
4.2E-3
3.^*105y
52,OUQy
32U0y
320y
ill
*.2y
Wl4
P-CNjGCaH4{0)H(h7-


2.3E1







C6Mj
1
-
320d
32d
3. 2d
7. 7h
44a
4.ba
iHKC
e-ClCfcrt4OC(0)N-










*fcN5
1
-
1.8E3
4.3d
llh
l.lh
6.4a
39mc
i.vacc
J.tkCC
p-N^CbniOCCOJSCH)-


2.7L3







C6n3
1
*
43a
4.3a
2«acc
^.ba«c
u. Ja«c
u.u3a«c
u.cujfc*
p-*O^L0n<>v»C{O)N-










(Otj;CeHj
1
-
«.Ut-4
47,3UOy
273Uy
27>y
^7.5y
i. 7y
luuo
1^0
i-C^K-wCCQWnJCrlj
1
-

2.3y
«34
t».3d
^Uh
kh
l.im
J«MC

1
l.Ht-Ll
-
l^UUy
12UUy
12uvy
l^UUy
UWJ
liiMy
iiOV,
i C/t i) jHtH2C»»2uC"
1
-

B.JilU^y
»S,UUUy
»4uuy
e>Uy
•>y
d.iy
JIUO











j*ch/;«20c-
1
"
V.4C-7
a.JiiiVy
i.iiiy6y
2.4ilu^y
IS(UVUy


iij
uHj> yttoH4oC(U)M-CM)


».7*-l






1.7k
Ch3
1
-
yjy
3.3y
12U0
It*
l.tc
l.tb
(CH£)ydt*H40C(0)N-
1
-
2.SE-4
B9,3UVy
7«S0y
783y
14.iy
J.ty

111
ClCM2CK2OC(0)li(fi)-










cj»Mi
1
-
1.6E-3
14,0W0y
14V0y
14Uy
i*y
1.4y
>04
ia
CUCHCH2lX:(0)NHC6Hi
CCl3CM20C(0)NhC6M5
1
-
3.0E-2
440y
44,
4.4y
16Ud
lbd
l.frd
j.vn
1
-
3.2E-1
69y
6.9y
232d
23d
2.3d
*>h
jfta
Cf3Crt20C(0)HMC6^3
1
-
1.0E-1
220y
22 f
2.2y
SOd
•4

2h
Ethyl carbaaate
2
-
-
-
-
ll.UOOy
•
•
-
-
C2Hy)C(0)MHCHj
2
-
-
-
-
JD.DUOy
-
-
-
-
C2H5OC(0)M(CH3)2
2




39,000y




»oc«: (U1 value# at 25' C. IUt« conatanti la »ic^.
a " alnutc
h " hour
a " day
y - y««r
j«ta iourc«»; I. Ma bey and Hill, l*7tt
i. iuadln« «t •!., i*/7
TABLE 4-9
HYDROLYSIS OF PHOSPHORIC AND PHOSPHONIC ACID ESTERS
1U at yM~
CNsP(0)(0C2H4)2
CMjP(0)(OCM(CHj)j)j
CHjKOXOCiKjMO-p-
c6h4mo2)
CjHjP(0)(0CM(CHj)j)j
C6H5P(0)(OC2H5)2
(Cll jO) jPO
(c2h<,o)3po
(CjHtS)3PO
«V»50)3PO
P<0)<-p-
Cfefi4H02)
< p-JK
(CHjOjjKS)?-
c^h4mo2
CH,OP(S)SCKgH-
t
I	p~
C.H.MO.)
1. JfeE-*
1.7E-9
6.4E-9
1.2 C-7
J.2E-9
1.1E-9
1.8E-8
4E-9
1.4E-9
2.7E-11
J.3E-*
l.QE-3
2.3K-3
2.2E-4
3.2E-7
4.0E-2
3.7E-®
3E-4
l.JE-4
«.2E-«
1.2E-2
1.7E-2
3.3E-1
3.43E-1
B7Wjr
U.UOUv
3.4*10>f
330y
*.9il03y
43,OOOy
l.Ijr
3.5y
l*y
3J0y
2d
12*
3y
7y
tHWy
99#Uy
2.3ilU*y
ft.2il(j'y
4400y
i.2y
3.3y
14y
112«
2d
12a
l#7d
7y
My
1.5y
l.*y
i.Jy
24
11a
l.M
7yr
204
>9,UWy
4.4y
l.ly
3.4y
IM
474
9h
9a
4.>h
*yr
iy
»yuy
>9tN)y
1MM
2 MM
4.ty
74
34
*7a
iili
i.iC
•Yy
4.»h
3»Uy
1M
344
l.#7
lbh
llh
22a
ZUmc
*ol«: Ail values reported tor T"*5°C. Kate cooataota 1b
• * aloute
h " hour
d • day
y ¦ y«*r
Aata Sourct: 1. Nabcy aad Hill, 1978
The contaminated surface waters were treated by aeration, solar
oxidation, evaporation and powdered activated carbon adsorp-
tion/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 migra-
tion. In addition, the black carbon absorbed solar radiation,
thereby elevating soil temperatures and catalyzing pesticide destruc-
tion (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 sedi-
ment 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
49
-------
TABLE 4-10
HYDROLYSIS OF MISCELLANEOUS COMPOUNDS (INCLUDING PESTICIDES)
t III .t
Cos?ound Data
Source

*N

J
6
J
0
9
10
11
beta*Proptolac:ob«
1
.
3.3E-3

i.b»
3.fe

3.5a
3.5*
3.Sa
3.
CN2C.1 jS ()
1

2.1SE-S
"
e.vh
8.9h
e.9h
6.9b
0.9h
0.^h
0.Vh
Olaethyl sulfate
1
-
1.66E-4
1.48E-2
1.2h
1.2h
l.2h
1.2h
1.2h
l.lh
l.lh
&la(cTtloro*etnyl) cthet
1
-
l.at-2
-
2>ttc

^M(
2>mc


2>MC
k-her.y LdlaethyltrlazlM
1
-
2.7JE-S
-
7h
7h
7h
7h
7h
7h
7h
fremoyl chloride
1
-

-
lbtcc
16MC
l^MC
Umc
Umc
IbMC
lowc
lUtj^NCU
1
-
I. St-3
-
4«
4a
4S
4a
4a
4a
4a
CHjuC(U)
i
-
>.*4t-4
-

2U»
2U*
2U*
2Ub
2Ua
2ua
Hett.otychlor
1
-
2.V¥t-0
3.*4fc-4
1704
4 7Ud
2 7U4
**7d
241d
l^ld
i kX3
Uptan
1
-

S.7fc2
lUh
0h
3h
2Ua
im
liMC
iMC
Atratloe
1
i.9E->
7.*E-i
-
*.5h
2.!>h
i.) h
*.>h
l.> ii
I.it
i.Sfi
flalathloo
I
4.at-*
7.7t-V
i.JEU
l.oU
12bd
I4d
l.Sd
J.lh
21a
im
Parathion
i
-
*.St-0
2.3t-2
1704
177a
l7Ud
llttd
2»d
J.*d
«.*n
faraoion
2
-
*.lE-0
1.3C-1
19>d
iWd
l4*d
~7d
ftd
l>h
l.)h
Dlazlnoc
2
2.IE-2
4.3E-0
J.3E-3
32 d
12Jd
17td
l*>d
14d
14d
l.io
Ulaxoioo
2
ft.4E-l
2.8E-7
7 .»E-6
l.i«
9d
23d
204
l* d

it a
ChlopyrlCos
2
-
l.E-7
lE-i
01X1
7¥d
73d
40d
7d
1th
2h
S«vlo
2
-
-
7.7WJ
i.9h

lUd
Id
2.ih
Lte
l.ia
Stvlr
2
-
-
3.4E0
«.Sy
23M
24d
2.4d
5.7h
34a
3.«a
&ay|oa
2
-
-
4.6E-1
48y
4.8y
174d
I7d
1. 7d
4.^h

Pyrolaa
2
-
-
1.1E-2
JOWjr
20Cy
IQj

73d
7.3d
10h
OlMtllan
2
-
-
5.7E-S
3.9*10'y
39.000y
3VU0y
390y
39 y
3.9y
14 la
P-Nlirophenyi-N-»«chyl
2
-
4E-J
3.QE3
Oh
Z.Bh
34a
3.9a
23mc
2.3mc
U.23tec
ca rbaaatc











2 ,4-D,a-butoiyethyletter
2
2.0E-J
2.QE-5
3.02E1
9.6h
9. 5h
8.4h
4,4h
36a
4a
23»cc
Vthoiychlor
2
-
2.8E-8
2.SE-4
28ftd
286d
2B6d
283d
2i2i
liid

UOT
2
-
1.9E-9
9.98-3
I2y
liy
7.6y
lAy
79d
fld
Uh
2,<>-3.aethyLes:er
2
-
"
1.7L1
1. 3yf
47d
4.7d
llhr
l.lhr
b.0alo
41mc
Sott: U1 values for	utt CDmcintl la
• * »V*-'JX«
h • hour
d • d*»
y • yea:
'j«ta Source*: 1. Habcy anti ri 111, JV?o
I. rwrrla, lial
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 addi-
tion 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 com-
pounds 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 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 of groundwater flow, adsorption of atmospheric
carbon dioxide, and other sources of acidity to the deposit are likely
to require periodic addition to 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:
•	Esters,
•	Amides,
•	Carbamates,
•	Phosphoric and Phosphoric Acid Esters, and
•	Certain Alkylating Agents and Pesticides.
Potential application of base-catalyzed hydrolysis for various
classes of compounds is summarized in Table 4-11.
References
1.	Fendler, E.J. and J.H. Fendler, 1970. Micellar Catalysts in
Organic Reactions: Kinetic and Mechanistic Implications.
Advances in Physical Organic Chemistry, 8:271-406.
2.	Harris, J.C. 1982. Rate of Hydrolysis. In: Handbook of
Chemical Property Estimation Methods (Chapter 7). Lyman,
W.J., W.F. Reehl and O.H. Rosenblatt (eds). McGraw Hill,
New York, NY.
3.	Mabey, W,, and T. Mill. 1978. Critical Review of Hydrolysis
and Organic Compounds in Water under Environmental Con-
ditions. J Phys Chem Ref Data, 7(2): 383-415.
4.	Mill, T. 1979. Structure Reactivity Correlations for Environ-
mental Reactions. EPA-560/11-79-012, U.S. Environmental
Protection Agency, Washington, D.C.
5.	Pettyjohn, W.A., J.R.J. Studlick, R.C. Bain, and J.H. Lehr.
1979. A Ground-Water Quality Atlas of the United States.
-------
TABLE 4-11
APPLICABILITY OF BASE CATALYZED HYDROLYSIS
AS A TREATMENT METHOD FOR ORGANIC COMPOUNDS
Class of Compounds
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
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
Application of
Base-Catalyzed
Hydrolysis Indicated
(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.
National Water Well Association, Worthington, OH.
6.	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, D.C.
7.	Ryckman, M.O. 1984. Detoxification of Soils, Water and Bum
Residues from a Major Agricultural Chemical Warehouse fire.
In: Proceedings of the 5th National Conference on Manage-
ment of Uncontrolled Hazardous Waste Sites. HMCRI, Silver
Spring, MD. pp 420-426.
8.	Stumm, W. and J.J. Morgan. 1980. Aquatic Chemistry. Wiley-
Interscience, New York, NY. 583 pp.
9.	Versar, Inc. 1979. Water Related Environmental Fate of 129
Priority Pollutants. Vols I and II. EPA/440-4-029 a and b,
U.S. Environmental Protection Agency, Washington, D.C.
10. 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.
Jl
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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 elec-
trons 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 +H202 = 2HCOOH + H2
HCOOH + H202 = C02 + 2H20
or the oxidation of phenol by ozone:
C6HsOH + 140 j = 6C02 + 3H20 + 1402
(5-1)
(5-2)
(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 perixide, 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 manu-
facturers 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 situa-
tions. 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.
5.1 HYDROGEN PEROXIDE
5.1.1 Properties of Hydrogen Peroxide
Hydrogen peroxide (H202) is 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 covaient
oxygen-oxygen bond. The principal reaction is oxidation, although
some applications involve decomposition, molecular additions,
substitutions and reductions. These reactions of hydrogen peroxide
can be expressed as:
(oxidation)	(5-4)
(decomposition) (5-5)
(molecular addition)(5-6)
Hj°2 + W = WO + H20
2H202
h2o2
Hj02 + RX= ROOH + HX (substitution) (5-7)
H202 + Z = ZH2 + 02	(reduction)	(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 decom-
posed to hydroxyl and perhydroxyl radicals. These are very power-
ful 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:
Fe2++H,0, = Fe3"+OH"+OH (hydroxyl radical) (5-9)
radical)	(5-10)
2 2
w
Fe3++H,0,= Fe2 ~ + H *" + H02 (perhydroxyl
= 2H202 + 02
= yh2o2
Hydrogen peroxide is a moderate strength chemical oxidant com-
pared 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 ppm 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 con-
ditions 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 oxi-
dizing 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, mer-
52
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TABLE 5-1
ORGANIC CHEMICAL CLASSES ABILITY TO REACT WITH HYDROGEN' PEROXIDE
YES NO UNKWOWW
1.	\iiphatic Hydrocarbooa
2.	Aityl Halldee
3.	Lct-.er*
4.	nalogeaated £ther* aad Epoxide*
5.	Alcohol*
fc.	Jiycole, Eposldee
7.	Aldehydea, Ketoaee
8.	Carboxyllc Acid*
9.	Aaldea
10.	Eater*
11.	Nlc rtlea
12.	Aalnee
13.	aio CoipouDdi, Hydras in*
Derivative*
14.	Mtroaulua
1). Tnlole
lb. Sulfidea, Dieulflde*
17.	2-lfoolc Adas, Sulfoxide*
Id.	serueoe and eubatltuted fteoreoe
19.	rUlogeoated Aroaetlc Coapouoda
20.	Aroaatlc Hltro Coapouoda
(Saturated iUimi uacaectlvei unsaturated coapeuuda for* epoxide* end poly
hydroxy coapeuoda).
ftaqulrea U catalyat; foraa acetic acid end CO. .
Nay require re catalyat and alkaline conditions (pfl ¥-11);
foraa organic aclda. Reaction tl^ • alnutea.
Forme aaldee.
Priaary ealaee react to fora hydrosylaalnca, Aso, Aioxy, nltroao aod nitro-
compounds; aecoadary aalaea react to fora dl~N-aub*tltuted (tjNOH) taydroyxl
aalaaa. Inaction tin* * alautee to hour*.
Hay require catalyat.
May require catalyat; aay require low pH, Pe cetalyat or elweted ceapera-
cur* 
-------
SOURCE: HANDBOOK OF OZONE TECHNOLOGY AMD
APPLICATIONS, R. RICE AND A. NETZER. 1M2
FIGURE 5-1
DECOMPOSITION RATES OF OZONE IN VARIOUS WATERS
water only 10% of the ozone is decomposed after 85 minutes (an
extrapolated half-life of about 9 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 sub-
stantial 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):
•	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.
TABLE 5-2
ORGANIC CHEMICAL CLASSES
ABILITY TO REACT TO OZONE
1.	Aliphatic Hydrocarbons
saturated
unsaturated
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
21.	Phenols
22.	Halogenated Phenolic Compounds
23.	Nitrophenolic Compounds
24.	Fused Polycyclic Hydrocarbons
25.	Fused Non-Aromatic Polycyclic
Hydrocarbons
26.	Heterocyclic Nitrogen Compounds
27.	Hetrocyclic Oxygen Compounds
28.	Hetrocyclic Sulfur Compounds
29.	Organophosphorus Compounds
YES \ O UXK.\0 K'Y
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
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 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 poten-
tial 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
•	Color, Taste and Odor Removal (by Oxidation of Organics)
•	Algae Removal
•	Oxidation of Organics (Phenols, Detergents, Pesticides)
•	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).
54
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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
•	Processing wastewaters of mixed ore
•	Oxidation of organic waste streams.
TABLE 5-3
OZONATION OF VARIOUS COMPOUNDS IN WATER

Iniiial
Ozone
Final


Concn.
Dose
Concn.
Percent
CompoundsJ
(mg/l')
(mg/1)
(mg/1)
Reduction
Petroleum
10
4.5
0.2-0.3
97-98
Gasoline
50
1.29
1
98


5.1
0.1
99.8
Benzene
200
20
5
97.5
Diethylbenzene
125-100
150-10
5-12.5
88-96
2,2,4-dinitro-
50
100
0.35
99.3
phenol
3
17
0.05
98.3


14
0
100
DDT
0.5
13.8
0.25
50
Malathion
10
3.5
2
80


9.8
1
90


2.6
0
100
Methyl-
10
4.5
0.5
95
parathion

9.5
0.1
99
Trichloro-
10
8-10
0.07
99.3
methyl
0.5-1
3.5-4.5
0
100
parathion




Dinitro-
10
5-6
0
100
orthocresol
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.
FIGURE 5-2
CITY WATERWORKS, KARLSRUHE, WEST GERMANY

Initial
Final


Concn.
Concn.
Percent
Compound(s)
(mg/l)
(mg/l)
Reduction
Methanol
2000
160
92
Ethanol
1000
90
91
Isoamyl alcohol
1000
80
92
Glycerine
1000
0
100
Hydrazine
100
0
100
Carbon Disulfide
100
0
100
Hydrogen Sulfide
10
0
100
Phenol
100
0
100
o-C resol
100
0
100
Hydroquinone
100
0
100
Salicyclic Acid
100
0
100
Gasoline
1000
0
100
Benzene
500
0
100
Toluene
500
0
100
Xylene
500
0
100
Acetone
100
30
70
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 ground-
water 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 has 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
55
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INFILTRATION
WELLS
FIGURE 5-3
BASIC FLOW DIAGRAM FOR OZONATION OF GROUNDWATER
AT KARLSRUHE
was produced on site from dry air at a concentration of 25 g/m3
(0.0016 lb/ft3) and was introduced to the contaminated ground-
water 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
m3/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 insufficient 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 ozona-
tion. 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.
The application reported in this case history diverges somewhat
from strictly in situ treatment for the following reasons:
•	The groundwater was collected and ozonized conventionally
(above ground) to reduce the organic content.
•	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 con-
centrations of dissolved oxygen. There was thus no in situ chemical
oxidation of organics by the ozone.
S.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 effec-
tiveness of subsurface (in situ) ozonation. The only study of the
effect of soil on oxidation of organics using ozone is an experi-
ment 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.
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 refrac-
tory organics of concern. In addition, it is important to design
sufficient oxidant into the process to accomplish 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 in-
nocuous compounds. An example of this would be the ozonation
of the pesticides parathion and malathion, which produces para-
oxon and maloxon, respectively (USEPA, 1984). These inter-
mediates 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.
56
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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)2) and
sodium hypochlorite (NaOCl). Sodium hypochlorite is usually pro-
duced 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 con-
taining 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 or 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 sub-
sequently 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.
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 com-
mon 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 chlorina-
tion of polynuclear aromatic materials (Liukonnen et al., 1983).
Ghanbari et al. (1983) reported incorporation of chlorine into fatty
acids, fatty acid esters and triglycerides.
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 degrada-
tion 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 treat-
ment (disinfection) of water, wastewater, sewerage and for cleaning
swimming pools. The potential use of hypochlorinates (in aqueous
solution) for in situ treatment of organic wastes is, at best,
extremely limited because the chief products of chlorination are
usually undesirable chlorinated organics (Table 5-4). The greatest
potential use of chlorination for organic waste treatment resides
with phenols and phenolic compounds, where documented
oxidative degradation to aliphatic acids have 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 some specific situations, use
of aqueous solutions of hypochlorites is not generally advisable
for in-situ treatment of organic chemicals due to the possible for-
mation of chlorinated organics, as well as the lack of available in-
formation 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 31,600 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" + CI + H20
3)	2CNO- + 30C1- + H20 = 2COz + N2 + 3CL" +
20H"
A total of 4360 kg (9600 lbs) of lime was spread over the area
first to raise the pH. Then a water solution containing 410 kg (900
lbs) of HTH was sprayed over the area, and an additional 180 kg
(400 lbs) 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).
References
1. Bieber, T.L. and M.L. Trehey. 1983. Dihaloacetonitriles in
Chlorinated Natural Waters. In: Water Chlorination: Environ-
mental Impacts and Health Effects, R.L. Jolley (Ed„), Ann
Arbor Science, Ann Arbor, MI. 4(1): 85-96.
57
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TABLE 5-4
ORGANIC CHEMICAL CLASSES
ABILITY TO REACT WITH HYPOCHLORITES
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, Substituted Benzene
19.	Halogenated Aromatic Compounds
20.	Aromatic Nitro Compounds
21.	Phenols
22.	Halogenated Phenolic Compounds
23.	Nitrophenolic Compounds
24.	Fused Polycyclic Hydrocarbons (PNAs)
25.	Fused Non-Aromatic Polycyclic Hydrocarbons
26.	Hetrocyclic Nitrogen Compounds
27.	Heterocyclic Oxygen Compounds
28.	Heterocyclic Sulfur Compounds
29.	Organophosphorous Compounds
YES
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
\0 USKNOWS
X
X
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 (CHClj)
Chlorinated byproducts possible
Forms chloramines, hydrolysis of C-N bond, possible NCI,
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 sulfones
Forms chlorinated aromatic
Forms chlorinated aromatic, possible oxidation
Forms chlorinated aromatic or chloramine
Forms chlorinated phenols, oxidized to aliphatic acid
Oxidized to aliphatic acid
Chlorination of aromatic ring
Chlorinated and oxidized products (e.g., phenols and
quinolines)
Chlorinated product formed
Chlorinated product formed
Chlorinated product formed
2.	Bower, E.J., M. Rein hard, 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.
3.	Carlson, R.M. and R. Caple. 1978. Organochemical Impli-
cations of Water Chlorination. In: Water Chlorination:
Environmental Impact and Health Effects, R.L. Jolley (Ed.).
Ann Arbor Science, Ann Arbor, MI. 1:65-75.
4.	Dorfman, L.M., and G.E. Adams. 1973. Reactivity of the
Hydroxyl Radical in Aqueous Solution. U.S. Department of
Commerce, National Bureau of Standards, Washington, D.C.
5.	E I DuPont de Nemours and Company, Inc. Hydrogen Per-
oxide, Waste Treatment Handbook. Wilmington, Delaware.
6.	Eisenhauer, H.R. 1964. Oxidation of Phenolic Wastes. Water
Pollution Control Fed. J. 36:1124.
7.	FMC Corporation, Industrial Chemical Group. Industrial
Waste Treatment with Hydrogen Peroxide. FMC Corp.,
Philadelphia, PA.
8.	Ghanbari, H.A., W.B. Wheeler, and J.R. Kirk. 1983.
Reactions of Chlorine and Chlorine Dioxide with Free Fatty
Acids, Fatty Acids 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.
9.	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 Chlorina-
tion: Environmental Impacts and Health Effects. R.L. Jolley
(Ed.), Ann Arbor Science, Ann Arbor, MI. 1:139-159.
10.	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.
11.	Jolley, R.L., G. Jones, W.W. Pitt, and James E. Thompson.
1978. Chlorination of Organics in Cooling Waters and Pro-
cess Effluents. In: Water Chlorination: Environmental Impact
and Health Effects. R.L. Jolley (Ed.), Ann Arbor Science, Ann
Arbor, MI. 1:105-138.
12.	Katz, J. 1980. Ozone and Chlorine Dioxide Technology for
Disinfection of Drinking Water. Pollution Technology Review
No. 67, Noyes Data Corp., Park Ridge, NJ.
13.	Kirk-Othmer. 1979. Encyclopedia of Chemistry and
Technology. 3rd Edition. Vol 5. J. Wiley and Sons, New York,
NY.
14.	Liukonnen, R.J., S. Lin, A.R. Oyler, M.T. Lukasewyz, D.A.
Cox, Z.J. Yu, and R.M. Carlson. 1983. Product Distribution
and Relative Rates of Reaction of Aqueous Chlorine and
Chlorine Dioxide with Polynuclear Aromatic Hydrocarbons.
In: Water Chlorination: Environmental Impacts and Health
Effects. R.L. Jolley (Ed.), Ann Arbor Science, Ann Arbor,
MI. 4(1): 151-0185.
15.	Masschelein, W. 1982. Ozonization Manual for Water and
Wastewater Treatment. John Wiley & Sons, New York, NY.
16.	Morris, J.C. 1978. The Chemistry of Aqueous Chlorine in
Relation to Water Chlorination. In: Water Chlorination:
Environmental Impact and Health Effects, R.L. Jolley (Ed.),
58
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Ann Arbor Science, Ann Arbor, MI. 1:21-35.
17.	Murphy, J. and J. Orr. 1975. Ozone Chemistry and
Technology, a Review of the Literature, 1961-1974. The
Franklin Institute Press, Philadelphia, PA.
18.	Nagel, G. 1982. Ozonizarion for Groundwater Protection at
the City Waterworks of Karlsruhe. In: Ozoaization Manual
for Water and Wastewater Treatment, W. Masschelein (Ed.),
John Wiley & Sons, New York, NY.
19.	Rice, R. and M. Browning, 1981. Ozone Treatment of
Industrial Wastewater. Pollution Technology Review No. 84,
Noyes Data Corp., Park Ridge, NJ.
20.	Rook, J.J. 1980. Possible Pathways for the Formation of
Chlorinated Degradation Products During Chlorination of
Humicacids and Resorcinol. In: Water Chlorination: Environ-
mental Impacts and Health Effects. R.L. Jolley, (Ed.), Ann
Arbor Science, Ann Arbor, MI. 3:85-98.
21.	Schumb, W.C., C.N. Satterfield and R.L. Wentworth.
Hydrogen Peroxide. Reinhold Publishing Co., New York, NY.
22.	Snider, E.H. and F.C. Albey. 1980. Kinetics of Biphenyl
Chlorination in Aqueous Systems in 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.
23.	Stevens, A.A., C.J. Slocum, D.R. Seeger, and G.G. Robeck.
In: Water Chlorination: Environmental Impact and Health
Effects, R.L. Jolley (Ed.), Ann Arbor Science, Ann Arbor,
MI. 1:77-104.
24.	U.S. Environmental Protection Agency. 1979. The Chemistry
of Disinfectants in Water Reactions and Products. NTIS
publication PB 292-776. Washington, DC.
25.	White, G.C. 1978. Current Chlorination and Dechlorination
Practices in the Treatment of Potable Water, Wastewater and
Cooling Water. In: Water Chlorination: Environmental Impact
and Health Effects, R.L. Jolley (Ed.), Ann Arbor Science, Ann
Arbor, MI. 1:1-18.
59
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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 remembered that:
•	The process technologies for in situ treatment have not been
established with confidence. In fact, there is very little documen-
tation 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).
•	The delivery/recovery systems for implementing in situ treat-
ment methodologies can probably be adapted from other existing
applications (i.e., wastewater treatment, ground water collec-
tion and alquifer management, construction site dewatering, sub-
surface injection of waters and grouts, irrigation engineering).
However, 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.
•	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 docu-
ment 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 re-
quired for the evaluation of biodegradation, surfactant-assisted
flushing, hydrolysis and oxidation applications for in situ treat-
ment 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 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 environ-
ment, 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 susceptible 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.
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
60
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TABLE 6-1
POTENTIAL APPLICATIONS OF TREATMENT METHODS TO WASTE CONTAMINANTS
Treitwnt Technology
Chemical
Class
Aliphatic Hydrocarbon*
Alkyl Hal idea
Ethers
Halogensted Ethers and Epoxide*
Alcohols
Clycola/fcpoxldes
Aldehydes, Ketones
Csrboxyllc Acids
A»ld«*
Esters
NicrUcs
Blodegradatl on
|4K-C«talyf
-------
Finally, it must be recognized that the waste may consist of a
mixture of compounds with varying treatability by these methods;
this 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 exten-
sive, 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 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 ground-
water 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 characteriza-
tion 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.
In addition, if the wastes are in solid or particulate form (e.g.,
spent resins), permeabilities of these wastes should also be deter-
mined via in situ methods or in the laboratory as appropriate. These
data are needed for designing the deli very/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 microbe
are available, this evaluation should move on to Step VI; other-
wise the research described in the following step would be needeci
to develop microbes capable of biodegrading the identify
organics.
6.3.1.5	CONDUCTING RESEARCH TO DEVELOP
NEW MICROBIAL STRAINS (STEP V)
Microbes with the required characteristics may be developed j,y
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 utiljZe
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 environ-
ment 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 con-
tamination and subsequent human health effects, and changes in
the local microbial community and soil conditions so that the
reclaimed land may not be capable of sustaining the original vegeta-
tion. 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 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 deter-
mine the maximum biodegradation rate which can be achieved
under simulated in situ conditions. In this simulation, proper
modifications of site conditions to improve the biodegradation rate
should also be considered. These modifications may include:
•	Addition of buffers to adjust pH,
•	Addition of fertilizers to provide adequate nutrients,
•	Addition of emulsifiers to solubilize the wastes,
•	Addition of water to adjust the moisture content, and
•	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:
•	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
62
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the waste organics are in fact biologically degraded, and whether
any other toxic intermediates of 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.
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 renova-
tion project.
6.3.1.7	ONSITE PILOT TEST (STEP VII)
In this step, a representative plot at the waste site (either sur-
face 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 descrip-
tion of pilot-scale testing, including sampling, analysis and moni-
toring 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.
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 pro-
duce 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 com-
mercially 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 sub-
sequent 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.
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 (Kow) of these
organics. If Table 3-2 does not contain the required data, supple-
mental literature surveys or laboratory tests should be conducted
to determine the solubilities and Kow of the identified organic
waste materials. If their solubilities are greater than 5 x 10"2M
and log Kow 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 (deter-
mined 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:
•	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
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/solubilization).
•	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 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) in
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 condi-
tions. In these tests, the optimum range of water or surfactant con-
centrations 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.
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 sur-
factant 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
63
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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 imple-
mentation 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:
•	Identifying the organic compounds in the waste susceptible to
hydrolysis (Step I)
•	Analyzing the effect of variables on hydrolysis rates and com-
pleteness,, and assessing the need for and method of accelerating
the hydrolysis rate (Step II)
•	Bench scale treatability studies (Step III)
•	Design and implementation of a field demonstration program
(Step IV)
•	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 accelera-
tion 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 nonhydrolyzable 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.
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
overally 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, extra-
polating laboratory temperature hydrolysis data to environmen-
tal 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 com-
munication). 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 con-
cern for implementaton 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.
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
FIGURE 6-2
EVALUATION OF FLUSHING AND SURFACTANTS
64
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OCOHTOAOlOGfCAL

CONDITIONS

FIGURE 6-3
EVALUATION OF HYDROLYSIS
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 produces (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 investiga-
tions of hydrolysis in soils have been conducted. There is no prac-
tical 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 pro-
duce 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.
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 decom-
position 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 decom-
position 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 determina-
tion 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 concen-
trations 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 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 treat-
ment 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 out-
side 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 (Sec-
tion 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.
6J
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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, 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 system 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 the Requirements of a Delivery/Recovery
System (Step I)
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 system 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 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.
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 deter-
mination 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
—	Goundwater 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 deter-
mined that the data provided in the RI are insufficient.
•tOPEGWAPMUm
MYQUOlYtH

HmWIMOrtOMACTAWT
DCTCftMMING
RfOVIRCMENTS
Of Of livid*/
mcovin*
APPLICATION
•	SUftFACf
AmiCATlOM
•	1UMUMACC
APPLICATION
FIGURE 6-4
SYSTEMATIC APPROACH TO DELIVERY/RECOVERY SYSTEM SELECTION
66
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6.4.3 Selecting the Deliver)' and Recovery Methods (Steps V)
The selection of the most appropriate deli very/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 sub-
surface geohydrologic characteristics of the materials surrounding
the waste deposit.
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.
1.	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.
2.	Contamination Present at the Surface—
This consideration may eliminate certain gravity based delivery
systems (such as the use of ditches, infiltration galleries and in-
filtration 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.
3.	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 syn-
thetic material.
4.	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.
5.	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.
6.	Hydraulic Conductivity of the Waste Deposit and Sur-
rounding 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"1 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 between 1 x 10"3 cm/sec and 1 x
10~4 cm/sec (2.8 to 0.28 ft/day). For a hydraulic conductivity less
than 1 x 10"4 cm/sec (0.28 ft/day), forced injection assisted by
electro-osmosis used as a recovery method may be the only effective
system.
7. 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 ft).
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 meters (16 feet) with hydraulic
conductivity greater than 1 x 10~3 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"4 cm/sec, although at present the applicability and
effectiveness of electro-osmosis has not been demonstrated. In
general, forced methods should be highly effective for waste with
hydraulic conductivities in the range of 1 x 10"' cm/sec to 10"4
cm/sec (280 to 0.28 ft/day).
6.4.3.2 Recovery Methods
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 suf-
ficiently 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 conseptual 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"3 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 hydraulic conductivity in the range
of 1 x 10~3 cm/sec to 1 x 10"4 cm/sec (2.8 to 0.28 ft/day-
The electro-osmosis method may be considered for low
permeability conditions (below 1 x 10 "4 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.
67
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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 feasi-
ble 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, specifica-
tion preparation, equipment procurement and installation of the
following facilities necessary to apply, distribute and collect treat-
ment solutions.
—	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 descrip-
tion 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 1987 contamination was observed in a small creek
that discharges into Allendale Brook in the Town of Waldwick,
NJ. Biocraft Laboratories, a small synthetic penicillin manufac-
turer is located on a 1.72 ha (4.3 acre) plot near the contaminated
creek within an industrial park in Waldwick. It was determined
that leakage had occurred 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, 1984c). 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:
dimethyl aniline
acetone
11925 kg
4840 kg
(26300 lbs)
(10890 lbs)
6.5.2 Remedial Investigation
The Biocraft site was well characterized both geohy drologically
and chemically. Figure 2-1 (in Section 2) is a site plan for th^
Biocraft site showing the approximate configuration of the con^
taminant plume prior to remediation. Figure 6-5 is the configura-.
tion of the water table at the site before implementation of remev
diation. Six groundwater monitoring wells were installed at the sit«
in January 1976, followed by 22 more wells in June 1976. Thes^
were used both to monitor and selectively pump contaminate*}
water. The basic results of the monitoring and testing program wer^
that:
1.	The contaminant plume (as measured by COD greater thaty
100 mg/liter) roughly followed the plume outline show it*
Figure 2-1 and was approximately 0.71 ha (1.75 acres) in area.
It was estimated that 9175 m3 (12,000 yd3) of soil wer^
contaminated.
2.	The contamination had not penetrated a semi-consolidate*}
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 si*
sampling wells at the site during Jan-June 1976.
Parameter
pH
BOD
COD
TOC
CI"
Concentration Range
5.2 - 7.5
2 - 21000 mg/1
8 - 31000 mg/1
2 - 9625 mg/1
5 - 6246 mg/1
An onsite test well sampled in 1981 (just before the biodegrada-
tion operation began at the site) revealed groundwater concentra-
tions 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
\0\ \ \| /
1*$.
8

"~\ot«r wtu -
\	143 *'
ISOURCI: JHAVERI AND M«Z»OC*. 1W3I
methylene chloride
n-butyl alcohol
82310 kg
30305 kg
(181500 lbs)
(66825 lbs)
FIGURE 6-5
GROUNDWATER SURFACE CONTOURS, BIOCRAFT SITE
68
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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 con-
ductivity of the glacial till layer ranges from 1.7 x 10"3 to 9.4 x
10"7 cm/sec (0.02 to 36 gallons/day per ft2). 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 Mazzacca (1983) for a complete description of the ex-
tent of contamination at the site, site characteristics (soil proper-
ties, 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 biodegrada-
tion 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 (GDS), which holds a patent on the system ultimately in-
stalled at the site, required approximately 2Vi years to proceed
through the research and development stage of the biodegrada-
tion 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:
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 sub-
surface 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), con-
taminated groundwater was inoculated with soil samples taken from
the Biocraft site itself as well as uncontaminated (control) samples
from the homes of various employees. Completion of this step in-
dicated 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 II of Figure 6-2. A shaker flask study using con-
taminated 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 effec-
tive 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 feasiblity 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
4)	Pilot plant studies conductedin 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 date 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 com-
pares 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 1 x 10"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 recir-
culating in situ and above ground (bioreactor) treatment system
illustrated in Figure 2-2. The operation of this system would pro-
vide both plume containment and 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" (infiltra-
tion 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 1/day (6850 gallons/day) per trench. Each trench
also have two monitoring wells (one at each end) which can also
be used to flush the trenches if sludge accumulates.
The recovery system, located approximately, 90 m (300 ft)
69
-------
northwest of the infiltration trenches, consists of a buried trench
and slotted pipe collection system (Figure 2-2). The trench is about
24 m 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 ground-
water flow from the groundwater mound (USEPA, 1984c; Jhaveri
and Mazzacca, 1983).
The in situ aeration system (Figure 2-2) consists of nine aera-
tion 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).
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 in-
terceptor 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 dif-
fuses at a rate of 0.8 m3 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°7o of the total
biodegradation of the contaminants takes place in the surface reac-
tors, 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 ter-
minated in 1986.
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 deci-
sions which have to be made, during the evaluation of in situ treat-
ment 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 2Vi years to complete. While
the percentage of the cost attributable to the feasibility study and
its duration would undoubtedly be reduced in future applications
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 R1 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 $ 50000
Total Feasibility Study Costs
3.	Implementation Costs
a.	Biostimulation Plant Design and Construction
i)	Engineering Design
—	Biocraft in-house (360 hrs @ S50/hr)
—	Engineering, Drafting
Total Biostimulation Plant Design
ii)	Masonry and Construction
iii) Equipment
Total Biostimulation Plant Costs
b.	Delivery/Rrecovery System Design and
Construction
i)	Design
—	Laboratory Testing
Labor (consultants $24673;
Biocraft in-house $26400)
Total Delivery/Recovery Design
ii)	Installation (All Contractor Costs)
—	Air and monitoring well points
—	Trenches, air well construction and site
work
—	Supervising Geohydrologist
—	Engineering
Total Delivery/Recovery Installation
Total Delivery/Recovery
Systems Costs
4. Operating Costs ($ per day)
a.	Utilities
—	Electricity (26.4 Kw, 24 hrs/day)
—	Steam (71 lbs/day @ 90 PSI)
Total Utilities
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 = SO.0165/gallon
Expenditltre
S 6874
S 2770]4
S_3937g
S 7394*
S 296280
S 40000
S 60Q0n
S 1000(X)
S 446280
18000
40400
S 58400
S 73975
S 88833
S 221207
S 10418
$ MOT*
S 61401
S 12740
S 80500
$ 21513
S 800Q
LJ22753
$ 184244
46.82
.58
S	47.40
S	24.40
S	97.10
S	20.26
S	17.14
S	158.90
S	19.20
S	225.50
70
-------
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 discus-
sion illustrates that application of an in situ treatment system is
likely to involve a significant research and development compo-
nent which will be site specific, possibly requiring a lengthy develop-
ment period and possibly involving expenditures which will be a
large fraction of the total capital cost.
REFERENCES
1.	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. En-
vironmental Protection Agency, Cincinnati, OH.
2.	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.
3.	Flathman, P. E., J. R. Quince and L. S. Bottomely. 1984.
Biological Treatment Ethylene Glycol-Contaminated Ground-
water at Naval Air Engineering Center, Lakehurst, NJ. In:
Proc. 4th Nat. Symp. on Aquifer Restoration and Ground-
water Monitoring. Nat. Water Well Assoc., Worthington, OH,
pp. 111-119.
4.	Flathman, P. E. and J. A. Caplan. 1985. Biological Cleanup
of Chemical Spills. In: Proceedings of Hazmacon 85. Assoc.
of Bay Area Governments, Oakland, CA., pp. 323-346.
5.	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.
6.	Little, A. D., 1983. Handbook for Evaluating Remedial Action
Technology Plans. EPA-600/2-83-076. Municipal Environ-
mental Research Laboratory, U.S. Environmental Protection
Agency, Cincinnati, OH.
7.	Repa, E. and C. Kufs, 1985. Leachate Plume Management.
Draft Report for Hazardous Waste Engineering Research
Laboratory, U.S. Environmental Protection Agency, Cincin-
nati, OH.
8.	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 Spring, MD. pp. 420-426.
9.	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.
10.	USEPA. 1984b. Caste Studies 1-23; Remedial Responses at
Hazardous Waste Sites. EPA-540/2-84-002b. MERL, U.S.
Environmental Protection Agency, Cincinnati, OH.
11.	USEPA. 1985. National Oil and Hazardous Substances Pollu-
tion Contingency Plan. Federal Register, 50(29):5862 Feb 12,
1985.
12.	Wetzel, R. S., S. M. Henry, P. A. Spooney, S. C. James and
E. Heyse. 1985. In Situ Treatment of Contaminated Ground-
water and Soils, Kelly Air Force Base, Texas. In: Land Disposal
of Hazardous Waste: Proceedings of the 11th Ann. Research
Symp. EPA/600/9-85/013, HWERL, U.S. Environmental
Protection Agency, Cincinnati, OH.
71
-------
APPENDIX A
BIOLOGICAL DEGRADATION OF
ORGANIC MATERIALS
-------
APPENDIX A
EVALUATION OF SYSTEMS TO ACCELERATE STABILIZATION OF WASTE PILES
OR DEPOSITS, BIOLOGICAL DEGRADATION OF ORGANIC MATERIALS
II010CICAL
ACTIVITY
IIOLOCICAL
ACPfT(S)
—snmr	
ENVIRONMENTAL
MQUlKEWOffS
-—TSTtTaI
S.UK»t*ATfc
LQNlbMTRAUUW
CONTACT
HKfc
Acyrllr hydrocarbon*
alkjnee
•r thane
(Cj to C44)
*-i)odfciw (C-I?)
•HlrldrciM (C-1J)
N-hriidrcinf (C-It)
N-itodfc*M ~
N-h#ied#rane
H-t Mdecsne ~
N-heaadrcene
N-octederene (C-10)
Oilrfatloa-
OlldMM end
dchydrogeneeee
Cralditloot while
irons oa aathoae
Coasldatloaa while
grown oa atthiM
Oildatloo of propaoa
Cooildatlooa wfctlo
(row on wthist
Coosldetiooe while
grow on ecetete
CO], N2O
NeMUtilaal oilditloei to prlury
alcohol*; to aldehyde*; and Co
•onocarboaylfc acid*
Acetic i(14, ethaool, teatlldihydt,
COj» call aaterlal aod
estracellular cooetltueoca
froplNlC OCld, ptopeool, UttODt.
CO}, coll aaterlal a ad
ntrac«ll«lir ceutltMiti
lutyrlc acid* 1-kutnol 2-butaoooe
1.12	dodecnnldoic *cld
1.13	trldecenldolc acid
1,1ft heiadecanldotc *cid
1,12 dodecanidofc acid ~
1,16 htndecMldolc Mid
13 trldecenldolc Mid ~
1,1* hesadecanldolc acid
Eatvra produced via
eoosldetton* of *n
n-AJkane eubetreta to
a hoMlo|Aui oiygonetrd
coopound without degradation
aslt-ai +ch.-
CO0CN2-(dl2>u-C^ (1T1 Slltur*)
Nethylotroplc bacteria
In *oll*
Mlcroorgealaaa
aerobic, landfllla, natural |*«
laak*
PnudoacMi oethanica	aerobic, growth oa Mthaa*
PituJoacMi oethanlca	aerobic, growth on aithatt
Mycobacteria vaccae	aerobic
Paeudoaoaaa —thaiIf	aerobic, growth oa Mtl«M
Candida cloacae
aerobic, growth on acetate
Micrococci cerlfecana aerobic, growth on aikane
40-Su* ~ ICt v/v
St ethane, 4)1 ne thane,
SOS air
301 propane, 401 •ethane,
301 air
51 V/V
>1 V/V
» V/V
n v/v
v/v
Saturated aolla
b. »1, 7S,
IQU. 114
79
79
t>*
3, b, 11)
N-ho«4«ciM (C-U)
R-tatraderane (C-U)
a-Oodec*oe (C-12)
g*aollne
OefOet ton, cooafdatfaa
beta, aooga, alpha
aaldarlon of aaao and
dlcerboeyllc *clda,
ring cleavage 1*1lowed
by oaidetlon, ate., of
acyclic Mncioie
Iwbt oil
12 I utl oil
f) Iwe I ell
heavy mIr dleeel
crude peiraleuo coadoaaato
linear poreffle*
ocn2-(CN2) u-cn,
C«)-(CM2>i«~GDOCM3-CCIl2>|2"cai3
Car(CM2)t«-00O-I (unknown)
CO}, MjO hydroiyperoildea. alcohole
eldehytee, WtoaM and eatera
COj, MjO, alcohol*, rarboayla,
aldehyde*, b*ton«* and eater*
Soli bacteria
foil bacteria
Salt wecar bacteria
Surfer* and aubeurfaca eoll and
groundwater cooioalnatloa, MjO}
Injection, aerobic and
anaerobic rveplratloo,
feraootetIon*
•eroblc and anaerobic
reeplration, feraontatloo
Oil aplll oa beech aaa t*ap*r*~
lure about IU*C air teoperature
ll*C aand epreyed with aaa water
(wice dally, aerobic* nitrogen
ead phoapfceroue nutrlanta
lb day*
j, t, bo, bn
11, 111, U)
114
114
114
114
114
114
• leltrenre* Hated following Sectloa 2 of Chla report
-------
BUBSTMTf
•IOLOCKAI
ACTIVITT
WOWICTI
Iniirkid 4UMII
iMfllMll llkaM«
pt IflMf
•aibyl braarhla§
leubutyl Uokvtyttl*
• UfMI
OMfl MiilltN
ta»|« •* ilfhi iilO*
I Ian; tr bat* «Hyt
|ro«y ruHil
Mcirtoiillt aclda
Mm* m4 4l«rloi)IU itUi
TkvImI tiHiilou Cirkeiyllt kU«, wnaatwri
ivkuralMl dehydroB*" fcydraij tcU «i4 il(okat
miIoa, i^lliilai «(
cirbu cMli
Cooaldatloa to corraa- 1,1'tpoiireciiM
aposlda
I hriidfcrfM
falsi(oleic t(14
1. No((Uit«f
4 I I•ubwifl*n*
aiuratad alhyl hill4«i
browchlorowthini
broaodIrhloroatthan*
chlorinate-d hydrocarbon!
carbon tetrerhlorlde
rhloroecettr acid
1. J-• paiyki m4icin
tpoaldalea Intaraal
double bo«d
CaaiHitioa
4ltp«i)rKiiat
ntfiKtlvi dechlorination
Partial ck«a. trmifor*
ait|«a (Br) ri<«ci|vt
4«chl«rlutlM
iathlarlM*
ttaa partial ekta.
Uaaiforaatlai (lr)
ri4«ctl«« iadilorlM-
(lot, au«raklt and
••rakle naflr«tlM
reductive ItchlorlM*
(loa,.auiraklc
raaplratloa e«oitdttl«n
Dihaleitnitiea,
oildation
CO] CM*. tt+. »r" and CI"
CO], CH4, K*. cr
COj CM4, U* and CI"
rl>loro«cetete
rh|nrob»tyrlc odd
Dehalogenat lM,
oildatton lata
Dehalofenetloo
oaldattoa
COj, cr
chlor«(»r»
2~cMoroproplonlc acid
4lbru«B(hloroMthaM
dlbroaoaathaaa
IlkfMMtktM
4trMorurihan*
reductive dachlorlna-
I Ion
tlon, anaerobic raaplra*
Ilea cooaldatton
Daha 1ogenat100•,
oildatlon
tlon anaerobic rea-
pfratlon, partial
ckaa. traMfanattoa
(IT) CMllteliM
Dakal*|Mitlea
Partial chM. traae-
r«n«ilM (Br),
anaerobic raaplrattoo
coasJdatloa
teductlve iatMoTlM-
tton, iMtrabtc rea-
plrattaa CMiU«iia«
CO}, C*4, M* and CI"
ItOUlilCAl

IPKCIAL
KMVIkUIMKHTAL
ttomwmw
liTm'AJ.
SUISTlUTt
gmcptTHATlmi
contact
tim
fcmJUMCU
toll htcutli
loll WclttU
MflllC
Ultklt
Pevudnaonaa ulfovurena
fWwdaaaMi airailaoM
i to 10
Parcaat (V/V)
Candida 1Ipolytlea.
Carina>act
lactllw ¦agatarlw
atraklt, M*C
Paeadwoaaa altawrtaa
Hwtaat bacteria
MtUm bactvrU
aarottr two pfcaae «ltb
(•llMUtlM rftltiMUM
tut aolor
cuavaraloa
114
114
Soil becterlua, oathano- • anaerobic, continuous flow	37 yf/1	2 day*	*7
B«nlc aliad cultura with fliad flla reactor
acatata aa eubatrata In anaerobic, contlnuoua How	M «|/1	2 day*	U.
imB« ilud|«	flaed Ml* reactor
It, W
Bacteria, i1*CHEH C«C-1,
•1-CHEM K>»-6 and
II-CM1N HP-7
anaerobic, aerobic
Mtliino|tnlt alsad cul-
tura with acetate aa
substrate
anaerobic, contlnuoua flo
flied flla raectore,
denltrlfIcatlon
Sewage baccarla
Psaudoaonaa ap.
aaroblc, H'C
Mathanoganlc elietf cul-
ture with acatata aa
eubetrat*
anaerobic, contlnuoua flo
filed flla reactor
Hfthanoganlc alied cul-	anaerobic, contlnuoua flo
tvre with acatata aa	fliad flla reactor,
eubatrata, coll bactarla,	deeltrlfIcatlon
eowaga sludge
>4 ug/1
lb daya
2. 11. 14,
2b.
iV. n
1J, 14
47 ug/al
*01 lu 2 &aya
10 daya
iO ug/mJ-40 ug/1
#3 ug/1
X) daya
351. 3 daya
2 daya
]), 1), 71
44
13, 14, IS
•all bacteria	Aaeereblc
Anaerobic, contlnuoua flow	j$ u »/i	j daya
ftaed flla rest tor.
danltrlfIcatlon
Soil battttli, •tth*no-
genlc alied culture with Anaerobic, contlnuoua flow
acatata aa ewbatrate	fliad flla reactor
-------
suismr*
BIOLOGICAL
ACTIVITY
WOPUCTt
dlchlofop-uethyl
••thy) chloride

proptrhl or
irU'(>lvrorth«nt
Seductive deeMorlne-
(Ion 4«e«rb0l)rl4lcMerMtli)fliM
1,7-rf (rMorofthjr|r. >'
1J, l>. if
enoilc condltlone
IJ, >7
iJ
Sewage aludge
gl-CNCM CEC-l
Sewaga aludge
Bl-CNEM GtC-)
Sewage aludge
Sewaga aludge
Aerobic	(f)
Aerobic	(?)
Aerobic	(?)
Aerobic (f)
>7. >1
n, w
iJ
>7
bacteria
PHENOIAC
Nethanogenlc cultura
Hethaoogealc culture
aeration baa la
Anaerobic
b week*
I weebe
114
U, 13. >7
1J, U, w
it
Soil bacteria
aaoslc conditions
114
>7
Sewaga aludge
facudoaooada ip,
Aerobic
Aaroblc
ActIvated aludge alero- Aerobic, JO'C; Aftaarablc, V*C MKl N>
orgialaai, dlgeater
¦Icraarpalaaa
Salt requiring becterlua Aerobic, JO'C	ttf.S g/l
T-H, ActlobatUf sp.,
CluconoWltr
AreaitobacTcr «p,
ActtviiH aludge alcro- Aarollc, )0*Cj Mwraklt, )'#C 1®® Pf*
•rgiilni, ilmttr
•IcfWffAlMI
3b (mure
lii
Hi
W
-------
IIOUXtCAl
activity
glyrprol
Pyruvitt to CO}, MjO,
fwlyrtliyt«f»* glycol
Alcohols
Mthinol
Polysthylcn* glycol
4«h)iirft|«nM(
C02, HjO
arcanol
liobulyl alcohol
lao-noayl alcohol
2-tthyl htKsnol
dacyl ilcohol
N-propyl alcohol
tori butyl Alcohol
•ctyl alcohol
Isopropyl alcohol
polyvinyl alcohol
Aldvhydea ««4 Katonas
Osldstlon
OiUitteo
Oildstto*
OtlMotion
Oaldatlon
OlKltlOD
OiUatlM
taldattoa
Oaldatloa
Oaldatloa
PVA oi(4«*« di|ridatl«n by
eucceostva naetori
asthyl alhyl batons
i COj
attliy 1-1 anbury l-k«tono
Carboayllc Aclda
m-etlf kI4
aimorarboiyl l( aclda to
N-CU
Short 1 loft* rk«ln sclda
J4"C|M carboyllc aclda
dUrfrkmyllr irlda
brinrhtd dlcsrbosyllc
ar Ids
p«iilinul( itld
hcsanolr ir14
h*pf*nolc irld
orianlc l(ld
mmanolr «r Id
propanediolt arid
butonodulc Kid
?-ae(bylbu(anldlolc acid
P*ntan#d!olc acid
hr>an*dfolc irld
iwpimdlolc acid
ori«n*4|o|c acid
««MKfdlaU acid
dfrinMloK Kid
•nrrfl ol c acid
cfId#ran#diole arid
l*tr
raCNOIAC	75
raeitoiAC	7)
PHKNOIAC	/J
PNCMOIAC	jj
PNEN01AC
PHCHOMC	j)
PawdosoMi put Ids	miA tectorial ayablotic relation* 93
Peoudoaonae sp. VMlSC	ship, a«roblc( 10*C
Spray and leschste reclrcule* 1*00 i
tloa, aoroblc, Uorsactor
Aerobic, digester, Isnd
feraiag
Aoroblc, digestion, land
plsnnlng
22 days
1 day
1 day
7>
75
UctarU
tecterle,
Paeudoaonads «p.
Irevebacterlua ap.
Flsvobsctrrlua ap.
Mycobacteria isccm
riwdOSOMI sp.
toll bactarla aa
lnnoculua
Aerobic
Aaroblc, snssroblc
Degradation of organic sclds
In ahalo oil retort water,
aerobic. M'C
1UU
*. w
7
>4
If
V
V
Begrodstl
shalo all
24 #C
ratort watsr, aoroblc,
JO
9
PP«
aa
C

lb
lb

py*
aa
C
reduction la


1
PP«
M
C
21 days
•b

2
ppa
aa
c

at

4

aa
c
•0-402
it

9
PP*
aa
c
reduction la
i»

t
PP«
aa
c
9 days
u


ppa
aa
c

M


PP»
sa
c

•b

«
PP*
aa
c

lb

1
PP«
aa
c

M
20

PI"
aa
c

H
-------
SHRKTlATI
BIOLOGICAL
ACTIVITY
fOPUCTS
¦Ml.yl «r*l4tc
vlhyl icr|«(i
• ¦yl acacat*
diethyl adlpata
'ik«alc aaCafa
iriMtitfiMiM
dlt tltyl«a|n»
OflhlMl l«ln«

C02
CO,
CO,
a-phthallc acid
e-pMliallc acid
p-phthallr acid
dlbwtyl phthatata
aonobutyl pktKiliti
urtphiliiltt acid
dl-tfkul)rlphthiU(<
dl-(2-athylhasyl)
pfctbalat*
n-d|bwtyl phlhiUt*
dl-n-octyl phthalata
dt**thyl phthalata
capabla oI aarabolslitf
ochar phthalata aataea
aa wall
dftthyi pfttlialata
d I butyl phthalata
t>u(ylg|yroryl
butylphthalai*
butylbanyl phihilili
dllsobutyl pkOialill
dl'lieniuiiflptiililili
Ml (ruMt(n«i
dla*thylAltroaa«lna
Sacoad ordar alcroblal
degradation rata eonatanta
Nontusorlgantc product
f-nl (roaedlathawlaalfta MJaaralltad
(Mbt U)
Met J4«AtJflad but iu||Mtid re ba
aon~c*rcaaofaale dlaara of NDCU
Thiol*
tlhinriklol
brnitM thiol
Cyrllc AUanti
cy< loaltanaa
llod«|n
HlMialliitlM
CO,
ryrluh*ian«
ryrlohaaaaa
K-hfUM
taldatloa	COj, NjO
Cooildatioaa,	CyrlohfiiiiMi, Nliklild to
rfliMMilln	CO} ~ MjO
CooildailMi,	CycIohfUMl ua#d by mcm4 or|«at*a
(OMMiiiin « io«fci •( («rb«i and turiy
SPECIAL	INiT 1*1.
• lOLOClCAl C NV1MOMCNTA L SUfcMkAit LUNiAU
ACEMT(S)	HEQtlHEHtlfTS	IWHWmATlU*	lIHfc	t*
•t-CMEM CCC-1
114
11*
114
114
114
Paaudoaowaa ip.	Aivnlc c«lt«r« IncukatH	11
hmJoaoniM ap.	aarobic, at 2VC, la I ha dark ?10 Uf/1	JO* la 4 da/a 11
PMW7, VI
113
14
Hind alcroblal
population
Hycobdctaflui vitra*.
Aarobic, farMaiatlon
Crowth oa praptaa tube
7V
Two fiawdoaonaa pp.
-------
SIHSTBATt
BIOLOGICAL
ACTIVITY
WOOUCTS
cyrloperaf f tna
cyclopropane
cyclopentane
••thyI eyelopentane
2-cyclopentene
cyclohetane
cycloacptane
cyclooctane
n-butyl cycloheianol
¦ltrllea. cyaaldae
acrylonltrlle
Cooxldatlone,
COtMUl 1 lu
hIth other toil
•Icroorganlaaa
utilise products ••
sources of carbon and
energy
Propaldahyde
Cyclopantone
Methyl cyclopantone
Cyclopentanone
Cycloheiienone
Cycloalptanon*
Cyclooctanone
Cyclohesaneacattc eel
acviontirll«
•rrylonltrllc
Degradation
Degradation
•rfl |hhi 11 r 11 v
pheoylnltr1Ir
phtha fonttfIle
hydrocyanic acid
Degradation
Degradation
Degradatlea
Degradation
d irhloropttrnyl Isocyanate Degradation
iy.«Md*a, cy*nat*s
alnti
im) lor
dlethylanalIne
o-cMoroenal Ine
p-chloroansl Ine
4-chloroen«llne
t r 1 rhl eroeaallna
Degradation
DachlorlnatIon
Dagradatlea
alkyl bensene aulfoftate Cooaldatlon In the
presence of glucose
followed by coapleta
aatabollaa of and
products
Heta (estradiol)
pathway * hydration or
hydrogenatian followed
by nonoildatIva ring
fission
growth aubatrstea
Ortho (Intradlol) path- CO]
way aoao A dlosyganaae
C02
nm—
SUBSTlUtU
COWCPfTlATlUH
BIOLOGICAL
ACtHTQ)
SKC1AI
ENVIRONMENTAL
KEQUUDgWTS
CONTACT
TUtt
other istla Icroorganlaaa
Worardla ap.
Crovth on n-alkan«a
79
79
79
79
19
79
79
79
Mlsad culture of ysaet	Aerobic
•old, prototoa bactprla;
act lvatad tludg*
fetant alcroorganlsa	20*C
Mutant alrToorKan1»a	20'C
Mutant alcroorganlaa	blotraat
Mutant alcroorganlaa	20*C
HitMt alcroorganlaa	10'C
Mutant ¦Icroorganiia	20'C
Mutant alcroorganlsa	20*C
Mutant alcroorganisa	20aC
II-CHPJ1 CS~ 1-9
>00 *«/l
100 ppv-1000
500 ag/1
500 mil
250 ag/1
250 ag/1
500 ag/1
10UZ Id J lira
«i. n, 114
100X In 3 ».rw 41, 75, 114
1 ppa, 3 aonthe bl. 75, 114
100Z Id 10 hra
1001 La lb bra
lOOt la 1 hra
lOOt In 1 hra
lOOt io • hra
~1. 75, 114
*1, '5, 114
*1, 75, 114
61. 75, 114
*1, 75, 114
Hwtaac Aerobacter ap. >0u ^/1 lOOt la 10 hra bl,	114
114
Mutant Aerobacter sp. ^/| joyj j0 jo hra bl,	?>, 114
Ibitant Aerobacter ap. yg ag/1 1001 to lb hf« bl,	7). 114
Sewage alcroflora ««
Mutant Aarabacter ap. ^ 1W1 u x hr, M§	jj4
b
Faeudoapnas ap.	Aerobic, pur* culture	1001 in 20 daya A3
Bacteria	Aerobic, aolacular Oj la	Ji
eaaentlal
Bacteria	Anaerobic	jj
Paaudoaonaa put Ida.	Aerobic
Sewage aludge	Aerobic
>7
Stabilisation poad	Aerobic (T)	..
•Icrobaa
-------
SltgsTMTT
SfOLOCtCAL
ACTIVITY
FtOOUCTS
PhotwtUb«ltia -	Platlat*
reduction by a novel
pathway leading to
rl«| fleeloe
Meadelece pathway	Succinate and aeetyl-CgA
by citric acid cycle to COj 4 water
Ntaerellaatloe
pate kcMMlt
pathway
Oaldetlea
Wncolc acid
dI vinyl benteae
liydrAiy benanate
p-hydrnay beaceate
p-bydroay benaoete
2,4 dlhydrosy benaoate
Microbial
degradetloa of oil
•kilt retort water
Aerobic aatahollea
4,$-*sygeeeee pathw
beta-kctoedlpete
pathwe f
rfcomMiM let lea
CO ,
Ca—a carbeayl-alpfce-hydraiy
MCMlt (MlalMyda a ad aalaajweat
break dam Involving pyrwvate
Succinate and ecetyl-Ca A
oetabellae ta COj aad WjO la
citric wU cycla
difcydroay boaaaate
J aethyl benceeta
3-oaeadlpete pathway
defeydresybeaeeete
debydrofeoeae
Mtyl foraate
c la-c lf-a«caaata
A-bytylbencena
ethylbenivne
ft*propylbeaten*
)~aethytauconlc add
p-t aopropy I toluene
pyrulidvna
toluene
Maadelate patfewey
Oitdetloa
aucoaa te-cyc loi eeoereec
Ceealdetloe
Cooiidatioa
Cooildatloa
IkKMItttycUIWMIIM
Cooaldatlee
beta-ketoadlpate parkway ta
acetyl~CoA and aucclaete
(Mtiklltn) tf tbe citric ul4 cycle
Pbeaylacetlc acid
Vfeeaylecetlc acid
Claaealc acid
p-leopropylbeasoate
Cletaalc add
o-aylene
p-«ylene
Coos Idatlea
Ceoildat loa
e-tolulc acid
p-tolulc acldlllHtliydroir
p-tolulc acid
tri-p-creeyl phoepfcate Ovldetloa
CO,
SPECIAL	INITIAI.
IIOtOClCAL ENV I ROWffNTA I	SUlSTMTt CONTACT
ACgHT(S)	KQUTKEHENTS	CUWCBITHATION	TlWt	WtKLMKMCES
>hotfopeeudoaonaa Anaerobic, photoeynthetic	20 u oolea Hi alnutee	29
paluatrla	condltlona
Faeudoooaaa putlda	Aerobic	7;
Freeh water bacteria.	Aerobic	w w	u „ ,	M
aewage	*
Feeudo^onae ap. Ill	Aerobic
Tranaconjwfant
Alcalljenoe ap. etraln
A 7-2 treaefer ef
halocatecfcol-degradlng
capacity froa
FaeudoaBaae ap. ID
Aerobic, lew aeneltlvlty to	91
pheaola,	II'C oa eheker culture
Nlied population af aoll Aerobic. phoapfcate euppleaeot 21 m* M cirhM	90S la 9 daya M
bacteria
•1-CMBf CtC-1	yt
thodaaaaudaaanaa ap. Aerobic	29
Peeudoeonaa (Jidda Aerobic	jj
irawtb « baa^te „ „ ^	2^0 nlaute. W
faendeadMi ap. II)	Aerobic
PaoMdoaonae ap., a wage Aerobic, plaaaid tranafer lad
ta tba ability to wee thla
coo pound aa a aala energy aad
carboa aoarca
>ae%*daoenea putlda	Aerobic
Khodopeeudooonae
paluatrla
Paeudooooaa ap.
Wocardla ap.
Wocardla ap.
Wocardla ap.
Fiewd— ow ep. 913
Wocardla op.
lac111ua ap.
iaclllue ap.
Feoudoooaea putlda
Wocardla ap.
Wocardla ap.
Aerobic, growth 00 p-hydroay
benioata
Aerobic
Crowtb on N-alkaaee
Growth on N-alkaaea
Crowtb oa M-alkaaea
Aerobic
Crowth oa W-alkeoae
Aaoarablc
Aerobic
29
90
A3
43
43
90
43
43
S7
4)
43
Activated return elwdge Aerobic, J1*C
1 g/al
70-901. 24 hra
-------
—rermr—
SlIISTMU
cowcnniATioii
•lOlOCtttl
ACTIVITY
IJOUCICAl
ACtMT(I)
—mem—
CNVIKONNCKTAI
MOmtEWEWTS
COtfTACT
TIKE
itramCB
Arooatlca
alpha cMorotoluona
chlorobontano
1-rhlurobantoaCt
p-; o-thloroboni
1,7-; 7,1-; 11 *•
dlrhloroboBt«na
1,4-;),5-
dlcfclorobootoata
BocbUrlMtto*
CO,
6»cU«rlMtU«
Dtchlorlaatloa, growth
D»(kl»rlMil«>, growth
•Icroflora
Nocorotrophlc kicttrli Aerobic, 2?'C to 4*C
Ntudwomi sp, 11)
AlciHttMi ap, AM
Nta*, 9/,
114
4V
-------
SIOLOCICAL
ACTIVITY
PRODUCTS
AruMilc Nitrogen
cuapounds
4~aalnobencene
hr**hydro-1,3,5-
trlni r ro-
ll 3,5-1 rial In*
Ring cltvifi, poiilbU
alnerellsatton
R1odegredetIon,
luccMlivi reduction of
ntcro group* to « point
where tfiitibillntion
and fragmentation of
the ring occur*.
Hon cyclic degradation
products aria* via
aubeeqoent reduction
and reerrengeaent of
reaction product*.
Meaehydro-I-nitroaoOtS-dinltro-
l,),}-irl«iine, he*ahydr&-1,3-
dlnlt roeo-5-nttro-l, 3,5-trlnttroeo-
l, 3,5-tries (ne .hydraslne, 1,1-
dlaethyl-hydrealne,l ,2,-dlaet hyl hyd-
razine. foraeldehyde, and aethanot
King clevege, poaalble Analine
alneraliiatlon
)- and 4-
nItrobcntolr acid
I. 2- and 1,)-
dlni t robentenes
•Inobensotc acid
1,\-dJ«l trohemolc acid
J,*~dlnitrotoluene
Aalnonltrobanaolc acid
2 ,* ,6-trlnltrotoluene
Aroull( Nitre coapounde wltb
Other Nnrtleul Croupe
Wli I uritbrnt un I trie#
t-thlofir ),i-
dlnltrobeniolc acid
4-rhIoro- ?,5-
dlnltrohensolc acid
3-and	4-nltrotoluene*
2,4-dInltrotoluene
4-tolutdtne
Phenol*
caaphor
Mineralisation
Mineralisation
Mineralisation
Dechlorination
Dehalogenatlon,
caaetibollia
Dehalogeutloo
Degradation
Monoosygeneaee
oildation
2-eaino-4,6-dinltrotoluene;
4 aalno-?,b-dlnttrotnlutne;
2,4-dlaalno*6-nltrotoluene;
2 ,HI«*ln(fWI trotoluene;
2,2,6,fc-tetranitro -4,4
eaosytoluene; 2.*,4,fc-
tet fan It ro-2,k-aioiy toluene
Alphe-hydrosyauconlc
eealaldehyde atnerellsed by
Straptoayces ap. „
Toluldlne
Aalnonltrotoluene
J,4,4; t rlaethy l-5-cerbo*ys*ethyl
delte 2-cyclopentone aetebollsed to
Isobutyreta, converted to leobutyl-
CoA which la Mtabollted «U valine
cetabollaa
CooxldatIon
hydrosylatlon
Clnerolone
ItOLOCICAL
ACENT(S)
SPECIAL
ENVIRONMENTAL
REQUIREMENTS
—initial—
SUBSTRATE
CONCENTRATION
CONTACT
TIME
Anaerobic, denltrlflcation
500-100 ug/»I
4 deye
Stabilisation pond
alcrobea
Sewage
Aerobic <1)
Aerobic
57
AO
blc
Sewage
Stabilisation pond
alcrobea
Thenaophl He
alcroorganleu
Anaerobic or aerobic
Anaerobic or earoblc
Aerobic (T)
Aerobic, ceapoetlng ayateaa,
ii'C, 60Z aolstufe
40
40
i7
VI day coapoetlng 51
Lake water bactarla
Lake water, eewage
Lake water or eewage
Sewage alcroflora
Chlaaydoaonaa ap. Al and In the light In the absence of
A2, imn|* ileroflore RutrUnti and In the dark with
acetate
Sewage
Sewage
Sewage
Anaerobic
Aerobic, anaerobic
Aerobic
I v g/ml
250 Rg/ftl or laaa
2S4 pg/al
500 ag/al'Vl.r"1 8V
590 pg/al^/l.r"1 IU4
10 houra	1U5
4)
4)
45
40
40
Pseadoaones put Ida
Aerobic, lnducable ansyae
¦yate*a
Aaperglllua nlger
-------
SUiSTiATK
BIOLOGICAL
ACTIVITY
ffOOOCTI
dlphaay) I
Nloorallaad
NlMraltsad
p,p-dlckl((^ipkH)rl*
MltMSI
Klaaraltaatlo*
CO} 1*4 CNt
CO,
p-chloropha«yUcotata
Uaad aa citIm «Mr|) tourci
CO,
Mthyl rrraoltMtl	NlMrillllttM
athoiylatvd fktnoli	HlurilluttM
phanollra	llodagradatloaa
Di|riiitlai ring
dlaruptioa
PirtMdtloa	To viltriu, kutyrm, propionate,
foraata iW b)r4re|tn.
Or|i«lc acids faraantad to COj 4 CH4
Phanol dagradatlon
affactad bjr Iron con*
ctntrailoa, bactarial
lnnoculua alta and pM
Mlnarallaatlon
Mineralisation
NlnirilUitlot
MlnaralltailOA
Mineralisation
Mineralisation
Degradation, ring	CO},
4lirur(lM
Oitdatloa
Mineralisation
Oalrfatlon, ortho	lata katoadlpate
pathway, phenol
hydroayloae, catechol
1, 2-oaygeaeee
Mlaerelliacloa	CO}
pti*nyl|rfteual
r««orcliMl
AroawMra will) italogeneted
*ld« Chain*
(rlchleroiliena
p, p-d1 rli I orod1 pfceoy 1-
¦••thane
1 ^.VtrUMorofhaMiy
•casta eel*
wiakliaa
Coaetabolia*
CaaiiaHlN
trlckUroproplonlc
a«U
torophany | ace ta ta
l.WlcHloroantechol
ncmc	wirm
•IOLOCICAI	KNVltONHEtrTAL	SUKTUTI	CONTACT
Actwrfi)	ttouumwTi	ccMcanuticii	Tim	Kfaacu
S«M|t a lodge	Anoaroblc	W
Crouad vittr bacteria,	JDS la MO bra *J
airaaa baeterla	^
Savage eludge	AeroMe
M»ai ap.	AaraMc
hmdaaQMi ap.	Aerobic
AurtoUilffiw pmUmUw
Adapted flu bacteria Anaerobic, sedlaent and water JO «/l	70 lira	99
of eetwerlae and urlna altee
10#C
Hlcroorgaaleae	]g
>1-CHEN CtC-1	*9
•I-CMCM T*X-4	>V
Hlcroorgenleee Mte, ponde,	lagooae eolle and ill
wad waiar
II-CHEM COC-2	»
Mutant PeoMdoaonas af, 50*C	aj( ;j( jol,
no, W*
•acterlal eonaortla	Anaerobic
froa a variety of	jj
•athaoofanlc ecoeyeteaa.
100-400 ag/llter	34 10 100 hour# 4)
Oil refinery aactllng	Aaroblc, ahakar culiura 2t C.
pond bacteria	Mailaua degradation rat* ia
100 ag/llter with a pH optlaua
batmn 7 and I. Contiawowe
light.
if
yecudoaonoe ap.,	Aaroblc
Vibrio ap.. Iptrlllua ap.
»;
lacUlua ap.,	Aerobic
Mocardla ap.
57
Chleaydoaonae	Aaroblc, light required
ulearonele
fhorldlua luveolaruy.	Aaroblc	jj
leaned Ieawa baecllleneee
tmleas gracllua	Aaroblc, light required	^
Corynabactarlun ap.	Aaroblc
Mitant Paeudoaonoa ap. Aaroblc, J0*C Inorganic	jqq m/i	»4 hra	41 A 1U«
lartillaara (N and P)
II > 114
Constructed atralna of Aaroblc, )0"C, coneinuMa
Peeudoaonae ap. etreln culture	„
>11. AlcoTTienes ap«
atraia A ?-!
Uka Mtar bactarla	Aaroblc, Ji'C, no lllualaatlon
""«• >*••"*	Tract 1-100 ut/ml	II
Trichoaporon	Aaroblc, 30°C growth la a
cutanly *0114	Marubiehi jar feraeatetioa 1)0 ^/1	|00t 14 hra	94
Preehwater and aawaga Aaroblc, pM 7.0 ,29*C
bactarla	irece	10)
II-CNIH CEC-1
BI-CMCM CtC-1	**
>*
Brawlbactarlw ap.
4)
Mydroaewoaowae ap.	Aaroblc
Hydroaanoionaa ap.	Aaroblc	41
-------
BIOLOGICAL
ACTIVITY
PBOOUCTS
Nalophanola
Oifdatloa uaa •• aola COj
aourca •! organic
ctrkoa aad M«r|y
Matabolliad 00 tola COj, C«U Mat
carboa «wre«
aonocMeropfi#nol	Coaccabollaa
4-rhloroph#no)	Coaatabollaa
4-rhloroph*nol	Sol* anargy and carbon
Okldation, aola carbon
1,2-dloayganatlon of
cMorocotachola
llo^t|riditlon
df eMorophtnyl laocyanata Dagradatloa and dliruptlon
2, S-d1 hrn»oph*nol
rbrnM|ihtAAl
p*n(ifltbr«ph«Ael
J, 1 ,S-i rlehlorophfiiol
?, 5-dicMoropttanol
p-ch)oropfc«nol
Mltrop*«nol •
nIt rophanol
p-nltrophano)
Dagradatloa and ring
dtaruptloa
Nltroraductloa
NlAoralltacloa
p-aalnophanol
fanltr«thlna
(0,IVd taatHyl»0(V-
m* t hjr 1-4-ft I tropbanyl |
phoapfcorothloaoto)
Coaatabollaa,
can to	oo tola
carboa aourco
PttanAl a-Olhydr Idaa,
Polyhydrldaa
t-bvtyl catacHol
cMorocataehol
4-chlorocatacbol
), VdlcMorocatocbol
1,4-rltlo
Mataboltaa
Matabollaa
Oagradatloa
CiMtaMiM, Mta
Coaatabollaa, Mta
d«a«a|i tgyiiMM
1,2 daoiy|aaaao
1(2 JoosjrgoMoo
lata katoadlpota, aiiaboUud fortbor
to iwcImu an4 it(()rl*C» A, thn
to CO] aad R]0
2-bydro*f-4-cbloro-a«coolc-
ooalaldohydo
3 kydreiy^Hichlsro^uctatc*
aoalaldohyda
BIOLOGICAL
ACENT(S)
—IKClkl	
environmental
hequihehents
JNJT1AJ.
SUiS'lMATt
LUkcUITUATlUH
CUM ACT
TlHt
II-CKEM CEC-1
BI-CHEM SUS-I, Otloia«l Aaroblc, Cootaalaatod soil
dapth to I lutai
•actarlwa ICC-J
Aaroblc, conttnwowo flow
aarlehaaat culture pN 7.3, 2J*C
25 ppa
200 ib/1
41
39
1 ppa, 21 day# 39
7JI la 24 lira 24
Soil alcrobaa
II-CMEM nr-7
Arthroboctar ap. atrola
NC
Mocardla op.,
Hycoboctorlua op.
Paaudoaooaa ap. §13
Paaudoaonaa ap.,
aawaga
Paaudoaonaa. ap.
atrala II),
Alcalltanaa ap. atraln
Aaroblc, ehaaoatat
Aaroblc, uaa of plaaald
tranafar to anhanca aaiabollaa
500 u »/*!
V7X raductloa
i?
59
101
HI
37
91
91
A7-2 autont bactorla
Hutant bactarla
Mutant Paaudoaonaa ap.
Mutaat hawdoadm ap.
Aaroblc, pond laachatta of
phanola and o-chlorophanola
Acroblr , eontaalnatcd soil and
pond apray Injartlon laachata
ayataa, blotr«aior pond
15.000 ppa
Aaroblc, 20*C
Aaroblc, )0*C
500 ^/1
*00 at/1
HJO */l
200 a«/l
200 */l
200 ag/l
200 */l
752 to M daya 110
1 ppa, 9 auntha 114
1001.
i houra
1001,	37 iMMira
1001, 29 Uowra
2bl, 120 t«wr«
1001, 50 bouta
1001, 40 Iwura
1001, 42 hour*
*1, 75, 104,
*1, 75, 100,
luaat alcroorganlaaa	Aaaaroblc	j;
Acclla.<<4 ..Hurl*. or	10-100 Mb	»~
aarlna factorial coaunl-
tlaa (]iMtka adaption)
Ukt Oatarlo aadlaaato, Anooroblc, cfdaaa	0.5-1 day	17
•ell 1 activated aludga faraantor
Aclwatoboctar op.,	Aaroblc	0.5-1 day	(7
Aroaonoo op.t loelllua ap.,
F>tdoa
-------
BIOLOGICAL
ACTIVITY
)*»«thylce(echo1
4-chlororoaorclool
protoriltckuit*
Foly5
31
Aclnetobacttr ap. P6
Arthroboctor ap. N)
AlcalUlnea faccalla
lotJarlnckla M/3i
Oaclllatorla ap..
P. pytlda
Port Valdac. AK
aaatNtir*) • depth
Aclnatoboctet ap. Pb
Aaroblc. plaaald
Involvoaent, 25*-30*C
Aaroblc. 30*C
Aarohlc
Aaroblc
10"C, no aeration
Aerobic, plaaald lavolveaaat
2S*-30*C
4.7-4.4 u aole/
liter
1-3 daya
31
a
if
if
9.3-%.g a molt/ Hi
liter/day
*1 MMgl
Arthrobactar ap. with
plaaald p AC3?-or PAC31-
harboring faewdoaooaa
pwltda
aaawatar 5 a depth
10*C, no aeration
1.5-4.3
Hear
aole/
MS la 1
1.2-4.1 a aole/ •)
liter/day
faeitdoaowaa ap., Vibrla
*P«» Spiralluo ap.,
flavobactar ap.
Achroaohatter ap.t
Chroaobacter. ap.,
laclllla »Pi. Wocardla ap.
fungi
Sewage aludga
Aerobic
-------
IIOLOCICAL
ACTIVITY	MODUCTS
COj» uoUHtififd j«llw coapeuN
Gentieete
l-niph(ho|( t-hydroijcl'Utrileni,
(U*Mph(lMUn* dihydrodlol.
t It-1,2-4 1 hydroxy-1,2-4! hydro
uphlKiltni; t-hydrosjr-l-'totrelone;
and l-aaptithol
Microbial	CO], call bound C
transformation
Sedtaent Absorbed
blodegredadon
Ability to breakdown
thla coo pound ccaaon
aaong il|ii
1-naphthol; cla-1, 2-dlhydroxyl-
1,2-dlhydronaphthalone;
4-hydroity-l-tetrelene
alpha-naphthol; b*te*naphtho);
trans* 1, 2-dehydroay*-l, 2-
dehydronepthalene; 4-hydrovyl
-I'tttnlinail.i-naphtKo qulnone
Cli-j. OCHj, n or NOj	Degradation
aubetltuted naphthalenes
rn?H, ni^rojH or S0)H Dagr
substituted naphtKaionaa
ntfptht ti«)«neatil fonafe
1-napht	halriitaulfonai*
2-n«phthalenesulfon«te
Degradation
Hydroiylatad In 7,1 poiltioni
and 4-subatituted aallcylatea ara
accumulated
HydroiylatIon In 1,2 position,
1,2-dehydrosy-l,2-dehydro
naphthelane~2-carboayllc acids ara
(orwd
Unidentified and products which ara
•etebollted by oihtr bacteria
Clo-),4-dehydroiy-3,4-dehydro
phananthracene
pelyrycllc aroaatlc	llodo|radatloo
hydrocarbons
Degradation
Blodegradatlori
pn I y.h i. .f i natrd	Btodegradetlo*
blphenyle
•ferfegredatioa appear*
to b* Imriljr related
to titiit of dtUrlif
tloa
—	SfECIAI.	1N1TIAI.
JI0I.OCICAL KNVIRONMENTAL SUBSTKATK CONTACT
ACEWT(S)	tEQUmPgNTS	COHCEWTRATIOW	T1HE
Naphthalene utilising
¦Icroarganleae fro* aoll
Faeudoaonee ap. A3
and Ntudaaowat ap. C22
Cyanobecterla and
alcroalgae
Oaclllatorla ap.
strata JCN
Aerobic, pH 7.0, shaker cult
A, 50®C, bioatac feraantore
Photoeutotrophic conditions
Photoautotrophfc conditions
•41, 4« lire
Streaa eedlaente dowir-
str«M fro* a coal
coking ueetevater die-
charge alt* aedlaent
Bactarla
c, 20*C
310 daye
BI-CHEM roc-3
Agntntllus ap.,
ftarlllatorla ap.,
Anabaana ap.,
Cynnlnghaaella
~lagans	~

Mlcrocolaua ap. ,
Moetoc ap.,
Coccochlorla ap.,
Aphanocapea ap.,
CnloraUa ap.,
Puna 11117 ap.,
dWlsaydaaonas op.
Aaphora ap.
Paeodoaonaa ap.,
Flavobactarlua ap.,.
AlcalIgenes ap.,
Corgiibacterlua ap,
Atroaonia ap.,
Moreardll ap.
Pseudoaonae ap. A)
Paeudoaonas ap. C22
Aerobic, light required
Aaroblc, 30*C, bloatat
faroentors
Paeudoaonae ap. A3
Peewdoaonao ap. C22
Aaroblc, )0*C, bloatat
Faeodoannaa ap. A3,
Paeudoaonae ap. C22.
naphthalene degrading
eewage bacteria
AlcalIgeneo faecalla,
he!jrrlnckla But,
Peewdoaonas SPHM
Marina aedlaent*
Beljerlnckla ap.
Aerobic, 28*C
Aerobic, 30"C
2-3 daya
14, i?
St
57
Indigenous reeervolr
•lcr6blal population
Microorganisms
21*C In dark
Pita, ponda, lifoona, lolll
and vsstewater streeas
100 u g/iuu *ole
Mil, ~ we oka
Nerlne aedlaent bacteria U'C, dark
Indlgentoua reservoir
Microbial populat ton
MtcroorgaaUa*
25*C. dark
Pita, ponda, lagoona, eolle and
Hiimtir *tr**a*
y««M
-------
IIOIOCICAL
ACTIVITY
F««r riA| Cuwd Pelycycllc
Nydrocarbone
Microbial trMif«rai(lM
dlWAiiiitlirtccnt
rtiryarn#	Degradation
polynurlear arooeflca
Ho4*|ri4itl»R
3,4-;lt*-;iO,U-Dihydrel»
Vtry a low Inatgalflceat breakdown
Five linn Nn4 Polcyelle
Hydrocarboaa
b«Aso{a)pyrefie
Treae- 7,A-dl hydroiy 7, WI hydra-
be«so(a)pyreae
hitfil Poljrcycllc
Hydrocarboaa
coal	Degradation,	Black liquid
llgala
Argaaophoaffcetea
a«pon
•MratblN
tteed aa eole pbor
phorowe aourcae-noee Diaatkyl pfcoaphata
a# tba trjaiM-plMc	diethyl pfeoepfcoretbleoete
pfcetee aervod a* • diethyl yhoiybontklMMM
eat boa aowrce	4lMtkjrkpfeoafliwa4itklMti
dletbylpfcoepfcorotbloete
4 la t by 1 phoopfcoro tbloet a
I lag cleeveje	00]
alaerelltetlao
Mydfolyeea	Oletbyltblopkoepfcorlc acid eod
oetbyl paratbloa
rmicldia and Harblcldaa
tr IcMorophanoayecettc
acid (i|int orange) Degredetlea
Total degraJatloa, aala
eource af carbaa. J,4,
J-T degredetloe paltanj
Dacblortaetioa,
oaldetlea
CI*, ckltrapbatoli ead related
caopouada
—nrrriAE—
SUMTKATC
COMCIilTltikTlON
•J0L0CICAL
ACW(t)
—mem	
ENVtKO (MENTAL
HtQUKEKIITS
CONTACT
TIKI
Strew aadlamla ti
vlcfolty of coal
coklDf witnMUr
dlecfcarfe lit*
Cm—1 mI>—alia alaaaaa
Activated atudga
Marina aedtaeeta
ii^chbn roc-3
lAdlgeoove reeervalr
bactarla
Aerobic, IS*C
Aerobic
Aerobic
IS#C, dark
34 day*
*7
i7
it
4 to • v«aka
Cuaalmb—elle eleaane Aaroblc
fixidai
letjerlacbla ap.
Aaroblc
Aaroblc
Polyporooa veralcola,
Perl*. io«tlc>U
Lignite coal,
21 *C, MX relative bualdliy
tactarla laalatad (n
aall aad eevage
¦oa Aaroblc, Jf*C ikakar <
5rl
"
•7
•7
•7
17
•7
•7
•7
•7
•7
•7
•7
•7
Kkyubma ap.. Qiloralla
iyrmr<»i.
aoll bacteria
Ilea rhltoepliere
Paewdowmee dlalawta
rioodad and ooa flooded
coodltlana
li daya
Aiihftieke alcreorgealeae
Agnatic olcroblel growth
attached to ewbaergod avrfacaa
ar ewepeadad la etreoaere or
rnwrfowMi one I#	Crovch la (all, JO'C, IV-iOl i _/_ MtJ	»>j, 1	2], 12, U
AC 1100	aolatwre
Nlcroorgaalaa frea waata	Plaaolda pAC 25, TOL, CAN, SAL |00 w |/i)	701, 7 daya Jl
duap altaa	MC21; cheaoetat eavlroaaeat
Feaadeao—e cepacia	Aerobic, 30*C ebefcer culture 1 ag/al
AC1 100
971, * dayi
>2
-------
00
SUBSTRATE
BIOLOGICAL
ACT1VITT
wopocrs
nethoaychlor	Dechlorination	CI", CD]
ItMccuwlitln
l,l-41cM»r«>2,2-kli (p-Mthoiy*
phenyl) ethylene: l,l-4t«Mar»-
2,hbli (p<«tciwi)r|lwit)f))ttht«*
cMorodlnefora
tatjmilc
degredetloe
D00
A-chloro-o-feraotoluldllene,
4-chloro-o-toluIdene, S-chloro
enthrenlllc tdd, n-for»yl-
3-chloroeachrantilc tell
Degradation	DOD, DOB, ktlthatt, DOF a«4 I
Reductive dechlorination DM
Muctlvi JichlsrlMtUi DOO
DOt
DOM, DONS, TDC, DOC
AMtreklc degradation
ftitlMliiltfrU,
<«riM

CklortlM, cbUN«H epovlde
Fbotodleldrln. eldren dtol
MtUrli
Aldrledlol, phetodleldrla
pentechloronltrobantene
phurti* iuKiIi
toMphta*
e tres Ine
NiictlUntoui H|4ro0 (MENTAL	SUBSTMTE	CONTACT
ACENT(S)	UOIIIPtlfft	COWCWmTlON	TIKE	Iter PEACES
toll bicttrU	teitrobic, denitrifying 1,000 ppa MX, J withi 37
condltloni
*»rob«ct«r Kfoitiwi.	Aerobic, ihtktr culture 21'C O.VS.O u g/lltor 30 alouui 44
bclllm
fcc«r<<« p.,	Airvklc 57
ItMKWf« ep.
Aerobecter urowxi	Aerobic, tnttrekic
Sit* MUr ilcrobii	Aerobic
Chlorelle, OidHitorii Aerobic, 2S*C, ^r»i«M« of light M0 u |/il	14 4tyi	57
Aauroklc
aludge
e, J5*C
• rclal jraiit
Anaerobic, >«ri than 20
bacteria epeclea are reported
to be able to K DeClj 00T
Aerobic, anaerobic ayvblotle
reletlonehlp
Klebetelle pnewonla.
Eacherlchla Coll
Aerobecter aerogonee.
FBcudoaooea cloitrHui,
Fiendaeoini »ulgerl>
Oceanic condltlone
Nlied cultere ol fungi
and becterle
Anaerobic dlgeater aludge Anaereblc, 35*C
Anaeyetle aldulaae.	Anaerobic
Agnenolow
...rillfllc.t.M.
raeudoaonee ap.
bMl flel7	Anaerobic
Actlneeycetee	Anaerobic, eeroblc
Ocean eedlaeat*
Anaerobic dlfeeter eludg#	Anaerobic, 33'C
Site Mtir alcrebee,
MM|< elnige
Sea weter, bettoa eedl-
neate free ocean and
eetunrlneo
Aapergellue nlger.
Fllaarlun eolanl,
tloaereila
tongaiato,
Helelntaoeporlun
vtctoreee.
Nyrotheu*,
Ponlcllllun ap.,
~rlchodeme
vorldeo
Soil bacteria
1 |/al
lUUS. 40 ««yo
2taC-27*C and pN 7.4-7.7
growth
Corynibacterlua pyrogeneB Aiueroble
ll-CMIM P0I-4	Detoieol aaturated aolla
39
4V
37
>7
7b
37
42
37
37
37
78
U
37
37
17
39
FHENOAAC,	Aeration tank, pM 5.>-4.0,
a alature of eeroblc
end facultative elcro-
orgenleae
100-400 ^/1 MO,	Plant under con- 114
3-73 a*/l	trol la 13 day*,
phenole	**-*3X BOO relet*
tl«a, 12 bra.
-------
BIOLOGICAL
ACTIVITY
•IOLOCICAL
ACMT(I)
—Irmxi	
INVIftOMCNTAL
uoummti
INITIAL
susroun
cowcomuTiow
cgnucT
Time
crud* petrolcua producti Blodegrodetloa
Moa-toilc krproducti
HlcrMr|Mt«M
Cmim
iivlr«
U, t*
detergaat* (nonloalc,
anionic, catloalc)
•thoaylated phonela
fern
h
-------
•IOLOCICAL
SUBSTRATE	ACTIVITY	KOOUCT8
refined petrolsua	Biadegradattoa	Moiftoslc byproduct!
producl•
¦ )u4|ii (p«p«r Industry Mil
end	procsssiag)
lulfur CMp«i»di
w«*c• oil	Dagradstloo
FiAttcklorocyelAlMUM
cyclohiiiM, alpha MC
1 ndua trial iwfUctuti	llo4«|ri4«l{M	Non-toxic byproducts
oil
Btodsgredatloo
organic eolvsnts	llodsgradatloa	Nontoxic byproducts
ptidcUta 4 bsrbicldss	Blod«grada«loa	Ho*-K«stc byproducts
petrotsuo distillate
phvnolics	llodaaradatloa	NotrtMlc byproducts
polytyrllc aroMtlc
hydrocarbons	llodsgrsdatloa	ion-coslc byproducts
polythlorinatad btphenyls Blodsgrsdatloa	Moo-toxlc byproducts
polynuflear arooatlea
refined petroleua	Blodsgradsttoa	Non-toxic byproducts
products
sludgsa (pap«r Industry fish
•nd vsgatsbls yr«cisst»|)
sulfur cosyouads
SPECIAL	INITIAL
BIOLOGICAL	ENV1RONMNTAL	SUBSTBATt	CONTACT
ACCNTtS)	UqumtXZHTS	CONCOfflUTlON	TIW	I EH* PICM
NlcroorisnlsM	ig, *»
»I-CHEW PAC-5	19
SI-OBN SUf-l	*
fNIMIAC,	4 alllton gallon	100 ppm	991, 9 Booths 114
PITVOBAC	lagoon FOLYIAC N (sacrotiw-
trlcnta nitrogen snd
phosphorous)
Anasroblc dlgsstsr aludga Anasroblc, 1S*C	jjq U>/«1	10UI 2> days	42
CMorslls vulgar!#,	Anssroblc (t)	5?
Chlanydaaonaa	Asroblc	j;
relnhardtll
Chloeterldlun sp,,	Anasroblc
Fsaudasoixs sp.
toil bactarls	Anasroblc	j;
tw«|« slud|s	Aassroblc	jj
Mlcroorgaalaas	Contaalnatod soils* aquatic	]|
systaaa
bi-cmim roe*)	i9
HSNOIAC	Aaratloa basts 0.5 MCO vast*- OU/grassa	IU 24 hra	114
water traataaat at Marios bulk MO to bOO ^/l	72 bra
liquid storage facility	raducad to 7 ^/1
Microorganism	Coataalnatad soils, aquatic	^
systaaa
Mlcroorganlaas	Coataalnatad soils, «|Mtl<
systaaa
ll-CHIH	4 icrtf c«auaJMl«4 Mil	1,200 ppm	I ppm, 21 days 59
Klcroorgaalaas	Coataalaatad soils, Hwilc	2|
ayatoaa
Bl-CNIM COC-2	M
Mtcroorfsalaaa	21
Mlcroorgsalsas	21
ll-CUM NH	59
Bl-CBEM rOC-6	a
Mleroorgsalsns	u, m
ll-CHIM PAC-5	59
•1-OBH fUf-l	»
NKNOWC,	4 Billion gallon	COO ppm	99S, 9 aoaths 114
ftTtOAAC	lagoon fOLYBAC N (ascrsBW-
trlents altrogaa snd
phosphorous)
-------