&EPA
United States
Environmental Protection
Agency
Handbook on Advanced
Nonphotochemical Oxidation
Processes
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EPA/625/R-01/004
July 2001
HANDBOOK ON ADVANCED NONPHOTOCHEM1CAL
OXIDATION PROCESSES
Center for Environmental Research Information
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
/T"V Recycled/Recyclable
Printed with vegetable-based ink on
paper that contains a minimum of
50% post-consumer fiber content
processed chlorine free.
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Notice
A,A haS bee" funded bV the U-S- Environmental Protection Agency under Contract
nH I Work Assignment No. 3-86 and its predecessors. Mention of trade names or commercial
products does not constitute endorsement or recommendation for use.
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Foreword
The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation's land, air, anc
water resources. Under a mandate of national environmental laws, the Agency strives to formulate and
implement actions leading to a compatible balance between human activities and the ability of natural systems
to support and nurture life. To meet this mandate, EPA's research program is providing data and technical
support for solving environmental problems today and building a science knowledge base necessary to
manage our ecological resources wisely, understand how pollutants affect our health, and prevent or reduce
environmental risks in the future.
The National Risk Management Research Laboratory is the Agency's center for investigation of technological
and management approaches for preventing and reducing risks from pollution that threaten j human health
and the environment. The focus of the Laboratory's research program is on methods and their cost-
effectiveness for prevention and control of pollution to air, land, water, and subsurface resources; protection
of water quality in public water systems; remediation of contaminated sites, sediments and groundwater;
prevention and control of indoor air pollution; and restoration of ecosystems. NRMRL collaborates with both
public and private sector partners to foster technologies that reduce the cost of compliance and to anticipate
emerging problems. NRMRL's research provides solutions to environmental problems by: developing and
promoting technologies that protect and improve the environment; advancing scientific and engineering
information to support regulatory and policy decisions; and providing the technical support and information
transfer to ensure implementation of environmental regulations and strategies at the national, state, and
community levels.
This publication has been produced as part of the Laboratory's strategic long-term research plan. It is
published and made available by EPA's Office of Research and Development to assist the user community
and to link researchers with their clients.
E. Timothy Oppelt, Director
National Risk Management Research Laboratory
in
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Abstract
The primary purpose of this handbook is to summarize commercial-scale system performance and cost data
for advanced nonphotochemical oxidation (ANPO) treatment of contaminated water, air, and soil. Similar
information from pilot- and bench-scale evaluations of ANPO processes is also included to supplement the
commercial-scale data. Performance and cost data is summarized for the following ANPO processes:
(1) Fenton, (2) ozone/hydrogen peroxide, (3) electrochemical oxidation, (4) supercritical water oxidation,
(5) cavitation (acoustic, electrohydraulic, and hydrodynamic), (6) electrical discharge-based nonthermal
plasma, (7) gamma-ray, (8) x-ray, and (9) electron-beam. This handbook is intended to assist environmental
practitioners in evaluating the applicability of ANPO processes and in selecting one or more such processes
for site-specific evaluation.
ANPO processes have been demonstrated to be effective in treating contaminated water, air, and soil to
varying degrees. Regarding contaminated water treatment, a number of ANPO processes have been
evaluated in terms of their effectiveness in treating various contaminants in groundwater, industrial
wastewater, municipal wastewater, drinking water, landfill leachate, and surface water. Of these processes,
the Fenton process has been evaluated for the most contaminant groups, and the electrohydraulic cavitation
and x-ray processes appear to have been evaluated for the fewest. Regarding contaminated air treatment,
only three ANPO processes have been evaluated in terms of their effectiveness in treating various
contaminants air stripper off-gas, industrial emissions, and automobile emissions. Of these processes, the
electrical discharge-based nonthermal plasma and electron-beam processes have been evaluated for the
most contaminant groups, and the gamma-ray process has. been evaluated for the fewest. Regarding
contaminated soil treatment, only two ANPO processes have been evaluated in terms of their effectiveness
in treating various soil contaminants. Of these processes, the Fenton process has been evaluated for more
contaminant groups than the gamma-ray process.
IV
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Contents
Notice .................... , .............................. • .......................... "
Foreword [[[ '"
Abstract [[[ iv
Tables [[[ viii
Figures [[[ ix
Acronyms, Abbreviations, and Symbols [[[ x
Glossary ................ [[[ Xl"
Acknowledgments .......................................... ........................ XVI
Executive Summary .................... • .......................................... ES~1
1 Introduction [[[ 1~1
1.1 Purpose and Scope [[[ 1~1
1 .2 Organization [[[ 14
1 .3 References ................................... • ....................... 1"4
O -1
2 Background [[[
2.1 ANPO Processes [[[ 2~1
2.1 .1 Fenton Process ................................ • • ................ 2"1
2.1 .2 O3/H2O2 Process ................................ ................ 2"3
2.1 .3 Electrochemical Oxidation Process .................... .............. 2-4
2.1 .4 SCWO Process ................................................. 2'4
2.1 .5 Cavitation Processes ............... . ............................. 2"5
2.1 .5.1 Acoustic Cavitation ........................................ 2~5
2.1.5.2 Electrohydraulic Cavitation . . ............................... 2'7
2.1 .5.3 Hydrodynamic Cavitation ........ ........................... 2~7
2.1.6 Electrical Discharge-Based Nonthermal Plasma Processes ............... 2-7
2.1 .7 Gamma-Ray Process ........................................... • 2~^
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Contents (Continued)
2.2.4 Chematur Aqua Critox® SCWO System 2-15
2.2.5 General Atomics SCWO System 2-15
2.2.6 OSI HYDROX Hydraulic Cavitation System ....'.'.'.'.'.'.'.'.'.'.'.'.'.'.'.'. 2-16
2.2.7 HVEA E-Beam Treatment System '.'.'.'.'. 2-17
2.3 ANPO System Design Considerations ' 2-18
2.4 References ] 2-19
3 Contaminated Water Treatment 3.-l
3.1 Contaminated Groundwater Treatment 3_1
3.1.1 VOC-Contaminated Groundwater 3_1
3.1.2 SVOC-Contaminated Groundwater 3.7
3.1.3 PCB-Contaminated Groundwater 3_-|4
3.1.4 Pesticide- and Herbicide-Contaminated Groundwater 3-14
3.1.5 Dioxin- and Furan-Contaminated Groundwater 3--I5
3.1.6 Explosive- and Degradation Product-Contaminated Groundwater '.'.'.'. 3-16
3.1.7 Humic Substance-Contaminated Groundwater 3-17
3.1.8 Inorganic-Contaminated Groundwater '' ]'_ 3.18
3.2 Industrial Wastewater Treatment 3_26
3.2.1 SVOC-Contaminated Industrial Wastewater 3-26
3.2.2 Dye-Contaminated Industrial Wastewater !!!!.'!!.' 3-27
3.2.3 Inorganic-Contaminated Industrial Wastewater 3-30
3.2.4 High-COD Industrial Wastewater '.'.'.'.'.'.'.'.'.'.'.'. 3-32
3.3 Municipal Wastewater Treatment 3,35
3.3.1 VOC-Contaminated Municipal Wastewater 3.35
3.3.2 Microbe-Contaminated Municipal Wastewater ', 3.35
3.4 Contaminated Drinking Water Treatment 3.37
3.4.1 VOC-Contaminated Drinking Water 3.37
3.4.2 SVOC-Contaminated Drinking Water '.'.'.'.'.'.'. 3-37
3.4.3 Humic Substance-Contaminated Drinking Water '. 3.37
3.4.4 Microbe-Contaminated Drinking Water '.'.'.'.'.'.'.'. 3-38
3.5 Landfill Leachate Treatment 3_40
3.6 Contaminated Surface Water Treatment 3-40
3.7 References ' 3.40
4 Contaminated Air Treatment 4.^
4.1 Air Stripper Off-Gas Treatment 4_1
4.1.1 VOC-Containing Air Stripper Off-Gas 4-1
4.1.2 SVOC-Containing Air Stripper Off-Gas '.'.'.'......... 4.4
VI
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Contents (Continued)
Section
4.2 Industrial Emission Treatment 4-6
4.2.1 NOX- and SOx-Containing Industrial Emissions .- 4-6
4.2.2 Metal-Containing Industrial Emissions 4-7
4.3 Automobile Emission Treatment 4-9
4.4 References 4-9
5 Contaminated Soil Treatment 5-1
5.1 VOC-Contaminated Soil 5-1
5.2 SVOC-Contaminated Soil 5-3
5.3 PCB-Contaminated Soil 5-7
5.4 Pesticide- and Herbicide-Contaminated Soil 5-7
5.5 Dioxin-Contaminated Soil 5-8
5.6 Explosive- and Degradation Product-Contaminated Soil 5-8
5.7 References 5-15
Appendix
Technology Vendor Contact Information A-1
VII
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Tables
Table Page
ES-1. Summary of Commercial-Scale ANPO Processes for Contaminated Water Treatment ES-7
ES-2. Summary of Commercial-Scale ANPO Processes for Contaminated Soil Treatment ES-9
1-1. Oxidation Potential of Several Oxidants in Water 1-1
1-2. Overall Rate Constants for O3 and -OH Reactions with Organic Compounds in Water 1-2
3-1. Summary of Contaminated Groundwater Treatment , 3-19
3-2. Summary of Industrial Wastewater Treatment 3.33
3-3. Summary of Municipal Wastewater Treatment 3-36
3-4. Summary of Contaminated Drinking Water Treatment 3-39
4-1. Summary of Air Stripper Off-Gas Treatment 4.5
4-2. Summary of Industrial Emission Treatment 4-8
5-1. Summary of Contaminated Soil Treatment 5-10
VIII
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Figures
Figure
1-1. Performance and cost data organization 1"3
2-1. Flow configuration in an in situ Geo-Cleanse® Fenton system 2-12
2-2. Flow configuration in an ex situ Geo-Cleanse® Fenton system 2-13
2-3. Flow configuration in an ISOTEC™ Fenton system 2-14
2-4. Flow configuration in a Chematur Aqua Critox® system 2-15
2-5. Flow configuration in a General Atomics SCWO system 2-16
2-6. Flow configuration in an OSI HYDROX hydraulic cavitation system 2-17
2-7. Flow configuration in an HVEA E-Beam system 2-18
ix
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Acronyms, Abbreviations, and Symbols
<
±
ug/kg
ADMI
ANPO
bgs
BOD
BTEX
c.u.
CaCO3
cfu/mL
Chematur
cm
cm2
CMHPOS
""Co
CO
CO2
COD
CP
DCA
DCE
DCP
DMMP
DNP
DOC
DRO
DSD
e",q
E-beam
EDTA
Fe(0)
Fe(ll)
Fe(lll)
FeCI2
FeSO4
FSR
Geo-Cleanse
GRO
H-
H&H
H2O
Greater than
Less than
Plus or minus
Microgram per kilogram
Microgram per liter
American Dye Manufacturers Institute
Advanced nonphotochemical oxidation
Below ground surface
Biochemical oxygen demand
Benzene, toluene, ethylbenzene, and xylene
Color unit ;
Calcium carbonate
Colony forming unit per milliliter
Chematur Engineering AB
Centimeter
Square centimeter
Complexed metal hydrogen peroxide oxidation system
Cobalt 60
Carbon monoxide
Carbon dioxide
Chemical oxygen demand ;
Chlorophenol
Dichloroethane
Dichloroethene
Dichlorophenol
Dimethyl methane phosphoric acid ester
Dinitrophenol '•
Dissolved organic carbon
Diesel range organics
1-Amino-8-naphthol-3,6-disulfonic acid
Aqueous electron
Electron beam
Ethylenediaminetetraacetic acid
Metallic iron
Ferrous iron
Ferric iron
Ferrous chloride
Ferrous sulfate
Fenton sludge recycling ;
Geo-Cleanse International, Inc.
Gasoline range organics
Hydrogen radical
H&H Eco Systems, Incorporated
Water molecule
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Acronyms, Abbreviations, and Symbols (Continued)
H202
H2S04
H3P04
HCI
HMX
HO2-
HVEA
ISOTEC
keV
kg
kHz
kPa
krad
kW
kWhr/L
L
L/hr
L/min
L/s
M
m
m3
m3/hr
m3/min
mA
mA/cm2
MeV
mg/kg
mg/L
ml
mM
MPa
Mrad
Mrad/min
MTBE
NORAM
NOX
NP
O&M
0(1D)
02
03
OCDD
•OH
Hydrogen peroxide
Sulfuric acid
Phosphoric acid
Hydrochloric acid
Octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine
Hydroperoxide ion
High Voltage Environmental Applications, Inc.
In-Situ Oxidative Technologies, Inc.
Kiloelectron volt
Kilogram
Kilohertz
kiloPascal
kilorad
Kilowatt
Kilowatt-hour per liter
Liter
Liter per hour
Liter per minute
Liter per second
Mole per liter
Meter
Cubic meter
Cubic meter per hour
Cubic meter per minute
Milliampere
Milliampere per square centimeter
Million electron volt
Milligram per kilogram
Milligram per liter
Milliliter
Millimole per liter
MegaPascal
Megarad
Megarad per minute
Methyl-tert-butyl ether
NORAM Engineering and Constructors Limited
Nitrogen oxides
Nitrophenol
Operation and maintenance
Singlet oxygen atom
Oxygen
Ozone
Octachlorodibenzo-p-dioxin
Hydroxyl radical
XI
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Acronyms, Abbreviations, and Symbols (Continued)
OH-
OSI
PAH
PCB
PCE
POP
PDU
ppmv
Pt/Co
RDX
RTD
scmm
SCWO
SO2
so,
ssco™
svoc
SW-846
TCA
TCDD
TCE
TCP
TIC
TiO2
TKN
TNT
TOC
U.S. EPA
UV
VC
VOC
W
W/cm2
WPO®
Hydroxide ion
Oxidation Systems, Inc.
Polynuclear aromatic hydrocarbon
Polychlorinated biphenyl
Tetrachloroethene
Pentachlorophenol i
Process Development Unit
Part per million by volume '
Platinum/cobalt
Hexahydro-1,3,5-trinitro-1,3,5-triazine
Resistance temperature device
Standard cubic meter per minute
Supercritical water oxidation
Sulfur dioxide
Sulfur oxides
Solid State Chemical Oxidation™
Semivolatile organic compound
"Test Methods for Evaluating Solid Waste"
Trichloroethane !
2,3,7,8-Tetrachlorodibenzo-p-dioxin
Trichloroethene ',
Trichlorophenol
Tentatively identified compound
Titanium dioxide
Total Kjeldahl nitrogen
2,4,6-Trinitrotoluene
Total organic carbon :
U.S. Environmental Protection Agency
Ultraviolet
Vinyl chloride
Volatile organic compound ,
Watt
Watt per square centimeter ;
Wet Peroxide Oxidation®
XII
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Glossary
Acoustic impedance. The ratio of the sound pressure on a given surface to the sound flux through that
surface. Acoustic impedance is expressed in acoustic ohms.
Batch reactor. A container in which a reaction is performed without any inflow or outflow of material during
the reaction
Bioassay test. A test for quantitatively determining the concentration of a substance that has a specific effect
on a suitable animal, plant, or microorganism under controlled conditions
Biochemical oxygen demand (BOD). The amount of dissolved oxygen consumed by microorganisms during
biochemical decomposition of oxidizable organic matter under aerobic conditions. The BOD test is widely
used to measure the pollution associated with biodegradable organic matter present in wastewaters.
Bioluminescence. The emission of visible light by living organisms
Capacitance. The ratio of the charge on one of the conductors of a capacitor to the potential difference
(voltage) between the conductors
Capacitor. A device that essentially consists of two conductors (such as parallel, metal plates) insulated from
each other by a dielectric and that introduces capacitance into a circuit, stores electrical energy, blocks the
flow of direct current, and allows the flow of alternating current to a degree dependent on the capacitor's
capacitance and the current frequency. A capacitor is also known as a condenser.
Catalyst. A substance that alters the rate of a chemical reaction and that may be recovered essentially
unaltered in form and amount at the end of the reaction
Chelating compound. An organic compound in which atoms form more than one coordinate bond with
metals in solution
Chemical oxygen demand (COD). A measure of the oxygen equivalent of organic matter that is susceptible
to oxidation by a strong chemical oxidant under acidic conditions. The COD test is widely used to measure
the pollution associated with both biodegradable and nonbiodegradable organic matter present in
wastewaters.
Complex. A chemical compound formed by the union of a metal ion with a nonmetallic ion or molecule called
a ligand or complexing agent
Congener. A chemical substance that is closely related to another substance, such as a derivative of a
compound or an element belonging to the same family as another element in the periodic table. For example,
the 209 polychlorinated biphenyls are congeners of one another.
Dielectric. A material that is an electric insulator or in which an electric field can be sustained with minimum
dissipation of power. Commonly used dielectrics include air, rubber, plastics, and oil.
Dielectric constant. For an isotropic medium, the ratio of the capacitance of a capacitor filled with a given
dielectric to that of the same capacitor having only vacuum as dielectric. Also known as relative permittivity
Dielectric strength. The maximum electrical potential gradient (voltage difference across a material per unit
length) that a material can withstand without destruction of the material itself, which causes arcing. Dielectric
strength is usually expressed in volts per millimeter of the thickness of the material and is also known as
electric strength.
Diff usivity. The quantity of heat passing normally through a unit area per unit time divided by the product of
the specific heat, density, and temperature gradient. Diffusivity is also known as thermometric conductivity.
XIII
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Glossary (Continued)
Electrical conductivity. A measure of the ability of a solution to carry an electrical current Electrical
conductivity varies with both the number and type of ions present in a solution.
Electrical potential. The amount of work that must be done against electric forces to bring a unit charge from
a reference point to the point in question. The reference point is located at an infinite distance, or for practical
purposes, at the surface of the earth or some other large conductor.
Electrolytic cell. A cell consisting of electrodes immersed in an electrolyte solution. Such a cell is used to
carry out electrolysis.
Electromagnetic radiation. A form of energy that appears to consist of both waves and particles called
photons. It includes visible light, ultraviolet radiation, radio waves, x-rays, and other forms of radiation
differentiated by their wavelengths and equivalent energies.
Electron volt. A unit of energy equal to the energy acquired by an electron when it passes throuqh a potential
difference of 1 volt in vacuum
Half-life. The time required for a given material to decrease to one-half of its initial amount durinq a chemical
reaction
Hydraulic retention time. The time spent by a unit volume of water in a reactor expressed as the ratio of the
reactor volume to the influent flow rate
Implicit price deflator. The ratio of the gross national product measured at current prices to the gross
national product measured at prices in some base year
Isotropic medium. A medium whose properties do not depend on the direction along which they are
measured y
Nonphotochemical oxidation. A chemical reaction that is not influenced or initiated by light and that
removes electrons from a compound or part of a compound
Oxidant. A chemical that decreases the electron content or increases the oxygen content of other chemicals
Oxidation potential. The difference in electrical potential between an atom or ion and the state in which an
electron has been removed to an infinite distance from the atom or ion
Pyrolysis. The chemical decomposition or change brought about by the use of heat in the absence of oxygen
Radical. An uncharged species containing one or more unpaired electrons; also known as a free radical
Saturated organic compound. An organic compound in which all the available valence bonds along the
carbon chain are attached to other atoms. Such compounds contain single bonds.
Scavenger. In advanced oxidation process chemistry, any compound or ion that is not a target contaminant
and that consumes primary reactive species (hydroxyl radicals). Carbonate and bicarbonate ions are two
examples of scavengers.
Singlet oxygen atom. Oxygen with no unpaired electrons. It is more reactive than triplet oxyqen foxvqen
with two unpaired electrons—the ground state).
Solvent polarity. The tendency of a solvent to promote ionization of a solute. Water has a hiqher polarity
than oil. * •
,xiv
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Glossary (Continued)
Specific heat. The ratio of the amount of heat required to raise a mass of material 1 degree in temperature
to the amount of heat required to raise an equal mass of a reference substance, usually water, 1 degree in
temperature, usually at constant pressure or constant volume
Stoichiometry. The numerical relationship of elements and compounds as reactants and products in
chemical reactions
Superfund. A program established in 1980 by the U.S. Environmental Protection Agency (U.S. EPA) to
identify abandoned or inactive sites where hazardous substances have been or might be released to the
environment in order to (1) ensure that the sites are.cleaned up by responsible parties or the government,
(2) evaluate damages to natural resources, and (3) create a claim procedure for parties that have cleaned up
the sites or spent money to restore natural resources
Superfund Innovative Technology Evaluation Program. A program established by U.S. EPA to encourage
development and implementation of innovative technologies for hazardous waste site remediation, monitoring,
and measurement
Synergistic effect. A condition in which the total effect of two or more active components in a mixture is
greater than the sum of their individual effects
Unsaturated organic compound. An organic compound that contains one or more double or triple bonds
Unsaturated zone. The zone between the land surface and the water table. The unsaturated zone is also
called the vadose zone or zone of aeration.
Viscosity. The resistance that a gaseous or liquid system offers to flow when the system is subjected to
shear stress. Viscosity is also known as flow resistance.
xv
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Acknowledgments
This handbook was prepared under the direction and coordination of Mr. Douglas Grosse and Ms. Norma
Lewis of the U.S. Environmental Protection Agency (U.S. EPA) National Risk Management Research
Laboratory (NRMRL) in Cincinnati, Ohio. Mr. Grosse served as the work assignment manager (WAM) and
Ms. Lewis served as the technical coordinator for the project. Contributors to and reviewers of this handbook
included Mr. Grosse and Ms. Lewis, Mr. Gary Peyton of the Illinois State Water Survey, Dr. John Roth of
Vanderbilt University, Dr. Sardar Hassan of the U.S. Air Force, and Dr. E. Sahle-Demessie of U.S. EPA
NRMRL. The handbook cover was designed by Mr. John McCready of U.S. EPA NRMRL. Dr. Jean Dye of
U.S. EPA NRMRL provided editorial assistance.
This handbook was prepared for U.S. EPA NRMRL by Dr. Kirankumar Topudurti, Mr. Eric Monschein, and
Ms. Suzette Tay of Tetra Tech EM Inc. (Tetra Tech) under a subcontract with Science Applications
International Corporation (SAIC). Ms. Virginia Hodge served as the SAIC WAM for this project. Special
acknowledgment is given to Ms. Jeanne Kowalski, Mr. Jon Mann, Mr. Stanley Labunski, Mr. Joseph Abboreno,
and Dr. Harry Ellis of Tetra Tech and Ms. Hodge of SAIC for their assistance during the preparation of this
handbook.
XVI
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Executive Summary
Over the past 2 decades, environmental regulatory
requirements have become more stringent because
of increased awareness of the human health and
ecological risks associated with environmental
contaminants. Therefore, various treatment
processes have been developed over the last 10 to
15 years in order to cost-effectively meet these
requirements. One such group of processes is
commonly referred to as advanced oxidation
processes. These processes generally involve the
generation and use of powerful but relatively
nonselective transient oxidizing species, primarily the
hydroxyl radical ('OH) and in some cases the singlet
oxygen atom. The 'OH can be generated by both
photochemical and nonphotochemical means to
oxidize environmental contaminants. This handbook
discusses the applicability of advanced
nonphotochemical oxidation (ANPO) processes for
treatment of contaminated water, air, and soil.
Similar information on advanced photochemical
oxidation processes is presented in a separate
document, "Handbook: Advanced Photochemical
Oxidation Processes," published by the U.S.
Environmental Protection Agency.
The primary purpose of this handbook is to
summarize commercial-scale ANPO system
performance and cost data for treatment of
contaminated water, air, and soil. In addition, it
presents similar information drawn from pilot- and
bench-scale evaluations of ANPO processes to
supplement the commerciaj-scale performance and
cost data. The handbook is intended to serve as an
ANPO reference document for remedial project
managers, on-scene coordinators, state and local
regulators, consultants, industry representatives,
and other parties involved in management of
contaminated water, air, and soil. Specifically, it
is designed to assist these intended users in
evaluating the applicability of ANPO processes
and in selecting one or more ANPO processes for
site-specific evaluation.
This handbook is not intended to summarize all the
ANPO system performance and cost data available
in the literature. Rather, it is intended to present
information on state-of-the-art ANPO processes for
treating contaminated environmental media.
Commercial-scale ANPO system performance and
cost data is presented in greater detail than
pilot-scale results because the handbook is intended
for environmental practitioners. Similarly, pilot-scale
results are presented in greater detail than
bench-scale results. In addition, pilot- and
bench-scale results are presented only where they
supplement commercial-scale results or where they
fill information gaps, such as those associated with
by-product formation.
This handbook presents an introduction (Section 1);
provides background information on various ANPO
processes, typical commercial-scale ANPO systems,
and system design considerations (Section 2); and
summarizes ANPO system performance and cost
data for treating contaminated water, air, and soil
(Sections 3, 4, and 5, respectively). References
cited in each section are listed at the end of the
section. Contact information for ANPO process
vendors is presented in an appendix.
This executive summary briefly describes the ANPO
processes and summarizes the commercial-scale
system performance and cost data for treatment of
contaminated water, air, and soil. Tables ES-1 and
ES-2 at the end of this executive summary present
commercial-scale performance and cost data for
contaminated water and contaminated soil treatment,
respectively, using various ANPO processes. At the
time of this handbook's preparation, no
commercial-scale ANPO systems for contaminated
air treatment were available.
ANPO Processes
ANPO processes can be broadly divided into the
following:(1) Fenton, (2) ozone (O3)/hydrogen
peroxide (H2O2), (3) electrochemical oxidation,
(4) supercritical water oxidation (SCWO),
(5) cavitation, (6) electrical discharge-based
nonthermal plasma, (7) gamma-ray, (8) x-ray, and
(9) electron-beam (E-beam). These ANPO
processes and their variations are briefly described
below.
Fenton Process
Decomposition of H2O2 using ferrous iron (Fe[ll]) or
ferric iron (Fe[lll]) under acidic conditions to yield
•OH is known as the classic Fenton process.
Several variations of the Fenton process have been
researched, and some of them have shown definite
advantages over the classic Fenton process. For
example, in both electro-Fenton and bio-Fenton
processes, at least one of the two reactants is
produced through an electrochemical process
(eiectro-Fenton) or a microbial process (bio-Fenton),
thus eliminating the need to continuously supply a
chemical reagent. In another variation of the Fenton
process known as complexed metal hydrogen
peroxide oxidation, soluble organoiron complexes
ES-1
-------
are used to carry out the Fenton reaction in both
acidic and neutral pH ranges. :
Commercial-scale Fenton systems are currently
available from the following vendors:
(1) Geo-Cleanse International, Inc. (Geo-Cleanse);
(2) In-Situ Oxidative Technologies, Inc. (ISOTEC);
(3) Mantech Environmental Corporation; and
(4) H&H Eco Systems, Incorporated (H&H). The
Gee-Cleanse®, ISOTEC™, and Mantech
Environmental Corporation CleanOX® systems can
be applied in situ to treat contaminated water and
soil. The Geo-Cleanse® system can also be applied
ex situ to treat contaminated soil. The H&H Solid
State Chemical Oxidation™ (SSCO™) system is
applicable for ex situ treatment of soil. At the time of
this handbook's preparation, no information was
available on the CleanOX® system.
Oj/W2O2 Process
The O3/H2O2 process, also known as the peroxone
process, has been used for treatment of
contaminated water. In this process, two -OH are
formed for each mole of H2O2 reacting with two
moles of O3. The only commercial-scale O3/H2O2
system available is the Applied Process Technology,
Inc., HiPOx™ water treatment system. At the time of
this handbook's preparation, no information was
available on the HiPOx™ system.
Electrochemical Oxidation Process
The electrochemical oxidation process has been
used for treatment of contaminated water. In this
process, electricity flows through an electrochemical
reactor consisting of electrodes separated by an
electrolyte. Oxidation and reduction reactions occur
on the surface of the electrodes at the
electrode-electrolyte interface. Multiple reaction
pathways have been proposed for -OH formation
during electrochemical oxidation. These pathways
include (1) oxidation of hydroxide ions to -OH in the
anodic region under alkaline conditions, (2) formation
of -OH when the water molecules or organic
compounds present in the waste stream that are
adsorbed on the anode surface are electrochemically
oxidized, and (3) formation of -OH as a result of
oxygen (O2) evolution in the anodic region. Also,
•OH may be produced electrochemically by using
metallic iron as a sacrificial anode and by producing
H2O2 at the cathode through electrolytic reduction of
O2 generated in the anodic region (electro-Fenton
process).
No commercial-scale electrochemical oxidation
systems are currently available. However,
pilot-scale applications indicate that the
electrochemical oxidation process has significant
potential for treating contaminated water.
SCWO Process
SCWO involves oxidation of organics in water at
temperatures and pressures above the critical point
of water (the critical temperature is 374 °C, and the
critical pressure is 22 megapascals) in the presence
of an oxidant. Two oxidants commonly used in
SCWO are O2 and H2O2. When O2 is the oxidant,
free radicals are initially formed by removal of a
hydrogen atom from the weakest C-H or O-H bonds
of organic compounds present in contaminated
water. This step is followed by several reactions
involving organic radicals and O2. H2O2 and organic
hydroperoxides formed in these reactions
decompose to form -OH. When H2O2 is the oxidant
used in SCWO, decomposition of H2O2 under
supercritical conditions yields «OH.
Commercial-scale SCWO systems currently
available fortreatment of contaminated water include
the General Atomics and Chematur Engineering AB
(Chematur) Aqua Critox® SCWO systems.
The literature search conducted in developing this
handbook revealed a single reference indicating that
•OH may also be produced under subcritical
conditions. However, this handbook does not
discuss this process, which is known as wet air
oxidation, because significant information on the
effectiveness of the process is readily available in
many environmental engineering books, as wet air
oxidation has been in use for almost 40 years.
Cavitation Processes
Cavitation processes have been used for treatment
of contaminated water. Cavitation refers to
formation, growth, and implosive collapse of gas- or
vapor-filled cavities (bubbles) in a liquid matrix.
Collapse of the cavities produces localized
high-temperature (about 5,000 °C) and
high-pressure (about 50 megapascals) hot spots.
The extreme conditions generated during Cavitation
result in «OH formation. Methods for inducing
Cavitation include ultrasonic irradiation of water
(acoustic cavitation or sonolysis), high-voltage
discharge in water (electrohydraulic cavitation), and
creating a pressure differential (from below vapor
pressure to above vapor pressure conditions) in a
flowing water stream (hydrodynamic cavitation).
No commercial-scale acoustic cavitation or
electrohydraulic cavitation systems are currently
available. However, at the bench-scale level, the
ES-2
-------
acoustic cavitation process has been shown to be
effective in treating contaminated water. The
electrohydraulic cavitation process is still in the
developmental stage. The only commercial-scale
hydrodynamic cavitation system available is the
Oxidation Systems, Inc. (OSI), HYDROX system for
treatment of contaminated water.
Electrical Discharge-Based Nonthermal
Plasma Processes
Electrical discharge-based nonthermal plasma
processes have been used for treatment of
contaminated air. A nonthermal plasma is a plasma
in which the mean electron kinetic energy, or
temperature, is significantly higher than that of the
molecules in the bulk gas, which are at ambient
temperature. Traditionally, nonthermal plasmas are
produced by a gas discharge under a strong electric
field. Under these conditions, both the electrons and
ions are accelerated to high energies (several
electron volts); however, because electrons have
longer mean free pathlengths and lighter mass, they
are typically accelerated to much higher energies
than the ions.
Nonthermal plasmas can be generated by E-beam
irradiation or electrical discharge. The main
difference between these two processes involves
the location where the high-energy electrons are
generated. In the E-beam process, high-energy
electrons are produced in an electron accelerator
and then injected into a reaction chamber. A plasma
is formed as the high-energy electrons collide with
the molecules in the bulk gas. In the electrical
discharge-based process, high-energy electrons are
produced by an electric field generated between
high-voltage electrodes in a reaction chamber.
Specifically, free electrons gain kinetic energy as
they drift along the high-voltage region between the
electrodes, resulting in production of high-energy
electrons. As in the E-beam process, a plasma
forms as the high-energy electrons collide with the
molecules in the bulk gas.
No commercial-scale electrical discharge-based
nonthermal plasma systems are currently available.
However, at the bench-scale level, the electrical
discharge-based nonthermal plasma processes have
been shown to be effective in treating contaminated
air.
Gamma-Ray Process
The gamma-ray process has been used for
treatment of contaminated water, air, and soil.
Gamma rays are high-energy photons
(electromagnetic radiation) emitted by excited atomic
nuclei in transition to a state of lower excitation.
When gamma rays collide with irradiated water,
high-energy (secondary) electrons are generated
along the trajectory of the photons. The high-energy
electrons generated can initiate several thousand
reactions as they dissipate energy in irradiated
•water. The reactions cause formation of three
primary reactive species that can destroy organic
compounds (-OH, aqueous electrons, and hydrogen
radicals), thus making the gamma-ray process more
similar to the E-beam process than to photochemical
processes.
Gamma rays have a high penetration depth within
irradiated water. For example, a water depth of
about 76 centimeters (cm) is required to absorb
90 percent of a gamma-ray energy level of
1.25 million electron volts (MeV). Therefore, the
gamma-ray process can be used to treat flowing
waste streams as well as containerized liquid
wastes.
No commercial-scale gamma-ray systems are
currently available. However, at the bench-scale
level, the gamma-ray process has been shown to be
effective in treating contaminated water, air, and soil.
X-Ray Process
The x-ray process has been used for treatment of
contaminated wastes. X-rays are high-energy
photons generated by accelerating high-energy
(incident) electrons in the form of an E-beam against
a material with a high atomic number. X-rays are
emitted when the high-energy electrons are
decelerated in the nucleus field of the target atom in
the solid material and when electrons in the target
atom fall from one atomic shell to another. As with
gamma rays, when x-rays collide with irradiated
water, high-energy (secondary) electrons are
generated along the trajectory of the photons. The
high-energy electrons generated can initiate several
thousand reactions as they dissipate energy in
irradiated water. The reactions cause formation of
three primary reactive species that can destroy
organic compounds ('OH, aqueous electrons, and
hydrogen radicals), thus making the x-ray process
more similar to the E-beam process than to
photochemical processes.
ES-3
-------
Like gamma rays, x-rays have a high penetration
depth within irradiated water. For example, a 1-MeV
x-ray has an effective water penetration depth of
about 27 cm. Therefore, the x-ray process can also
be used to treat flowing waste streams and
containerized liquid wastes.
The x-ray process is still in the developmental stage,
so no commercial-scale x-ray systems are currently
available.
E-Beam Process
The E-beam process involves irradiation of water or
air with a beam of high-energy electrons produced
by an electron accelerator. Within the electron
accelerator, an electric current (beam current) is
passed through a tungsten filament in a vacuum to
produce a stream of electrons. This electron stream
is accelerated by applying an electric field at a
specified voltage and is focused into a beam by
collimating devices. In the E-beam process, the
mechanism of 'OH formation is determined by the
medium being irradiated. E-beam irradiation of
water causes formation of three primary reactive
species that can destroy organic compounds ('OH,
aqueous electrons, and hydrogen radicals). E-beam
irradiation of air causes formation of a nonthermal
plasma when high-energy electrons in the beam
react with the bulk gas molecules.
High-energy electrons generated in the plasma react
with wet air to form -OH. In dry-air applications,
high-energy electrons in the nonthermal plasma
react with O2 to form oxygen radicals. These
oxidizing radicals play an important role in
generation of O3 and the initial decomposition of
some types of organics.
The depth to which an E-beam can penetrate
irradiated water is significantly less than the depths
associated with gamma rays and x-rays. For
example, a 1-MeV electron deposits its energy in
water within a depth of 4 millimeters. As a result,
E-beams are typically used to treat contaminated
water of relatively shallow depths.
A commercial-scale E-beam system called the High
Voltage Environmental Applications, Inc. (HVEA),
E-beam treatment system is currently available for
treatment of contaminated water. In addition, the
E-beam process has been shown to be effective in
treating contaminated air at the pilot-scale level.
Contaminated Water Treatment
ANPO processes have been demonstrated to be
effective for treatment of contaminated water. Water
matrices to which ANPO has been applied include
(1) contaminated groundwater, (2) industrial
wastewater, (3) municipal wastewater,
(4) contaminated drinking water, (5) landfill leachate,
and (6) contaminated surface water. As shown
below, a number of ANPO processes have been
evaluated in terms of their effectiveness in treating
various waterborne contaminants. Of these
processes, the Fenton process has been evaluated
for the most contaminant groups. The electro-
hydraulic cavitation and x-ray processes appear to
have been evaluated for the fewest.
ES-4
-------
Contaminant Group
Volatile Organic
Compounds (VOC)
Semivolatile Organic
Compounds (SVOC)
Polychlorinated
Biphenyls (PCB)
Pesticides and
Herbicides
Dioxins and Furans
Explosives and Their
Degradation Products
Humic Substances
Inorganics
Dyes
ANPO Process Status for Gontamin^ted Water Treatirient
---.•••
Fenton
*
*
Q
O
Q
*
Q
O
•
O3/H202
•
O
Q
O
Q
Q
•
O
O
o
Electro-
chemical
Oxidation
Q
O
Q
Q
Q
Q
Q
O
•
Q
seWQ
Q
•
01
Q
Q
*
Q
*
O
Q
Acoustic
Cavitation
O
O
Q
O
Q
Q
O
O
o
Q
Hydro-
dynamic
Cavitation
*
*
*
Q
,*
Q
Q
Q
Q
Q
Gamma-
Ray
Q
O
Q
Q
Q
O
Q
Q
Q
•
E-Beam
*
Q
Q
Q
Q
*
Q
Q
•
•
Notes: * = Commercial-scale, • = Pilot-scale, O = Bench-scale, Q = Developmental
Table ES-1 at the end of this executive summary
presents commercial-scale performance and cost
data for contaminated water treatment using various
ANPO processes. This table shows that the Fenton
and hydrodynamic cavitation processes have been
found to be effective in treating various
contaminants. Other ANPO processes, including the
SCWO and E-beam processes, have also been
found to be effective, but for fewer contaminant
groups. The treatment costs vary widely depending
on the types and concentrations of contaminants
treated and the ANPO system used for treatment.
The information sources cited in this handbook
should be carefully reviewed before a cost
comparison is made because the cost estimates-
presented in the literature were not developed using
a consistent set of assumptions.
Contaminated Air Treatment
ANPO processes have been demonstrated to be
effective for treatment of contaminated air. Air
matrices to which ANPO has been applied include
(1) air stripper off-gas, (2) industrial emissions, and
(3) automobile emissions. As shown below, only a
limited number of ANPO processes have been
evaluated in terms of their effectiveness in treating
various airborne contaminants. Of these processes,
the electrical discharge-based nonthermal plasma
and E-beam processes have been evaluated for the
most contaminant groups, and the gamma-ray
process has been evaluated for the fewest.
Contaminant Group
VOCs
SVOCs
Nitrogen Oxides (NOx)
Sulfur Oxides (SOx)
ANPO Process Staftis for Contaminated Air Treatment
Nonthermal Plasma
O
O
0
Q
o
Gamma-Ray
O
Q
Q
Q
Q
E-Beam .
•
Q
•
•
Q
Notes: • = Pilot-scale, O = Bench-scale, Q = Developmental
ES-5
-------
No commercial-scale performance and cost data is
available for contaminated air treatment using AN PO
processes. However, at the pilot-scale level, the
E-beam process has been found to be effective in
treating various contaminant groups.
Contaminated Soil Treatment
ANPO processes have been demonstrated to be
effective for treatment of contaminated soil. As
shown below, only two ANPO processes have been
evaluated in terms of their effectiveness in treating
various soil contaminants. Of these processes, the
Fenton process has been evaluated for more
contaminant groups than the gamma-ray process.
Contaminant Group.
VOCs
SVOCs
PCBs
Pesticides and Herbicides
Dioxins
Explosives and Their
Degradation Products
ANPO Process Status for Contaminated Soil Treatment
Fenton
*
*
O
•
«
•
Gamma-Ray
Q
Q
O
Q
Q
Q
• = Pilot-scale, O = Bench-scale, Q = Developmental
Table ES-2 at the end of this executive summary
presents commercial-scale performance and cost
data for contaminated soil treatment using trie
Fenton process. This table shows that the Fenton
process has been found to be effective in treating
various contaminants.
ES-6
-------
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ES-9
-------
-------
Section 1
Introduction
Overthe past two decades, environmental regulatory
requirements have become more stringent because
of increased awareness of the human health and
ecological risks associated with environmental
contamination resulting from improper waste
management practices. In many cases,
conventional treatment processes, such as air
stripping, carbon adsorption, biological treatment,
and chemical oxidation using ozone (O3) or hydrogen
peroxide (H2O2), have limitations. For example,
stripping and adsorption merely transfer
contaminants from one medium to another, whereas
biological treatment and conventional chemical
oxidation have low removal rates for many
environmental contaminants, including chlorinated
organics. Therefore, various alternative treatment
processes have been developed over the last 10 to
15 years in order to cost-effectively meet
environmental regulatory requirements. One such
group of processes is commonly referred to as
advanced oxidation processes.
Advanced oxidation processes generally involve
generation and use of powerful but relatively
nonselective transient oxidizing species, primarily the
hydroxyl radical (-OH); in some vapor-phase
advanced oxidation processes, singlet oxygen atoms
(O[1D]) and oxygen radicals have also been
identified as the dominant oxidizing species.
Table 1-1 shows that -OH has the highest
thermodynamic oxidation potential, which is perhaps
why -OH-based oxidation processes have gained the
attention of many advanced oxidation process
developers. In addition, as shown in Table 1-2, most
environmental contaminants react 1 million to
1 billion times faster with -OH than with O3, a
conventional oxidant. -OH can be generated by both
photochemical processes (for example, ultraviolet
[UV] radiation in combination with O3, H2O2, or a
photosensitizer) and nonphotochemical processes
(for example, electron beam [E-beam] irradiation, O3
in combination with H2O2, or Fenton's reagent). This
handbook discusses the applicability of advanced
nonphotochemical oxidation (ANPO) processes for
treatment of contaminated water, air, and soil.
Similar information on advanced photochemical
oxidation processes is presented in a separate
document (EPA/625/R-98/004) published by the
U.S. Environmental Protection Agency (U.S. EPA)
(1998).
Table 1-1. Oxidation Potential of Several Oxidants in
Water
Oxidant Oxidation Potential (electron volt)8
•OH
0(1D)
03
H202
Perhydroxy radical
Permanganate ion
Chlorine dioxide
Chlorine
Oxygen (O2)
2.80
2.42
2.07
1.77
1.70
1.67
1.50
1.36
1.23
Note:
Source: CRC Handbook of Chemistry and Physics, 1985
This section discusses the purpose and scope
(Section 1.1) and organization (Section 1.2) of this
handbook.
1.1 Purpose and Scope
The primary purpose of this handbook is to
summarize commercial-scale ANPO system
performance and cost data for treatment of
contaminated water, air, and soil. In addition, it
presents similar information drawn from pilot- and
bench-scale evaluations of ANPO processes to
supplement the commercial-scale performance and
cost data. The handbook is intended to serve as an
ANPO reference document for remedial project
managers, on-scene coordinators, state and local
regulators, consultants, industry representatives, and
other parties involved in management of
contaminated water, air, and soil. Specifically, it
assists these intended users in evaluating the
applicability of ANPO processes and in selecting one
or more ANPO processes for site-specific evaluation.
For the purposes of this handbook, commercial-,
pilot-, and bench-scale systems are defined as
follows:
1-1
-------
Table 1-2. Overall Rate Constants for O3 and -OH
Reactions with Organic Compounds in Water
Overall Rate Constant ;
(liter per mole-second)°'b!
Compound Type
Acetylenes
Alcohols
Aldehydes
Alkanes
Aromatics
Carboxylic acids
Chlorinated alkenes
Ketones
Nitrogen-containing organics
Olefins
Phenols
Sulfur-containing organics
03
50
10-2to1
10
10'2
1 to 102
lO-MolO-2
10-1to103
1
10to102
1 to450x103
103
10to1.6x103
•OH
10Bto109
108to109 •
109
106to109
108to1010
107to109
109to1011
109to1010
108to10to
109to10"
109to1010
109to10'°
Note:
Sources: Cater and others, 1990; Dussert, 1997
Because the overall rate constants may not actually reflect
the effectiveness of the oxidant in question, the reader
should calculate the oxidant's effectiveness by taking
project-specific reaction conditions into consideration (for
example, water chemistry parameters such as pH).
A commercial-scale system is a system
manufactured by an ANPO process vendor and
available for purchase or leasing from the
vendor.
A pilot-scale system is a system designed and
fabricated by an engineering firm to (1) estimate
the performance and cost of a particular ANPO
process, (2) identify field operational problems of
the process and their resolutions, and
(3) evaluate scale-up requirements for
implementing the process. A commercial-scale
system is selected after the pilot-scale system
proves to be successful. !
A bench-scale system is a system that (1) is of
much smaller scale than commercial- and
pilot-scale systems, (2) is used to evaluate the
feasibility of a particular ANPO process, (3) is
used to gain insight into the process kinetics and
mechanisms, and (4) may be used to generate
a preliminary cost estimate for comparison with
the costs of alternative processes. A pilot-scale
evaluation of a system may follow successful
performance by a particular ANPO process at
the bench-scale level.
This handbook is not intended to summarize all the
ANPO performance and cost data available in the
literature. Rather, it is intended to present
information on state-of-the-art ANPO processes for
treating contaminated environmental media.
Commercial-scale ANPO system performance and
• cost data is presented in greater detail than
pilot-scale results because the handbook is intended
for environ mental practitioners. Similarly, pilot-scale
results are presented in greater detail than
bench-scale results. In addition, pilot- and
bench-scale results are presented only where they
supplement commercial-scale ANPO system
evaluation results or where they fill information gaps,
such as those associated with by-product formation.
This handbook does not address nonenvironmental
ANPO process applications. For example, it does
not discuss ANPO process applications in industrial
operations, such as use of an ultrasonic irradiation
process for surface cleaning or use of an
electrochemical process for electroplating.
Finally, the information included in this handbook is
derived from an extensive literature review, and thus
the level of detail presented varies depending on the
information sources available. Specifically, the
treatment costs included should be considered only
order-of-magnitude estimates because most of the
information sources do not state the assumptions
made in estimating treatment costs. To facilitate
quick ANPO process comparisons, cost estimates
from the literature were adjusted for inflation using
implicit price deflators for gross national product and
are presented in 2000 U.S. dollars herein. This
approach has been proposed by the U.S.
Department of Commerce and is used to estimate
financial assurance requirements under Resource
Conservation and Recovery Act Subtitle C as
documented in 40 Code of Federal
Regulations 264^42(b). Cost estimates reported in
currencies other than U.S. dollars were converted to
U.S. dollars using the exchange rates for the
appropriate years before adjusting the estimates for
inflation.
1.2 Organization
This handbook is divided into five sections and
one appendix. Section 1 presents an introduction to
the handbook. Section 2 provides background
information on various ANPO processes, typical
commercial-scale ANPO systems, and system
design considerations. Sections 3, 4, and 5
summarize ANPO system performance and cost
data for treating contaminated water, air, and soil,
respectively. References cited in each section are
; 1-2
-------
listed at the end of the section. The appendix
contains ANPO process vendor contact information.
To facilitate user access to information, the
handbook presents performance and cost data for
each environmental medium by matrix, contaminant
group, scale of evaluation, process evaluated, and
process vendor or proprietary system (see
Figure 1-1). For example, where performance and
cost data for water (the medium) is summarized,
groundwater (matrix 1) is discussed before other
matrices. For the groundwater matrix, volatile
organic compounds (VOC) or contaminant group 1
is discussed before other contaminant groups.
For the VOC contaminant group, commercial-scale
applications are summarized before pilot- and
bench-scale evaluations. Similarly, the
commercial-scale applications are organized by
ANPO process and by vendor or proprietary process.
If performance and cost data for a given process is
available for both in situ and ex situ applications, in
situ applications are discussed before ex situ
applications.
If bench-scale results for a particular contaminant
were derived using a synthetic matrix (for example,
distilled water spiked with target contaminants), the
results are included under the matrix that is
described first. For example, in general, bench-scale
results derived using synthetic wastewater are
presented under the groundwater matrix because the
groundwater matrix is the first matrix discussed in
Contaminated Water Treatment (Section 3).
However, bench-scale results for dye removal from
synthetic wastewater are not presented under the
groundwater matrix because no commercial- or
pilot-scale results are available for dye removal from
groundwater. Therefore, bench-scale results for dye
removal from synthetic wastewater are appropriately
presented under the industrial wastewater matrix.
Environmental
Medium
(Water)
I
\
Matrix 1
(Groundwater)
Matrix 2
(Industrial
Wastewater)
Matrix 3
(Municipal
Wastewater)
/
|\
/
Contaminant
Group 1
(VOCs)
Contaminant
Group 2
(Pesticides
and
Herbicides)
Contaminant
Group 3
(Inorganics)
/
V
\
Commercial
Scale
Pilot
Scale
Bench
Scale
{\
\
\
ANPO
Process 1
(Fenton)
ANPO
Process 2
(C-3/HA)
ANPO
Process 3
(Electrochemical
Oxidation)
I Vendor/Proprietary Process 1
Vendor/Proprietary Process 2
Note: The information in parentheses represents a typical example.
Figure 1-1. Performance and cost data organization.
1-3
-------
1.3 References
Cater, S.R., K.G. Bircher, and R.D.S. Stevens.
1990. "Rayox®: A Second Generation Enhanced
Oxidation Process for Groundwater
Remediation." Proceedings of a Symposium on
Advanced Oxidation Processes for the
Treatment of Contaminated Water and Air.
Toronto, Canada. June. i
CRC Handbook of Chemistry and Physics. 1985.
Edited by R.C. Weast, M.J. Astle, and W.H.
Beyer. CRC Press, Inc. Boca Raton, Florida.
Dussert, B.W. 1997. "Advanced Oxidation."
Industrial Wastewater. November/December.
Pages 29 through 34.
U.S. Environmental Protection Agency (U.S. EPA).
1998. "Handbook: Advanced Photochemical
Oxidation Processes." Office of Research and
Development, Washington, DC.
EPA/625/R-98/004. December.
1-4
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Section 2
Background
This section provides background information on
ANPO processes (Section 2.1), commercial-scale
ANPO systems (Section 2.2), and ANPO system
design considerations (Section 2.3). The level of
detail included in this section facilitates
understanding of the performance and cost data
included in Sections 3, 4, and 5 of this handbook.
For additional information, the references listed in
Section 2.4 should be consulted.
2.1 ANPO Processes
As described in Section 1, ANPO processes use
•OH generated by nonphotochemical means to
oxidize environmental contaminants. ANPO
processes can be broadly divided into the following:
(1) Fenton, (2) O3/H2O2, (3) electrochemical
oxidation, (4) supercritical water oxidation (SCWO),
(5) cavitation, (6) electrical discharge-based
nonthermal plasma, (7) gamma-ray, (8) x-ray, and
(9) E-beam. These ANPO processes and their
variations are briefly described below.
The literature search conducted in developing this
handbook revealed a single reference indicating that
•OH may be produced under subcritical conditions as
opposed to supercritical conditions (SCWO).
However, this handbook does not discuss this
process, which is known as wet air
oxidation, because significant information on the
effectiveness of the process is readily available in
many environmental engineering books, as wet air
oxidation has been in use for almost 40 years.
2.1.1 Fenton Process
The Fenton process has been used for treatment of
contaminated water and soil. The dark reaction of
ferrous iron (Fe[ll]) with H2O2 known as Fenton's
reaction (Fenton, 1894), which is shown in Equation
2-1, has been known for over a century.
Fe(ll)+ H202 -» Fe(lll)+ OH' + .OH (2-1)
The *OH thus formed can react with Fe(ll) to produce
ferric iron (Fe[lll]) as shown in Equation 2-2.
•OH+Fe(ll)-»Fe(IM) + OH-
(2-2)
Decomposition of H2O2 is also catalyzed by Fe(lll)
(Walling, 1975). In this process, H2O2 is
decomposed to the water molecule (H2O) and O2,
and a steady-state concentration of Fe(ll) is
maintained during the decomposition. This process
is shown in Equations 2-3 and 2-4.
Fe(lll)+H202 o [Fe(lll)...02H]2++H*
o Fe(ll)+H02'+H+
H02< + Fe(IH)-»Fe(ll)+H++02
(2-3)
(2-4)
Alternatively, the «OH can react with and initiate
oxidation of organic pollutants present in a waste
stream.
The Fe(ll) ions react with H2O2to generate -OH (see
Equation 2-1), which, then reacts with organic
pollutants. However, the initial rate of removal of
organic pollutants by the Fe(lll)/H2O2 reagent is
much lower than that for the Fe(ll)/H2O2 reagent,
perhaps because of the lower reactivity of Fe(lll) with
H2O2 (Pignatello, 1992).
Although -OH is generally believed to be the primary
oxidant in the Fenton process, some researchers
argue that oxidation may not always involve -OH.
For example, Wink and others (1994) indicate that a
metallo-oxo species such as the solvated ferryl ion is
the primary oxidant in the Fenton process-mediated
oxidation of N-nitrosodimethylamine.
The pH of the reaction system is an important
parameter for the Fenton process. The classic
Fenton process is most effective at a pH level of
about 3 (Walling, 1975; Pignatello, 1992). Although
the process may be carried out at higher pH levels
(up to a pH of 6), the reaction rates are generally
expected to decrease with increasing pH
(Lipczynska-Kochany and others, 1995). In addition,
increasing the pH from 3 to 6 decreases the solubility
of iron species, which in turn decreases the amount
of iron available for the Fenton process and results
in significant sludge generation problems.
A variation of the classic Fenton process known as
the electro-Fenton process has been examined by
many researchers. The electro-Fenton process
combines an electrochemical process with the
Fenton process and involves reactions between
Fe(ll) and H2O2, in which at least one of the two
reactants is produced electrolytically. Chou and
others (1999) summarize four methods by which the
electro-Fenton process can be carried out. In the
first method, H2O2 is added to a system, and Fe(ll) is
2-1
-------
produced by oxidation of metallic iron (Fe[0]) at a
sacrificial anode. In the second method, Fe(ll):is
added to the system, and H2O2 is electrogenerated
at the cathode by reducing the dissolved 02
produced at the anode through electrolysis of H2O.
The third method involves a Fenton sludge recycling
(FSR) system, which contains two reactors. In this
method, both H2O2 and Fe(ll) are added to the first
reactor. The ferric (III) hydroxide sludge generated
in the first reactor is electrolytically reduced to Fe(ll)
in the second reactor. Fe(ll) is then recycled to the
first reactor, thereby eliminating the need for
continuous addition of Fe(ll). The fourth method is
similar to the third method except that (1) Fe(lll) is
added to the system instead of Fe(ll) and (2) both
the classic Fenton process and sludge recycling are
carried out in one reactor. '.
Another variation of the classic Fenton process is
studied by Lin and others (1997). The study shows
that combining the Fenton process with ultrasonic
irradiation process yields greater removal and
mineralization of 2-chlorophenol (CP). The
researchers, however, do not offer a theoretical
basis for combining the Fenton and ultrasonic
irradiation processes. It is unclear whether the
greater removal and mineralization are associated
with synergistic effects or with simple additive effects
of the Fenton and ultrasonic irradiation processes.
Falcon and others (1995) identify a Wet Peroxide
Oxidation® (WPO®) process in which the Fenton
reaction occurs at temperatures slightly below
100 °C. Regarding oxidation of carboxylic acids, the
researchers observed that the WPO® process could
be catalyzed using a combination of the transition
metal ions copper (II), manganese (II); and Fe(ll).
The research results show the significant synergistic
effect ofthe three transitional catalysts.
The ability of a process to overcome pH constraints
is important in environmental applications because
highly acidic pH conditions may not occur in the
environment and because adjusting the pH to very
low levels for in situ soil treatment could damage the
subsurface microbial community. Tachiev and
others (1998) identify a modified Fenton process
called the complexed metal hydrogen peroxide
oxidation system (CMHPOS) that can be
successfully used up to a pH of 9. The CMHPOS
uses soluble iron complexes to carry out the Fenton
reaction in both the acidic and neutral pH ranges.
Specifically, the researchers show that ligands such
as ethylenediaminetetraacetic acid (EDTA),
ethylenebis(oxyethylenenitrilo)tetraacetic acid, and
diethylenetriaminepentaacetic acid form soluble
complexes with Fe(ll) and Fe(lll) and therefore do
not require highly acidic conditions for the Fenton
process to be effective. McKinzi and Dichristina
(1999) identify a microbially driven Fenton process
that may be used under neutral pH conditions. In
this process, the facultative anaerobe Shewanella
putrifaciens strain 200 is used as a catalyst for both
Fe(lll) reduction and H202 production as the process
alternates between anaerobic and aerobic
conditions. Fe(lll) is reduced under anaerobic
conditions to produce Fe(ll), and H2O2 is produced
as a metabolic by-product under aerobic conditions,
thus eliminating the need for a continuous supply of
Fe(ll) and external addition of H2O2.
Some compounds commonly present in water may
react with the reactive species formed by the Fenton
process, thereby exerting an additional demand for
reactive species on the system. These compounds
are called scavengers, and they can potentially
impact system performance. A scavenger is defined
as any compound in- water other than the target
contaminants that consumes reactive species.
Carbonate and bicarbonate ions are examples of
•OH scavengers found in most natural waters and
wastewaters. Therefore, alkalinity is an important
system operating parameter. If alkalinity is high,
influent pH adjustment may be required to shift the
carbonate-bicarbonate equilibrium from carbonate (a
scavenger) to carbonic acid (not a scavenger).
According to Lipczynska-Kochany and others (1995),
several anions such as phosphate, sulfate, and
chloride, which are typically present in groundwater
and surface water, could interfere with the
effectiveness of the Fenton process. Specifically,
(1) phosphate acts as a -OH scavenger; (2) sulfate,
which reacts with Fe(ll) to form an ion pair (FeSO4°),
inhibits -OH formation; and (3) chloride, which also
reacts with Fe(ll) to form an ion pair (FeCI+ or
FeCI2°), acts as both a -OH formation inhibitor and a
•OH scavenger. However, the radicals that are
formed upon reaction of «OH with phosphate and
sulfate are also fairly reactive with organic
contaminants.
Commercial-scale Fenton systems are currently
available from the following vendors:
(1) Geo-Cleanse International, Inc. (Geo-Cleanse);
(2) In-Situ Oxidative Technologies, Inc. (ISOTEC);
(3) Mantech Environmental Corporation; and
(4) H&H Eco Systems, Incorporated (H&H). The
Geo-Cleanse®, ISOTEC™, and Mantech
Environmental Corporation CleanOX® systems can
be applied in situ to treat contaminated water and
soil. The Geo-Cleanse® system can also be applied
ex situ to treat contaminated soil. The H&H Solid
State Chemical Oxidation™ (SSCO™) system is
applicable for ex situ treatment of soil. At the time of
2-2
-------
this handbook's preparation, no information was
available on the CleanOX® system.
Geo-Cleanse offers an ex situ soil treatment system
and an in situ soil and groundwater treatment
system. The Geo-Cleanse® systems are based on
the classic Fenton process in which Fe(ll) is added
in the form of ferrous sulfate (FeSO4). Sulfuric acid
(H2SO4) (66 percent technical grade) and phosphoric
acid (H3PO4) (85 percent technical grade) may be
added to adjust the pH to the range of 5.5 to 6.0. A
stabilizer (calcium phosphate) is also added to delay
formation of free radicals.
The ISOTEC™ system is an in situ soil and
groundwater treatment system based on a modified
Fenton process. Specifically, ISOTEC uses
proprietary catalysts composed of active
components that chelate iron and keep it in dissolved
form as an organometallic complex. The form of iron
and the chelating components used are site-specific
and are determined through bench- and pilot-scale
evaluations. According to ISOTEC, its modification
to the classic Fenton process allows the system to
effectively treat organic contaminants in the
subsurface over a pH range of 2.5 to 8.5. ISOTEC
also uses proprietary stabilizer and mobility control
agents that control formation of «OH and dispersion
of its precursors.
The H&H SSCO™ system is an ex situ soil treatment
system based on a modified Fenton process. In the
SSCO™ process, Fe(0), which is available as a
waste product from many industries, is applied in
powder form to soil in order to carry out the Fenton
reaction. According to H&H, Fe(0) is a more
economical form of iron than Fe(ll) or Fe(lll). The
SSCO™ process is carried out at pH values ranging
from 5 to 7.
+ ->HO
2.1.2
O/H2O2 Process
The O3/H2O2 process, also known as the peroxone
process, has been used for treatment of
contaminated water. H2O2 can initiate decomposition
of O3 with the hydroperoxide ion (HO2~) as shown in
Equations 2-5 and 2-6 (Pedit and others, 1997).
H,0,
HO2-+O3
• O3"+HO2'
(2-5)
(2-6)
The products of Equation 2-6 participate in -OH
formation as shown in Equations 2-7a and 2-7b and
2-8a through 2-8d.
HO
H02* .
O2" +O3
HO3'
'~ +O
(2-7a)
(2-7b)
(2-8a)
(2-8b)
(2-8c)
(2-8d)
These equations may be combined to represent the
overall reaction between H2O2 and O3 that yields
•OH as shown in Equation 2-9.
H2O2 + 2O3 -> 2 «OH + 3O2
(2-9)
Equation 2-9 shows that two -OH are formed for
each mole of H2O2 reacting with two moles of O3.
However, H2O2 present in quantities significantly
exceeding stoichiometry is known to scavenge -OH
and thereby reduce the overall effectiveness of the
O3/H2O2 process (Glaze and others, 1987).
Alkalinity is an important parameter in the O3/H2O2
process. If alkalinity is high, influent pH adjustment
may be required for irradiated matrices to shift the
carbonate-bicarbonate equilibrium from carbonate (a
scavenger) to carbonic acid (not a scavenger).
Additional, detailed information on alkalinity effects
is provided by Peyton and others (1998).
Processes that use O3 or H2O2 alone to produce -OH
have also been studied by several researchers. -OH
has been proposed as an intermediate in the
base-catalyzed decomposition of O3 (Weiss, 1935).
Such an O3 process is typically carried out at high pH
because increasing the pH is expected to increase
•OH formation. However, increasing the pH results
in a shift in bicarbonate to carbonate species in
water containing bicarbonate or carbonate alkalinity.
Because the rate constant for the reaction of *OH
with the carbonate ion is about 28 times greater than
that with the bicarbonate ion (Buxton and others,
1988), the O3 process is not considered to be a
viable alternative. Urs and Hoigne (1998) identify a
process known as the carbozone process that is
based on activated carbon catalyzed conversion of
O3 to form 'OH. Their study shows that the
stoichiometric yield of -OH is comparable to that of
non-activated carbon-catalyzed processes and that
the carbozone process proceeds at a greater rate.
Martinez and others (1993) study the effect of adding
H2O2 to an air stream at high temperatures (greater
than [>] 600 °C). The study shows that H2O2 can be
thermally dissociated in air streams at such
temperatures to form -OH.
2-3
-------
The only commercial-scale O3/H2O2 system available
is the Applied Process Technology, Inc., HiPOx™
water treatment system. At the time of this
handbook's preparation, no information was
available on the HiPOx™ system. However, several
pilot-scale studies indicate that the O3/H2O2 process
has significant potential for treating contaminated
water.
2.1.3 Electrochemical Oxidation
Process
The electrochemical oxidation process has been
used for treatment of contaminated water. In this
process, electricity flows through an electrochemical
reactor consisting of electrodes separated by an
ionic conductor (electrolyte). Oxidation and
reduction reactions are carried out on the surface of
the electrodes at the electrode-electrolyte interface.
The electrode at which oxidation occurs is called the
anode, whereas the reduction reaction occurs at the
cathode.
Multiple reaction pathways have been proposed for
•OH formation during electrochemical oxidation.
According to Huang and Chu (1991), hydroxide ions
(OH~) are oxidized to -OH in the anodic region under
alkaline conditions. Polcaro and others (in press)
and Comninellis and Pulgarin (1993) report that -OH
are formed when the organic compounds present in
the waste stream or H2O adsorbed on the anode
surface are electrochemically oxidized. In addition,
•OH can be formed as a result of O2 evolution in the
anodic region (Polcaro and Palmas, 1997). -OH may
also be produced electrochemically by using Fe(0)
as a sacrificial anode and producing H2O2 at the
cathode through electrolytic reduction of O2
generated in the anodic region. This process is
referred to as the electro-Fenton process. Variations
of the electro-Fenton process are described in
Section 2.1.1.
Recent studies indicate that -OH formation depends
on the chemical or electrochemical characteristics of
the anode. According to Rodgers and others (1999),
oxidation by -OH occurs when tin dioxide and iridium
dioxide anodes are used, whereas direct electron
transfer oxidation occurs when a lead dioxide anode
is used. Koppang and others (1999) also report that
•OH formation occurs during the initial stage of O2
evolution at the nondiamond carbon impurity sites of
a diamond thin-film electrode.
Commercial-scale electrochemical oxidation systems
are not currently available. However, pilot-scale
applications show that the electrochemical oxidation
process has significant potential for treating
contaminated water.
2.1.4 SCWO Process
As the name implies, the SCWO process has been
used for treatment of contaminated water. SCWO
involves oxidation of organics in water at
temperatures and pressures above the critical point
of water (critical temperature = 374 °C and critical
pressure = 22 megaPascals [MPa]) in the presence
of an oxidant such as O2 or H2O2. In practice,
SCWO is typically performed at 400 to 650 °C and at
25 MPa (Schwinkendorf and others, 1995b). Below
the critical point, the liquid and gas phases of water
can coexist in equilibrium. Above the critical point,
water exists in only one phase, which is referred to
as supercritical water.
The physical and chemical properties of supercritical
water, including its de.nsity, viscosity, diffusivity, ion
mobility, and dielectric constant, are markedly
different from those of water under standard
conditions (Gloyna and Li, 1998). For example, at
25 MPa the density of water decreases from 0.507 to
0.0786 gram per cubic centimeters as the
temperature increases from 375 to 550 °C; similarly,
the viscosity of water decreases from 10,000 to
597 micropoises as the temperature increases from
375 to 450 °C. As a result, diffusivity and ion
mobility are higher under supercritical conditions. In
addition, as the density of water decreases,
hydrogen bonding and the solvent polarity of water
decrease. The dielectric constant (relative
permittivity), which is a measure of hydrogen
bonding and reflects the polarizability of molecules,
decreases from 78.5 at standard temperature and
pressure to about 5 under supercritical conditions.
Supercritical water thus exhibits properties similar to
those of a nonpplar organic solvent. As a result,
sparingly soluble, nonpolar organic compounds and
oxidants become highly soluble in or even miscible
with supercritical water. Consequently, an SCWO
system is not mass transfer-limited and has high
reaction rates.
SCWO of organics proceeds by means of a
free-radical reaction mechanism. Because of the
elevated temperatures that occur under supercritical
conditions, pyrolysis can occur; however, pyrolysis is
not likely to be a major pathway for contaminant
destruction because its reaction rates are orders of
magnitude lowerthan those of oxidation (Gloyna and
Li, 1995). Two oxidants commonly used in SCWO
are O2 and H2O2 (Gloyna and Li, 1998). When O2 is
the oxidant, free radicals are initially formed by
removal of a hydrogen atom from the weakest C-H
2-4
-------
or O-H bonds of the organic compounds present in
the contaminated water. This step is followed by
several reactions involving organic radicals and O2
as shown in Equations 2-10 through 2-13 (Duffy and
others, 2000).
->R'+H02*
R'+O2-»ROO'
RH+HOZ* -»R'+H2O2
RH+ROO- ->R'+ROOH
(2-10)
(2-11)
(2-12)
(2-13)
H2O2 and organic hydroperoxides formed in these
reactions decompose to form -OH as shown in
Equations 2-14 through 2-16.
H202 + (C)-» 2.0H+(C) (2-14)
H2O2 + Mn+ -4 «OH+ OH- + M(n+1)+ (2-15)
ROOH + (C)->RO'+OH+(C) (2-16)
The collision partner (C) may be a homogenous (for
example, H2O) or heterogenous (for example, the
surface of the reactor or the soil or sediment)
material. The influence of the transition metal (M) in
•OH formation is represented in Equation 2-15.
When H2O2 is the oxidant used in SCWO, 'OH
formation may also involve homogenous or
heterogenous material as depicted in Equations 2-14
and 2-15. Additionally, the thermal decomposition of
H2O2 that yields -OH may be significant under
supercritical conditions.
As discussed in Section 2.1.1, alkalinity is an
important operating parameter in oxidation
processes involving -OH. Therefore, if alkalinity is
high, influent pH adjustment may be required in
SCWO applications to shift the
carbonate-bicarbonate equilibrium from carbonate (a
scavenger) to carbonic acid (not a scavenger).
However, when the impact of -OH scavengers is
estimated and the influent characteristics are
adjusted to minimize the scavenger impact, the
influence of supercritical conditions (high
temperature and pressure) on all relevant chemical
reactions should be considered.
Use of catalysts to improve the efficiency of the
SCWO process has been studied by several
researchers. According to Gloyna and Li (1998), use
of catalysts has resulted in enhanced oxidation of
complex organic compounds at relatively low SCWO
temperatures (380 to 450 °C) and short reactor
residence times (typically less than [<] 30 seconds).
Frisch and others (1994) report higher removals for
dimethyl methane phosphoric acid ester (DMMP),
hydroxymethylprogesterone acetate, and methanol
in the presence of a platinum catalyst. Catalysts are
coated on support materials in SCWO applications
because of the extreme conditions involved.
Therefore, the stability and binding characteristics of
the catalyst are important design considerations.
Catalyst support materials studied by Frisch and
others (1994) that are stable under supercritical
water conditions include zirconia, titania, hafnia, and
alpha-aluminum.
Commercial-scale SCWO systems currently
available for treatment of contaminated water include
the General Atomics and Chematur Engineering AB
(Chematur) Aqua Critox® SCWO systems.
2.1.5 Cavitation Processes
Cavitation processes have been used for treatment
of contaminated water. Cavitation refers to
formation, growth, and implosive collapse of gas- or
vapor-filled cavities (bubbles) in a liquid matrix.
Collapse of the cavities produces localized
high-temperature (about 5,000 °C) and
high-pressure (about 50 MPa) hotspots (Suslick,
1990). The extreme conditions generated during
cavitation result in «OH formation. Methods for
inducing cavitation include ultrasonic irradiation
(acoustic cavitation or sonolysis), high-voltage
discharge (electrohydraulic cavitation), and pressure
differential (hydrodynamic cavitation). These
methods are discussed below.
2.1.5.1 Acoustic Cavitation
Ultrasound spans the frequencies from about 20
kilohertz (kHz) to 10 megahertz; human hearing has
an upper limit of 18 kHz. Ultrasound travels through
water as a cycle of expansion (negative pressure)
and compression (positive pressure) waves induced
in the molecules. When water containing small, solid
particles with gas-filled crevices is exposed to a
negative pressure cycle, the reduced pressure
induces the gas to expand until a small, gas-filled
cavity is formed and released into the surrounding
water.
The cavities grow as they absorb energy from
alternating compression and expansion waves. As
energy is absorbed, the surface area of a cavity
increases during expansion cycles and decreases
during compression cycles. Because the amount of
gas that diffuses into or out of the cavity depends on
the surface area of the cavity, diffusion into the cavity
during expansion cycles exceeds diffusion out of the
cavity during compression cycles. Over successive
2-5
-------
cycles, the cavity eventually reaches a critical size at
which it can no longer absorb energy efficiently
enough to sustain itself. At this point, the cavity
cannot withstand the net external pressure of the
surrounding water, and the cavity implodes in a very
short time (about 10 microseconds).
When the cavity implodes, the gas inside the cavity
is compressed, generating intense heat. This heat
raises the temperature and pressure of the gas
phase in the collapsing cavity and the water
immediately surrounding it. After the violent
collapse of the cavity, the surrounding water rushes
into the cavity. Suslick (1989) reports that the
gaseous contents of a collapsing cavity reached a
temperature of about 5,500 °C and that the water
immediately surrounding the cavity reached a
temperature of 2,100 °C. Because the cavities are
small (10 to 200 microns in diameter) and have short
lifetimes (about 10 microseconds), the temperature
and pressure conditions of the bulk water are
relatively unaffected (Cropek and Kemme, 1996).
As a result of the localized temperature and pressure
conditions produced during cavitation, chemical
reactions are believed to occur in two distinct
regions: (1) the gas phase in the center of a
collapsing cavity and (2) a thin layer of supercritical
water (see Section 2.1.4) surrounding the cavity
(Hoffman and others, 1997). Three thermal
destruction pathways are believed to occur in a
system subjected to ultrasonic irradiation: oxidation
by -OH formed by thermolysis of H2O, SCWO, and
pyrolysis (Cropek and Kemme, 1996). These
pathways are described below.
The primary pathway for destruction of organic
compounds by cavitation is believed to be oxidation
by «OH. Generation of -OH by thermolysis of H2C>
in the gas phase is described by Equation 2-17 (Hua
and Hoffman, 1997). ;
(2-17)
Because -OH are highly reactive, these radicals do
not have a long travel pathlength into bulk water. As
a result, destruction of organic compounds by -OH
mainly occurs in the cavity or near the cavity surface.
However, Petrier and others (1992) suggest that at
higher frequencies, -OH are ejected from the
collapsing cavity before they can recombine in the
gas phase because the collapse times at higher
frequencies are shorter. Because -OH can escape
into the bulk water and because contaminant
oxidation could occur in the bulk water at a short
distance from the cavity, alkalinity is an important
acoustic cavitation parameter. If alkalinity is high,
influent pH adjustment may be required to shift the
carbonate-bicarbonate equilibrium from carbonate (a
scavenger) to carbonic acid (not a scavenger).
As discussed in Section 2.1.4, supercritical water
conditions increase the solubility of organic
compounds, and if an oxidant is present in the
supercritical water shell surrounding the cavity,
SCWO of organics occurs. However, because the
volume of the cavity is estimated to be about 20,000
times greater than the volume of the thin supercritical
water shell, the value of SCWO in acoustic cavitation
may be limited to increasing the solubility of the
organic contaminant near the cavity surface for OH
attack (Cropek and Kemme, 1996).
Pyrolysis is defined as thermal destruction of organic
compounds in the absence of O2. According to
Cropek and Kemme (1996), for pyrolysis to occur
during acoustic cavitation, organic contaminants
must be present in the gas phase inside the cavity.
Organic contaminants with higher vapor pressures
(such as VOCs relative to semivolatile organic
compounds [SVOC]) will be present at higher
concentrations inside the cavity. Therefore, pyrolysis
is expected to be more prevalent as the vapor
pressure of an organic compound increases. In
addition, according to Hoffman and others (1997),
pyrolysis is the predominant pathway of organic
destruction in waste streams with high contaminant
levels, whereas -OH is the predominant pathway in
waste streams with low contaminant levels.
An alternative but less accepted explanation of
acoustic cavitation is based on several electrical
theories. According to these theories, charges could
build up on opposite faces of a cavity as it is formed.
The cavity undergoes pulsations and then
fragmentation, during which intense electrical fields
are generated. Electrified sprays of the surrounding
liquid are injected into the cavity during its collapse
(Lepoint and Mullie, 1994). Because of the intensity
of the electrical phenomenon, partial ionization (-OH
formation) of the cavity's contents is believed to
occur. This partial ionization is similar to cold
plasma chemistry in that aqueous electrons (e~aq),
hydrogen radicals (H-), and -OH will likely be formed.
However, Misik and Riesz (1997) recently
demonstrated that e~aq are not formed during
sonolysis of water.
The effect of adding saturating gases to the acoustic
cavitation process has been studied by several
researchers. Saturating gases are generally used to
produce acoustic cavitation at lower acoustic power
than would otherwise be necessary. Hua and
Hoffman (1997) have studied the effects of four
2-6
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different saturating gases: krypton, argon, helium,
and O2. Of the four gases, krypton is the least
heat-conducting and most soluble; therefore,
collapsing cavities containing krypton produce the
most extreme transient temperatures, resulting in the
highest production of H2O2 and -OH.
The combination of acoustic cavitation with
ozonolysis, which is commonly referred to as the
sonozone process, has been studied by many
researchers (Olson and Barbier, 1994; Weavers and
others, 1998). During acoustic cavitation, O3 is
thermolytically decomposed in the gas phase of a
cavity to form two 'OH. In addition, because
sonication of ozonated water (relative to oxygenated
water) produces H2O2, -OH also form near the
surface of the cavity and in bulk water as a
result of the reaction between O3 and H2O2 (see
Section 2.1.2).
Commercial-scale acoustic cavitation systems are
not currently available. However, acoustic cavitation
has been shown to be effective at the bench-scale
level in treating contaminated water.
2.1.5.2 Electrohydraulic Cavitation
Electrohydraulic cavitation is produced by injecting
energy into a liquid matrix through a plasma channel.
A plasma is a highly ionized gas composed of a
nearly equal number of positive and negative free
charges (positive ions and electrons). The plasma
channel is formed by sending short pulses (of about
20 microseconds) of a high-current, high-voltage
electrical discharge between two electrodes
submerged in water (Lang and others, 1998). As the
plasma channel expands, it produces a
high-pressure (>1,400 MPa) Shockwave (Hoffman
and others, 1997). If the Shockwave is reflected
back from a material in the reaction vessel with a
different acoustic impedance, cavitation will occur.
The pathways for destruction of organic compounds
by acoustic cavitation discussed in Section 2.1.5.1
also apply to electrohydraulic cavitation.
Additional reaction mechanisms result from
formation of the plasma channel. According to
Robinson and others (1973), the plasma channel can
reach temperatures exceeding 13,000 °C. As a
result, organic compounds can be destroyed directly
by pyrolysis within the plasma channel. However,
because the plasma channel occupies only a small
volume (about 1 to 3 milliliters [mL]), pyrolysis is not
considered to be a primary reaction pathway (Lang
and others, 1998). The extreme temperatures that
exist in the plasma channel also cause the channel
to function as a black body radiation source with
maximum emittance in the vacuum UV region of the
light spectrum (Robinson and others, 1973). As a
result, soft x-rays and high-energy UV radiation
emitted from the plasma channel into the bulk liquid
serve as a source of -OH (Hoffman and others,
1997). Steam bubbles also form as thermal energy
is transferred from the plasma channel to the bulk
liquid (Buntzen, 1962). These bubbles, which exhibit
the extreme temperature and pressure conditions
observed in supercritical' water (Ben'Kovskii and
others, 1974), serve as an additional location for
SCWO to occur.
Hoffman (1992) briefly describes two additional
approaches to inducing electrohydraulic cavitation:
(1) spark-gap discharge in water and (2) pulsed or
continuous ultrasonic irradiation. In spark-gap
discharge, energy stored in a condenser is released
within several microseconds; this release of energy
results in an arc across the electrode gap that
vaporizes the water around the path of the arc,
establishing a rapidly expanding plasma. Formation
of plasma using ultrasonic irradiation is explained in
Section 2.1.5.1.
Commercial-scale electrohydraulic cavitation
systems are not currently available. However, the
electrohydraulic cavitation process has been shown
to be effective at the bench-scale level in treating
contaminated water.
2.1.5.3 Hydrodynamic Cavitation
When water flows into a region where the pressure
is at or below the vapor pressure of water
(2.5 kiloPascals [kPa] at 20 °C), gas cavities form as
the water starts to vaporize (Hicks and Edwards,
1971). When the cavities move into a region where
the pressure exceeds the vapor pressure of water,
the cavities collapse. The contaminant destruction
pathways associated with hydrodynamic cavitation
are expected to be similar to those described for
acoustic cavitation in Section 2.1.5.1.
A commercial-scale hydrodynamic cavitation system
called the Oxidation Systems, Inc. (OSI), HYDROX
system is currently available for treatment of
contaminated water.
2.1.6 Electrical Discharge-Based
Nonthermal Plasma Processes
Electrical discharge-based nonthermal plasma
processes have been used for treatment of
contaminated air. A nonthermal plasma is a plasma
in which the mean electron kinetic energy, or
temperature, is significantly higher than that of the
2-7
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molecules in the bulk gas, which are at ambient
temperature. Traditionally, nonthermal plasmas are
produced by a gas discharge under the application
of a strong electric field. Under these conditions,
both the electrons and ions are^accelerated to high
energies (several electron volts); however, because
electrons have longer mean free pathlengths and
lighter mass, they are typically accelerated to much
higher energies than the ions (Vercammen and
others, 1997).
The high-energy electrons in the nonthermal plasma
react with wet air to form -OH by (1) electron
attachment to H2O, (2) direct dissociation of H2O by
electrons, (3) dissociation of H2O by O(1D), and
(4) dissociation of H2O by excited oxygen atoms
(Penetrante, 1993a). These reactions are describee!
in Equations 2-18 through 2-25.
Electron attachment to H2O:
e"+H2O-»H~+»OH
Direct dissociation of H2O by electrons:
e~+H2O-»e-+»H+»OH
Dissociation of H2O by O(1D):
(2-18)
O(1D)+H2O->2«OH
(2-20)
(2-2t)
Dissociation of H2O by excited oxygen atoms: ,
e- + 02 -> 2e- + 02+ (2-22)
(H30)+ + 02 + .OH (2-24)
O2*(H2o)+H20-» H30+(OH)+02 (2-25a)
or :
H30+(OH)+H20-»H3O++H20 + .OH ,
(2-25bj
!
In dry-air applications, high-energy electrons in the
nonthermal plasma react with O2 to form oxygen
radicals. These oxidizing radicals play an important
role in generation of O3 and the initial decomposition
of some types of organics. '
Nonthermal plasmas can be generated by E-beam
irradiation or electrical discharge. The main
difference between these two processes is where the
high-energy electrons are generated. In the E-beam
process, high-energy electrons are produced in an
electron accelerator and then injected into a reaction
chamber. A plasma is formed as the high-energy
electrons collide with the molecules in the bulk gas.
The E-beam process is discussed in more detail in
Section 2.1.9. In the electrical discharge-based
process, high-energy electrons are produced by an
electric field generated between high-voltage
electrodes within a reaction chamber. Specifically,
free electrons gain kinetic energy as they drift along
the high-voltage region between the electrodes,
resulting in production of high-energy electrons. As
in the E-beam process, a plasma forms as the
high-energy electrons collide with the molecules in
the bulk gas.
There are several types of electrical discharge-based
reactors that are distinguished by their electrode
configuration and electrical power supply. Types of
electrical discharge-based reactors commonly used
to treat contaminated air include pulsed corona,
dielectric-barrier discharge, surface discharge, and
ferroelectric packed bed reactors. These reactors
are briefly described below. More information is
provided by Vercammen and others (1997) and
Penetrante (1993b).
In a pulsed corona reactor, at least one of the
electrodes is a thin wire, is a needle, or has a sharp
edge. The other electrode can be a plate or a
cylinder. Short electrical pulses (of <1 microsecond)
of high voltage are sent between two electrodes to
produce electrical discharges (coronas) (Vercammen
and others, 1997), which in turn produce short-lived
plasmas.
In dielectric-barrier discharges reactors, also known
as silent discharge reactors, one or both of the
electrodes are covered with a thin, dielectric
material. Dielectric materials are electrical insulators
in which an electrical field can be sustained with
minimum dissipation of power. Materials with high
dielectric strength and a high dielectric constant,
such as glass and aluminum, are used (Vercammen
and others, 1997). A high-voltage alternating current
is applied between electrodes to produce electrical
discharges. When the electrical potential across the
discharge gap reaches breakdown voltage, the
dielectric material acts as a stabilizer, leading to
formation of a large number of micro-discharges of
short pulses that are spread over the discharge gap.
The discharge-generated ions traverse the gap in a
pulse and are stored at the surface of the dielectric
material. In the pulsed corona method, the transient
behavior of the plasma is controlled by the applied
voltage pulse, whereas the plasma in the
dielectric-barrier discharge method self-extinguishes
2-8
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when charge buildup on the dielectric material
reduces the local electric field.
A surface discharge reactor contains strip-like
electrodes on a ceramic, tubular or planar surface
and a film-like counterelectrode embedded inside the
ceramic. A high-frequency, high-voltage alternating
current is applied to generate a surface discharge
that appears at the sides of the strip-like electrode
and uniformly covers the ceramic surface.
The ferroelectric packed bed reactor is a tubular
reactor with high-dielectric ceramic pellets packed
between two metal mesh electrodes. When
high-voltage alternating current is applied, the pellets
are polarized, and the high-strength electric fields
that develop in the interstitial spaces between the
pellets form many pulsed discharges. Dielectric
materials commonly used for the pellets include
barium-titanium trioxide, strontium-titanium trioxide,
or lead-titanium trioxide (Penetrante, 1993b).
Commercial-scale electrical discharge-based
nonthermal plasma systems are not currently
available. However, the electrical discharge-based
nonthermal plasma processes have been shown to
be effective at the bench-scale level. in treating
contaminated air.
2.1.7 Gamma-Ray Process
The gamma-ray process has been used for
treatment of contaminated water and soil. Gamma
rays are high-energy photons (electromagnetic
radiation) emitted by excited atomic nuclei in
transition to a state of lower excitation. The most
common source of gamma rays is radioactive decay
of radioisotope cobalt 60 (60Co), which emits gamma
rays at energies of 1.17 and 1.33 million electron
volts (MeV) as it decays to nickel 60 and has a
half-life of 5.27 years. When gamma rays collide
with irradiated water, high-energy (secondary)
electrons are generated along the trajectory of the
photons through three processes: photoelectric
absorption, Compton scattering, and
electron-positron pair production. Information on
these processes is provided by Swallow (1973).
High-energy (secondary) electrons generated from
gamma rays can initiate several thousand reactions
as they dissipate energy in irradiated water. The
reactions cause formation of three primary reactive
species responsible for organic compound
destruction (-OH, e~aq, and H-). The overall reaction
for irradiation (gamma-ray, x-ray, and E-beam)
processes is described by Equation 2-26 (Cooper
and others, 1993).
H2O + high energy e ->e aq(2.6)
• H (0.55) +.OH (2.7)+
H2(0.45)+H202(0.71) +
H30+(2.7)
(2-26)
The relative concentrations of the reaction products
are presented as the "G values" shown in
parentheses in Equation 2-26. The G values
represent the numbers of free radicals, ions, or
molecules formed in water absorbing 100 electron
volts of energy. The G values in Equation 2-26
indicate the relative reaction product concentrations
10~7 second after high-energy electron impacts the
water. Because strong oxidizing species (*OH) and
strong reducing species (e~aq and -H) are formed in
about equal concentrations, multiple mechanisms for
organic compound destruction are provided by
irradiation processes. In this way, irradiation
processes differ from other processes that involve
free radical chemistry and that typically rely on a
single organic compound destruction mechanism,
usually involving -OH.
Alkalinity is an important parameter in the
gamma-ray process. If alkalinity is high, pH
adjustment may be required for the irradiated water
to shift the carbonate-bicarbonate equilibrium from
carbonate (a scavenger) to carbonic acid (not a
scavenger).
Gamma rays have a high penetration depth within
irradiated water. According to Gray and Cleland
(1998), studies indicate that a water depth of about
76 centimeters (cm) is required to absorb 90 percent
of a gamma-ray energy level of 1.25 MeV.
Therefore, the gamma-ray process can be used to
treat flowing waste streams as well as containerized
liquid wastes.
A variation of the gamma-ray process known as
radiocatalysis has been studied by Su and others
(1998). The study shows that the presence of
titanium dioxide (TiO2) catalyst and O2 increases
removal of EDTA. The increased removal of EDTA
is facilitated by electron/hole pair formation in TiO2
induced by gamma radiation, which is similar to an
advanced photocatalytic oxidation process such as
UV/Ti02.
Commercial-scale gamma-ray systems are not
currently available. However, the gamma-ray
process has been shown to be effective at the
bench-scale level in treating contaminated water and
soil.
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2.1.8 X-Ray Process \
i
The x-ray process has been used for treatment of
contaminated wastes. X-rays are high-energy
photons (electromagnetic radiatjon) generated by
accelerating high-energy (incident) electrons in the
form of an E-beam against a material with a high
atomic number. X-rays are emitted when the
high-energy (incident) electrons are decelerated in
the nucleus field of the target atom in the solid
material (bremsstrahlung x-rays) and when electrons
in the target atom fall from one atomic shell to
another (characteristic x-rays) (Swallow, 1973).
The energy of x-rays emitted is a function of the
energy of the high-energy (incident) electrons.
According to Bailey and Lackner (1995), in the x-ray
process, incident electron energy ranges from 8 tb
10 MeV in order to avoid nuclear activation and tb
maximize the conversion to x-ray energy. However,
the conversion is not efficient. Bailey and Lackner
(1995) report conversion efficiencies of 1.6,8.4, and
16.6 percent for 1-, 5-, and 10-MeV incident
electrons, respectively. i
As with gamma rays, when x-rays collide with
irradiated water, high-energy (secondary) electrons
are generated along the trajectory of the photons
through three processes: photoelectric absorption,
Compton scattering, and electron-positron pair
production. Information on these processes is
provided by Swallow (1973).
High-energy (secondary) electrons generated frorh
x-rays can initiate several thousand reactions as they
dissipate energy in the irradiated medium. The
reactions cause formation of three primary reactive
species responsible for organic compound
destruction ('OH, e~aq, and H-). The overall reaction
for irradiation (gamma-ray, x-ray, and E-beam)
processes is described by Equation 2-26. !
Alkalinity is an important parameter in the x-ray
process. If alkalinity is high, pH adjustment may be
required for irradiated matrices to shift the
carbonate-bicarbonate equilibrium from carbonate (d
scavenger) to carbonic acid (not a scavenger). i
Like gamma rays, x-rays have a high penetration
depth within irradiated water. According to Bailey
and Lackner (1995), a 1-MeV x-ray has an effective
water penetration depth of about 27 cm. Therefore,
the x-ray process can also be used to treat flowing
waste streams and containerized liquid wastes. •
Commercial-scale x-ray systems are not currently
available. However, the x-ray process has been
shown to be effective at the bench-scale level in
treating contaminated water.
2.1.9 E-Beam Process
The E-beam process involves irradiation of water or
air with a beam of high-energy electrons produced
by an electron accelerator. Within the electron
accelerator, an electric current (beam current) is
passed through a tungsten filament in a vacuum to
produce a stream of electrons. This electron stream
is accelerated by applying an electric field at a
specified voltage and is focused into a beam by
collimating devices. The applied voltage determines
the energy (speed) of the accelerated electrons,
which affects the depth to which the E-beam
penetrates the medium being irradiated. The
number of electrons emitted per unit time is
proportional to the beam current; therefore, the
E-beam power is the product of the beam current
and accelerating voltage. The E-beam produced is
injected into a reaction chamber through a thin foil
titanium window that serves as the vacuum seal
required for high-energy electron conversion.
Electron accelerators can provide electron energies
in the range of 0.1 to 10 MeV. High-energy (about
2-MeV) E-beams are used for irradiation of water,
whereas medium-energy (about 0.2-MeV) E-beams
are adequate for irradiation of air (Schwinkendorf
and others, 1995a). The depth to which an E-beam
can penetrate irradiated water is significantly less
than the depths associated with gamma rays and
x-rays (see Sections 2.1.7 and 2.1.8, respectively).
According to Bailey and Lackner (1995), a 1-MeV
electron deposits its energy in water within a depth
of 4 millimeters. As a result, E-beams are typically
used to treat contaminated water of relatively shallow
depths.
In the E-beam process, the mechanism of OH
formation is determined by the medium being
irradiated. E-beam irradiation of water causes
formation of three primary reactive species
responsible for organic compound destruction (OH,
e~aq, and H-). The reaction for irradiation
(gamma-ray, x-ray, and E-beam) processes is
described by Equation 2-26.
Alkalinity is an important parameter in the E-beam
process. If the alkalinity is high, influent pH
adjustment may be required for irradiated water to
shift the carbonate-bicarbonate equilibrium from
carbonate (a scavenger) to carbonic acid (not a
scavenger).
2-10
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E-beam irradiation of air causes formation of a
nonthermal plasma when high-energy electrons in
the beam react with the bulk gas molecules.
High-energy electrons generated in the plasma react
with wet air to form -OH by (1) electron attachment
to H2O, (2) direct dissociation of H2O by electrons,
(3) dissociation of H2O by O(1D), and (4) dissociation
of H2O by excited oxygen atoms (Penetrante,
1993a). These reactions are described for
nonthermal plasma processes in Equations 2-18
through 2-25. In dry-air applications, high-energy
electrons in the nonthermal plasma react with O2 to
form oxygen radicals. These oxidizing radicals play
an important role in generation of O3 and the initial
decomposition of some types of organics.
A commercial-scale E-beam system called the High
Voltage Environmental Applications, Inc. (HVEA),
E-beam treatment system is currently available for
treatment of contaminated water. In addition, the
E-beam process has been shown to be effective at
the pilot-scale level in treating contaminated air.
2.2 Commercial-Scale ANPO Systems
This section describes typical commercial-scale
ANPO systems for treatment of contaminated water
and soil. No commercial-scale ANPO systems are
available for treatment of contaminated air. The
information included in this section was obtained
from ANPO process vendors and from published
documents. The level of detail provided varies
depending on the source of information used.
The commercial-scale ANPO systems described in
this section include the (1) Geo-Cleanse® Fenton
system, (2) ISOTEC™ Fenton system, (3) H&H
SSCO™ Fenton system,(4) Chematur Aqua Critox®
SCWO system, (5) General Atomics SCWO system,
(6) OSI HYDROX hydraulic cavitation system, and
(7) HVEA E-beam treatment system.
2.2.1 Geo-Cleanse® Fenton System
Geo-Cleanse in Kenilworth, New Jersey, U.S.A.,
developed and manufactures an in situ system for
treatment of contaminated groundwaterand soil and
an ex situ system for treatment of contaminated soil.
Both the in situ and ex situ treatment systems are
based on the classic Fenton process. These
systems are described below.
2.2.1.1 In Situ System
A typical Geo-Cleanse® in situ treatment system
consists of a site-specific, horizontal and vertical
array of injectors that encompass contamination in
the subsurface and a mobile trailer containing the
equipment necessary to inject Fenton's reagents into
the subsurface. A simplified process flow diagram
for one injector is shown in Figure 2-1. The system
is pressurized and does not depend on diffusion of
reagents into the subsurface. Instead, the system
depends on a pressure gradient and subsurface
permeability to distribute reagents. The pressure
gradient is produced by chemical reactions that
occur during the treatment process—specifically,
degassing of carbon dioxide (CO2) and O2.
Backpressure at the mobile trailer ranges from about
240 to 310 kPa for most system applications. The
pressure required depends on the depth of the
contamination, the permeability of the subsurface
being treated, and contaminant concentrations.
An injector unit is composed of an injector mixing
head fitted on top of an injector well. Each injector
well is constructed of a 5-cm-diameter,
0.25-millimeter slot screen assembly attached to the
bottom of a 3.2-cm-diameter, 0.6- to 2-meter
(m)-long, stainless-steel casing. Above the casing,
3.2-cm-diameter, black iron riser pipe (not shown in
Figure 2-1) can be attached if necessary. The
injector mixing head, which is made of stainless
steel, is attached to the top of the injector well at
ground level. The mixing head, which is
hollow-bodied, contains a mixing chamber and
injection ports for supply lines carrying air, catalyst
solution (FeSO4), and 50 percent H2O2. The mixing
head is designed to prevent the catalyst solution and
H2O2 from mixing together in the injector until they
reach the screened interval in the injector well; as a
result, «OH formation occurs only in the subsurface.
The injector mixing head is connected by supply
lines to Fenton's reagents housed in the mobile
trailer. The trailer contains two 660-liter (L) storage
tanks for catalyst solution, one 660-L storage tank for
H2O2, one 10-horsepower air compressor, and a
control panel. The control panel is used to regulate
the flow of air and Fenton's reagents to the injectors.
The air compressor maintains a positive pressure on
the Fenton's reagents in the injector mixing head
and supplies air to the air-driven feed pumps. The
combined flow rate in the injector can range from
0.95 to 11 L per minute (L/min), depending on
subsurface geological characteristics.
In a typical system application, air and catalyst
solution are initially injected into the subsurface to
verify that the injector is open to the subsurface.
Added to the catalyst solution is a stabilizer (calcium
phosphate) that delays formation of free radicals.
H2SO4 (66 percent technical grade) and H3PO4
(85 percent technical grade) may also be injected
2-11
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Catalyst Solution Catalyst Solution j H202
Storage Tank 1 Storage Tank 2 Storage Tank
Injector
- Mixing
Head
Injector
Well
Not to Scale
Figure 2-1. Flow configuration in an in situ Geo-Cleanse® |Fenton system.
with the catalyst solution to adjust the groundwater
pH to a range of 5.5 to 6.0, within which iron will be
present in the Fe(ll) state and will remain dissolved
in solution. When acceptable flow and pH conditions
are established, H2O2and additional catalyst solution
are simultaneously injected into the subsurface. The
catalyst solution-acid-stabilizer mixture may be
added throughout the injection process to maintain
the groundwater pH within the range of 5.5 to 6.0. i
2.2.1.2 Ex Situ System
The Geo-Cleanse® ex situ treatment system is a
continuous-feed mixing unit that rests on a 9-m-long,
self-powered trailer. The mixing unit consists of a
soil hopper, catalyst hopper, conveyor belt, and
screw-type mixing chamber. A simplified process
flow diagram for the system is shown in Figure 2-2.
The system is powered by two generators. In a
typical system application, excavated soil with a
particle diameter of <3.2 cm is placed in the soil
hopper using a front-end loader. Soil is dispensed
through the bottom of the soil hopper onto the left
side of the conveyor belt, which proceeds to the
catalyst hopper. Catalyst (FeSO4) in powdered form
is dispensed through the bottom of the catalyst
hopper onto the right side of the conveyor belt
adjacent to the soil. The conveyor belt carrying the
soil and catalyst then proceeds to the rear of the
trailer, where the mixing chamber is located. At the
inlet of the mixing chamber, H2O2 is sprayed over the
soil and catalyst. The H2O2 is supplied by a tank
located on a separate trailer. Within the mixing
chamber, the Fenton process occurs as the soil,
catalyst, and H2O2 are mixed. Stabilizers, H2SO4, or
H3PO4 may also be added for pH adjustment. The
ex situ system is capable of processing about 19 or
46 cubic meters (m3) of soil per hour.
;2-12
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Contaminated Soil
Not to Scale
Figure 2-2. Flow configuration in an ex situ Geo-Cleanse® Fenton system.
2.2.2 ISOTEC™ Fenton System
The ISOTEC™ system was developed and is
manufactured by ISOTEC in West Windsor, New
Jersey, U.S.A. The system is an in situ soil and
groundwater treatment system based on a modified
Fenton process. In the ISOTEC™ process, the
components of a proprietary catalyst chelate iron and
keep it in dissolved form as an organometallic
complex. The form of iron and the chelating
components used are site-specific and are
determined through bench- and pilot-scale
evaluations. According to ISOTEC, its modification
of the classic Fenton process allows the system to
effectively treat organic contaminants in the
subsurface within a soil pH range of 2.5 to 8.5.
ISOTEC also uses proprietary stabilizer and mobility
agents to control formation and dispersion of -OH.
A typical ISOTEC™ treatment system consists of a
site-specific, horizontal and vertical array of injection
wells that encompass a contaminant plume and the
equipment necessary to inject modified Fenton's
reagents into the subsurface. The system, whose
flow configuration is shown in Figure 2-3, relies on
hydrostatic pressure or a pressure gradient to
distribute reagents throughout the subsurface. The
pressure gradient is produced by capping each
injection well and allowing the chemical reactions
(degassing of CO2 and O2) that occur during the
process to proceed. Pressure may also be
externally applied to inject reagents into the
subsurface. Backpressure at the delivery pump can
range from 14.0 to 790 kPa during such applications.
The pressure needed depends on the depth of
contamination, the permeability of the subsurface
being treated, and contaminant concentrations.
The injection wells are constructed of 5- to
10-cm-inside diameter, polyvinyl chloride material
with a 2.5-cm slot screen. The proprietary catalyst,
proprietary stabilizer, and oxidizer are kept in 570-L
storage tanks.
Reagents are delivered from storage tanks to the
injection wells by a delivery pump, an air
compressor, and supply lines. The air compressor
maintains a positive pressure on the reagents in the
injection wells and supplies air to the air-driven
delivery pump. Therefore, flow to the injection wells
is regulated by air pressure generated at the pump.
The combined flow rate in the injector can range
from 0.4 L/min under gravity conditions to 19 L/min
when applied pressure is used.
In a typical system application, catalyst is injected
into the subsurface, the supply lines are washed with
water, and then the oxidizer is applied. The oxidizer
is prepared by diluting 35 percent H2O2 with water
and ISOTEC's proprietary stabilizer. The treatment
process is repeated as necessary until cleanup goals
are met.
2.2.3 H&H SSCO™ Fenton System
The H&H SSCO™ system was developed and is
manufactured by H&H in North Bonnevilie,
Washington, U.S.A. This system is an ex situ soil
treatment system based on a modified Fenton
process. In the SSCO™ process, Fe(0) is applied in
powder form to soil in order to carry out the Fenton
reaction. According to the vendor, Fe(0) is a more
economical form of iron than to Fe(ll) or Fe(lll).
The SSCO™ process is carried out using the
Microenfractionator™, which is a 1,700-kilogram (kg)
soil mixing machine that is driven through windrows
(soil piles). This machine is self-propelled with
four-wheel drive. In the center of the
Microenfractionator™ is a counter-rotating drum unit
(5.5-m-long, 0.71-m diameter) with 74 sets of
fan-knife blades (0.25-m-long) attached to the
outside of the drum. The blade sets are positioned
2-13
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Delivery Pump
and Air Compressor
Stabilized H2O2
(5 to 20 percent)
Injection
Well
Not to Scale :
Figure 2-3 Flow configuration in an ISOTEC™ Fenton system.
on the drum at various angles relative to the drum
unit. The drum and blades are made of a proprietary
steel alloy and are powered by a diesel engine.
As the Microenfractionator™ is driven through
windrows, the drum unit rotates at about 600
revolutions per minute. Soil particles contacted by
the fan-knife blades become entrained in the air
vortices produced by the rotation of the drum unit.
The mixing action simultaneously homogenizes the
soil and coats soil particles with oxidizer (50 percent
H2O2) reagent, which is pumped directly into the
microenfraction chamber through a spray nozzle.
The Microenfractionator™, which can be equipped
with a 1,900- to 19,000-L H2O2 storage tank, can
supply H2O2 at a rate of up to 450 L/min. According
to the vendor, the Microenfractionator™ can
displace about 1,130 cubic meters per minute
(nrYmin) of air during the process and can achieve
up to 95 percent soil homogeneity. Furthermore,
the Microenfractionator™ can process up to
1,100 cubic meters per hour (m3/hr) of soil. ,
In a typical system application, excavated soil is
placed in windrows that are 4.9 m wide and 2.0 m
tall. The number of windrows and their length can
vary based on the amount of soil that needs to
be treated. At a minimum, two passes are
made through each windrow with the
Microenfractionator™. The first pass is used to
apply the catalyst (Fe[0]). To apply the catalyst, 23-
to 36-kg bags containing catalyst in dry powder form
are placed on top of the windrows. The amount of
catalyst added to the soil typically corresponds to 0.1
to 1 percent of the total weight of the soil to be
treated. The Microenfractionator™ is then used to
microenfractionate the catalyst into the soil. After
a 24-hour stabilization period, the second pass
is made, during which the oxidizer is
microenfractionated into the soil. Additional oxidizer
may be microenfractionated into the soil during
subsequent passes to meet cleanup goals.
2-14
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2.2.4 Chematur Aqua Cr/tox® SCWO
System
The Chematur Aqua Critox® system was developed
by ECO Waste Technologies and is manufactured by
Chematur in Karlskoga, Sweden. This treatment
system is designed to remove organic contaminants
from water using SCWO. A simplified process flow
diagram for the system is shown in Figure 2-4. In
this process, contaminated water is pressurized to
about 25 MPa using a positive-displacement,
high-pressure feed pump and then flows through a
heat exchanger and into a trim heater. The
temperature of the feed stream on the outlet side of
the trim heater (360 to 377 °C) is near or above the
supercritical temperature of water (374 °C). An
oxidant (O2) is added to the feed stream before it
reaches the reactor. Within the reactor, SCWO
occurs as temperatures reach 540 to 650 °C. The
reactor effluent flows through a heat exchanger shell,
a waste heat recovery-boiler, and an effluent air
cooler. A pressure letdown control valve lowers the
effluent pressure to atmospheric pressure, and a
liquid-gas separator separates the effluent into liquid
and gas phases. The liquid effluent consists of water
saturated with CO2, and the gaseous effluent
consists of CO2 and O2.
Contaminated Water
Figure 2-4. Flow configuration in a Chematur Aqua Critox® system.
Liquid O2 Supply
2.2.5 General Atomics SCWO System
The General Atomics system was developed and is
manufactured by General Atomics in San Diego,
California, U.S.A. This system is designed to treat
contaminated water using SCWO. A simplified
process flow diagram for the system is shown in
Figure 2-5. In this process, a transfer pump supplies
contaminated water at low pressure (about 100 kPa)
to a high-pressure feed pump. In addition, transfer
pumps supply dilution water and auxiliary fuel at a
pressure of about 380 kPa to their respective
high-pressure feed pumps. An oxidant (O2) is
supplied to the system by a high-pressure
(2.3 nrrVmin at 34 MPa) air compressor. A liquid O2
system can be used as an alternative source of O2.
Contaminated water and auxiliary fuel flow from the
high-pressure feed pumps directly into the reactor.
The dilution water and oxidant flow through a
titanium-lined preheater before entering the reactor.
2-15
-------
Dilution Water
Auxiliary Fuel
Contaminated Water
High-Pressure
Air Compressor
Back- Capillary
Pressure Pressure
Control Letdown
Valve Chamber
Gas Analysis
System
^. Gas Sample
Collection
_>. Release
I I Release
1 | °f Gases
Charcoal
Filter
Liquid
Sampling Tank'
Station
Not to Scale
Figure 2-5. Flow configuration in a General Atomics SCWO systen
Within the reactor, SCWO occurs as the temperature
and pressure climb as high as 650 °C and 30 MPa,
respectively. Multiple reactor designs are available
for the system, including a titanium-lined reactor and
a reactor fitted with salt and solids transport
equipment. Quench water is mixed with the reactor
effluent. The effluent is then cooled in ja
titanium-lined heat exchanger and reduced in
pressure by a pressure letdown system. The
pressure letdown system includes capillaries and
pressure control valves that can be used separately
or in combination.
.lid
Collection
Tank
(Multiple)
The process effluent consists of gaseous, liquid, and
solid components that flow into liquid collection
tanks. The gaseous effluent is vented from the
tanks, and most of the gas passes through a
charcoal filter before it is released. A portion of the
gas undergoes continuous analysis for O2 and
carbon monoxide (CO) content in the gas analysis
system. At predetermined intervals, a sample of the
liquid effluent, including any entrained salts or other
solids, is manually collected at the liquid sampling
station. i
2.2.6 OSI HYDROX Hydraulic
Cavitation System
The OS I HYDROX system was developed and is
manufactured by OSI in Arcadia, California, U.S.A.
This system is designed to treat contaminated water
using hydraulic cavitation. A typical OSI HYDROX
system consists of a feed pump, a HYDROX reactor,
interconnected piping and valves, and a sampling
port as shown in Figure 2-6. The feed pump
supplies contaminated waterto the HYDROX reactor
at about 340 to 590 kPa. In the reactor, a proprietary
nozzle on the inlet side of the reactor lowers the
pressure of the system to a level at or below the
vapor pressure of water (2.5 kPa at 20 °C), causing
cavities to form in the water. As the water, flows
through the reactor, the system pressure increases
to atmospheric pressure (about 100 kPa), causing
•OH formation. The water leaving the reactor can be
discharged from the system or recycled to the
suction side of the feed pump. Diverting all or part of
the system flow into the recycling unit increases the
retention time of the contaminated water in the
HYDROX reactor. Skid-mounted HYDROX systems
with treatment rates of up to 470 L/min are readily
available. HYDROX systems with treatment rates of
>950 L/min usually need to conform to site-specific
design criteria and require on-site assembly.
2-16
-------
Contaminated Water
Recycled Water
Recycle Valve
HYDROX
Reactor
Treated Water
Outlet Valve
Inlet Valve
Feed Pump
Not to Scale
Sampling Port
Figure 2-6. Flow configuration in an OSI HYDROX hydraulic cavitation system.
2.2.7 HVEA E-Beam Treatment System
The HVEA E-beam system was developed and is
manufactured by HVEA in Miami, Florida, U.S.A.
This system is designed to treat contaminated water
using the E-beam process. A simplified process flow
diagram for the system is shown in Figure 2-7. The
system is housed in a mobile trailer and is rated for
a maximum flow rate of 190 L/min. The E-beam
system includes the following components: a strainer
basket, an influent pump, the E-beam unit, a cooling
air processor, a blower, and a control console (not
shown in Figure 2-7). These components are
situated in three separate rooms: the pump room,
process room, and control room. The pump room
contains all the ancillary equipment required by the
E-beam unit for both water and air handling, the
radiation-shielded process room contains the
E-beam unit itself, and the control room contains the
control console where system operating conditions
are monitored and adjusted.
The E-beam unit is made up of the following
components: an electron accelerator, a scanner, a
contact chamber, and lead shielding. The electron
accelerator produces the E-beam. Within the
electron accelerator, a stream of electrons is emitted
when an electric current (beam current) is passed
through a tungsten wire filament. The electron
stream is accelerated by applying an electric field at
a specified voltage and is focused into a beam by
collimating devices. The electron accelerator can
generate an accelerating voltage of 500 kiloelectron
volts (keV) and a beam current of between 0 and
42 milliamperes (mA).
A pyramid-shaped scanner located beneath the
electron accelerator deflects the E-beam, causing it
to scan the contaminated water (E-beam scanner
operation is similar to that of the scanner in a
television set). Contaminated water is pumped
through the contact chamber, which is located
beneath the scanner. The scanner is operated in
such a way that the E-beam contacts the entire
surface of the water flowing through the contact
chamber.
A titanium window separates the scanner from the
contact chamber. This window is necessary to
maintain a vacuum in the scanner; the vacuum is
required to minimize E-beam energy losses. As the
E-beam passes through the titanium window, some
of the E-beam's energy is absorbed by the window.
This energy absorption is manifested in the form of
heat. The titanium window is cooled by passing
cooling air through the contact chamber. Cooling air
exiting the contact chamber flows through a cooling
air processor and is returned to the contact chamber
by a blower.
The cooling air processor includes an air filter, a
carbon adsorber, and an air chiller. The air filter is
used to remove particulates from the cooling air.
The carbon adsorber is used to destroy O3 that is
formed in the cooling air when it is exposed to the
E-beam in the contact chamber. O3 must be
removed from the cooling air to prevent corrosion in
the air lines and the blower. According to HVEA, any
VOCs present in the cooling air as a result of
incidental VOC volatilization in the contact chamber
are destroyed by the E-beam. Vapor-phase VOCs
that are not destroyed by the E-beam may be
removed by the carbon adsorber. The air chiller is
used to cool the air. According to HVEA, under
normal operating conditions, cooling air is
recirculated in a closed loop through the contact
2-17
-------
Control Room
E-beam Unit
Lead Shielding
Scanner
Contact Chamber
Influent Pump
Strainer Basket
Contaminated
Water
Not to Scale
Cooling Air Processor
Figure 2-7. Flow configuration in an HVEA E-Beam system.
chamber and cooling air processor. When the
E-beam system is operated, both the influent pump
and the blower run continuously. If either water flow
or cooling air flow stops, the system automatically
shuts down. j
i
Lead shielding surrounds the E-beam unit to preverjt
x-ray emissions. X-rays are formed when the
E-beam contacts various internal, stainless-steel
surfaces. As an added safety measure, the process
room is inaccessible during system operation. !
Resistance temperature devices (RTD) are used t<5
measure the temperature of water before and after
treatment. HVEA uses the change in water
temperature to estimate the E-beam dose based on
established equations defining the relationship
between dose and temperature change (Nickelsen
and others, 1992). The HVEA E-beam system is
configured with two RTDs immediately upstream and
two RTDs immediately downstream from the contact
chamber. Output from the RTDs is fed into a
computer in the control room for processing and
recordkeeping. !
The contaminated water flow rate is monitored at a
point upstream from the contact chamber. The flow
rate is manually adjusted in the pump room and is
measured by a flow meter.
2.3 ANPO System Design
Considerations
ANPO system design has not reached a stage where
systems can be designed based on mechanistic
models because of the complex nature of (1) the
process chemistry involved, including the generation
of primary reactive species and the destruction of
contaminants; (2) matrix effects; and (3) waste
streams with multiple contaminants. As a result,
ANPO system design typically follows a scale-up
approach.
Pedit and others (1997) propose that a model to
facilitate the design of ANPO systems should (1) be
based on process chemistry, (2) account for mass
transfer of compounds between gas and liquid
phases, (3) be applicable to a variety of system
configurations, (4) predict transient behavior in
response to changes in influent concentrations and
other transient phenomena, (5) account for
pH-controlled speciation, (6) be applicable to a
variety of micropollutants and background water
characteristics, and (7) be able to incorporate
modifications in reaction mechanisms or reaction
2-18
-------
rates. The model developed by Pedit and others is
used to describe data from a full-scale O3/H2O2
demonstration plant and is shown to provide a
reasonably accurate representation of the data.
However, because of the complexity and uncertainty
associated with the input requirements for the model,
the model has not yet been used to facilitate the
design of ANPO systems.
Models have also been developed to assist in ANPO
process optimization. Peyton (1996) describes a
semi-empirical model with which the overall
efficiency of an ANPO process can be evaluated as
the product of efficiencies for the various steps that
occur during radical generation, propagation, and
attack on a contaminant. The model can thus be
used to assist in ANPO process optimization by
identifying steps for which the efficiency can be
improved.
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Section 3
Contaminated Water Treatment
ANPO processes have been demonstrated to be
effective for treatment of contaminated water. Water
matrices to which ANPO has been applied include
(1) contaminated groundwater, (2) industrial
wastewater, (3) municipal wastewater,
(4) contaminated drinking water, (5) landfill leachate,
and (6) contaminated surface water. Collectively,
ANPO has been applied to the following types of
waterborne contaminants: VOCs, SVOCs,
polychlorinated biphenyls (PCB), pesticides and
herbicides, dioxins and furans, explosives and their
degradation products, humic substances, dyes,
inorganics, and microbes.
To assist an environmental practitioner in the
selection of an ANPO system to treat contaminated
water, this section includes (1) commercial-scale
system evaluation results for Fenton, SCWO,
hydrodynamiccavitation, and E-beam processes and
(2) pilot-scale system evaluation results for Fenton,
O3/H2O2, electrochemical oxidation, SCWO, and
E-beam processes. This section also summarizes
supplemental information available from bench-scale
studies of ANPO processes.
As described in Section 1.2, this handbook organizes
performance and cost data for each water matrix first
by contaminant group, then by scale of evaluation
(commercial, pilot, or bench), and finally by ANPO
system or process. In general, commercial- and
pilot-scale applications are discussed in detail. Such
discussions include, as available, a system
description, operating conditions, performance data,
and system costs presented in 2000 U.S. dollars.
Bench-scale studies of ANPO processes are
described in less detail and only if they provide
information that supplements commercial- and
pilot-scale evaluation results. The level of detail
provided for bench-scale studies varies depending
on the source of information used. For example,
percent removals and test conditions are not
specified for some bench-scale studies because
such information is unavailable in the sources.
At the end of each matrix section, a table is provided
that summarizes operating conditions and
performance results for each commercial- and
pilot-scale application discussed in the text. The
references cited in Section 3 are listed in
Section 3.7.
3.1 Contaminated Groundwater
Treatment
The effectiveness of ANPO processes in treating
contaminated groundwater has been evaluated for
various contaminant groups, including VOCs,
SVOCs, PCBs, pesticides and herbicides, dioxins
and furans, explosives and their degradation
products, humic substances, and inorganics. This
section describes ANPO process effectiveness with
regard to each of these contaminant groups. The
operating conditions and performance results for
each commercial- and pilot-scale application
discussed in Section 3.1 are summarized in
Table 3-1 at the end of the section.
3.1,1 VOC-Contaminated Groundwater
This -section discusses treatment of VOCs in
groundwater using the Fenton, hydrodynamic
cavitation, and E-beam processes on a commercial
scale. Additional information on VOC removal using
the (1) O3/H2O2 process at the pilot scale and (2)
Fenton, O3/H2O2, and acoustic cavitation processes
at the bench scale is also included.
Commercial-Scale Applications
This section summarizes the effectiveness of the
Geo-Cleanse® Fenton, ISOTEC™ Fenton, OSI
HYDROX hydrodynamic cavitation, and HVEA
E-beam treatment systems in removing the following
VOCs from contaminated groundwater.
ANPO Process
Fenton
Hydro-
dynamic
cavitation
VOCs Removed
1,2-Dichloroethene (DCE);
methyl-tert-butylether
(MTBE); 2-methyl-
naphthalene; naphthalene;
tetrachloroethene (PCE);
1,1,1-trichloroethane
(TCA); trichloroethene
(TCE); vinyl chloride (VC);
benzene, toluene, ethyl-
benzene, and xylene
(BTEX)
2-Methylnaphthalene;
naphthalene
3-1
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ANPO Process
• E-beam
VOCs Removed
Carbon tetrachloride; I
chloroform; 1 ,2-DCE, 1 ,2-
dichloroethane (DCA); :
MTBE;PCE;1,1,1-TCA;
TCE; BTEX
Geo-CIeanse® Fenton System
A Geo-CIeanse® in situ Fenton system was used „
remove 1,2-DCE; PCE; and TCE from groundwater
under former waste lagoons at Letterkenny Arrpy
Depot in Chambersburg, Pennsylvania, U.S.A.
Groundwater in the limestone bedrock aquifer at this
site contained 860 micrograms per liter (ug/L) of
1,2-DCE; 220 ug/L of PCE; and 2,100 ug/L of TCE.
The Geo-CIeanse® Fenton system treated about
10,000 m3 of groundwater. About 48,000 L of
50 percent H2O2; 190 L of 66 percent H2SO4; 38 L of
85 percent H3PO4; and 17 kg of FeSO4 were injecte!d
at six locations in the treatment zone over
8 consecutive days. As part of the treatment,
Geo-CIeanse also used calcium phosphate as a
stabilizer for the reaction; however, the amount of
stabilizer used is unknown. After a treatment time of
60 days, 1,2-DCE; PCE; and TCE concentrations
had been reduced by about 40 percent each. In
addition, no tentatively identified compounds (TIC)
were detected when groundwater samples were
analyzed using U.S. EPA's "Test Methods fair
Evaluating Solid Waste" (SW-846) Method 8260
(U.S. EPA, 1996). According to Geo-CIeanse, the
total treatment cost was about $170,000, which
includes equipment rental, reagent, mobilization, and
labor costs (Geo-CIeanse, 2000).
In another in situ application at Letterkenny Army
Depot, a Geo-CIeanse® Fenton system was used to
remediate contaminated groundwater in a shale
bedrock aquifer beneath an oil burn pit and fire
training area. The primary contaminant of concern
at the site was 1,1,1-TCA, which was present at 'a
maximum concentration of 280 milligrams per liter
(mg/L). The Geo-CIeanse® system treated about
640 m3 of groundwater in a 15- by 15- by 9-m areai
About 7,600 L of 50 percent H2O2; 110 L of
66 percent H2SO4; 38 L of 85 percent H3PO4; and
18 kg of FeSO4 were injected at four locations in the
treatment zone for 7 days over 4 weeks. The
amount of calcium phosphate stabilizer used is
unknown. Four months after the last injection, the
TCA concentration had been reduced by 84 percent
at the maximum concentration location. No TICs
were detected when groundwater samples were
analyzed using SW-846 Method 8260. The
vendor-estimated treatment cost was about
$230,000, which includes equipment rental, reagent,
mobilization, and labor costs (Geo-CIeanse, 2000).
In 1998, a Geo-CIeanse® in situ Fenton system was
used at Pensacola Naval Air Station in Florida,
U.S.A., to treat TCE-contaminated groundwater
under a sludge drying bed at a wastewater treatment
plant. The initial TCE concentration averaged 1,500
ug/L. The system treated about 62 m3 of
groundwater. About 38,000 L of 50 percent H2O2;
170 L of 66 percent H2SO4; 57 L of 85 percent
H3PO4; and 4.5 kg of FeSO4 were injected at eight
locations in the treatment zone for a total of 10 days
over 1.5 months. The amount of calcium phosphate
stabilizer used is unknown. After a treatment time of
60 days, the system removed 98 percent of the TCE
in the groundwater. No TICs were detected when
groundwater samples were analyzed using SW-846
Method 8260. According to Geo-CIeanse, the total
treatment cost was about $160,000, including
equipment rental, reagent, mobilization, and labor
costs (Geo-CIeanse, 2000).
A Geo-CIeanse® in situ Fenton system was used to
treat groundwater contaminated with PCE and its
natural degradation products (1,2-DCE; TCE; and
VC) at Kings Bay Naval Submarine Base in Georgia,
U.S.A. The contaminants, which originated from the
former Camden County landfill, were present in the
groundwater at an average total of 5,000 ug/L. The
system treated about 380 m3 of groundwater. About
46,000 L of 50 percent H2O2; 450 L of 66 percent
H2SO4; 110 L of 85 percent H3PO4; and 54 kg of
FeSO4 were injected at 23 locations in the treatment
zone for 18 consecutive days. The amount of
calcium phosphate stabilizer used is unknown. After
a treatment time of 60 days, the system achieved an
average of 89 percent removal of the contaminants
in the groundwater. No TICs were detected when
groundwater samples were analyzed using SW-846
Method 8260. The vendor estimates the total
treatment cost to be $210,000, which includes
equipment rental, reagent, mobilization, and labor
costs (Geo-CIeanse, 2000).
At Wright Army Airfield in Fort Stewart, Georgia,
U.S.A., a Geo-CIeanse® in situ Fenton system was
used to remediate contaminated groundwater in a
sand aquifer beneath a helicopter refueling area.
The primary contaminants of concern at the site
were petroleum hydrocarbons, specifically
2-methylnaphthalene (640 ug/L), naphthalene
(640 ug/L), benzene (600 ug/L), toluene (23 ug/L),
ethylbenzene (670 ug/L), and xylene (1,500 ug/L).
The system treated about 350 m3 of groundwater.
About 9,500 L of 50 percent H2O2; 83 L of 66 percent
i3-2
-------
H2SO4; 8 L of 85 percent H3PO4; and 54 kg of FeSO4
were injected at two locations in the treatment zone
over 4 consecutive days. The amount of calcium
phosphate stabilizer used is unknown. After a
treatment time of 6 days, the system achieved the
following removal efficiencies: 2-me.thylnaphthalene,
88 percent; naphthalene, 86 percent; benzene,
96 percent; toluene, 18 percent; ethylbenzene,
78 percent; and xylene, 62 percent. The in situ
system not only achieved significant source
reduction but also allowed the airfield to remain
active during remediation because no excavation
was necessary. No TICs were detected when
groundwater samples were analyzed using SW-846
Method 8260. The vendor-estimated treatment cost
was $75,000, which includes equipment rental,
reagent, mobilization, and labor costs (Geo-Cleanse,
2000).
Groundwater beneath a dry-cleaning facility in
Houston, Texas, U.S.A. was treated in situ using a
Geo-Cleanse® Fenton system. The chemicals
1,2-DCE, P'CE, TCE, and VC were present in the
groundwater at average concentrations 'of 930;
1,200; 600; and 17 ug/L, respectively. The system
treated about 4,100 m3 of groundwater. About
32,000 L of 50 percent H2O2; 190 L of 66 percent
H2SO4; 40 L of 85 percent H?PO4; and 45 kg of
FeSO4 were injected at 40 locations in the treatment
zone over 9 consecutive days. The amount of
calcium phosphate stabilizer used is unknown. As a
result of the treatment, the concentrations of all four
contaminants fell below detection limits in all the
monitoring wells sampled 30 days after the last
injection. No TICs were detected when groundwater
samples were analyzed using SW-846 Method 8260.
The vendor-estimated treatment cost was $320,000,
including equipment rental, reagent, mobilization,
and labor costs (Geo-Cleanse, 2000).
ISOTEC™ Fenton System
An ISOTEC™ in situ Fenton system was used to
treat MTBE- and BTEX-contaminated groundwater
in a silty sand aquifer beneath a warehouse facility in
northern New Jersey, U.S.A. Groundwater at the
site contained >54 mg/L of MTBE and BTEX total.
The groundwater pH before treatment was nearly
neutral. About 13,000 L of Fenton's reagent was
used to treat a plume spread over an area 38 m
long, 15m wide, and 4 m deep. The treatment zone
was 3 to 7 m below ground surface (bgs). Reagent
injections were performed at six locations throughout
the plume for a total of 16 days over a 7-month
period. The system achieved >99 percent MTBE
and BTEX removal. All the treatment
goals—70 ug/L for MTBE; 1 ug/L for benzene;
.1,000 ug/L for toluene; 700 ug/L for ethylbenzene;
and 40 ug/L for total xylene—were met, and the site
was issued a no further action letter by the New
.Jersey Department of Environmental Protection
4 months after the final round of injection activities.
According to ISOTEC, the total treatment cost was
about $110,000; no breakdown of the cost was
provided by the vendor (ISOTEC 2000).
Groundwater in a fine-grained, sandy silt aquifer
contaminated with 1,2-DCE; PCE; and TCE at a
manufacturing facility in northern New Jersey,
U.S.A., was treated using an ISOTEC™ in situ
Fenton system. The chlorinated VOCs were present
in the groundwater at a total of >150 mg/L. The
groundwater pH before treatment was nearly neutral.
About 58,000 L of Fenton's reagent was used to
treat a plume spread over an area 53 m long, 14m
wide, and 11m deep. Reagent injections were
performed at four shallow locations (3 m bgs) and 13
deep locations (12 m bgs) for 32 days over a
6-month period. The system achieved >99 percent
removal for each contaminant. Most of the
groundwater plume met the cleanup criteria of
70 ug/L for 1,2-DCE; 1 ug/L for PCE; and 1 ug/L for
TCE. According to ISOTEC, system operating costs
totaled $180,000; no breakdown of the cost was
provided by the vendor (ISOTEC, 2000).
OSI HYDROX Hydrodynamic Cavitation
System
Groundwater containing furans, polynuclear aromatic
hydrocarbons (PAH), and VOCs was treated in a
7,600-L/min OSI HYDROX hydrodynamic cavitation
system. Concentrated creosote chemicals were
collected from the bottom of wells in Visalia,
California, U.S.A., and were mixed with well water for
treatment. The addition of the chemicals rendered
the well water opaque. The chemicals 2-methyl-
naphthalene and naphthalene were present at initial
concentrations of 260 and 340 mg/L, respectively.
About 41 kilowatts (kW) of power was consumed in
treating the groundwater for 9 hours. Treatment
continued until the water became clear and
contaminant concentrations reached acceptable
discharge levels. Hydrodynamic cavitation reduced
the concentrations of 2-methylnaphthalene and
naphthalene by 78 and 75 percent, respectively.
According to OSI, the total treatment cost was about
$200,000, which includes the capital cost and an
operation and maintenance (O&M) cost of $0.65/m3
of groundwater treated. The estimated cost savings
associated with treating the contaminated
groundwater using ANPO rather than disposing of it
by conventional methods were >$500,000 (OSI,
2000).
3-3
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HVEA E-Beam System
A Model M25W-48S HVEA E-beam system was
demonstrated in September and November 1994
under the U.S. EPA Superfund Innovative
Technology Evaluation Program. This
demonstration involved removing VOCs .from
groundwater in the M-area at the U.S. Department of
Energy Savannah River Site in Aiken, South
Carolina, U.S.A. (Alvarez and others, 1998). I
;
PCE and TCE were the primary groundwater
contaminants at the site, with concentrations ranging
from 9,200 to 12,000 ug/L and 25,000 jo
30,000 ug/L, respectively. In addition, 1,2-DCE was
present in groundwater at low concentrations ranging
from 40 to 43 ug/L. Thirteen test runs of the system
were conducted in five phases: four test runs used
raw groundwater, and the remaining nine test runs
used groundwater spiked with carbon tetrachloride
(150 to 400 ug/L); chloroform (240 to 650 ug/L);
1,2-DCA (210 to 840 ug/L); 1,1,1-TCA (200 to 5QO
ug/L); benzene (220 to 550 ug/L); toluene (170 to
360 ug/L); ethylbenzene (95 to 250 ug/L); and xylene
(85 to 200 ug/L). Unspiked Run 3 and spiked Run
13 had flow rates of 150 and 76 L/min, respectively.
During both runs, groundwater was irradiated with a
beam current of 42 mA. In Run 13, in which the
effect of alkalinity on VOC removal was tested, the
influent pH and alkalinity were 7.6 and 500 mg/L as
calcium carbonate (CaCO3), respectively.
The system treated about 260 m3 of
VOC-contaminated groundwater at the U.S.
Department of Energy Savannah River Site. For the
unspiked groundwater (Run 3) in which the alkalinity
was <5 mg/L as CaCO3, removal efficiencies for
1,2-DCE; PCE; and TCE were >90, 96, anjd
97 percent, respectively. For the spike|d
groundwater (Run 13) in which the alkalinity wa's
increased to 500 mg/L as CaCO3, removal
efficiencies for 1,2-DCE; PCE; and TCE were >91;
99, and 99 percent, respectively. In addition, carboh
tetrachloride and BTEX were removed with
efficiencies of >98 percent during Run 13.
Chloroform; 1,2-DCA; and 1,1,1-TCA concentrations
were reduced by 80, 57, and 81 percent,
respectively. \
Bioassay tests of freshwater test organisms showed
that E-beam treatment of groundwater increased
acute toxicity for the fathead minnow (Pimephales
promelas) but not for the water flea (Ceriodaphnia
dubia). The increased toxicity for the fathead
minnow was attributed primarily to formation of toxic
by-products in the effluent, including haloacetic acids
and aldehydes (Topudurti and others, 1998). i
Several VOCs present in the contaminated
groundwater were volatilized and detected in the
cooling air, including carbon tetrachloride;
chloroform; 1,2-DCA; PCE; 1,1,1-TCA; TCE; and
BTEX. These VOCs were converted to CO2, CO,
and phosgene. Hydrochloric acid (HCI), nitrous
oxide, and O3 were also formed in the air phase
(Topudurti and others, 1998)
Groundwater treatment costs were estimated for two
scenarios, in each of which the HVEA E-beam
system was assumed to treat about 1,200,000 m3 of
contaminated groundwater. In Case 1, the
groundwater was assumed to contain unsaturated
VOCs (PCE and TCE). In this scenario, the
groundwater treatment costs directly associated with
the E-beam system were $1.21/m3 of water treated
for a system with a beam power of 21 kW operating
at a 150-L/min flow rate for 15 years. In Case 2, the
groundwater was assumed to contain saturated
VOCs (carbon tetrachloride; chloroform; 1,2-DCA;
and 1,1,1 -TCA) and aromatic VOCs (BTEX) as well
as PCE and TCE. In this scenario, the groundwater
treatment costs were $1.77/m3 of water treated for a
system with a beam power of 21 kW operating at a
76-L/min flow rate for 30 years. Treatment costs
included treatability study and system design,
mobilization and startup, one-time system capital,
labor, utility, equipment maintenance, residual waste
shipping and handling, and site demobilization costs
(Alvarez and others, 1998).
In another application, an HVEA E-beam system was
tested to examine its removal efficiency for MTBE in
groundwater. Experiments were conducted using a
mobile system housed in a 15- by 2.4-m semi trailer
containing the E-beam unit with an accelerating
voltage of 500 keV and a beam power of 20 kW.
Groundwater samples with a pH of between 7.7 and
8.8 were collected for the study. The initial MTBE
concentration in the samples was 170 ug/L. E-beam
irradiation at a dose of 250 kilorads (krads) resulted
in a 92 percent reduction in MTBE concentration. At
a dose of 500 krads, the MTBE concentration fell to
1.4 ug/L (99 percent removal). Tert-butyl alcohol
and tert-butyl formate were identified as MTBE
irradiation by-products. MTBE removal was affected
by the presence of carbonate alkalinity, which
created competition for «OH (Cooper and others,
2000).
Pilot-Scale Applications
VOCs in groundwater have been removed using
ANPO processes on a pilot scale. This section
presents pilot-scale evaluation results for the
3-4
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O3/H2O2 process in terms of its removal of the
following VOCs from groundwater.
ANPO Process
• 03/H202
VOCs Removed
• PCE; TCE
An O3/H2O2 system was pilot-tested at a
semiconductor manufacturing facility in Essex
Junction, Vermont, U.S.A. The system was used to
treat groundwater contaminated with PCE and TCE.
The system contained a metering pump that added
35 percent H2O2 to groundwater and O3 generators
that injected ozonated air into the system influent
through four Venturi-type injectors. Groundwater
from an on-site bedrock production well was pumped
into the system at a flow rate of about 900 L/min. A
static mixer was used to agitate the influent during
treatment. Treated groundwater was pumped to two
carbon adsorption units for final treatment. The
initial groundwater concentrations of PCE and TCE
were 3.2 and 0.17 mg/L, respectively. The O3 dose
was 1 to 2 mg/L, and the H2O2 dose was 0.5 to
1 mg/L. The retention time was 15 to 20 minutes.
Under these conditions, the O3/H2O2 system
achieved 91 percent removal of PCE, and TCE
concentrations fell below the detection limit. The
estimated treatment cost was about $11,000; no
breakdown of the cost is available (Clancy and
others, 1996).
A pilot-scale O3/H2O2 system was field-tested at the
Department of Water and Power in Los Angeles,
California, U.S.A. The system was used to treat
PCE- and TCE-contaminated groundwater from a
drinking water well. The system consisted of an O3
contactor column reactor, a four-stage turbine mixer,
and an eight-element static mixer (into which H2O2
was injected). The influent PCE and TCE
concentrations were about 12 and 100 ug/L,
respectively. The flow rate was maintained at
0.25 liter per second (Us), corresponding to a
retention time of 15 minutes. The optimum H2O2:O3
dose ratio was 0.4 to 0.5 by weight. At an O3 dose
of 9.5 mg/L and an H2O2 dose of 3 mg/L, 88 and
99 percent of the PCE and TCE were removed,
respectively. The estimated treatment cost was
$38/L of groundwater treated based on a volume of
7,570 L. This cost includes capital, O&M, and
granular activated carbon replacement costs (Aieta
and others, 1988).
Bench-Scale Studies
This section summarizes the results of bench-scale
studies of the effectiveness of ANPO processes for
VOC removal from groundwater. The bench-scale
results are summarized only for studies that provide
information beyond that discussed above for the
commercial- and pilot-scale applications. This
section presents bench-scale study results for the
use of the Fenton, O3/H2O2, and acoustic cavitation
processes to remove the following VOCs from
groundwater and synthetic wastewater.
ANPO Process
Fenton
• O3/H2O2
Acoustic
cavitation
VOCs Removed
Benzene; chloroform;
TCE; toluene;
1 ,2,3-trichloropropane
Chloroform; 1-chloro-
pentane; 1,2-DCA;
1,1-dichloropropene; PCE;
TCE; toluene
Ethane; methane; TCE
Fenton
Vella and Veranda (1993) have studied the
degradation of TCE (10 mg/L) in synthetic
wastewater by the Fenton process. The reaction
was maintained at a pH between 3.9 and 4.2 and at
an Fe(ll):H2O2 molar ratio of 0.2. The kinetics of the
reaction were studied using H2O2 doses of 53 and
75 mg/L. At H2O2 doses of 53 mg/L or more, over
80 percent of the TCE was removed within
2 minutes. The authors suggest that pH adjustment
to about 4 is required for the Fenton process to be
an effective treatment approach.
Fenton's reagent treatment of 1,2,3-trichloropropane
in synthetic wastewater has also been studied. The
initial concentration of the contaminant was
150 mg/L. Chloroaceticacid; 1,3-dichloropropanone;
and formic acid were identified as reaction
by-products. In the study, a pH range of 2.0 to 3.3
was required for optimum destruction efficiency. In
addition, increasing the Fe(ll) concentration,
temperature, or both increased the rate of
1,2,3-trichloropropane destruction (Hunter, 1996).
The Fenton process has been used in bench-scale
treatment of organics in groundwater. Benzene,
chloroform, and toluene were present at initial
concentrations of 2, 5.8, and 43 mg/L, respectively.
Other parameters studied included chemical oxygen
demand (COD) (present at 490 mg/L) and total
organic carbon (TOG) (present at 120 mg/L).
Optimum conditions were found to include a
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Fe{ll):H2O2 molar ratio of 0.25 and a pH of 4. Under
these conditions, the following removal efficiencies
were achieved: benzene, 82 percent; chloroform,
>91 percent; and toluene, 78 percent. In addition,
COD and TOC were reduced by 63 and 71 percent,
respectively. At a pH above 4.5, the TOC removal
efficiency stabilized, and significant precipitation of
ferrous and ferric hydroxides was observed
(Kochany and Lugowski, 1998). I
Chloroform removal from groundwater at Rocky
Mountain Arsenal in Commerce City, Colorado,
U.S.A., was studied using the O3/H2O2 process in a
bench-scale study conducted by Zappi and others
(1992). Groundwater samples were collected from
two locations at the arsenal called Basin A and
South Plants.' Chloroform was present in
groundwater at Basin A and South Plants at 33i4
and 0.738 mg/L, respectively. During the study, O3
was continuously sparged into a 1-L batch reactor at
a rate of 3 mg/L. For the treatment of Basin A
groundwater, the optimum H2O2 dose for the amount
of O3 added was in the range of 0.25 to 1 mg/L.
About 90 percent of the chloroform was removed
during 20 minutes of batch treatment. For the
treatment of South Plants groundwater, the optimum
H2O2 dose for the amount of O3 added was in the
range of 1 to 10 mg/L. Under these conditions,
about 90 percent of the chloroform was removed
within 30 minutes. ;
i
Another bench-scale study examined degradation of
chlorinated hydrocarbons, including 1-chloropentane;
1,2-DCA; 1,1-dichloropropene; and TCE, using the
O3/H2O2 process. Although the combination of the
two oxidants enhanced production of secondary
oxidants, resulting in an accelerated process, a
significant amount of the O3 applied was consumed
by natural .OH scavengers, including dissolved
organic matter, bicarbonate ion, and carbonate ion.
The researchers suggest that such problems could
be overcome by introducing an ozonation step to
eliminate the -OH scavengers before using O3 and
H2O2 to treat the more recalcitrant compounds
(Masten and Hoigne, 1992).
An O3/H2O2 process was evaluated in a bench-scale
study using PCE- and TCE-contaminated
groundwater in Oitti, Finland. The pH of the
groundwater from the contaminated well was 6.8,
and the alkalinity was 55 mg/L as CaCO3.
Groundwater samples were diluted to establish PCE
and TCE concentrations equivalent to those detected
in fhe drinking water (1 00 to 200 ug/L). Experiments
were performed using an O3 dose of 7 mg/min and a
H2O2:O3 dose ratio of 0.7 on a mass basis. Under
these conditions, 92 and 96 percent of the PCE and
TCE, respectively, were removed within 5 minutes
(Hirvonen and others, 1996).
Glaze and Kang (1988) conducted batch studies of
O3/H2O2 degradation of PCE and TCE in
groundwater collected from two wells owned and
operated by the Department of Water and Power in
Los Angeles, California, U.S.A. The groundwater
alkalinity in both wells ranged from 200 to 300 mg/L
as CaCO3, and the pH ranged from 7.2 to 7.4. The
groundwater also contained a TOC concentration of
1.1 mg/L. During a run with the initial concentrations
of PCE and TCE at 55 and 475 ug/L, respectively,
405 ug/L of chloride ions was formed in 20 minutes,
and the PCE and TCE concentrations decreased to
5 and 8 ug/L, respectively. These findings indicated
that at least 97 percent of the PCE and TCE
degraded to chloride ions during the process. High
levels of bicarbonate ions in the groundwater
significantly decreased the efficiency of PCE and
TCE removal, suggesting that softening of the
groundwater prior to oxidation might improve the
process.
The O3/H2O2 process has also been evaluated in
terms of oxidation of toluene in synthetic waste-
water. A total of 91 experiments were conducted
using a H2O2 dose range of 0.0002 to 0.02 mole per
liter (M), a toluene concentration range of 0.00075 to
0.0015 M, and a pH range of 3 to 11. The
experiments were designed to provide a molar ratio
of H2O2 to O3 up to about 120. Under acidic
conditions (a pH of 3), the process is controlled by
direct oxidation of O3 molecules, resulting in low
reaction rates. Under alkaline conditions (a pH of 10
or above), formation of .OH primarily determines the
rate of oxidation (Kuo and Chen, 1996).
Acoustic Cavitation
Acoustic cavitation was evaluated in terms of
oxidation of ethane and methane under an argon
atmosphere. A frequency of 300 kHz and an
intensity of about 2 watts per square centimeter
(W/cm2) were used to induce acoustic cavitation in
synthetic wastewater. The primary by-products of
methane oxidation included acetylene, ethane,
ethylene, and hydrogen. Butadiene, n-butane,
1-butene, n-butyne, 2-methyl-butene,
2-methyl-propane, propane, propene, and propyne
were among the other by-products observed. The
decomposition products of ethane were the same as
those of methane but had higher yields (Hart and
others, 1990).
!3-6
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In batch experiments, acoustic cavitation was
evaluated in terms of mineralization of TGE in
synthetic wastewater. A frequency of 20 kHz and a
600-watt (W) power source were used. An initial
solution pH between 3 and 11 did not affect TCE
removal, and bicarbonate concentrations of up to
10 millimoles per liter (mM) had no significant effect
on TCE destruction. Higher bicarbonate
concentrations (about 100 mM) caused a significant
reduction in the destruction rate, suggesting that
radical-mediated degradation of organics is
important in acoustic cavitation. The presence of
metal/metal oxides and H2O2 (at 1 mM) did not
enhance TCE destruction rates (Cyr and others,
1999).
In another study, acoustic cavitation was used for
TCE (3.34 mM) at frequencies of 20 and 520 kHz.
The reaction temperature was maintained at about
32 °C in the 20-kHz reactor and about 30 °C in the
520-kHz reactor. TCE degradation was energetically
more efficient at 520 kHz. The primary volatile
compounds identified as by-products include
chloroacetylene; dichloroacetylene; 1,4-dichloro-
1,3-butadiene; hexachlorobutadiene; PCE; and
pentachlorobutadiene (Drijvers and others, 1996).
Drijvers and others (1999) have used a modified
acoustic cavitation process that couples ultrasound
with H2O2 and copper oxide to degrade TCE
(3.34 mM). Acoustic cavitation was performed at a
frequency of 520 kHz using a 12-W power source
and 100 mM of H2O2 and 1 milligram per milliliter of
copper oxide were applied. The temperature was
maintained at 30 °C during acoustic cavitation. TCE
.removal was not enhanced by the presence of H2O2
and copper oxide.
3.1.2 SVOC-Contaminated Groundwater
SVOCs have been removed from groundwater using
the Fenton and hydrodynamic cavitation processes
at the commercial scale. The SCWO process has
been used to treat SVOCs in synthetic wastewater at
the pilot scale. Oxidation of SVOCs by the Fenton,
O3/H2O2, electrochemical oxidation, acoustic
cavitation, and gamma-ray processes has been
evaluated at the bench-scale level. This section
describes the results of these applications.
Commercial-Scale Applications
SVOCs in groundwater have been treated using the
Fenton and hydrodynamic cavitation processes at
the commercial scale. This section presents
performance data from field studies that involved use
of the Geo-Cleanse® Fenton and OSI HYDROX
hydrodynamic cavitation treatment systems to
remove the following SVOCs from groundwater.
ANPO Process
Fenton
Hydrodynamic
cavitation
SVOCs Removed
• PAHs
Acenaphthene;
anthracene; benzo(a)-
anthracene; chrysene;
fluoranthene; fluorene;
phenanthrene; pyrene
Geo-Cleanse9 Fenton System
In 1999, the Geo-Cleanse® in situ Fenton system
was used at a manufactured gas plant in Burlington,
Wisconsin, U.S.A., to treat PAH-contaminated
groundwater. The initial total PAH concentration in
groundwater collected from two monitoring wells,
MA/-4 and M/V-5, was 630 and 400 ug/L,
respectively. The system treated about 100 m3 of
groundwater. About 30,000 L of 50 percent H2O2;
570 L of 66 percent H2SO4; 45 L of 85 percent
H3PO4; and 23 kg of FeSO4 were injected at nine
locations in the treatment zone over 8 consecutive
days. As part of the treatment, Geo-Cleanse used
calcium phosphate as a stabilizer for the reaction;
however, the amount of stabilizer used is unknown.
The total treatment time is unknown. The
concentration of total PAHs in M/V-4 and M/V-5 was
reduced by 99 and 90 percent, respectively. No
TICs were detected when groundwater samples
were analyzed using SW-846 Method 8260.
According to Geo-Cleanse, the total treatment cost
was about $130,000, which includes equipment
rental, reagent, mobilization, and labor costs
(Geo-Cleanse, 2000).
OSI HYDROX Hydrodynamic Cavitation
System
Groundwater contaminated with furans, PAHs, and
VOCs was treated in a 7,600-L/min OSI HYDROX
hydrodynamic cavitation system. Concentrated
creosote chemicals were collected from the bottom
of wells in Visalia, California, U.S.A., and were mixed
with well water for treatment. The addition of the
chemicals rendered the well water opaque. The raw
system influent contained the following PAHs:
acenaphthene (300 mg/L), anthracene (21 mg/L),
benzo(a)anthracene (24 mg/L), chrysene (20 mg/L),
fluoranthene (100 mg/L), fluorene (120 mg/L),
phenanthrene (250 mg/L), and pyrene (160 mg/L).
During system operation, 41 kW of power was
consumed in treating the groundwater for 9 hours.
3-7
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Acenaphthene, anthracene, benzo(a)anthracene,
chrysene, fluoranthene, fluorene, phenanthrene, and
pyrene concentrations were reduced by 92, >99,
>99, >99, 70, 91, 55, and 83 percent, respectively.
According to OSI, the total treatment cost was about
$200,000, which includes the capital cost and an
O&M cost of $0.65/m3 of groundwater treated. The
estimated cost savings associated with treating the
contaminated groundwater using ANPO rather than
disposing of it by conventional methods were
>$500,000 (Pisani and others, 1997; OSI, 2000).;
Pilot-Scale Application
Li and others (1994) have studied removal of acetic
acid, n-octanol, and phenol from synthetic
wastewater at the pilot-scale level using the SCWO
process. Wastewaterwas run through a 6.1-m-long,
concentric tube reactor at flow rates of 0.75 to
1.9 L/min. Under temperature conditions of 420 to
440 °C and a pressure of 25 MPa, an initial
concentration of 4,300 mg/L of acetic acid was
reduced by >92 percent after 90 seconds of
treatment. Under similar temperature and pressure
conditions, an initial concentration of 290 mg/L of
phenol was reduced by >98 percent after
11 seconds of treatment, and an initial concentration
of 660 mg/L of n-octanol was reduced by
>99 percent after 85 seconds.
Bench-Scale Studies
This section summarizes the results of bench-scale
studies on the effectiveness of ANPO processes in
removing SVOCs from groundwater. Thb
bench-scale results are summarized only for studie|s
providing information beyond that discussed above
for the commercial- and pilot-scale applications.
This section presents bench-scale study results for
the use of the Fenton, O3/H2O2, electrochemical
oxidation, acoustic cavitation, and gamma-ray
processes to remove the following SVOCs from
groundwater and synthetic wastewater. '
ANPO Process SVOCs Removed
Fenton
03/H202
Electro-
chemical
oxidation
Acoustic
cavitation
Acenaphthene; aniline;
benzothiazole; carboxin;
chlorobenzene; 2-chIoro-
benzoate; 3-chloro-
benzoate; 4-chloro-
benzoate; p-chloro-
biphenyl; 2-CP; 3-CP;
4-CP; o-cresol; 2,3-
dichlorophenoi (DCP);
2,4-DCP; 2,5-DCP;
2,6-DCP; 3,4-DCP; 2,4-
dinitrophenol (DNP);
1,4-dioxane; fluorene;
hexamine; 2-mercapto-
benzothiazole; 2-methyl-
1,3-dioxolane; nitro-
benzene; 2-nitrophenol
(NP); 4-NP; N-nitroso-
dimethylamine; N-nitroso-
diphenylamine; phenan-
threne; phenol; 2,4,5-tri-
chlorophenol (TCP);
2,4,6-TCP
Acetone; chlorobenzenes;
EDTA; (1-hydroxyethyl-
idene)biphosphonic acid;
methanediphosphonic acid
organophosphoric acid
triesters; tetrahydrofuran-
tetracarboxylic acid; tri-
ethylene glycol dimethyl
ether
Alkylbenzene sulfonate;
chlortetracycline; 2-CP;
4-CP; 2,6-DCP; EDTA;
fatty alcohol ethoxylate
F1416-7; lignin; linear
alkylbenzene sulfonate;
2-(4-methylphenoxy)-
ethanol; nonylphenol
ethoxylate N9; phenols;
tannic acid
Acenaphthylene;
L-ascorbic acid;
chlorobenzene;
1,4-dioxane; 4-CP;
Freon-113;
2-methyl-1,3-dioxolane;
nitrobenzene; 4-NP;
sodium
pentachlorophenate
3-8
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ANPO Process
Gamma-ray
SVOCs Removed
• 2-CP; 3-CP; 4-CP;
2,4-DCP; EDTA;
nitrilotriacetic acid;
2,4,6-TCP
Fenton
Lipczynska-Kochany and others (1995) have studied
the use of Fenton's reagent to degrade 4-CP
(5 x 1CT4 M) in synthetic wastewater at the bench
scale. Experiments were conducted under the
following conditions: a FeSO4 dose of 4 x 10"5 M, a
H2O2 dose of 4 x 10-3 M, an initial pH of 7.1 to 8.2, a
contact time of 20 minutes, and a flow rate of
1 mL/min. The chemicals 4-chlorocatechol and
hydroquinone were identified as by-products.
Bicarbonate and phosphate anions slowed the
degradation process. To optimize treatment
efficiency, the researchers suggest reducing the
inhibiting influence of bicarbonate and phosphate
anions on the degradation of CPs by conducting
Fenton oxidation at a moderately low pH.
In another bench-scale study, Fenton's reagent was
applied to degrade 2,4-DNP; nitrobenzene; 2-NP;
and 4-NP in aerated synthetic wastewater. NP
degradation by the Fenton reaction was faster than
that of nitrobenzene. When the NP:ferrous chloride
(FeCI2) molar ratio was kept at about 3 (specifically,
1 x 10" M of NPs to 3.5 x 1CT5 M of FeCI2),
90 percent of the 2,4-DNP; 2-NP; and 4-NP was
destroyed in 3, 4, and 3.5 hours, respectively.
Nitrobenzene was degraded by 90 percent in
15 hours. Degradation rates increased and aromatic
oxidation by-products for all chemicals were
removed when the contaminant to FeCI2 ratio was
changed to about 0.3 (specifically, 1 x 10~* M of NPs
to 3 x 10"* M of FeCI2). An excess of H2O2 was used
in all cases (the contaminant:H2O2 molar ratio was
0.056) (Lipczynska-Kochany, 1991).
Tachiev and nthers (1998) have compared the
oxidation kinetics of 2,4-DCP; 2,4-DNPi and phenoJ
using the Fenton process and a CMHPOS in
completely mixed batch reactors. Fenton's reagent
uses Fe(ll) in free form, and CMHPOS uses Fe(lll) in
complexed form, involving DTPA, EDTA, and
nitrilotriacetic acid as chelating ligands. Reaction
rates for the Fenton process were strongly affected
by pH; the process was active only in the low acidic
range (pH <4). A pH of 3.5 facilitated optimum
oxidation. CMHPOS, however, could be applied in
the neutral pH range (pH 6 to 9) and acidic pH
ranges. Like Fenton oxidation, CMHPOS can
completely oxidize susceptible organic compounds
including anilines, CPs, NPs, and phenols.
The Fenton process was used to oxidize three PAHs
(acenaphthene, fluorene, and phenanthrene) in
organic-free water. Studies were conducted under
the following initial conditions: an Fe(ll) dose of (0 to
2 x 10-4 M), an H2O2 dose of 1Q-5 to 10~1 M), a pH of
2 to 12, a bicarbonate ion concentration up to 10"3 M,
and concentrations of commercial humic substances
up to 25 mg/L. Fe(ll) and H2O2 played a double role
during oxidation. At low concentrations, they were
initiators of -OH, but at high concentrations, they
slowed the oxidation rate. The highest efficiency
was achieved at doses of 10'3 M for H2O2 and 7 x
10~5 M for Fe(ll). The researchers found that the
process was optimal at a pH of 7; at lower pH,
formation of «OH diminished and at higher pH,
oxidation by dissolved O2 was more significant than
that by H2O2. The presence of bicarbonate ions and
humic substances in the water slowed the oxidation
rate. The researchers suggest that for natural
waters containing significant concentrations of
bicarbonates, Fenton oxidation of PAHs could be
accomplished after a decarbonation step was
performed to reduce the presence of the -OH
scavengers. The following process by-products were
identified: acenaphthene; 3,4-dihydro-2(H)
1-benzopyran-2-one; dibenzofuran; 2-ethyl-
1-naphthaphenol; fluorene; 9-fluorenol; 9-fluorenone;
0-hydroxybiphenyl; phenanthrene; 9-phenanthrenol;
and phthalic anhydride (Beltran and others, 1998b).
Groundwater containing 1,4-dioxane and
2-methyl-1,3-dioxolane was treated using Fenton
and electro-Fenton processes. Optimum oxidation
of 1,4-dioxane (1Q-3 M) in the Fenton process
occurred with an H2O2 dose of 2.5 x 10~2 M and
addition of over 10~3 M Fe(ll). Under these
conditions, 90 percent of the 1,4-dioxane was
oxidized in 10 minutes. In the electro-Fenton
process, in which H2O2 is produced by reduction of
dissolved O2 at the cathode, about 90 percent of the
1,4-dioxane (10~3 M) and 2-methyl-1,3-dioxolane
(10"3 M) was destroyed after 90 minutes of reaction,
and complete removal was observed shortly
thereafter (Takiyama and others, 1994).
A modified Fenton process involving acoustic
cavitation was used to remove 2-CP (100 mg/L) from
synthetic wastewater. Acoustic cavitation was
conducted at 20 kHz with a power input of 160 W.
The H2O2 and Fe(ll) concentrations tested ranged
from 0 to 500 and 0.5 to 10 mg/L, respectively. The
temperature in the continuously mixed reactor was
maintained at 25 °C, and the solution pH was 3.
Reactions were allowed to proceed for 100 minutes.
The percent removal increased when a higher
3-9
-------
concentration of Fenton's reagent was applied. With
Fe(ll) at 10 mg/L and H2O2 at 500 mg/L, nearly
100 percent of the 2-CP was decomposed, and
>80 percent was mineralized based on the TOG
removal. The chemical 2-chloro-p-benzoquinone
was the major process by-product (Lin and others,
1998a). i
I
Fenton's reagent treatment of phenol in synthetic
wastewaterwas studied. The initial concentration of
the contaminant was 2 x 10"3 M. The solution
temperature and pH were maintained at about 25 °C
and 4, respectively. Reaction times ranged from 5 to
60 minutes. The study showed that a pH range of 2
to 4 provided optimum destruction efficiency. In
addition, at a constant pH, an increase in ferrous
ammonium sulphate dose increased the rate of TOC
removal. When a solution of polyethyleneimine was
added to initial phenol oxidation product
(1,2-dihydroxybenzene), a precipitate formed
immediately. The researchers suggest that
precipitation of early oxidation products provides an
alternative to complete chemical oxidation
(conversion of phenol to CO2) as a means of
detoxifying phenolic effluents (Land and Ellis, 1982).
Barbeni and others (1987) have studied the
degradation of CPs—specifically, 2-CP; 3-CP; 4-CR;
3,4-DCP; and 2,4,5-TCP—using Fenton's reagent.
The study showed that increasing the concentration
of Fe(ll) enhanced the decomposition process and
that increasing the concentration of perchlorate ion
inhibited the reaction. In the absence of Fe(ll),
Fe(lll) with H2O2 had no effect on CP degradation.
Synthetic wastewater containing o-cresol (200 mg/L)
and 2,4-DNP (100 mg/L) was treated using the
Fenton process to evaluate the potential for using
chemical .oxidation to enhance anaerobic
biodegradability and reduce toxicity. The
experiments were conducted in batch stirred
reactors. The reaction pH was maintained between
3 and 4. Reaction by-products formed during
o-cresol oxidation include acetic acid, formic acid,
and oxalic acid. No by-product information is
available for 2,4-DNP oxidation. The treatment was
effective in generating biodegradable products and
reducing methanogenic toxicity (Wang, 1991).
Koyama and others (1994) have evaluated the
effectiveness of the Fenton process in treating
synthetic wastewater containing chlorinated aromatic
compounds, including 2-chlorobenzoate; 3-chloro-
benzoate;4-chlorobenzoate; p-chlorobiphenyl; 2-CPj
3-CP; 4-CP; 2,3-DCP; 2,4-DCP; and 2,5-DCP. Each
chlorinated compound was added to 100 mL of
0.3 percent H2O2 and 0.02 percent FeCI2 such that
the chlorinated compound's concentration was!
between 10 and 500 mg/L. The reaction mixture
was then incubated for 1 hour in 300-mL flasks at
60 °C and gently stirred. All the chemicals except
p-chlorobiphenyl were reduced to concentrations
below detection limits. The researchers are unsure
of the reasons for p-chlorobiphenyl's persistence but
postulate that the compound's low solubility may
have hindered oxidation. The by-products consisted
primarily of formate and oxalate.
Chou and others (1999) have studied the batch
electro-Fenton treatment of hexamine-contaminated
synthetic wastewater. When the initial pH exceeded
2.5, current efficiency dramatically decreased
because of formation of ferric hydroxide. When the
initial Fe(ll) ion concentration was about 3,000 mg/L,
the initial current efficiency of Fe(ll) generation was
almost constant (85 to 87 percent). At an initial
Fe(ll) ion concentration of about 1,000 mg/L,
however, current efficiency dropped sharply to
39 percent. The COD removal efficiency was
>94 percent after 5 hours of treatment. By-products
of hexamine oxidation included ammonium,
formaldehyde, formate, methanol, and nitrate. The
electro-Fenton process generated limited sludge
because after coagulation and pH adjustment, ferric
hydroxide sludge was reused to produce Fe(ll).
The Fenton process has been used in bench-scale
treatment of organics and inorganics in groundwater.
The groundwater contained aniline (13 mg/L);
benzothiazole (1.8 mg/L); carboxin (6.5 mg/L);
chlorobenzene (9.5 mg/L); 2-CP (0.095 mg/L);
cresols (0.12 mg/L); 2,4-DCP (0.39 mg/L); 2,6-DCP
(0.13 mg/L); 2-mercaptobenzothiazole (12 mg/L);
N-nitrosodimethylamine (0.32 mg/L); N-nitroso-
diphenylamine (4.2 mg/L); phenol (0.12 mg/L);
2,4,5-TCP (0.18 mg/L); and 2,4,6-TCP (0.04 mg/L).
Other parameters studied included COD (490 mg/L)
and TOC (120 mg/L). Optimum conditions required
a Fe(ll):H2O2 molar ratio of 0.25 and a pH of 4.
Under these conditions, the following removal
efficiencies were achieved: aniline (71 percent);
benzothiazole (87 percent); carboxin (97 percent);
chlorobenzene (95 percent); 2-CP (73 percent);
cresols (52 percent); 2,4-DCP (94 percent); 2,6-DCP
(85 percent); 2-mercaptobenzothiazole (97 percent);
N-nitrosodimethylamine (15 percent);
N-nitrosodiphenylamine (78 percent); phenol
(84 percent); 2,4,5-TCP (>97 percent); and
2,4,6-TCP (>88 percent). In addition,
trichloropropane and TOC concentrations were
reduced by 63 and 71 percent, respectively. At pHs
above 4.5, TOC removal stabilized, and significant
precipitation of ferrous and ferric hydroxides was
observed (Kochany and Lugowski, 1998).
3-10
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O/H2O2
The effectiveness of the O3/H2O2 process along with
that of other advanced oxidation processes was
examined in treating acetone (5 mg/L.) in a 1-L batch
reactor. H2O2 was added to acetone solutions at
doses of 9.4, 46, 100, and 860 mg/L within buffer
solutions. For each H2O2 concentration, ozonated air
with an O3 concentration of 2.3 percent was sparged
into the reactor. Increasing the H2O2 concentration
while using a constant O3 concentration improved
overall degradation of acetone to a certain extent. A
relatively high H2O2 concentration (860 mg/L) had an
adverse effect on acetone degradation, probably
because of the scavenging effect of excess H2O2 on
•OH. Reducing the O3 concentration reduced the
efficiency of the process. Acetone degradation was
maximized when an H2O2 dose of 10 mg/L (added
semicontinuously) and an O3 dose of 2 percent by
weight were combined. The acetone concentration
was reduced to below the detection limit in
60 minutes of treatment (Toro and others, 2000).
Cortes and others (1996) have investigated use of
O3/H2O2 to remove chlorobenzenes (1 to 10 mg/L)
and TOG (1,000 mg/L). Because of the electrophilic
character of.OH, chlorobenzene was oxidized faster
than dichlorobenzene, tetrachlorobenzene, and
trichlorobenzene. The halogen groups deactivated
the aromatic ring attack by the O3 molecule. As a
result, the more chlorine atoms the compounds had,
the lower the removal efficiency. The chemicals
2-CP and 4-CP were identified as the first oxidation
by-products.
Synthetic wastewater containing chelating agents,
including EDTA, (l-hydroxyethylidene)biphosphonic
acid, methanediphosphonic acid, and
tetrahydrofurantetracarboxylic acid was treated
O3/H2O2 in'a bench-scale study. Although EDTA
could be destroyed with O3 alone, removal of the
intermediates required the presence of H2O2.
Oxidation decreased with increasing pH, probably
because of scavenging of -OH by carbonate ions. At
low pHs, however, the overall reaction slowed down.
Hence, in any practical application, it is necessary to
choose an optimal pH for the reaction (Appelman
and others, 1996).
Echigo and others (1996) have compared the
effectiveness of several advanced oxidation
processes, including the O3/H2O2 process, in treating
organophosphoric acid triesters (20 mg/L) in
synthetic wastewater. Wastewater samples were
recirculated in a semi-batch system at a flow rate of
3.6 to 4.0 L/min. The treatment involved an O3 dose
rate of 0.62 mg/L-min and an initial H2O2
concentration of 3.4 mg/L. Decomposition rates
increased linearly with the O3 concentration and
decreased with an increase in the initial contaminant
concentration. Nitrate ions at low concentrations did
not affect oxidation rates.
Beschkov and others (1997) have evaluated the
effectiveness of the O3/H2O2 process in treating
synthetic wastewater contaminated with triethylene
glycol dimethyl ether, a typical component of oil
reclamation wastewaters. Solutions containing
10 mg/L of triethylene glycol dimethyl ether and
50 mg/L of humic acid were treated in a completely
mixed batch tank. H2O2 addition increased the
efficiency of organic solute removal by ozonation.
Continuous addition of 7 percent H2O2 resulted in
removal of triethylene glycol dimethyl ether to
nondetectable concentrations in 5 minutes and
removal of >85 percent of the initial TOC
concentration in 80 minutes.
Electrochemical Oxidation
Synthetic wastewater containing chlortetracycline,
EDTA, lignin, and tannic acid, each at a
COD-equivalent concentration of 2,500 plus or minus
(±) 200 mg/L, was treated using electrochemical
oxidation. Experiments were conducted in a 600-mL
batch electrolytic cell with lead dioxide-coated
titanium as the anode and a steel plate as the
cathode. The study showed that 5,000 mg/L of
sodium chloride at a current density of
7,500 milliamperes per square centimeter (mA/cm2)
was a better supporting electrolyte than sodium
sulfate and sodium nitrate at the same concentration
and current density. In addition, COD and color
removal was improved by increasing the sodium
chloride concentration and current density. COD
removals for chlortetracycline, EDTA, lignin, and
tannic acid were 92, 66, 79, and 89 percent,
respectively, when sodium chloride was the
supporting electrolyte. EDTA is colorless, but color
removals for chlortetracycline, lignin, and tannic acid
were 98, 95, and 91 percent, respectively, in the
presence of sodium chloride. Also, microtox test
results also showed that electrochemical oxidation
reduced the toxicity of the refractory compounds
(Chiang and others, 1997).
Huang and Chu (1991) have evaluated the
effectiveness of the electrochemical oxidation
process in treating synthetic wastewater containing
phenolic compounds. The batch reactions took
place in a 600-mL beaker at a constant temperature
of 25 °C. Most of the phenolic compounds with a
carboxylic group or another hydroxyl group were
readily oxidized on the platinum electrode.
Phenolate anions and phenoxium cations were
formed. Anodic currents were higher and lasted
3-11
-------
longer in alkaline solutions than in acid solutions.
The removal rates for organic compounds were
proportional to the anodic current. Among the
phenolic compounds tested, the phenols with the
greatest number of hydroxy functional groups were
the easiest to oxidize; the greater the number of
hydroxy groups, the higher the anodic current.
Electrochemical removal of 2-CP and 2,6-DCP was
investigated using porous carbon felt for the anodes.
Organic solutions were circulated through ah
electrolytic cell for 2 to 3 hours. The first oxidation
intermediates of the 2-CP and 2,6-DCP solutions
were 2-chlorobenzoquinone and 2,6-dichloro-
benzoquinone, respectively. The flow rate did not
significantly affect reaction kinetics (Polcaro and
Palmas, 1997).
Boudenne and Cerclier (1999) have examined use of
carbon black as a catalyst in electrochemical
oxidation of 4-CP. Carbon, black was selected
because of its electrical conductivity, resistance to
corrosion, low price, chemical inertia toward many
chemical products, and high adsorption capacity for
phenolic compounds. The reactor consisted of
titanium/platinum electrodes in a closed, 2-L cell.
The temperature was maintained at 25 °C, the flow
rate of the solution through the system was
7.5 L/min, and the current intensity was 15,000 mA.
When 0.5 gram of carbon black was used,
by-products detected included benzoquinone,,
catecnol, 2-CP, fumaric acid, hydroquinone, maleic
acid, phenol, and o-quinone. The number qf
by-products detected decreased when a larger
amount of carbon black was used.
Treatment of synthetic surfactant wastewater
containing fatty alcohol ethoxylate F1416-7,
2-(4-methylphenoxy)-ethanol, and nonylphenol
ethoxylate N9 was studied using a batch reactor with
a lead dioxide anode. The overall surface area of
the titanium-supported electrodes was 500 square
centimeters (cm2). The initial surfactant
concentration was 100 mg/L. The temperature was
maintained at 22 °C, and the pH was maintained at
7.7 to 7.9. After a 1-hour oxidation period with a
current density of 10 mA/cm2, ethylene glycql
hexadecanoate, ethylene glycol tetradecanoate,
hexadecanoicacid, hexadecanol, tetradecanoicacid,
and tetradecanol were identified as by-products of
fatty alcohol ethoxylate F1416-7 oxidation. The
chemical 4-methylphenol and its degradation product
4-methyl-4-hydroxycyclohexa-2,5-dienon were
identified as by-products of 2-(4-methylphenoxy)-
ethanol. The by-products of nonylphenol ethoxylate
N9 oxidation were not clearly identified.
Leu and others (1998) have studied enhanced
electrochemical oxidation of two anionic surfactants,
alkylbenzene sulfonate (10 mg/L) and linear
alkylbenzene sulfonate (10 mg/L), with addition of
H2O2. The electrolytic cell was a 1.5-L beaker
containing cast-iron plate electrodes with a total
surface area of 22.6 cm2. Table salt was added to
elevate the conductivity of the sample surfactant
solution. Addition of 100 mg/L of H2O2 was found to
be highly beneficial for alkylbenzene sulfonate
oxidation; the optimum amount of H2O2 added was
150 mg/L for linear alkylbenzene sulfonate oxidation.
Other optimum operating conditions included
addition of 0.05 M sodium chloride; a current density
of 16.8 mA/cm2, a pH of 7, and 8 minutes of
treatment time.
Treatment of synthetic wastewater containing phenol
(1,000 and 2,520 mg/L of COD) by electrochemical
oxidation was studied. The cathode and anode
consisted of four pieces of cast iron situated about
1.5 cm apart in the synthetic wastewater. The total
effective surface area of the cast-iron electrodes was
324 cm2. Salinity was adjusted using table salt (up
to 3 percent). COD removal reached about
34 percent after 60 minutes of treatment. Addition of
a small amount of H2O2 greatly enhanced COD
removal; after 60 minutes of electrochemical
treatment with 60 mg/L of H2O2, COD removal
increased to 70 percent. A peak COD removal
efficiency of about 70 percent occurred at a pH of
about 3. Specific information on phenol removal
efficiency is not available (Lin and others, 1998c).
Acoustic Cavitation
Acoustic cavitation decomposition of acenaphthylene
was evaluated at a frequency of 20 kHz and a power
input of 400 W. Aqueous extracts of contaminated
soil collected from an old gasoline station site in
Germany were treated using acoustic cavitation.
The primary by-products identified were 1,1,2-TCA
and 1,1,2,2-tetrachloroethane (Leonhardt and Stahl,
1998).
Gonze and others (1999) have evaluated acoustic
cavitation as a preoxidation treatment to support
biological degradation of sodium penta-
chlorophenate. Acoustic cavitation was conducted
at a frequency of 500 kHz, a power input of 0 to 100
W, and a duration of up to 10 hours. The pH of the
sodium pentachlorophenate solution was adjusted
to the 6.8 to 7.5 range. The reaction temperature
was maintained at about 20 °C. Degradation
by-products included tetrachlorobenzoquinone and
tetrachlorohydroquinone. Toxicity effects on marine
bacteria (Vibrio fischeri) and daphnids (Daphnia
magna) were evaluated after acoustic cavitation to
13-12
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measure the acute toxicity of organic compounds.
Bioluminescence inhibition increased with the
pollutant concentration. Toxicity was lowest after 2
and 4 hours when the power density was 220 and
110 kilowatts per cubic meter, respectively.
Hirai and others (1996) have examined
decomposition of Freon-113 (100 mg/L) in synthetic
wastewater by acoustic cavitation at a frequency of
200 kHz and a power density of 6 watts per cubic
centimeter. Decomposition rates increased with
increasing chlorofluorocarbon concentration.
Decomposition efficiency decreased with increasing
acoustic cavitation time, particularly after 30 minutes;
this trend was likely due to the rise in temperature
and the degassing effect of acoustic cavitation. The
decomposition rate was minimally affected by
addition of tert-butyl alcohol, a known *OH
scavenger, suggesting that the primary pathway in
acoustic cavitation decomposition of
chlorofluorocarbons is not the reaction with -OH but
the high-temperature pyrolysis in the cavitation
bubbles.
Barbier and Petrier (1996) have conducted a study
of 4-NP degradation by acoustic cavitation combined
with addition of O3. Reactions were conducted in a
continuous feed reactor at a pH of 2 and a
temperature of 20 °C. The TOC concentration was
more dramatically reduced during acoustic cavitation
at a frequency of 500 kHz than at 20 kHz. This
phenomenon correlates with higher yields of H2O2
formed during acoustic cavitation at the higher
frequency that in turn correspond to a higher rate of
•OH escape from the cavitation bubbles.
Takiyama and others (1994) have evaluated removal
of 1,4-dioxane and 2-methyl-1,3-dioxolane from
groundwater using acoustic cavitation alone and
coupled with H2O2 addition. Acoustic cavitation was
conducted at a power intensity of 400 W, and the
temperature was maintained at 40 °C. About
50 percent removal of each compound (which had
an initial concentration of 10"3 M) was achieved after
1 hour of treatment. When H2O2 (5 x 1Q-3 M) was
introduced into the system, the oxidation process
was improved because H2O2 initiates increases in
the concentration of -OH in solution. When H2O2
was added, 70 percent removal of each compound
was achieved after 1 hour of treatment.
The acoustic cavitation process was used to degrade
L-ascorbic acid in double-distilled water. Acoustic
cavitation was conducted at a frequency of 800 kHz.
The primary by-products observed were
L-erythro-2,3-hexodiulosonic acid and
L-glycero-4-hexulos-2-enonic acid. The secondary
by-products were glyceric acid, 4-pentulos-2-enonic
acid, 2-pentulosonic acid, and tetronic acid. The
tertiary by-product was tetraric acid (PortenlSnger
and Heusinger, 1992).
Drijvers and others (1998) have studied organic
intermediate formation under saturated air and argon
conditions during acoustic cavitation degradation of
chlorobenzene. -The contaminant was present in
synthetic wastewater at 1.72 mM. Acoustic
cavitation was performed at a frequency of 520 kHz
using a power input of 14.23 W, and the reaction
temperature was maintained at about 30 °C.
Solutions were treated at pH levels of 4.7,7, and 10.
No pH effect was observed. With the addition of the
•OH scavenger benzoate, no significant degradation
took place. Several organic by-products of
chlorobenzene degradation were identified under
air-saturated conditions, including acetylene,
benzene, butadiene, butenyne, CPs, methane, and
phenylacetylene. Under argon-saturated conditions,
the same by-products were identified exceptforCPs.
Removal of aromatic compounds such as 4-CP,
nitrobenzene, and 4-NP has been studied using
acoustic cavitation combined with ozonolysis.
Acoustic cavitation was performed at frequencies of
20 kHz (56.1 W of power) and 500 kHz (48.3 W of
power). The solution temperature was maintained at
about 25 °C. In the 20-kHz reactor, nitrobenzene
degraded fastest and 4-NP degraded slowest, but in
the 500-kHz reactor, 4-CP degraded fastest while
4-NP degraded slowest. When acoustic cavitation
and ozonolysis were combined, there was a
degradation enhancement at 20 kHz, but at 500 kHz,
the process slowed down. Treatment resulted in
contaminant mineralization in 3 hours with the
20-kHz reactor and in 6 hours with the 500-kHz
reactor. Aromatic intermediates observed during
acoustic cavitation with O3 were as follows: (1) for
nitrobenzene—4-NP, 3-NP, and 4-nitrocatechol; (2)
for 4-NP-^-nitrocatechol; and (3) for 4-CP—none
(Weavers and others, 1998).
Gamma-Ray
Zona and others (1999) have evaluated
detoxification of several CPs (6 to 10 mg/L) in
solution by the gamma-ray process, including 2-CP;
3-CP; 4-CP; 2,4-DCP; and 2,4,6-TCP. Gamma
irradiations were conducted using a 60Co source and
dose rates ranging from 21 to 183 megarads (Mrad)
per minute (Mrad/min). A radiation dose of 50 Mrads
resulted in concentrations below detection limits for
all the compounds except 2-CP, for which 60 Mrads
was required to achieve this result. No influence of
the dose rate on the degradation rate was observed
in the dose rate range studied. To determine the
acute toxicity of the aqueous solutions, a
3-13
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luminescent bacteria test was conducted. Total
detoxification was achieved for all compounds
studied at a dose of 500 Mrads.
Colloidal TiO2 was used to catalyze the degradation
of EDTA by the gamma-ray process in a bench-scale
study. Gamma irradiations were conducted using a
""Co source, at room temperature, and at a dose rate
of about 4.7 Mrad/min. Aqueous solutions of 0.02 M
EDTA were used; these solutions had an adjusted
pH between 4 and 5. Colloidal TiO2 was added to
the solutions at concentrations ranging from 0.01 to
0.1 gram per milliliter. Results showed that the
presence of the catalyst increased EDTA
degradation by up to 30 percent. The study also
showed that dissolved O2 is critical to tliie
degradation process, indicating that the process
involves electron/hole pairs (Su and others, 1998).
A bench-scale study conducted by Toste (199J3)
involved gamma-ray destruction of nitrilotriacetic acid
(12 mM), which was added to an inorganic waste
matrix. Almost 100 percent of the nitrilotriacetic acid
was destroyed after 100 hours of irradiation ata
dose rate of 75,000 rad/hr in a 60Co source; the
operating temperature was 90 to 95 °C. Under these
conditions, about 91 percent of the TOC was
removed. Degradation by-products identified include
N-[N'-amino(2-iminoethyl)]-iminodiacetic acid,
ethanedioic acid, hexanoic acid, N-hydroxymethyl-
N-methyliminoacetic acid, N-(methylamine)imino-
diacetic acid, and N-(methyl-N,N'-(dimethylamine)-
ethylenediamine)N'-acetic acid.
3.1.3 PCB-Contaminated Groundwater
PCBs have been removed from groundwater using
the commercial-scale OSI HYDROX hydrodynamic
cavitation treatment system. Groundwater
containing about 3 mg/L of PCBs was treated in a
230-L/min OSI HYDROX system in Munich,
Germany. The raw groundwater pH was 8.2. About
7 kW of power was consumed in treating the
groundwater for 8 hours. Hydrodynamic cavitation
reduced the PCB concentration by >99 percent and
achieved the treatment goal of 20 ug/L. According to
OSI, the total treatment cost was about $65,000,
which includes capital cost and an O&M cost of
$0.07/m3 of groundwater treated (OSI, 2000).
3.1.4 Pesticide-and
Herbicide-Contaminated
Groundwater
This section presents information on removal of the
following pesticides and herbicides from aqueous
solutions using the Fehton, O3/H2O2, and acoustic
cavitation processes at the bench-scale level. No
information is available on commercial- or pilot-scale
application of ANPO processes to treat pesticide- or
herbicide-contaminated groundwater.
ANPO Process
Fenton
O3/H2O2
Acoustic
cavitation
Pesticides and
Herbicides Removed
Chlorophenoxyacid
herbicides
Aldicarb; aldrin;
alpha-endosulfan;
atrazine;
hexachlorobenzene;
isoproturon; lindane;
linuron; malathion;
monocrotophos;
m-parathion; terbutryn
Alachlor; atrazine;
chlorpropham; parathion
Fenton
Oturan and others (1999) have used the
electro-Fenton process to degrade five
Chlorophenoxyacid herbicides: 2-(4-chloro-2-methyl-
phenoxy)propionic acid (1.0 mM); 2-(4-chloro-
phenoxy)-2-methylpropionic acid (1.0 mM); 2,4-
dichlorophenoxyacetic acid (1.0 mM); 2-(2,4-
dichlorophenoxy)propionic acid (1 .0 mM); and 2,4,5-
trichlorophenoxyacetic acid (0.5 'mM). Fe(ll) was
present at 2 mM. -OH were electrogenerated in two
steps: (1) electrochemical reduction of O2 to
superoxide ions and formation of H2O2 in an acidic
medium and (2) reaction of the H2O2 with Fe(Il) ions,
yielding -OH (the Fenton reaction). Fe(lll) ions were
electrochemically reduced to Fe(ll) ions. The
electro-Fenton process allowed the aromatic ring of
the Chlorophenoxyacid herbicides to be destroyed
following rapid polyhydroxylation, resulting in
formation of less toxic aliphatic products. The
researchers state that the modified Fenton process
may provide a low-cost, nonpolluting method for
destroying pesticide residues.
A bench-scale study was conducted to examine
degradation of atrazine and its primary by-products,
deethylatrazine and deisopropylatrazine, using the
O3/H2O2 process. Optimum oxidation of these
compounds took place at a pH of 7, a temperature of
20 °C, and an initial H2O2 concentration of 10'3 M.
Increasing the H2O2 concentration above 10"2 M led
to a significant decrease in the oxidation rate. The
!3-14
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optimum molar ratio of O3:H2O2 was 0.33 (Beltran
and others, 1998a).
Ku and Wang (1999) have studied O3/H2O2
decomposition of monocrotophos in deionized water.
Results showed that the herbicide was >95 percent
decomposed by the O3/H2O2 process within
20 minutes. CO2, nitrate, and phosphate formed as
by-products. Monocrotophos removal increased with
a decrease in pH, possibly because of the breakage
of the carbon-carbon double bond to form various
organic intermediates. However, mineralization of
organic intermediates was promoted under alkaline
conditions because of formation of -OH. The mode
of H2O2 addition was observed to affect degradation
of monocrotophos. In the batch mode, increased
H2O2 dose reduced the monocrotophos
decomposition rate, possibly because the large
amounts of H2O2 added initially served as a
scavenger for -OH. In the continuous mode, the
increased H2O2 dose accelerated the decomposition
rate until an optimum molar ratio of O3 to H2O2 was
reached; the.optimum ratio was pH-dependent.
The influence of O3 and H2O2 dose was studied
under batch conditions for the degradation of 11
pesticides, including aldicarb (9.5 ug/L), aldrin
(13 M9/L), alpha-endosulfan (8.7 ug/L), atrazine
(7.3 ug/L), hexachlorobenzene (1 M9/L), isoproturon
(10 ug/L), lindane (10 ug/L), linuron (11 ug/L),
malathion (11 ug/L), m-parathion (11 ug/L), and
terbutryn (10 M9/L) At an H2O2:O3 mass ratio of 0.4,
an initial pH of 8.3, and a retention time of
10 minutes, the highest percent removals for the
pesticides at an O3 dose of 5 mg/L were as follows:
aldicarb, >99 percent; aldrin, >99 percent;
alpha-endosulfan, 49 percent; atrazine, 95 percent;
hexachlorobenzene, 48 percent; isoproturon,
>99 percent; lindane, 21 percent; linuron, 99 percent;
malathion, >99 percent; m-parathion,>99 percent;
and terbutryn, >99 percent (Roche and Prados,
1995).
Volk and others (1993) have examined the effects of
the O3/H2O2 process on formation of biodegradable
dissolved organic carbon during treatment of
atrazine-contaminated synthetic water. Based on
dissolved organic carbon removal results, the study
showed that an H2O2:O3 mass ratio of 0.35 to 0.45
was optimal for oxidation of organic matter. As the
ratio increased to 0.4, the amount of biodegradable
dissolved organic carbon formed was found to
diminish.
Acoustic Cavitation
David and others (1998) have used acoustic
cavitation to degrade chlorpropham in synthetic
wastewater. Acoustic cavitation was performed at
two frequencies, 20 and 482 kHz. Treatment at the
higher frequency was determined to be much more
efficient for chlorpropham destruction, as the
herbicide completely degraded after 45 minutes. At
a frequency of 20 kHz, about one-third of the
chlorpropham remained after 60 minutes of acoustic
cavitation. By-products of chlorpropham degradation
included chloride ions, 3-chloroaniline, CO, CO2, and
formic acid. Formation of chlorohydroquinone was
also observed during degradation of 3-chloroaniline.
The kinetics of acoustic cavitation decomposition of
alachlor (0.7 mg/L) and atrazine (0.8 mg/L) have
been evaluated by Koskinen and others (1994).
Continuous acoustic cavitation of the herbicides in
solution was conducted using 70 to 80 W of power at
a frequency of 20 kHz. Acoustic cavitation did not
decrease the solution pH for either herbicide as
reported by Kotronarous and others (1992a) for
parathion decomposition. Koskinen and others
(1994) showed that the half-lives of alachlor and
atrazine were 330 and 86 minutes, respectively.
Kotronarous and others (1992a) have examined
parathion removal from synthetic wastewater by
acoustic cavitation at a frequency of 20 kHz and a
power intensity of 75 W/cm2. The temperature was
maintained at 30 °C. Nitrate, nitrite, p-NP, oxalate,
phosphate, and sulfate were identified as
by-products of parathion degradation. The pH of the
solution changed during acoustic cavitation; after
30 minutes, the pH dropped from 6.1 to 4.1 and then
stabilized. The observed decrease in pH was
attributed to acid formation during acoustic
cavitation.
3.1.5 Dioxin- and Furan-Contaminated
Groundwater
Dioxins and furans have been removed from
groundwater using the commercial-scale OSI
HYDROX hydrodynamic cavitation treatment system.
Groundwater containing furans, PAHs, and VOCs
was treated in a 7,600-L/min OSI HYDROX system.
Dibenzofuran was present in raw water at 110 mg/L.
Concentrated creosote chemicals were collected
from the bottom of wells in Visalia, California, U.S.A.,
and were mixed with well water for treatment. The
addition of the chemicals rendered the well water
opaque. About 41 kW of power was consumed in
treating the groundwater for 9 hours. Treatment
continued until the water became clear and the
3-15
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dibenzofuran concentration reached an acceptable
discharge level. Hydrodynamic cavitation reduced
the concentration of the dibenzofuran by 91 percent.
According to OSI, the total treatment cost was abdut
$200,000, which includes the capital cost and an
O&M cost of $0.65/m3 of groundwater treated. The
estimated cost savings associated with treating the
contaminated groundwater using ANPO rather than
disposing of it by conventional methods were
>$500,000 (Pisani and others, 1997; OS! 2000). i
3.1.6 Explosive-and Degradation \
Product-Contaminated \
Groundwater |
Explosives and their degradation products 'in
groundwater have been removed using the Fentoi,
SCWO, and E-beam processes at tre
commercial-scale level. Use of the Fenton and
gamma-ray processes to remove such contaminants
in synthetic wastewater has been studied at thje
bench-scale level. The results of the
commercial-scale applications and bench-sea e
studies are discussed below.
Commercial-Scale Applications
This section summarizes the effectiveness of the
Geo-Cleanse® Fenton, General Atomics SCWO, arid
HVEA E-beam treatment systems in removing the
following explosives and their degradation products
from contaminated groundwater.
ANPO Process
Fenton
• SCWO
E-beam
Explosives and
Their Degradation
Products Removed
Hexahydro-1 ,3,5-trinitro-
1,3,5-triazine(RDX);
octahydro-1 ,3,5,7-tetra-
nitro-1 ,3,5,7-tetrazocine
(HMX); 2,4,6-trinitro- ;
toluene (TNT)
CYH propellant containing ;
HMX
• DMMP
Geo-C/eanse® Fenton System I
A Geo-Cleanse® in situ Fenton system was used to
treat contaminated groundwater under an outwash
area of a production line at the Milan Army
Ammunition Plant in Tennessee, U.S.A. The primary
contaminants of concern in the groundwater were
explosives; specifically HMX, RDX, and TNT were
present at a total of 140 ug/L. The system treated
about 1,100 m3 of groundwater. About 6,800 L of
50 percent H2O2; 76 L of 66 percent H2SO4; 15 L of
85 percent H3PO4; and 23 kg of FeSO4 were injected
at three locations in the treatment zone over
3 consecutive days. As part of the treatment,
Geo-Cleanse used calcium phosphate as a stabilizer
for the reaction; however, the amount of stabilizer
used is unknown. After a treatment time of 60 days,
the treatment reduced the total explosives
concentration by 80 percent. Nitrate was identified
as a treatment by-product. According to
Geo-Cleanse, the total treatment cost was about
$97,000, which includes equipment rental, reagent,
mobilization, and labor costs (Geo-Cleanse, 2000).
General A tomics SCWO System
A General Atomics SCWO system installed at a
Thiokol site near Brigham City, Utah, U.S.A., was
demonstrated for the treatment of CYH propellant
containing HMX. Treatment was performed in two
test runs involving up to 315 kg of propellant, which
contained up to 21 percent hydrolyzed solution
during a continuous 34-hr run. Treatment flow rates
ranged from 1.1 to 1.7 L/min, and reaction
temperatures ranged from 450 to 580 °C. System
pressure was maintained at about 28 MPa.
Removals were measured based on the amount of
TOC destroyed. At a temperature of 575 °C and a
retention time of 33 seconds, >99 percent of the
TOC was destroyed (Spritzer and others, 1995).
HVEA E-Beam System
In full-scale experiments, HVEA E-beam treatment
system was used to remove DMMP from solutions
with pHs of 4 and 9. Wastewater samples were
recirculated at 400 L/min for about 2 hours. The
delivered radiation dose was constant at 0.550 Mrad
per pass; the final absorbed doses were 3 (pH of 4)
and 2 (pH of 9) Mrads. DMMP removal was more
effective at a pH of 4 than at a pH of 9 because of
reduced radical scavenging by carbonate species.
Diphosphonic acid, methane, and phosphate ions
were formed as by-products. At a pH of 4, an
influent DMMP concentration of 47 mg/L, and an
absorbed dose of 3 Mrads, the removal efficiency
was 89 percent. At a pH of 9, an influent DMMP
concentration of 62 mg/L, and an absorbed dose of
2 Mrads, the removal efficiency was 74 percent
(Nickelsen and others, 1998).
Bench-Scale Studies
This section summarizes the results of bench-scale
studies on the effectiveness of the Fenton and
3-16
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gamma-ray processes in removing the following
explosives and their degradation products from
synthetic wastewater.
- ; . '. .
ANPO Process
Fenton
• Gamma-ray
Explosives and
Their Degradation
Products Removed
2,4-Dinitrotoluene; RDX;
TNT
• DMMP
Fenton
Bier and others (1999) discuss Fenton treatment in
synthetic solutions containing RDX. Initial RDX
concentrations ranged from 20 to 35 mg/L. About
2 ml of 30 percent H2O2 and 0.5 mL of 143 mM
Fe(ll) were added to about 48 mL of RDX solution.
Under these conditions, complete RDX destruction
was observed within 24 hours, with RDX
transformation occurring most rapidly within the first
300 minutes at a pH of 3. The primary by-products
observed were ammonium ions, CO2, formic acid,
methylene dinitramine, and nitrate ions.
In a related study, Li and others (1997) examined
removal of TNT by Fenton oxidation. Treatment of
synthetic wastewater containing TNT (70 mg/L) with
Fenton's reagent (1 percent H2O2 and 80 mg/L of
Fe[ll]) at a pH of 3 resulted in complete destruction
of the explosive within 24 hours; 40 percent
mineralization was also observed. The chemicals
1,3,5-trinitrobenzene and 2,4,6-trinitrobenzoic acid
were formed after 15 minutes of treatment. Oxalic
acid was the primary organic by-product of the
oxidation. Increasing the initial H2O2 concentration,
Fe(ll) concentration, or both increased both TNT
removal and mineralization rates. The initial TNT
concentration (6.2 to 65.5 mg/L) had little influence
on degradation rates. The optimum pH for 'the
reaction was 3.
In another bench-scale study, Fenton's reagent was
used to oxidize 2,4-dinitrotoluene. At H2O2:dinitro-
toluene:Fe(ll) molar proportions of 20:1:2.5, the
compound was completely removed in 5 hours at
21 °C. Reaction by-products included benzoic acid;
1,1 '-biphenyl,3,3',4,4'-tetramethyl; 1,3-dinitro-
benzene; 1-isocyanato,3-nitrobenzene; and nitro-
aniline. At higher temperatures, H2O2 was observed
to deplete more quickly, and the TOC removal rate
was enhanced. The required retention time
decreased when the temperature was increased to
about 30 °C and when Fe(lll) was used in
conjunction with Fe(ll). Aeration of the reaction
mixture decreased the concentrations of dimers in
the reaction by-products (Mohanty and Wei, 1993).
Gamma-Ray
Nickelsen and others (1998). treated synthetic
wastewater spiked with DMMP in a bench-scale
study using the gamma-ray process and a 60Co
source. Treatment was conducted at pHs of 4 and
9. DMMP removal was more effective at a pH of 4
because of reduced -OH scavenging by carbonate
species. At a pH of 4, an initial DMMP concentration
of 1.4 mg/L, and an absorbed dose of 0.27 Mrad, the
removal efficiency was 86 percent. At a pH of 9, an
initial DMMP concentration of 35 mg/L, and an
absorbed dose of 1 Mrad, the removal efficiency was
65 percent.
3.1.7 Humic Substance-Contaminated
Groundwater
No information is available on commercial-scale
ANPO processes used for removing humic
substances from groundwater. However, the
O3/H2O2 process has been evaluated at the pilot
scale, and O3/H2O2 and acoustic cavitation
processes have been evaluated at the bench scale.
The results of these evaluations are summarized
below.
Pilot-Scale Application
Groundwater containing high color levels (up to 25
color units [c.u.]) from redwood chips was treated
using the O3/H2O2 process in a pilot-scale study in
southern California, U.S.A. To achieve the U.S. EPA
secondary standard of 15c.u.,the Huntington Beach
Water Department in Orange County, California,
U.S.A., used a 76-L/min pilot-scale O3 generator with
a retention time of 13 minutes. The H2O2:O3 mass
ratios used ranged from 0.2 to 0.5. Color reduction
of >90 percent was achieved at an average transfer
O3 dose of 6.5 mg/L and an H2O2 dose of about
3.3 mg/L. On a larger scale, the Mesa Consolidated
Water District in Orange County, California, U.S.A.,
operated a 12,000-L/min O3 treatment system with a
retention time of 16 minutes. At an applied O3 dose
of 7 mg/L and an H2O2:O3 mass ratio of 0.35, the
system reduced the color level (initially 65 c.u.) in
groundwater by 85 percent (Tan and others, 1990).
Bench-Scale Studies
This section summarizes the results of bench-scale
studies of the effectiveness of the O3/H2O2 and
acoustic cavitation processes in removing the
3-17
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following humic substances from synthetic
wastewater and groundwater. I
ANPO Process
03/H202
Acoustic
cavitation
Humic Substances
Removed
Glycine
Humic acid; 3-hydroxy-
benzoic acid; purified
fulvic acid; unspecified,
naturally occurring
organic carbon
Begnerand others (1999) have studied oxidation 6f
glycine in synthetic wastewater by the O3/H2O2
process in a 4-L, cylindrical batch reactor.
By-products included ammonium ions, formic acid,
oxalic acid, and oxamic acid.
Acoustic Cavitation
Nagata and others (1996) studied the decomposition
of humic acid and 3-hydroxybenzoicacid in synthetic
wastewater by the acoustic cavitation process as
well as the potential for chloroform formation.
Wastewater samples were treated at a frequency of
200 kHz with a power input intensity of 200 W.
Decomposition of 3-hydroxybenzoic acid was almost
completely inhibited by addition of butanol, which is
an effective «OH scavenger. The potential for
chloroform formation decreased with increasing
treatment time, but the decrease in the chloroform
formation potential did not correspond to the
decrease in humic acid and 3-hydroxybenzoic acid
concentrations on a one-to-one basis.
Acoustic cavitation was combined with use of O3 to
mineralize natural organic matter in a bench-scale
batch reactor. Study samples consisted of synthetic
wastewater containing purified fulvic acid (10 mg/L)
and untreated natural groundwater (with TOG
concentrations ranging from 2 to 8 mg/L) drawn from
an aquifer in Orange County, California, U.S.A,
About 91 percent of the TOC was removed from the
synthetic wastewater samples in 60 minutes of
continuous O3 application and acoustic cavitation at
a frequency of 20 kHz and a power input of 55 W;
Acoustic cavitation and O3 treatment of natural
groundwater samples (the pH was adjusted to 4 in
order to remove bicarbonate) resulted in TOC
removal of 98 percent after 40 minutes. The O3
decomposition rate was significantly increased by
acoustic cavitation power of up to 70 kW; no further
rate increase was observed at higher powers (Olson
and Barbier, 1994).
3.1.8 Inorganic-Contaminated
Groundwater
This section discusses removal of inorganics from
groundwater and synthetic wastewater at the
bench-scale level using the Fenton and O3/H2O2
processes. No information was available on use of
commercial- or pilot-scale ANPO processes to treat
inorganic-contaminated groundwater.
ANPO Process
Fenton
• 03/H202
Inorganics Removed
Ammonia
Ammonia .
Fenton
The Fenton process was used for bench-scale
treatment of groundwater containing inorganics and
organics. Ammonia was present at a concentration
of 290 mg/L. Other parameters studied include COD
(490 mg/L) and TOC (120 mg/L). Optimum
treatment conditions included a Fe(ll):H2O2 molar
ratio of 0.25 and a pH of 4. Under.these conditions,
ammonia, COD, and TOC were reduced by 10, 63,
and 71 percent, respectively. At pHs above 4.5,
TOC removal stabilized, and significant precipitation
of ferrous and ferric hydroxides was observed
(Kochany and Lugowski, 1998).
0/W202
Kuo and others (1997) have compared the O3/H2O2
process to traditional ozonation in terms of removal
of ammonia from synthetic wastewater. Studies
were conducted using a stopped-flow spectra-
photometer system operating at a temperature of
25 °C and aqueous solutions with pHs varying from
8 to 10. Study results indicated that O3/H2O2
oxidation was mediated primarily by -OH oxidation of
ammonia and that the O3 depletion rate was not
significantly influenced by the ammonia
concentration. The researchers suggest that adding
H2O2 to an ozonation process conducted at a pH of
11 or less would be economical for treating
wastewaters containing high concentrations of
ammonia.
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3.2 Industrial Wastewater Treatment
The effectiveness of ANPO processes in treating
industrial wastewater has been evaluated for various
contaminant groups, including SVOCs, dyes,
inorganics, and COD. This section discusses ANPQ
process effectiveness with regard to each of these
contaminant groups. The operating conditions and
performance results for each commercial- and
pilot-scale application discussed in Section 3.2 are
summarized in Table 3-2 at the end of the section^
3.2.1 SVOC-Contaminated Industrial \
Wastewater
i
This section discusses industrial wastewater
treatment for SVOCs using the SCWO process at
the pilot-scale level and the O3/H2O2, SCWO, and
acoustic cavitation processes at the bench-scafe
level. •
Pilot-Scale Application '
i
NORAM Engineering and Constructors Limited
(NORAM) used a pilot-scale SCWO reactor to treat
nitrophenolates in real and synthetic industrial
wastewaters. The NORAM pilot plant is located at
the University of British Columbia in Vancouver^,
British Columbia, Canada. The system can operate
at a flow rate of 0.5 to 2 L/min and can
accommodate temperatures up to 600 °C. The
SCWO reactor is tubular; is 120 m in length; has an
inside diameter of 0.62 cm; and has a volume of
3.5 L. Nitrophenolates were present in the
wastewater at 10,000 mg/L. Treatment tests were
conducted on about 200 L of wastewater at a flow
rate of 1 L/min. The retention time in the reactor was
10 to 60 seconds. Treatment resulted in >99 percent
removal of nitrophenolates (NORAM, 2000; Rehmajt
and others, 2000). !
Bench-Scale Studies
This section summarizes information on removal of
the following SVOCs from actual and simulated
industrial wastewaters using the O3/H2O2, SCWOJ
and acoustic cavitation processes at the bench-scale
level.
ANPO Process
• 03/H202
SVOCs 'Removed
5-Chloro-2-nitroaniline;
thiophenol
ANPO Process
• SCWO
• Acoustic
cavitation
SVOCs Removed
Acetic acid; 2,4-DCP;
2,4-dichlorophenoxyacetic
acid methyl ester;
pentachlorophenol (PCP);
pyridine
Porphyrin
The O3/H2O2 process was used to remove
5-chloro-2-nitroaniline and thiophenol from industrial
and synthetic wastewaters. About 200 mL of
solution was treated at a pH range of 1 to 13.
Effective oxidation of thiophenol and elimination of its
odor were achieved when H2O2 (0.25 M) was added
before ozonation. However, complete discoloring of
5-chloro-2-nitroaniline solution occurred when H2O2
was continuously fed to the ozonated wastewater at
a 0.25 molar ratio of H2O2 to O3 (Wiktorowski and
others, 1991).
SCWO
Lin and others (1998b) have studied the degradation
of 2,4-DCP in supercritical water. Experiments were
conducted in a high-pressure batch reactor (20 ml)
and in an isothermal, isobaric flow reactor (15 mL).
Several by-products, including 2,7-dichloro-dibenzo-
dioxin and trichlorophenols, were generated by
SCWO of 2,4-DCP at 400 °C and 25 MPa.
According to the researchers, removal of chlorine
species from 2,4-DCP by sodium or Fe(lll) ions in
supercritical water may suppress formation of dioxin
by-products and enhance contaminant degradation.
Corrosive iron and nickel were present in the effluent
from a Stainless Steel 316 reactor. The corrosion
rate was suppressed by about three orders of
magnitude in a quartz-lined reactor.
The U.S. EPA Office of Research and Development
funded a 5-year project to develop kinetic -models
and reaction pathways for SCWO of selected priority
pollutants, namely acetic acid; 2,4-DCP;
2,4-dichlorophenoxyacetic acid methyl ester; PCP;
and pyridine. As discussed below, the project
consisted of batch and continuous-flow studies,
corrosion studies, and chromium speciation and
separation studies (Gloyna and Li, 1993).
Batch and continuous-flow studies showed that
refractory and chlorinated organic compounds were
effectively destroyed by the SCWO process. For a
given temperature, the highest corrosion rate
3-26
-------
occurred at the lowest pH within the test conditions;
the pH ranged from 2.1 to 8.6. For a given pH,
higher corrosion rates were observed at 300 and
500 °C as compared to 400 °C. Chromium species,
especially hexavalent chromium, precipitated
effectively because of the limited solubility of
chromate salts in supercritical water.
The researchers recommend that selection of reactor
materials be based on the anticipated chloride
concentration and pH of the waste to be treated. For
a waste with low chloride concentrations and a
neutral pH, Stainless Steel 316 may be used as an
SCWO reactor material. If higher chloride
concentrations are present, a nfckel-chrome alloy
such as Hastelloy C-276 is recommended.
Monel 400 was not acceptable as a possible SCWO
reactor material because of selective leaching.
Finally, the chromium speciation and separation
study showed that the pH of the effluent was an
important factor in determining the oxidation state of
chromium corrosion. At a pH below 7, trivalent
chromium was the only chromium corrosion species
generated. At a pH above 7, both trivalent and
hexavalent chromium corrosion species were
generated. The removal of hexavalent chromium
was temperature-dependent. Greater amounts of
chromate salts were removed at highertemperatures
(Gloyna and Li, 1993).
Acoustic Cavitation
Yasuda and others (1999) have examined the effect
of mixing on porphyrin decomposition by acoustic
cavitation at a frequency of 23 kHz. The rotation
speed of the stirrerorthe liquid recirculation flow rate
was varied in the batch reactor. Decomposition was
observed to increase with increased mixing. In
addition, increased sonoluminescence of the
reaction mixture was observed, indicating that a
greater number of cavities were being formed.
3.2.2 Dye-Contaminated Industrial
Wastewater
Dyes have been removed from industrial wastewater
using the Fenton, electrochemical oxidation, and
E-beam processes in pilot-scale applications. The
Fenton, O3/H2O2, SCWO, acoustic cavitation, and
E-beam processes have also been used to treat
dye-contaminated water in bench-scale studies.
This section describes these applications and
studies.
Pilot-Scale Applications
This section discusses removal of the following dyes
from industrial wastewater at the pilot-scale level
using the Fenton, electrochemical oxidation, and
E-beam processes.
ANPO Process
•
*
•
Fenton
Electrochemical
oxidation
E-beam
byes 'Removed,. , ',
• Ming Jade; Navy 106;
unidentified azo-red
dye; unidentified
orange dye
Unidentified dyes
Reactive Red 120
Fenton
A pilot-scale Fenton system was installed at a textile
mill in Martinsville, Virginia, U.S.A., to pretreat a
concentrated jet-dye wastewater stream by removing
color and dissolved organic carbon (DOC). The
three primary dyes present in the wastewater were
a phthalocyanine-based dye called Ming Jade
(110,000 American Dye Manufacturers Institute
[ADMI] c.u.), Navy 106 (48,000 ADMI c.u.), and an
unidentified azo-red dye (60,000 ADMI c.u.). In the
pilot system, the jet-dye wastewater was pumped
into a 210-L feed tank, where the pH was adjusted to
3. The solution was stirred and heated to 65 °C.
H2O2 and Fe(ll) were added to the solution at
concentrations of 6,000 and 300 mg/L, respectively.
The flow rate of the pilot system was 1.1 L/min. After
about 120 minutes, the reduction in color and DOC
for the three dyes was as follows: for Ming Jade, 99
and 19 percent, respectively; for Navy 106, 98 and
3 percent, respectively; and for the azo-red dye, 90
and 19 percent, respectively. The estimated costs of
the reagents associated with removal of Ming Jade,
Navy 106, and the azo-red dye were $2.30/m3,
$2.10/m3, and $2.20/m3 of wastewater treated,
respectively (Price and others, 1994).
Francoisse and Gregor (1996) describe a pilot-scale
application of a modified Fenton process (the FSR
Process®) to remove COD and TOG from orange
dye-contaminated wastewater. The FSR Process®'
differed from the traditional Fenton process in that
the sludge was recycled. At the end of Fenton
treatment, the sludge was separated, concentrated,
dissolved, and electrochemically reduced to form
Fe(ll) salts that could be reused. COD and TOC
were initially present in the colored wastewater at
3,400 and 850 mg/L, respectively. The pH of the
3-27
-------
wastewater was 12. Concentrated (96 percent)
H2SO4 was used to adjust the pH and dissolve the
sludge. After wastewater, treatment in the FSp
Process® pilot plant, COD and TOG concentrations
were reduced by 82 percent, and toxicity wgs
sufficiently reduced to allow1 activated sludge
treatment as the final process. The estimated cost
of a full-scale, 250-L/min plant is $2,900,000/year,
which includes raw material, maintenance, labor, and
electricity costs. i
Electrochemical Oxidation
In a pilot study, the electrochemical oxidation
process was applied to effluent from the dyeing and
finishing process of a textile mill in Thrace, Greece.
The laboratory pilot plant consisted of an electrolytic
cell, recirculation reactor, wastewater feed system,
and cooling system. Titanium/platinum was used for
the anode and Stainless Steel 304 was used for the
cathode. The electrodes were operated at 20 volts
direct current and 50 A. The recirculation reactor
included a 5-L vessel and a peristaltic pump that
recirculated the wastewater at a 10-L/min flow rate.
The cooling system maintained the temperature of
the wastewater at 42 °C. The wastewater contained
5-day biochemical oxygen demand (BOD)
(450 mg/L), COD (1,200 mg/L), color (3,400 ADMi
c.u.), and total Kjeldahl nitrogen (TKN) (34 mg/L).
The temperature and pH were maintained at 42 °C
and 10, respectively. Addition of greater initial
amounts of 1 percent sodium chloride and
36 percent HCI resulted in greater COD and color
removal. The COD concentration was reduced by
93 percent after 40 minutes of treatment at a current
density of 890 mA/cm2 using 2 mL each of 1 percent
sodium chloride and 36 percent HCI. In addition, the
ADMI c.u., the 5-day BOD concentration, and the
TKN concentration were reduced by 96, 92, and
>99 percent, respectively (Vlyssides and Israilides,
1998).
E-Beam
The E-beam process was used at the pilot scale to
treat synthetic wastewater containing Reactive Red
120. Experiments were conducted in the Miami
Electron Beam Research Facility at the Miami-Dade
Central Wastewater Treatment Plant in Florida;
U.S.A. The treatment system had a horizontal;
1.5-MeV electron accelerator capable of delivering 0
to 50 mA of E-beam current. The absorbed radiation
doses were designed to reach as high as 800 Mrads
at a flow rate of 450 L/min. The E-beam was
scanned at 160 by 60 hertz to cover a
1,200-cm-wide by 5-cm-thick waste stream.
Reactive Red 120 was present in the wastewater at
50 mg/L. Influent pH was adjusted to 5. At an
irradiation dose of 450 Mrads, color was removed by
>95 percent. The study also showed that addition of
Fe(ll) did not improve decolorization efficiency and in
some cases decreased the removal efficiency to
some extent. Increasing the irradiation dose
increased color removal (Kurucz and others, 1998).
Bench-Scale Studies
This section summarizes information on removal of
the following dyes from actual and simulated
industrial wastewaters at the bench-scale level using
the Fenton, O3/H2O2, SCWO, acoustic cavitation,
and E-beam processes.
ANPO Process Dyes Removed
Fenton
03/H202
SCWO
Acoustic
cavitation
E-beam
Acid Blue 264; Acid Red
337; Acid Yellow 222;
Active Yellow Lightfast 2
KT; 1-amino-8-naphthol-
3,6-disulfonic acid (DSD);
Basic Blue 3; Basic Red
18:1; Basic Violet 7;
Basilen Red EB; 4,4'-
diaminostilbene-2,2'-
disulfonic acid; Direct
Black 112; Direct Blue 86;
Direct Blue G; Direct Violet
47; Disperse Blue 139;
Disperse Red 60;
Dispersol Black D-2B;
Erioglaucirie; Levafix Red
4BA; Procion Navy HEXL;
Reactive Black B; Reactive
Blue 15; Reactive Blue 71;
Reactive Blue 172;
Reactive Red 141;
Remazol Black B;
unidentified dyes
Acid Black 52; Direct Blue
80
Unidentified dyes
Remazol Black B (or
Reactive Black 5)
Acid Red 265
Fenton
Reactive dye wastewater containing Basilen Red EB
(70 mg/L; 1,800 platinum/cobalt [Pt/Co] c.u.), Levafix
Red 4BA (62 percent; 4,000 Pt/Co c.u.), Procion
3-28
-------
Navy HEXL (70 mg/L; 1,300 Pt/Co c.u.), and
Remazol Black B (60 percent; 1,800 Pt/Co c.u.) was
treated using the Fenton process. Fe(ll) and
30 percent H2O2 were used in all the experiments.
The pH of the solution was adjusted to between 4
and 5 using acetic acid. Color reduction of
>50 percent was observed to occur instantaneously.
Fenton treatment of individual dyes resulted in
reduction of Basilen Red EB Pt/Co c.u. by
93 percent, Levafix Red 4BA Pt/Co c.u. by
95 percent, Procion Navy HEXL Pt/Co c.u. by
92 percent, and Remazol Black B Pt/Co c.u. by
75 percent. The corresponding Fe(ll):H2O2 mass
ratios for each of the treatments were 0.05,0.2,0.05,
and 0.2, respectively. It was observed that darker
colors required higher Fe(ll):H2O2 mass ratios than
lighter colors. The treated solution was found to be
nontoxic. The study showed that an Fe(ll):H2O2
mass ratio of 0.25 was optimal to achieve dye
removal with the least amount of sludge generation.
Under the assumption that settled iron sludge would
be reused in the Fenton process, the material cost of
the treatment was estimated at about $0.23/m3 of
wastewater treated (Lev and Deshpande, 1.996).
Lin and Chen (1997) have evaluated the
effectiveness of the Fenton process in treating
effluent from the secondary wastewater treatment
plant of a dyeing and finishing mill in northern
Taiwan. The treatment process consisted of the
Fenton process, chemical coagulation, and an ion
exchange process. The Fenton reaction took place
over 2 hours. A reaction temperature of 30 to 40 °C
was determined to be optimal. The study showed
that an Fe(ll):H2O2 mass ratio of 0.75 yielded good
results for COD and color removal. A large number
of small floes generated in the Fenton process were
removed by chemical coagulation using 50 mg/L of
powdered activated carbon and 1 mg/L of polymer.
At the end of the three-step treatment process, the
COD concentration (initially 130 mg/L) was reduced
by 93 percent. Similarly, color was >99 percent
removed, and turbidity was reduced by 95 percent.
Five types of simulated dye wastewater were treated
with Fenton's reagent. The wastewaters contained
the following dyes: (1) Reactive Blue 71, Blue 172,
and Red 141; (2) Dispersol Black D-2B, Disperse
Blue 139, and Disperse Red 60; (3) Direct Black 112,
Blue 86, and Violet 47; (4) Basic Blue 3, Red 18:1,
and Violet 7; and (5) Acid Blue 264, Red 337, and
Yellow 222. The dyes, which were prepared in equal
concentrations of 300 mg/L, were mixed in equal
dye:Fe(ll):H2O2 proportions of 1:1:1. The treatment
results showed that the best pH value for
decolorization was below 3.5. The most effective
amount of H2O2 was determined to be 580 mg/L for
reactive dyes; 880 mg/L for acid dyes; and 290 mg/L
for disperse, direct, and basic dyes. Higher FeSO,,
doses resulted in better treatment. Lower
temperatures correlated with longer decolorization
times. Under optimal conditions, the average
percent removal of COD was about 90,
the transparency of the wastewater was above
25 cm, and the average' percent decolorization was
above 97 (Kuo, 1992). .
Jank and others (1998) have studied Fenton
treatment of synthetic wastewater containing
erioglaucine, a widely used blue acid
aminotriphenylmethane dye. The dye was prepared
in deionized water at 2,000 mg/L. Reagents used
included H2SO4, H2O2, iron sulfate, and activated
carbon with a specific surface area of 1,000 square
meters per gram, which was used as a catalyst for
activation of the H2O2. The reaction was optimal at
a pH of 3. Decolorization was achieved within
24 hours at a pH of 3 and at a temperature between
20 and 40 °C.
The Fenton process was used in batch studies to
remove COD in and decolorize simulated desizing
wastewater containing 100 mg/L each of 0.2 percent
polyvinyl alcohol (160 mg/L of COD), Direct Blue G
(71 mg/L of COD), and Reactive Black B (100 mg/L
of COD). The pH was adjusted to between 2 and 5
before the Fenton reaction, and the treatment time
was about 1 hour. The study showed that the
Fenton reaction was optimal at a pH of 3 and a
temperature of 30 °C (Lin and Lo, 1997).
A pretreatment process involving Fenton's reagent
and coagulation has been studied by Zhu and others
(1996). Wastewater containing 1-amino-8-naphtho.l-
3,6-disulfonic acid was treated in batch reactors for
1 hour using H2O2 and FeSO4. COD removal was
greatest under acidic conditions (a pH of 2 to 4). At
a pH above 8, Fe(ll) ions begin to form floe and
precipitate. The optimal Fe(ll) dose was 3.6 mM.
The Fenton process improved the effects of
coagulation. When the concentration of Fe(ll) was
3.60 mM, the concentration of H2O2 was 0.088 M,
and the ferric chloride dose of two-stage coagulation
treatment was 0.092 and 0.03 M, the overall COD
removal was at least 90 percent.
Yu and others (1998) have investigated pretreatment
and biodegradability enhancement of DSD-acid
manufacturing wastewater using the Fenton process.
The wastewater, which was taken from a typical
chemical plant in China, contained-22,000 mg/L of
COD and 2.2 x 105 c.u. of multiple dyes. Under
optimal conditions, that is, with addition of 150 mg/L
of Fe(ll) and 2,000 mg/L of H2O2 followed by
3-29
-------
two-stage coagulation of 5,000 and 2,000 mg/L ferric
chloride — about 90 percent of the COD and
95 percent of the color were removed in 1 hour of
treatment. The Fenton process was observed to
improve wastewater biodegradability because of the
conversion of nonbiodegradable organic compounds
to more biodegradable substances.
The Fenton process was used to treat wastewater
containing an azo dye called Active Yellow Lightfast
2 KT. Initial dye concentrations varied from 20 to
160 mg/L. FeSO4 was present at 14 mg/L, and the
pH was 3. At a dose of 17 mg/L of H2O2, color was
95 to 97 percent removed. An increase in reaction
temperature increased the color removal efficiency
(Solozhenko and others, 1995).
Flaherty and Huang (1992) have evaluated Fenton
treatment of refractory textile wastewater that
primarily contained Reactive Blue 15 dye. The raw
wastewater had a pH of 12; an alkalinity of
21,000 mg/L as CaCO3; a COD concentration of
2,100 mg/L; and a total copper concentration of
14 mg/L. The Fe(ll) concentration was maintained at
2x 10'2 M, and the pH of the influent was adjusted to
3.5. In continuous-flow experiments, the reactioh
took place in a 1-L, continuously stirred tank reactor
for about 2 hours. About 70 percent of the COD was
removed from the wastewater. After a 24-hr settling
period, the total copper concentration in the
supernatant had decreased to <1 mg/L, which
corresponds to 93 percent removal. i
Adams and others (1995) have evaluated O3/H2O2
treatment of two metal-complex azo dyes, Acid
Black 52 and Direct Blue 80, which contain
chromium and copper, respectively. The initial dye
solutions contained 340 mg/L of Acid Black 52 and
260 mg/L of Direct Blue 80. The temperature in the
semi-batch reactor was maintained at 20 °C. Ah
H2O2:O3 molar ratio of 0.5 was used. The study
showed that inorganic copper was readily released
during the decolorization of Direct Blue 80,
potentially affecting the toxicity of treated wastewater
associated with this dye. The decolorization of Acid
Black 52, however, did not result in appreciable
chromium in the treated wastewater.
SCWO I
Four colored spotting dyes from the U.S. military
stockpile were treated in a SCWO tubular reactor at
Sandia National Laboratories in Livermorej
California, U.S.A. The reactor is made of
Inconel 625 and consists of two sections. The first
section is 260 cm long and is equipped with heaters
capable of adding power at a rate of 17 watts per
centimeter. The second section is 490 cm long, is
powered at a rate of 5.4 watts per centimeter, and is
maintained under isothermal conditions. At
temperatures above 550 °C, the dyes and partial
oxidation products were destroyed in <10 seconds
with a removal efficiency of >91 percent based on
TOC. Formation of sulfate salt deposits within the
flow reactor and corrosion of the construction
materials emerged as issues for further research.
Specifically, "sticky" salts that adhered to the walls of
the reactor plugged the system and caused a rise in
pressure. Corrosion in the reactor resulted in
chromium and nickel concentrations in the effluent
(Lajeunesse and Rice, 1995).
Acoustic Cavitation
Vinodgopal and others (1998) discuss the removal of
a reactive textile azo dye, Remazol Black B (or
Reactive Black 5), from an aqueous solution using
acoustic cavitation at a frequency of 640 kHz and an
output power of 240 W. The solution was constantly
bubbled with a stream of O2 gas. The absorbance
measurement in the visible range of light showed
that after 90 minutes, the dye was no longer present
in the sample. The TOC measurements showed that
treatment for 6 hours resulted in about 60 percent
mineralization of the dye. By-products of the
reaction included oxalate, nitrate, and sulfate ions.
E-Beam
Water containing the azo dye Acid Red 265 was
treated by E-beam oxidation. Experiments were
performed at ambient temperature and at initial dye
concentrations ranging from 50 to 400 mg/L. The
flow rate was varied between 1.5 and 10 L/min,
mean E-beam dose rates ranged from 0.4 to 160
Mrads per second, and beam currents ranged from
1 to 3 mA. Under oxygenated conditions, 95 percent
decolorization was achieved using 0.2-Mrad
irradiation and an initial dye concentration of
100 mg/L (Kawakami and others, 1978).
3.2.3 Inorganic-Contaminated
Industrial Wastewater
This section discusses treatment of inorganic-
contaminated industrial wastewater using the SCWO
process at the commercial scale. Information is also
included on inorganic-contaminated industrial waste-
water treatment using the electrochemical oxidation
and acoustic cavitation processes at the bench-scale
level.
£-30
-------
Commercial-Scale Application
Electrochemical Oxidation
In 1994, the Chematur Aqua Critox® SCWO
treatment system was installed at the Huntsman
Corporation facility in Austin, Texas, U.S.A. The
SCWO system contains a positive-displacement
pump that pressurizes the wastewater feed stream
to between 25 and 28 MPa, a double-pipe heat
exchanger that channels the feed stream into a
process heater, a heat exchange shell, a heat
recovery boiler, and an effluent cooler. The feed
stream temperature at the exit point of the process
heater ranges from 360 to 380 °C. When O2 is
added to the heated feed stream, temperatures
within the reactor rise to 530 to 650 °C. After the
solution passes through the effluent cooler, a control
valve lowers the pressure to atmospheric pressure,
and a liquid-gas separator separates the effluent into
liquid and gas phases. The reactor is made of a
high-grade alloy to prevent stress corrosion resulting
from chlorides in the feed stream.
A 48-hour demonstration run of the Aqua Critox®
system was conducted in May 1994. The 19-L/min
system treated facility process and washdown water
containing 6,900 mg/L of ammonia and 50,000 mg/L
of TOC among other constituents. The system
achieved >99 percent removal of ammonia during
the demonstration. A study of air emissions revealed
no problems with nitrogen oxide (NOX) or sulfur oxide
(SOX) emissions. The total regulated exhaust
emissions from the process were <260 kg/year,
about 7.2 percent of the quantity allowed under
Texas' air permit exemption requirements. No costs
have been reported for the demonstration run
(Griffith, 1995).
Bench-Scale Studies
This section summarizes the results of bench-scale
studies of the effectiveness of the electrochemical
oxidation and acoustic cavitation processes- in
removing the following inorganics from industrial
wastewater.
AN PO Process
Electrochemical
oxidation
Acoustic
cavitation
Inorganics Removed
Ammonium; cyanide
Hydrogen sulfide
Using electrochemical oxidation, Chiang and others
(1995a) achieved almost .complete removal of
ammonium and 90 percent removal of COD in
2 hours while treating coke plant wastewater from a
steel manufacturing company. Electrolysis took
place on a lead dioxide-coated titanium anode.
Ammonium and COD. were present in the
wastewater at initial concentrations of 760 and
2,100 mg/L, respectively. The pH of the wastewater
ranged from 6.9 to 7.5, and the temperature ranged
from 50 to 65 °C. The effect of various anode
materials was tested using graphite, binary
rubidium-TiO2-coated titanium, tertiary
tin-lead-rubidium oxide-coated titanium, and lead
oxide/titanium. Graphite, rubidiuni-TiO2-coated
titanium, and tin-lead-rubidium oxide-coated titanium
were not found to be suitable anode materials for
electrolysis of coke plant wastewater because of
their poor stability and performance. Lead
oxide/titanium provided the best removals as an
anode material. The current efficiency improved with
increasing current density and chloride dose. In
addition, ammonium and COD removals improved at
pHs above 7.
A bipolar trickle tower electrochemical reactor
consisting of graphite Raschig rings was evaluated
in terms of its removal of cyanide from wastewater
effluent. During continuous reactor operation,
cyanide concentrations were reduced from 1,500 to
<60 mg/L (>96 percent removal) and from 1,000 to
<30 mg/L (>97 percent removal) with an energy
consumption of 18 to 27 kilowatt-hours per kilogram
of cyanide removed. At an initial cyanide
concentration of 300 mg/L, complete removal of
cyanide was achieved with an energy consumption
of 78 kilowatt-hours per kilogram of cyanide
removed. The rate of removal decreased as the
cyanide concentration decreased with time.
Although the initial pH of the raw water was 11, the
pH of the treated water was between 8 and 9,
indicating that pH may not have to be adjusted
before additional treatment by conventional
processes (Ojjutveren and others, 1999).
Acoustic Cavitation
Kotronarou and others (1992b) have studied
oxidation of hydrogen sulfide by acoustic cavitation
at a frequency of 20 kHz. The ultrasonic intensity
was about 75 W/cm2, which corresponds to a power
input of about 85 W. The temperature inside the
batch reactor used was maintained at 25 °C. At a
pH of 10, treatment by-products included sulfate,
sulfite, and thiosulfate.
3-31
-------
3.2.4 High-COD Industrial Wastewater
\
This section discusses COD removal from industrial
wastewater using the Fenton and O3/H2O2 processes
at the bench scale. Removal of TOG is also
discussed.
i
Fenton
Bozarslan and others (1997) have applied the
Fenton process to wastewaters from cigarette
factories in Izmir, Turkey. Jar test studies showed
that the optimum doses of FeSO4 and H2O2 were 600
and 80 mg/L, respectively. Under these conditions,
COD removals reached 96 percent after Fenton and
biological treatment.
Horng and others (1998) have studied Fenton
treatment of wastewater from a brewery plant.
Treatment of wastewater containing 188 mg/L of
COD using 500 mg/L of Fe(ll) and 250 mg/L of H2O2
resulted in 45 percent removal of COD.
The Fenton process was applied to wastewater from
a petroleum production facility in order to reduce
TOC. The initial TOC concentration range was 25 to
50 mg/L. Experiments were conducted using 0 to
500 mg/L of H2O2, 0 to 400 mg/L of Fe(ll), and a pH
range of 5 to 6. Optimum TOC reduction was
achieved with 125 mg/L of H2O2 and 105 mg/L of
Fe(ll) at a pH of 5 to 6. No specific percent removals
have been reported. Reduction of the bicarbonate
concentration in the wastewater was necessary
during pretreatment. The chemical oxidation cost
for the Fenton process used is estimated at
$0.04/barrel of water treated; the volume of a barrel
of water is not specified. Sludge generation
associated with the Fenton process is assumed to
have increased the cost.
0/«202
The effect of the O3/H2O2 on the chemical
degradation and biodegradability of debittering table
olive industry wastewaters was studied. Samples of
two kinds of wastewater were used: (1) synthetic
wastewater prepared by treating green olives in
10 percent sodium hydroxide and (2.) real
wastewater collected from the olive industry. COD
concentrations in both kinds of wastewater ranged
from 19 to 25 mg/L. The temperature was
maintained at about 20 °C during the O3/H2O2
process. Use of O3 and H2O2 at weights of about 3.5
and 2.4 gram, respectively, resulted in COD
reductions of up to 90 percent (Beltran and others,
1999).
Beltran and others (1997) have investigated O3/H2O2
treatment of distillery and tomato processing
wastewaters. The semi-batch experiments involved
use of a peristaltic pump that recirculated the
solutions at a rate of 30 liters per hour (L/hr). The
study showed that the O3/H2O2 process reduced
COD in tomato processing wastewater by 86 percent
at a pH of 6. The process, however, was not as
effective in treating distillery wastewater, as the
results obtained were similar to-those associated
with ozonation alone.
•3-32
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3.3 Municipal Wastewater Treatment
Gamma-Ray
The effectiveness of ANPO processes in treating
municipal wastewater has been evaluated to a
limited extent. This section describes pilot-scale
studies involving use of the gamma-ray and E-beam
processes to treat VOC- and microbe-contaminated
municipal wastewater. The operating conditions and
performance results for each pilot-scale application
discussed in Section 3.3 are summarized in
Table 3-3 at the end of the section.
3.3.1 VOC-Contaminated Municipal
Wastewater
Cooper and others (1990) describe the use of a
pilot-scale E-beam system to treat
VOC-contaminated municipal wastewater. The
E-beam system, which is located in the Electron
Beam Research Facility at the Virginia Key
Wastewater Treatment Plant, in Miami, Florida,
U.S.A., has the capacity to treat 450 L/min of
wastewater: The voltage required for the electron
accelerator is 1.5 MeV. The beam current can be
varied from 0 to 50 mA, providing doses of 0 to
850 krads. Influent streams of drinking water,
secondary wastewater effluent, and anaerobically
digested sewage sludge were used in the
application. The secondary wastewater was the
effluent of an extended aeration process. The
effluent was chlorinated up to 1 minute prior to
E-beam treatment. Chloroform, TCE, and toluene
were removed from the wastewater with efficiencies
of about 85, 97, and >95 percent, respectively. The
total cost to construct the E-beam system was
estimated to be $2.8 million. No breakdown of the
cost is available.
3.3.2 Microbe-Contaminated Municipal
Wastewater
This section describes pilot-scale evaluations of
removal of the following microbes using the
gamma-ray and E-beam processes.
ANPO Prpeeiss
• Gamma-ray
E-beam
Microbes Removed
Coliform bacteria
• Coliform; coliphage; total
bacteria
In a pilot-scale application, Rawat and others (1998)
have examined gamma-ray treatment of water
samples containing coliform bacteria that were
obtained from the Gajarawadi municipal sewage
treatment plant in Boroda, Gujarat, India. The
evaluation was conducted using the Sludge
Hygenization Research Irradiator, which was
commissioned by India's Department of Atomic
Energy and installed adjacent to the treatment plant.
The effluents used in experiments included (1) raw
sewage from the grit chamber of the treatment plant,
(2) effluent from the plant's primary settling tank, and
(3) effluent from the plant's final settling tank. A 60Co
source was used to treat samples in batches of 3 m3.
The total coliform level in the raw sewage was about
4 x 105 colony-forming units per milliliter (cfu/mL).
An irradiation dose of 200 Mrads reduced the
coliform population by >99 percent. The effluent
from the primary settling tank had about
3 x 104 cfu/mL of coliform. A dose of 40 Mrads
reduced the coliform count by >99 percent. The
coliform level in the effluent from the final settling
tank was about 2 x 103 cfu/mL An exponential
decrease in the coliform population was observed
when the dose was increased from 20 to 80 Mrads.
In general, aeration of each of the three effluents
prior to irradiation did not affect removal efficiencies.
Gamma irradiation resulted in an approximately
20 percent reduction in the BOD levels of the primary
and final settling tank effluents.
E-Beam
Researchers at the Electron Beam Research Facility
at the Virginia Key Wastewater Treatment Plant in
Miami, Florida, U.S.A., have studied removal of
coliphage, coliform, and total bacteria from raw
sewage and secondary effluent. The E-beam
system had a 1.5-MeV, 50-mA electron accelerator
capable of treating wastewater at a flow rate of 8 Us.
About 99.9 percent reduction in coliphage, coliform,
and total bacteria was observed at an absorbed dose
of 500 krads. Coliphage appeared to be more
resistant to irradiation than the coliform and total
bacteria. Better inactivation was observed in the raw
sewage samples than in the secondary effluent
samples. The researchers explain that the better
inactivation in raw sewage may be due to its greater
density of microorganisms and thus its greater target
density (Farooq and others, 1993).
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3.4 Contaminated Drinking Water
Treatment
The effectiveness of O3/H2O2 process in treating
contaminated drinking water has been evaluated for
various contaminant groups, including VOCs and
humic substances at the pilot-scale level, and
SVOCs, humic substances, and microbes at the
bench-scale level. This section discusses the
effectiveness of O3/H2O2 with regard to each of these
contaminant groups. The operating conditions and
performance results for each pilot-scale application
discussed in Section 3.4 are summarized in
Table 3-4 at the end of the section.
3.4.1 VOC-Contaminated Drinking
Water
Myers (1990) discusses a pilot-scale application of
O3/H2O2 treatment of trihalomethane-contaminated
water from two small water utilities managing 5,300
to 10,000 L/min in Macon (Plant 1) and Moberly
(Plant 2), Missouri, U.S.A. The mobile treatment
plant was housed in a 2.4- by 10.4-m trailer. Water
from the treatment plant's raw water supply was
piped to the plant's O3 contact chamber, where it
flowed by gravity. All chemicals except O3 were fed
to the treatment system in liquid form. O3 was
generated in the trailer and added through a porous
diffuser stone to the bottom of a cylindrical contact
column. Raw water was fed to the top of the column
and flowed countercurrent to the O3 for efficient
mixing and contact.
The O3 dose was 2 mg/L, and the transfer efficiency
was estimated at 70 percent. The H2O2 dose was
based on a mass ratio of H2O2:O3 of 0.5, so 1 mg/L
of H2O2 was used. The H2O2 solution was injected
into the raw water line upstream of the O3 contactor.
The retention time and temperature in the reactor
were 3 days and 25 °C, respectively. The raw water
pH values for Plants 1 and 2 were 8.0 and 7.5,
respectively.
The study showed a 30 percent trihalomethane
reduction at Plant 1 as a result of using O3 and H2O2
rather than O3 alone. However, the trihalomethane
concentration increased with use of O3 and H2O2 at
Plant 2; no explanation for this result has been
reported. The treatment cost was estimated at
$0.06/m3, including amortized capital (10 years at
8 percent) and operating costs. Capital cost
estimates are for a water plant designated for a
maximum flow rate of 11,000 L/min. O&M cost
estimates are based on an average-day demand of
3,900 L/min and are for the costs of additional
chemicals, energy, and equipment O&M only.
3.4.2 SVOC-Contammated Drinking
Water
No commercial- or pilot-scale information is available
on the effectiveness of ANPO processes in treating
drinking water contaminated with SVOCs. A
bench-scale study involving use of the O3/H2O2
process to treat SVOC-contaminated drinking water
is summarized below.
An O3/H2O2 system designed by Duguet and others
(1989) was used to treat water samples collected
from a drinking water well. The influent pH was 6.7,
and the alkalinity was 225 mg/L as CaCO3. The
primary pollutant in the water was
o-chloronitrobenzene, which was present at an
average concentration of 1,500 ug/L. Threshold
odor numbers were above 20. The system operated
at a maximum flow rate of 7.5 L/min. The system
consisted of two O3 contactors, an O3 generator, a
dosage system for H2O2, and sand and granular
activated carbon filters running in parallel to reduce
biodegradable oxidation by-products. Optimal
o-chloronitrobenzene removal was obtained when
the H2O2:O3 mass ratio was about 0.4. The O3 and
H2O2 doses used were 8 and 3 mg/L, respectively.
The retention time was 20 minutes. Treatment
resulted in 99 percent removal of the
o-chloronitrobenzene. The study showed that
increasing the retention time to 30 minutes did not
improve removal efficiency. Bacterial regrowth in the
distribution network was hindered by sand or
granular activated carbon filtration. Taste and odor
in granular activated carbon-filtered water diminished
after 3 months, with the threshold odor numbers
falling to 3.
3.4.3 Humic Substance-Contaminated
Drinking Water
No evaluations of commercial-scale ANPO
processes for removing humic substances from
drinking water have been reported. However, one
ANPO process (O3/H2O2) has been evaluated at the
pilot and bench scale in terms of removal of humic
substances. The results of these evaluations are
summarized below.
An O3/H2O2 system was pilot-tested by the
Metropolitan Water District of Southern California at
its facility in La Verne, California, U.S.A. The system
was used to treat raw water contaminated with
odor-producing compounds, including geosmin (an
earthy-smelling organic compound) and 2-methyl-
isoborneol. The two sources of raw water were
Colorado River Water and State Project Water. The
average concentration of geosmin in the untreated
3-37
-------
raw water was 0.1 |jg/L. The O3/H2O2 system
achieved 98 percent removal of the geosmin.
Acetaldehyde (2 to 5 ug/L) and formaldehyde (9 to
18 ug/L) were treatment by-products. It was
observed that greater bactericidal action occurred
when the ratio of H2O2 to O3 was decreased.
Effective geosmin removals were achieved
regardless of the contact time. The capital cost
estimated for implementing the O3/H2O2 process pt
five existing water treatment facilities was $200
million (McGuire and others, 1990). Additional
performance data regarding geosmin and
2-methylisoborneol removal using the O3/H2O2
process is presented in a document recently
prepared by the Metropolitan Water District bf
Southern California (2000).
Bench-Scale Study
Duguetand others (1989) discuss O3/H2O2 treatment
of synthetic wastewater contaminated with geosmjn
and 2-methylisoborneo! in a semi-batch reactor.
Geosmin and 2-methylisoborneol were present in the
wastewater at concentrations of 0.3 to 0.5 ug/L. The
O3 dose was 0.2 to 0.3 mg/L, the H2O2 dose was
about 0.1 mg/L, and the influent pH was 7.5. The
geosmin concentration fell below 0.01 ug/L in
<5 minutes of treatment; similar results were
observed for 2-methylisoborneol removal.
Increasing the H2O2 concentration had no effect on
removal rates. Tests were conducted at various
bicarbonate ion concentrations (0 to 400 mg/L) to
evaluate the influence of -OH scavenging by
bicarbonate ions on geosmin and 2-methylisobornepl
oxidation. Under the study conditions, bicarbonates
were not found to have a significant effect on
removal efficiency.
i
3.4.4 Microbe-Contaminated Drinking
Water
i
No evaluations of commercial- or pilot-scale ANPO
processes for removing microbes from drinking water
have been reported. However, one ANPO process
(O3/H2O2) for microbe removal from drinking water
has been evaluated at the bench scale. The results
of the evaluations are summarized below. ]
Miettinen and others (1998) have studied O3/H2O2
treatment of drinking water collected from the Kuopip
City Water Works in Finland. Raw water from Lake
Kallavesi was pretreated by bank filtration and was
then chemically purified using aluminum sulfate
coagulation, clarification, and sand filtration before it
was used in the bench-scale experiments.
The concentrations of nonpurgeable organic carbon
and assimilated organic carbon in the test water
were about 3.3 mg/L and 110 ug/L, respectively.
The pH of the test water was 6.5. The treatment
plant had an O3 generator, an upflow bubble
polyvinyl chloride column (O3 contactor), and a
nanofilter. The volume of the reactor was 100 L, the
flow rate was maintained at 3.3 L/min, the O3 dose
was 2.7 mg/L, and the retention time was
30 minutes. H2O2 was injected into the system
upstream of the O3 contactor and immediately prior
to the in-line static mixer. The H2O2 concentration
was adjusted to maintain an O3:H2O2 mass ratio of
0.4.
Microbial growth patterns in water samples were
monitored for 10 days. These patterns were
measured based on heterotrophic plate counts,
acridine orange direct counts, bacterial biomass, and
bacterial production. Bacterial colonies were
enumerated after 3 days of incubation at 22 °C. The
heterotrophic plate count was 60 cfu/mL, and the
acridine orange direct count was 142,000 cells/mL.
Immediately after O3/H2O2 treatment of the water, the
results were as follows: the nonpurgeable organic
carbon was 2.6 mg/L, the assimilated organic carbon
was 840 ug/L, the heterotrophic plate count was
2 cfu/mL, and the acridine orange direct count was
67,000 cells/mL. The heterotrophic plate count was
decreased by 97 to >99.9 percent with the
combination of O3 and H2O2. Hbwever, after the
acute toxicity produced by the oxidants disappeared,
regrowth of temporarily inactivated heterotrophic
microbes began quickly because O3 has no residual
bactericidal effect. Degradation of organic matter
increased the concentration of assimilable organic
carbon, thereby promoting microbial regrowth after
O3/H2O2 treatment (Miettinen and others, 1998).
In a related bench-scale study, Tuhkanen and others
(1994) applied the O3/H2O2 process to the water
described above. Optimum mutagenicity removal for
a given O3 dose was achieved using an H2O2:O3
mass ratio of 0.7. Ozonation combined with H2O2
treatment was significantly more effective than use
of O3 alone in removing mutagenicity and advanced
oxidation precursors. However, ozonation was more
effective in removing color and turbidity than
ozonation combined with H2O2 treatment.
i3-38
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3.5 Landfill Leachate Treatment
Fenton
In a bench-scale study, COD in landfill leachate was
removed using the electrochemical oxidation
process. Landfill leachate from a sanitary landfill site
in southern Taiwan had CO'D levels between
4,100 and 5,000 mg/L and a 5-day BOD
concentration of <1,000 mg/L. The 5-day BODiCOD
ratio of about 0.2 indicated that the organic
compounds in the landfill leachate were minimally
biodegradable. Ammonium and chloride were also
present in the leachate at high
concentrations—specifically, 2,100 to 3,000 mg/L
and about 2,500 mg/L, respectively. Several types
of anodes, including a graphite anode, a lead
oxide/titanium anode, a binary rubidium-TiO2-coatefJ
titanium anode, and an tin-lead-rubidium
oxide-coated titanium anode, were studied. At a
current density of 15,000 mA/cm2 and an additional
chloride concentration of 7,500 mg/L, leachate
treatmentfor240 minutes using the tin-lead-rubidium
oxide-coated titanium anode resulted in 92 percent
COD removal. A competition between COD removal
and ammonium removal was observed. It was also
found that treatment efficiency was proportionally
related to hypochlorite ion production efficiency. Qf
the four anode types investigated, the
tin-lead-rubidium oxide-coated titanium anode
exhibited the highest electrocatalytic activity and
achieved the best efficiencies for hypochlorite iori
production and leachate treatment (Chiang and
others, 1995b). ;
3.6 Contaminated Surface Water
Treatment !
The effectiveness of ANPO processes in treating
contaminated surface water has been evaluated for
pesticfdes and herbicides using the Fenton and
O3/H2O2 processes. This section discusses
bench-scale evaluations of these processes with
regard to removal of the following pesticides and
herbicides from surface water.
ANPO Process
. Fenton
. O3/H2O2
Pesticides and
Herbicides Removed
. Atrazine
. Atrazine; benazolin;
imazapyr; triclopyr
Atrazine oxidation in natural waters was studied in
jar tests using the Fenton process. Water samples
collected from the Seine River in Paris, France, had
an alkalinity of 200 mg/L as CaCO3. Atrazine was
present in the samples at an initial concentration of
3.5 ug/L. With an applied H2O2 dose of 5 mg/L and
an Fe(ll) dose of 10 mg/L, the atrazine removal
efficiencies were 52 and 29 percent at pHs of 5 and
5.5, respectively (Prados and others, 1995).
0/H202
Prados and others (1995) have also studied
degradation of atrazine in Seine River water by the
O3/H2O2 process. To achieve at least 90 percent
removal of atrazine, doses of 4.5 mg/L of O3 and
1.8 mg/L of H2O2 were used. The retention time in
the batch reactor used was 10 minutes.
In another study, O3/H2O2 degradation of atrazine,
benazolin, imazapyr, and triclopyr in raw lowland
surface water was compared to their oxidation by
ozonation. The initial concentration of each chemical
was 2 ug/L. Experiments were conducted at a pH of
7.5. With applied O3 and H2O2 doses of 3 mg/L, the
atrazine, benazolin, imazapyr, and triclopyr removals
achieved by the O3/H2O2 process were greater than
those achieved by ozonation alone by 20, 11, 18,
and 33 percent, respectively (Lambert and others,
1996).
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3-40
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3-41
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Clancy, Peter B.. Jennifer Armstrong, Michelle
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Section 4
Contaminated Air Treatment
ANPO processes have been demonstrated to be
effective for treatment of contaminated air at the
pilot- and bench-scale levels. No evaluations of
commercial-scale ANPO processes for treatment of
contaminated air have been reported. Matrices to
which ANPO has been applied include (1) air stripper
off-gas, (2) industrial emissions, and (3) automobile
emissions. Collectively, ANPO has been applied to
the following types of airborne contaminants: VOCs,
SVOCs, NOX, SOX, and metals.
To assist environmental practitioners in selecting
ANPO systems to treat contaminated air, this'section
includes pilot-scale system evaluation results for the
electrical discharge-based nonthermal plasma,
gamma-ray, and E-beam processes. This section
also presents supplemental information from
bench-scale studies of ANPO processes.
As described in Section 1.2, Section 4 organizes the
performance and cost data for each matrix first by
contaminant group, then by scale of application (pilot
or bench), and finally by ANPO system or process.
In general, pilot-scale applications are discussed in
detail. Such discussions include, as available, a
system description, operating conditions,
performance data, and system costs presented in
2000 U.S. dollars. Bench-scale studies of ANPO
processes are described in less detail and only if
they provide information that supplements pilot-scale
evaluation results. The level of detail provided for
bench-scale studies varies depending on the source
of information used. For example, percent removals
and test conditions are not specified for some
bench-scale studies because such information is
unavailable in the sources.
At the end of each matrix section, a table is provided
that summarizes operating conditions and
performance results for each pilot-scale application
discussed. The references cited in Section 4 are
listed in Section 4.4.
4.1 Air Stripper Off-Gas Treatment
The effectiveness of ANPO processes in treating air
stripper off-gas has been evaluated for VOCs and
SVOCs. This section discusses ANPO process
effectiveness with regard to both of these
contaminant groups. The operating conditions and
performance results for the pilot-scale application
discussed in Section 4.1 are summarized in
Table 4-1 at the end of the section.
4.1.1 VOC-Containing Air Stripper
Off-Gas
This section discusses treatment of VOC-containing
air stripper off-gas using the E-beam process on a
pilot scale. Information is also included on treatment
of VOC-containing air using the electrical
discharge-based nonthermal plasma, gamma-ray,
and E-beam processes at the bench-scale level.
Pilot-Scale Application
A mobile, pilot-scale E-beam system called the
AGATE-M system was field-tested using air stripper
off-gas at a groundwater remediation site in
Dusseldorf, Germany (Prager and others, 1998).
The E-beam system is housed in a 9-m-long
container and is rated for a maximum flow rate of
1,200 m3/hr. The air stripper off-gas passes through
an inlet filter and an optional ultrasonic humidifier
before entering the E-beam reaction chamber. The
E-beam is generated by a low-energy accelerator.
The window separating the accelerator from the
reaction chamber is made of titanium foil. The
accelerator is housed in a compact, lead-shielded
box with labyrinths for incoming and outgoing ducts.
Treated off-gas from the E-beam system is
dry-scrubbed using a compact, high-efficiency filter
filled with lime granulate. O3 formed during E-beam
irradiation is catalytically decomposed by passing the
scrubbed gas through an activated carbon filter.
During the field test, VOC contaminants in the air
stripper off-gas included the following: cis-
1,2-DCE (1.2 parts per million by volume [ppmv]);
1,1-DCA (0.016 ppmv); 1,1-DCE (0.002 ppmv);
PCE (0.58 ppmv); 1,1,1-TCA (0.34 ppmv); TCE
(0.73 ppmv); and trans-1,2-DCE (0.002 ppmv). The
total VOC concentration was about 3 ppmv. The
electron accelerator in the system generated an
accelerating voltage of 190 keV and a beam power
of 11 kW. The absorbed dose was 1.6 Mrads. The
flow rate in the system was 1,000 m3/hr, and the
mean residence time was 0.2 second.
The AGATE-M system achieved about 90 percent
removal of total VOCs. Higher removal efficiencies
were achieved for olefinic VOCs than for saturated
VOCs. Chloroacetyl chlorides and phosgene were
identified as reaction by-products.
The estimated treatment cost including capital,
depreciation,, labor, electricity, activated carbon, and
4-1
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dry-sorption material costs, based on a flow rate of
100 rrrVhr, was $0.28/m3 of air stripper off gas
treated; the cost of generating the off-gas has not
been reported.
Bench-Scale Studies
This section summarizes bench-scale study results
for use of the electrical discharge-based nonthermal
plasma, gamma-ray, and E-beam processes tp
remove the following VOCs from gas stream^.
Unless otherwise noted, the source of the gas
stream used for a given study is unspecified. !
ANPO Process VOCs Removed
Electrical
discharge-
based
nonthermal
plasma
Gamma-ray
E-beam
Acetone; benzene; carbon
tetrachloride; cis-1,2-DCE;
1,1-DCA;1,1-DCE;1,2-
dichlorobenzene; ethyl-
benzene; ethylene;
methanol; o-xylene; PCE;
1,1,1-TCA;TCE; toluene;
1,2,4-trimethylbenzene;
total xylenes
Benzene; chlorobenzene;
DCA; DCE; o-xylene;
PCE; TCE; toluene; VC
Benzene, carbon tetra-
chloride; chlorobenzene,
cis-1,2-DCE; DCA; DCE;
ethylene; Freon 113;
methanol; o-xylene; PCE;
1,1,1-TCA; TCE; toluene;
VC
Electrical Discharge-Based Nonthermal
Plasma
Removal of VOCs using a bench-scale silent
discharge plasma system developed by the
U.S. Department of Energy Los Alamos National
Laboratory has been studied as part of the Strategic
Environmental Research and Development
Program's National Environmental Technology Test
Sites program (Chapman and others, 1997). Forthis
study, 40 pyrex glass chambers were used as the
dielectric material in the reactor. The contaminants
in the influent included acetone (120 ppmv);
cis-1,2-DCE (2.4 ppmv); 1,1-DCA (3.3 ppmv);
1,1-DCE (4.4 ppmv); 1,2-dichlorobenzene
(31 ppmv); ethylbenzene (2.8 ppmv); PCE
(76 ppmv); 1,1,1-TCA (160 ppmv); TCE (83 ppmv);:
1,2,4-trimethylbenzene (10 ppmv); and total xylened
(16 ppmv). The gas stream was dehumidified, and
hydrogen gas was added at a ratio of one part
hydrogen gas to one part VOCs. Removals ranged
from >95 percent (cis-1,2-DCE) to >99 percent (total
xylenes). HCI, hydrofluoric acid, phosgene, O3, and
dioxin were identified as reaction by-products.
Penetrante and others (1997) have compared the
energy efficiencies associated with removing VOCs
(benzene, ethylene, o-xylene, and toluene) using
(1) an electrical discharge-based nonthermal plasma
process and (2) an E-beam process. Specifically,
the energy efficiencies of a pulsed corona reactor
capable of delivering 15 to 35 kilovolts of output in
100-nanosecond pulses at repetition rates ranging
from 100 hertz to 1 kHz were compared to those of
an E-beam reactor with an accelerating voltage of
125 keV. The study was carried out using 100 ppmv
of each VOC in dry air streams (20 percent O2,
80 percent nitrogen) at a temperature of 25 °C.
For the pulsed corona process, the study showed
that energy densities of 1.4 x 10~4,
1.0 x 10-5, 4.5 x 10-5, and 2.7 x 1Q-S kilowatt-hours
per liter (kWhr/L) were required to achieve
63 percent removal of benzene, ethylene, o-xylene,
and toluene, respectively. For the E-beam process,
the study showed that energy densities
of 1.1 x 1Q-5, 1.9 x 1Q-6, 1.4 x 10-6, and
4.2 x 10"6 kWhr/L were required to achieve
63 percent removal of benzene, ethylene, o-xylene,
and toluene, respectively. Thus, the energy
densities required by the E-beam process to achieve
63 percent removal were 92 percent lower for
benzene, 81 percent lower for ethylene, 97 percent
lower for o-xylene, and 96 percent lower for toluene.
In an earlier study, Penetrante and others (1996a)
compared the energy efficiencies associated with
removing carbon tetrachloride and TCE using (1) an
electrical discharge-based nonthermal plasma
process and (2) an E-beam process. The test
conditions were the same as those used later by
Penetrante and others (1997). For the nonthermal
plasma process, the study showed that the energy
densities required to achieve 63 percent removal of
carbon tetrachloride and TCE were 1.5 x 10"4 and
4.4 x 10"6 kWhr/L, respectively. For the E-beam
process, the study showed that the energy densities
required to achieve 63 percent removal of carbon
tetrachloride and TCE were 2.5 x lO"6 and
8.3 x 10~7 kWhr/L, respectively. Thus, the energy
densities required by the E-beam process to achieve
63 percent removal were 98 percent lower for carbon
tetrachloride and 81 percent lower for TCE:
Penetrante and others (1996b) have also compared
the energy efficiencies of removing methanol using
an (1) electrical discharge-based nonthermal plasma
4-2
-------
process and (2) an E-beam process. Specifically,
the energy efficiencies of a pulsed corona reactor
and a dielectric-barrier discharge reactor were
compared to the energy efficiency of an E-beam
reactor. The study was carn'ed out using test
conditions similar to those discussed above for
Penetrante and others (1997, 1996a). The study
showed that energy densities of 1.2 x 1 Q~* and 4.2 x
1CT6 kWhr/L were required to achieve 90 percent
removal of methanol by the electrical
discharge-based nonthermal plasma and E-beam
processes, respectively. Thus, the energy density
required to achieve 90 percent removal of methanol
by the E-beam process was 97 percent lower. No
significant difference was observed in the
performance of the pulsed corona reactor and the
dielectric-barrier discharge reactor.
Amirov and others (1997) have evaluated the
effectiveness of removing toluene using a
ferroelectric bed reactor packed with barium
titanate-based ceramic beads at the bench scale.
Reactor flow rates were varied from 5 to 20 L/min.
The study showed that the removal efficiency for
toluene increased from near zero at an initial
concentration of 200 ppmv to about 8 ppmv per
watt-hour per cubic meter at an initial concentration
of 700 ppmv.
Gamma-Ray
Hakoda and others (1998) have compared removals
of VOCs (benzene, chlorobenzene, DCA, DCE,
o-xylene, PCE, TCE, toluene, and VC) from a
synthetic, gaseous matrix using (1) the gamma-ray
process and (2) the E-beam process. For this study,
a 60Co gamma-ray irradiation system was compared
to an E-beam irradiation system having an electron
accelerator with an accelerating voltage of 3 MeV
and a beam current of 25 mA. For the gamma-ray
test, VOCs in the humid feed stream were present at
the following concentrations: benzene, 95 ppmv;
chlorobenzene, 130 ppmv; DCA, 320 ppmv; DCE,
320 ppmv; o-xylene, 93 ppmv; PCE, 320 ppmv; TCE,
320 ppmv; toluene, 140 ppmv; and VC, 50 ppmv.
For the E-beam test, VOCs in the humid feed stream
were present at the following concentrations:
benzene, 89 ppmv; chlorobenzene, 100 ppmv; DCA,
320 ppmv; DCE, 320 ppmv; o-xylene, 93 ppmv;
PCE, 320 ppmv; TCE, 320 ppmv; toluene,
150 ppmv; and VC, 51 ppmv. The feed streams
were maintained at a flow rate of 2 L/min.
For the gamma-ray process, the study showed that
chloroethenes were more easily removed than the
aromatic VOCs. The highest removals for benzene
(58 percent), chlorobenzene (40 percent), o-xylene
(80 percent), and toluene (72 percent) were
achieved at a dose of about 1.1 Mrads. However,
90 percent removals were achieved at smaller doses
for DCE (0.5 Mrad), PCE (0.5 Mrad), TCE
(0.48 Mrad), and VC (0.32 Mrad). The highest DCA
removal (about 30 percent) was achieved at a dose
of about 1.1 Mrads. For the E-beam process, the
study showed that chloroethenes were also more
easily removed than aromatic VOCs. The highest
removals for benzene (45 percent), chlorobenzene
(34 percent), o-xylene (80 percent), and toluene
(53 percent) were achieved at a dose of 1.3 Mrads.
However, 90 percent removal was achieved at
smaller doses for DCE (0.57 Mrad), PCE
(0.22 Mrad), TCE (0.35 Mrad), and VC (0.40 Mrad).
The highest DCA removal (about 15 percent) was
achieved at a dose of 1.3 Mrad.
ErBeam
Removal of VOCs using Zapit Technologies'
bench-scale E-beam system was studied as part of
the Strategic Environmental Research and
Development Program's National Environmental
Technology Test Sites program (Chapman and
others, 1997). For this study, E-beam doses ranging
from 0.015 to 0.14 kilowatt-hour per kilogram were
evaluated along with H2O2 additions ranging from 0
to 1.2 grams per minute. Contaminants in the inlet
air stream included cis-1,2-DCE; Freon 113; PCE;
1,1,1-TCA; TCE; and VC. The inlet contaminant
concentrations have not been reported. The flow
rate of the system was 1 m3/min. VOC removals
ranged from 48 to >99 percent, depending on the
specific test conditions involved. The highest total
VOC removal (>98 percent) was achieved when the
irradiation dose was 0.036 kilowatt-hour per kilogram
and H2O2 was added at 1.2 grams per minute.
Penetrante and others (1997, 1996a, 1996b) have
compared the energy efficiencies associated with
removing VOCs (benzene, carbon tetrachloride,
ethylene, methanol, o-xylene, and toluene) using
(1) the E-beam process and (2) the electrical
discharge-based nonthermal plasma process.
Hakoda and others (1998) have compared removals
of VOCs (benzene, chlorobenzene, DCA, DCE,
o-xylene, PCE, TCE, toluene, and VC) using (1) the
E-beam process and (2) the gamma-ray process.
The results of these studies are presented above in
the electrical discharge-based nonthermal plasma
process and gamma-ray process sections,
respectively.
4-3
-------
4.1.2 SVOC-Containing Air Stripper
Off-Gas
No evaluations of commercial- or pilot-scale ANPO
processes for treating SVOC-containing air stripper
off-gas have been reported. However, SVOC-
containing gas streams have been treated using an
electrical discharge-based nonthermal plasma
process at the bench-scale level. Removal of total
SVOCs using a bench-scale silent discharge plasma
system developed by the U.S. Department of Energy
Los Alamos National Laboratory has been studied as
part of the Strategic Environmental Research arid
Development Program's National Environmental
Technology Test Sites program (Chapman and
others, 1997). For this study, 40 pyrex glass
chambers were used as the dielectric material in the
reactor. The total SVOC concentration in the gas
stream was 0.47 ppmv. The individual SVOCs
present and their concentrations have not been
reported. Also, the source of the gas stream used
for the study is unspecified. The gas stream was
dehumidified, and hydrogen gas was added at a ratio
of one part hydrogen gas to one part SVOCs.
Removals of >99 percent were achieved. HCI,
hydrofluoric acid, phosgene, O3, and dioxin were
identified as reaction by-products.
4-4
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4.2 Industrial Emission Treatment
The effectiveness of ANPO processes in treating
industrial emissions has been evaluated for NOX and
SOX and for metals. This section discusses ANPO
treatment process effectiveness with regard to these
contaminant groups. The operating conditions and
performance results for each pilot-scale application
discussed in Section 4.2 are summarized In
Table 4-2 at the end of the section.
4.2.1 NOX- and SOx-Containing
Industrial Emissions
This section discusses treatment of NOX- and
SOx-containing industrial emissions using trie
E-beam process on a pilot-scale level. Information
on sulfurdioxide (SO2)-containing airtreatment using
the E-beam process at the bench-scale level is also
included.
Pilot-Scale Applications
A pilot-scale E-beam system called the Process
Development Unit (PDU) was developed by Ebara
International Corporation (a predecessor company to
Ebara Environmental Corporation) under a
cost-sharing program with the U.S. Department of
Energy. The PDU was field-tested using a
slipstream of flue gas from Indianapolis Power and
Light Company's E. W. coal-fired utility plant in
Indianapolis, Indiana, U.S.A. (Frank and Hirano,
1993). The PDU consists of the following
components: a spray cooler, E-beam system,
retention chamber, baghouse, electrostatic
predpitator, by-product storage system, and
ammonia vaporizer system. Flue gas entering the
PDU first passes through the spray cooler, where the
gas is cooled and humidified. Ammonia gas is then
injected into the gas stream by the ammonip
vaporizer system, which converts liquid ammonia to
the gas phase. After ammonia injection, the- gab
stream enters the E-beam system, which consists of
a reaction chamber with two electron accelerators
positioned on opposing sides of the chamber. The
gas stream is then directed to the retention chamber
in order to increase the time for chemical reactions
to take place. By-products generated by the
reactions are removed from the gas stream by a
baghouse or by a combination of a baghouse and
electrostatic precipitator.
During the field test, the primary contaminants in the
flue gas included NOX (220 to 540 ppmv) and SO2
(400 to 2,800 ppmv). The molar ratio of ammonia to
NOX and SO2 was 0 to 1.2. Each electrorj
accelerator in the E-beam system generated an
accelerating voltage of 800 keV. The beam power
rating was 160 kW. In the reaction chamber, the
dose rate ranged from 0 to 3.0 Mrads, the
temperature ranged from 54 to 150 °C, and the flow
rate ranged from 6.2 to 420 standard cubic meters
per minute (scmm).
The PDU achieved about 95 percent SO2 removal
when the reaction chambertemperature ranged from
73 to 77 °C, the dose exceeded 0.9 Mrad, and the
flow rates ranged from 110 to 130 scmm. NOX
removals of about 90 percent were achieved at
higher temperatures (79 to 85 °C), higher doses
(above 1.8 Mrads), and at flow rates ranging from
130 to 150 scmm. A reaction by-product mixture of
ammonium nitrate and ammonium sulfate in powder
form was observed. The by-product concentrations
have not been reported.
Another pilot-scale E-beam system developed by
Ebara Environmental Corporation and the Research
Association for Abatement and Removal of NOX in
the Steel Industry (the association's managing
company is Nippon Steel Corporation) was
field-tested using exhaust gas from a sintering
machine in a steel plant (Kawamura and others,
1980). The E-beam system consisted of a
2.6-m-diameter, vertical, hollow cylinder reaction
chamber with two electron accelerators positioned
on opposing sides of the chamber. The windows
separating the accelerators from the contact
chamber were made of a titanium-palladium alloy.
The windows were cooled by passing cooling air
through the contact chamber. At the chamber inlet,
an impeller was used to rotate the gases in order to
promote homogeneous distribution of the electron
dose in the chamber. The system was designed to
treat exhaust gas at flow rates ranging from 3,000 to
10,000 rrvYhr.
During the field test, the system was continuously
operated for 1 month. Typical components of the
exhaust gas during that period included NOX
(190 ppmv), SOX (200 ppmv), O2 (16 ppmv), CO2
(16 ppmv), H2O (10 ppmv), and dust particles
(40 milligrams per cubic nanometer). The molar ratio
of ammonia to NOX and SOX was 1.0. Each electron
accelerator in the system generated an accelerating
voltage of 600 keV and a beam current of 17 mA,
resulting in a total beam power of 20 kW. In the
reaction chamber, the dose was 1.5 Mrads; the
temperature was 80 °C; and the flow rate was
3,000 m3/hr.
NOX removal of about 80 percent and SOX removal of
>95 percent were simultaneously achieved during
the test period. In addition, the ammonia
concentration was maintained between 10 and
50 ppmv. A reaction by-product mixture of
4-6
-------
ammonium nitrate and ammonium sulfate in powder
form was observed. The by-product concentrations
have not been reported.
A pilot-scale E-beam system was developed by
Research Cotrell under contract to the U.S.
Department of Energy. This system was field-tested
using a slipstream of electrical utility flue gas from
coal-fired boilers at the Tennessee Valley Authority
Shawnee Steam Plant in Paducah, Kentucky, U.S.A.
(Helfritch, 1993). The purpose of the pilot program
was to study the effects of using an alkali-slurry
spray of hydrated lime to neutralize sulfuric and nitric
acids generated when SO2 and NOX were irradiated
with an E-beam. Components of the E-beam system
included a spray dryer, E-beam unit, and fabric filter.
Flue gas first passed through the spray dryer, where
the gas was cooled and humidified by slurry
evaporation and hydrated with lime in the form of fine
particulate that became entrained in the gas. SO2
reacted with calcium hydroxide in the spray dryer to
form calcium sulfite. The gas was subsequently
irradiated by an E-beam unit containing two electron
accelerators, resulting in conversion of SO2 and NOX
to sulfuric and nitric acids, respectively. The gas
then passed through the fabric filter, where the
sulfuric and nitric acids reacted with calcium
hydroxide to produce calcium sulfate and nitrate.
The primary contaminants in the flue gas included
SO2 (400 to 2,500 ppmv) and NOX (300 ppmv). In
addition, the molar ratio of calcium hydroxide to (SO2
+ 1/zNOx) ranged from 0.75 to 1.25. Each electron
accelerator in the system generated an accelerating
voltage of 750 keV and a beam current of 50 mA,
resulting in a total beam power of 75 kW. In the
reaction chamber, the dose ranged from 0 to
1.5 Mrads; the temperature ranged from 10 to 20 °C;
and the flow rate was maintained at 6,800 m3/hr.
Removals of SO2 achieved by the spray dryer
ranged from 50 to 80 percent. SO2 removals in the
E-beam unit ranged from 30 to 90 percent; higher
SO2 removals were observed with higher radiation
doses and lower inlet SO2 concentrations. The
highest SO2 removal (90 percent) was achieved
when the inlet SO2 concentration was 80 ppmv and
the irradiation dose was 1.5 Mrads. SO2 removal in
the fabric filter ranged from 0 to 60 percent. NOX
removal occurred primarily in the E-beam unit. NOX
removals increased with increasing radiation doses
and increasing SO2 concentrations. The highest NOX
removal (80 percent) was achieved when the inlet air
stream SO2 concentration was 1,100 ppmv and the
irradiation dose was 1.5 Mrads. Overall, the gas
temperature, the gas water vapor content, and the
molar ratio of lime to SO2 and NOX had minimal
effect on E-beam removal of SOX and NOX. Reaction
by-product information has not been reported.
Bench-Scale Study
Removal of high-concentration SO2 (up to
15 percent) from a simulated flue gas by the E-beam
process has been studied by Chmielewski and
others (1996). For this study, a 0.2-m-diameter
reactor chamber was used with an accelerating
voltage of 770 keV and a beam current of 1.4 mA.
The inlet SO2 concentration was 30 ppmv.
Additional components of the air stream included
CO2 (6.8 percent), nitrogen (76 percent), O2
(8.7 percent), H2O (7.5 to 8.5 percent), and NOX
(57 ppmv). Prior to irradiation, the following
components were added to the air stream: SO2 (up
to 15 percent), ammonia (up to 30 percent), and H2O
(up to 20 percent). In the reactor, the temperature
ranged from 105 to 118 °C, and the flow rate was
maintained at 5 m3/hr. SO2 removal of 95 percent
was achieved for -the gas stream containing
10 percent SO2, whereas 90 percent removal was
achieved for the gas stream containing 15 percent
SO2. Also, SO2 removals increased with increasing
irradiation doses, ammonia doses, and humidity. No
significant temperature effect was observed.
Ammonium sulfate was identified as a reaction
by-product.
4.2.2 Metal-Containing Industrial
Emissions
No evaluations of commercial- or pilot-scale ANPO
processes for treating metal-containing industrial
emissions have been reported. However,
metal-containing gas streams have been treated
using an electrical discharge-based nonthermal
plasma process at the bench-scale level. Removal
of mercury vapor from a simulated flue gas by the
corona discharge process has been studied by
Helfritch and others (1996) at the bench-scale level.
This study compares the mercury vapor removals
achieved using a corona reactor with pulsed
energization and with direct current energization.
The air stream treated contained 30 micrograms per
cubic meter of mercury vapor and up to 15 percent
water vapor. More than 99 percent removal of
mercury vapor was achieved using pulsed
energization at an applied corona power of about
350 watts per cubic meter per minute, while
90 percent removal was achieved with direct current
energization at the same energy density. Use of
pulsed energization is believed to generate higher
voltages and higher electric fields across the
electrode gap because the short voltage pulses
suppress formation of sparks.
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4.3 Automobile Emission Treatment
No evaluations of commercial- or pilot-scale ANPO
processes for treating automobile emissions have
been reported. However, automobile emissions
have been treated using the electrical
discharge-based nonthermal plasma process at the
bench-scale level. Fanick and Bykowski (1994)
evaluate the effectiveness of simultaneously
removing NOX, particulate matter, and CO from the
exhaust of two light-duty, diesel-engine vehicles
using the electrical discharge-based nonthermal
plasma process. An older, indirect-injection truck
and a newer, direct-injection truck were used. The
plasma was generated using a ferroelectric bed
reactor packed with barium titanate ceramic beads.
Reactor flow rates were varied from 1 to 15 L/min
(space velocities of 1,400 to 20,000 per hour). NOX
removal efficiencies exceeded 40 percent at most
space velocities below 7,000 per hour. Particulate
removal efficiencies exceeded 60 percent in most
cases but decreased with increasing space velocity.
CO was produced by reactions in the reactor; the
amount of CO produced decreased with increasing
space velocity.
Shimizu and Oda (1997) have studied the
effectiveness of removing NOX at high temperatures
(400 to 500 °C) using the electrical discharge-based
nonthermal plasma process combined with catalysts.
The effect of hydrocarbon addition on NOX removal
was also evaluated. The catalysts tested included
copper-doped zeolite; vanadium pentoxide-doped
TiO2; and a conventional, three-way, honeycomb
catalyst (palladium- and rhodium-doped aluminum
coated on cordierite, honeycomb mesh). The initial
composition of the synthetic flue gas used for the
study was as follows: NOX (400 ppmv), O2
(10 percent), and CO2 (10 percent). The gas flow
rate was maintained at 2 L/min. The highest removal
efficiency for NOX (>99 percent) was achieved using
the zeolite catalyst at 300 °C with hydrocarbon
(ethylene 0.5 percent) addition and an input power of
about 10 W. Even without nonthermai plasma, NOX
removal efficiencies of >80 percent were achieved at
about 300 °C with hydrocarbon addition.
4.4 References
Amirov, R.H., E.I. Asinovsky, and I.S. Samoilov.
1997. "Ferroelectric Packed Bed Reactor for
Non-Thermal Plasma Treatment of Effluent
Gas." Proceedings, 14th Symposium on Plasma
Processing. Nara, Japan. January. Pages 208
through 210.
Chapman, T.E., P.H. Mook, Jr., and K.B. Wong.
1997. "A Comparison of Innovative Air Pollution
Control Technology Demonstrations at
McClellan Air Force Base." Proceedings, Air
and Waste Management Association's 90th
Annual Meeting and Exhibition. Toronto,
Ontario, Canada. June 8 through 13. Pages 1
through 15.
Chmielewski, A.G., Z. Zimek, S. Bulka, J. Licki, G.
Piderit, L. Villanueva, and L. Ahumada. 1996.
"Electron Beam Treatment of Flue Gas with High
Content of SO2." Journal of Advanced Oxidation
Technologies. Volume 1, Number 2. Pages 142
through 149.
Fanick, E.R., and B.B. Bykowski. 1994.
"Simultaneous Reduction of Diesel Particulate
and NOX Using a Plasma." Proceedings, Society
of Automotive Engineers (SAE) International
Fuels and Lubricants Meeting and Exposition.
Baltimore, Maryland. October 17 through 20.
Volume 1053. Pages 239 through 246.
Frank, N.W., and S. Hirano. 1993. "The History of
Electron Beam Processing for Environmental
Pollution Control and Work Performed in the
United States." Nonthermal Plasma Techniques
for Pollution Control, Part A: Overview,
Fundamentals, and Supporting Technologies.
NATO AS1 Series, Series G: Ecological
Sciences. Volume 32, Part B. Edited by B.M.
Penetrante and S.E. Schwtheis.
Springer-Verlag, Berlin, Germany. Pages 1
through 26.
Hakoda, T., M. Yang, K. Hirota, and S. Hashimoto.
1998. "Decomposition of Volatile Organic
Compounds in Air by Electron Beam and
Gamma Ray Irradiation." Journal of Advanced
Oxidation Technologies. Volume 3, Number 1.
Pages 79 through 86.
Helfritch, D.J. 1993. "SO2 and NOX Removal from
Flue Gas by Means of Lime Spray Dryer
Followed by Electron Beam Irradiation."
Nonthermal Plasma Techniques for Pollution
Control Part A: Overview, Fundamentals, and
Supporting Technologies. NATO ASI Series,
Series G: Ecological Sciences. Volume 32,
Part B. Edited by B.M. Penetrante and S.E.
Schwtheis. Springer-Verlag, Berlin, Germany.
Pages 33 through 46.
Helfritch, D.J., G. Harmon, and P. Feldman. 1996.
"Mercury Vapor Control by Means of Corona
Discharge." Emerging Solutions to VOC & Air
4-9
-------
Toxics. Proceedings of Specialty Conference
Sponsored by Air & Waste Management
Association. Clearwater Beach, Florida.
February 28 through March 1. Pages 277
through 288. '
Kawamura, K., S. Aoki, H. Kimura, K. Adachi, K.
Kawamura, T. Katayama, K. Kengaku, and Y.
Sawada. 1980. "Pilot Plant Experiment on the
Treatment of Exhaust Gas from a Sintering
Machine by Electron Beam Irradiation."
Environmental Science & Technology.
Volume 14, Number 9. Pages 288 through 294.
Penetrante, B.M., M.C. Hsiao, B.T. Merritt, G E
Vogtlin, A. Kuthi, C.P. Burkhart, and J.R.
Bayless. 1997. "Comparison of Pulsed Corona
and Electron Beam Processing of Hazardous Air
Pollutants." Journal of Advanced Oxidation
Technologies. Volume 2, Number2. Pages 299
through 305.
Penetrante, B.M., M.C. Hsiao, J.N. Bardsley, B."
Merritt, G.E. Vogtlin, and P.H. Wallman. 1996a.
"Electron Beam and Pulsed Corona Processing
of Volatile Organic Compounds in Gas Streams,"
Pure and Applied Chemistry. Volume 68,
Numbers. Pages 1083 through 1087.
Penetrante, B.M., M.C, Hsiao, J.N. Bardsley, B T
Merritt, G.E. Vogtlin, P.H. Wallman, A. Kuthi,
C.P. Burkhart, and J.R. Bayless. 1996b.
"Comparison of Non-Thermal Plasma
Techniques for Abatement of Volatile Organic
Compounds and Nitrogen Oxides." Emerging
Solutions to VOC & Air Toxics. Proceedings of
Specialty Conference Sponsored by Air& Waste
Management Association. Clearwater Beach,
Florida. February ; 28 through March 1
Pages 240 through 251.
Prager, L, R. Mehnert, A. Sobottka, H. Langguth, W
Baumann, H. Matzing, H.-R. Paur, J. Schubert,
R. Rashid, K.M. Taba, H.-P. Schuchmann, and
C. von Sonatag. 1998. "Electron Beam
Degradation of Chlorinated Hydrocarbons
Air-stripped from Polluted Ground Water: a
Laboratory and Field Study." Journal of
Advanced Oxidation Technologies. Volume 3,
Number 1. Pages 87 through 97.
Shimizu, K., and T. Oda. 1997. "DeNOx Process in
Flue Gas Combined with Non-thermal Plasma
and Catalyst." Proceedings, 32nd IEEE Industry
Applications Conference. Part 3 (of 3). New
Orleans, Louisiana. October 5 through 9
Pages 1942 through 1949.
4-10
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Section 5
Contaminated Soil Treatment
ANPO processes have been demonstrated to be
effective in treating contaminated soil. ANPO has
been applied to only one solid matrix, contaminated
soil; no information on ANPO process application to
other solid matrices such as contaminated sediment
or ash is available in the literature. Collectively,
ANPO has been applied to the following types of
soil-bound contaminants: (1) VOCs, (2) SVOCs,
(3) PCBs, (4) pesticides and herbicides, (5) dioxins,
and (6) explosives' and their degradation products.
This section discusses ANPO treatment process
effectiveness with regard to each of these
contaminant groups.
To assist the environmental practitioner in selection
of an ANPO system to treat contaminated soil, this
section includes (1) commercial-scale system
evaluation results for the Fenton process and (2)
pilot-scale evaluation results for the Fenton process.
This section also presents supplemental information
from bench-scale studies of ANPO processes.
As discussed in Section 1.2, Section 5 organizes
performance and cost data for the contaminated soil
matrix first by contaminant group, then by scale of
evaluation (commercial, pilot, or bench), and finally
by ANPO system or process. In general,
commercial- and pilot-scale applications are
discussed in detail. Such discussions include, as
available, a system description, operating conditions,
performance data, and system costs presented in
2000 U.S. dollars. Bench-scale studies of ANPO
processes are described in less detail and only if
they provide information that supplements
commercial- and pilot-scale evaluation results. The
level of detail provided for bench-scale studies varies
depending on the information source. For
example, percent removals and test conditions are
not specified for some of the bench-scale studies
because such information is unavailable in the
sources.
At the end of this section, a table is provided that
summarizes results for each commercial- and
pilot-scale application discussed in the text. The
references cited in Section 5 are listed in
Section 5.7.
5.1 VOC-Contaminated Soil
This section discusses treatment of VOC-
contaminated soil using the Fenton process on a
commercial scale. Information is also included on
VOC-contaminated soil treatment using the Fenton
process at the bench-scale level.
Commercial-Scale Applications
This section summarizes the effectiveness of the
Geo-Cleanse® in situ and ex situ Fenton treatment
systems in removing the following VOCs from
contaminated soil.
ANPO Process
Fenton
VOCs Removed
1 ,2-DCE; methylene
chloride; 2-methyl-
naphthalene; naphthalene;
PCE; petroleum hydro-
carbons (gasoline range
organics [GRO]); TCE; VC;
BTEX
Geo-Cleanse® In Situ Fenton System
A Geo-Cleanse® in situ Fenton treatment system
was used to treat VOC-contaminated sandy soil
beneath a helicopter refueling area at the Wright
Army Airfield in Fort Stewart, Georgia, U.S.A. The
system consisted of six injector wells: four in the
unsaturated zone (about 3.6 to 4.9 m bgs) and two
in the saturated zone (about 6.1 to 7.3 m bgs).
Pretreatment soil samples were collected from two
locations: S-1 and S-2. The primary VOC
contaminants at S-1 included 2-methylnaphthalene
(290 micrograms per kilogram [ug/kg]), naphthalene
(210 ug/kg). petroleum hydrocarbons (GRO)
(58,000 ug/kg), benzene (440 ug/kg), toluene
(210 ug/kg), ethylbenzene (3,000 ug/kg), and
xylenes (10,000 ug/kg). The primary VOC
contaminants at S-2 included 2-methylnaphthalene
(190 ug/kg), petroleum hydrocarbons (GRO)
(270,000 ug/kg), toluene (92 ug/kg), ethylbenzene
(390 ug/kg), and xylenes (100 ug/kg). SVOC
contaminants present in site soil included petroleum
hydrocarbons (diesel range organics [DRO]); system
performance results for these SVOCs are discussed
in Section 5.2. About 9,500 L of 50 percent H2O2;
83 L of 66 percent H2SO4; 8 L of 85 percent H3PO4;
and 54 kg of FeSO4 were injected into the
subsurface over a 4-day period. The amount of
stabilizer (calcium phosphate) used is unknown. The
subsurface pH was 6.0.
5-1
-------
The Geo-Cleanse® system treated about 210 m3 bf
contaminated soil. After a treatment time of 6 days,
the removal efficiencies achieved by the system
indicated that treatment had been locally effective.
At location S-1, the following removal efficiencies
were achieved: 2-methylnaphthalene, >99 percent;
naphthalene, >99 percent; petroleum hydrocarbons
(GRO), 99 percent; benzene, >99 percent; toluene,
>99 percent; ethylbenzene, 83 percent; and xylenes,
86 percent. At location S-2, the following removal
efficiencies were achieved: petroleum hydrocarbons
(GRO), 23 percent; toluene, >99 percent;
ethylbenzene, >99 percent; and xylenes, 32 percent.
However, the 2-methylnaphthalene concentration at
S-2 increased by 560 percent, and naphthalene,
which was not detected at S-2 in the pretreatment
sample analysis, was present at 380 ug/kg in treated
soil samples. The limited data available is
inadequate to explain why soil treatment was more
effective at S-1 than at S-2. The estimated treatment
cost was about $74,000, which includes equipment
rental, reagent, mobilization, and labor costs
(Geo-Cleanse, 2000).
In another application, a Geo-Cleanse® in situ
Fenton treatment, system was used to treat
VOC-contaminated soil (predominantly clay) beneath
former waste lagoons at the Anniston Army Depot in
Anniston, Alabama, U.S.A. The system consisted of
255 injectors positioned horizontally across a 3-acre
area and vertically in four different layers of the
subsurface.
The primary contaminant in the soil was TCE, which
was present at 22,000 milligrams per kilogram
(mg/kg). About 500,000 L of 50 percent H2O2;
1,100 L of 66 percent H2SO4; and 360 kg of FeSO4
were injected into the subsurface over a period of
155 days. H3PO4 was not injected during this
application. The amount of stabilizer (calcium
phosphate) used is unknown. The subsurface pH
was 5.5.
The Geo-Cleanse® system treated about 33,000 m3
of contaminated soil. The system achieved
>99 percent removal of TCE. The total treatment
time is unknown. The estimated treatment cost was
about $2 million, which includes equipment rental,
reagent, mobilization, and labor costs (Geo-Cleanse
2000).
Geo-Cleanse® Ex Situ Fenton System
A Geo-Cleanse® ex situ Fenton treatment systerrj
was used to treat VOC-contaminated soil
(predominantly clay) from the Anniston Army Depot
in Anniston, Alabama, U.S.A. Geo-Cleanse treated
75 55-gallon drums (about 16 m3) of soil collected
from drill cuttings. Based on analytical results for
seven pretreatment soil samples collected from
seven drums, the primary contaminants in the soil
included 1,2-DCE (<5 to 6,300 ug/kg, or an average
of 1,800 ug/kg); methylene chloride (27 to
20,000 Mg/kg, or an average of 3,500 ug/kg); PCE
(<5 to 4,300 ug/kg, or an average of 690 Mg/kg); TCE
(3 to 200,000 Mg/kg, or an average of 43,000 Mg/kg);
and VC (<10 to 40 Mg/kg, or an average of 15 Mg/kg).
About 2,600 L of 50 percent H2O2; 38 L of 66 percent
H2S04; 38 L of H3PO4; and 4.5 kg of FeSO4 were
applied over a 3-day period.. The amount of
stabilizer (calcium phosphate) used is unknown. The
soil pH was 5.5.
The system achieved >99 percent removals for all
the contaminants. The total treatment time is
unknown. These results are based on a comparison
of the average pretreatment sample concentrations
and the contaminant concentrations detected in one
composite soil sample collected from eight random
locations in a borrow pit where the treated soil was
placed. The estimated treatment cost was about
$39,000, which includes equipment rental, reagent,
mobilization, and labor costs (Geo-Cleanse, 2000).
Bench-Scale Studies
This section summarizes bench-scale study results
for use of the Fenton process to remove the
following VOCs from synthetic soil matrices.
ANPO Process
Fenton
VOCs Removed
PCE; TCE
=]
Leung and others (1992) have evaluated the
effectiveness of the Fenton process in removing
PCE at the bench-scale level. For the study,
3.5 grams of sand was spiked with PCE at a
concentration of 1,000 mg/kg. The PCE was
completely mineralized in 3 hours; at that time
17.5 mL of 2.1 M H2O2 and 1.4 mL of 0.005 M FeSO4
were added to the spiked sand, and the initial pH
was adjusted to 3.
Wang and Brussea (1998) have evaluated the ability
of pyrophosphate to maintain iron in solution and
thus enhance removal of PCE by the Fenton
process. For this study, 20 grams of sandy soil was
spiked with PCE at a concentration of 80,000 rng/kg.
With addition of 2 ml of 30 percent H2O2,100 mL of
10 mM FeSO4, and 100 mL of 10 to 30 mM
pyrophosphate, and at a pH ranging from about 5 to
6, a PCE removal of about 18 percent was achieved.
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In the absence of pyrophosphate, minimal PCE
removal (<5 percent) was observed.
Ravikumar and Gurol (1992) have evaluated the
effect of external FeSO4 addition when
TCE-contaminated soil was treated using the Fenton
process in a column study. The soil used .for the
study was coarse-grained sand with a low organic
content (400 mg/kg) and an iron content of
800 mg/kg. The column was about 4 cm in diameter
and 85 cm long. The volume of soil in the column
was about 1 L. With an initial TCE amount of
1.1 millimoles in the soil, 65 percent removal of TCE
was achieved using 2.65 millimoles of H2O2 alone;
however, use of 1.76 millimoles of H2O2 and
0.1 millimole of FeSO4resulted in 77 percent removal
of TCE. The pH of the system has not been
reported. On a molar basis, H2O2 treatment
generated 0.70 mole of chloride per mole of H2O2
and Fenton's reagent generated 1.32 moles of
chloride per mole of H2O2. The greater amount of
chloride liberation resulting from treatment with
Fenton's reagent is a strong indication that Fenton's
reagent is more effective in oxidizing TCE in soil than
H2O2 alone.
5.2 SVOC-Contaminated Soil
This section discusses treatment of
SVOC-contaminated soil using the Fenton process
on a commercial scale. Information is also included
on SVOC-contaminated soil treatment using the
Fenton process at the pilot- and bench-scale levels.
Commercial-Scale Applications
This section summarizes the effectiveness of the
Geo-Cleanse®, ISOTEC™, and H&H SSCO™
Fenton treatment systems in removing the following
SVOCs from contaminated soil.
ANPQ Process
Fenton
SVOGs Removed
• PAHs; PCP; petroleum
hydrocarbons
Geo-Cteanse® In Situ Fenton System
A Geo-Cleanse® in situ Fenton treatment system
was used to treat SVOC-contaminated sandy soil
beneath a helicopter refueling area at the Wright
Army Airfield in Fort Stewart, Georgia, U.S.A. The
system consisted of six injector wells: four in the
unsaturated zone (about 3.6 to 4.9 m bgs) and two
in the saturated zone (about 6.1 to 7.3 m bgs).
Pretreatment soil samples were collected from two
locations: S-1 and S-2. The primary SVOC
contaminants at S-1 and S-2 were petroleum
hydrocarbons (DRO), which were present at 270,000
and 100,000 ug/kg, respectively. VOC contaminants
present in the soil included 2-methylnaphthalene,
naphthalene, petroleum hydrocarbons (GRO), and
BTEX; system performance results for these VOCs
are discussed in Section 5.1. Forthe demonstration,
about 9,500 L of 50 percent H2O2; 83 L of 66 percent
H2SO4; 8 L of 85 percent H3PO4; and 54 kg of FeSO4
were injected into the subsurface over a 4-day
period. The amount of stabilizer (calcium
phosphate) used is unknown. The subsurface pH
was 6.0.
The Geo-Cieanse® system treated about 210 m3 of
contaminated soil. After a 6-day treatment period,
the removal efficiencies achieved indicated localized
treatment effectiveness. At location S-1, the removal
efficiency for petroleum hydrocarbons (DRO) was
about 98 percent. However, at location S-2, the
petroleum hydrocarbon (DRO) concentration
increased by 450 percent. The limited data available
is inadequate to explain why soil treatment was more
effective at S-1 than at S-2. The estimated treatment
cost was about $74,000, which includes equipment
rental, reagent, mobilization, and labor costs
(Geo-Cleanse, 2000).
In another application, a Geo-Cleanse® in situ
Fenton treatment system was used to treat
SVOC-contarpinated soil (predominantly glacial till)
beneath a manufactured gas plant in Burlington,
Wisconsin, U.S.A. The system had nine injectors
positioned in the subsurface, and the primary
contaminant in the soil was PAH.
The initial PAH concentrations in soil at two
locations, MA/-4 and M/V-5, were 26 and 72 mg/kg,
respectively. About 30,000 L of 50 percent H2O2;
570 L of 66 percent H2SO4; 45 L of 85 percent
H3PO4; and 23 kg of FeSO4 were injected into the
subsurface over a 8-day period. The amount of
stabilizer (calcium phosphate) used is unknown. The
subsurface pH was 7.0.
The Geo-Cleanse® system treated about 450 m3 of
contaminated soil. The system achieved total PAH
removals of 98 and 96 percent at M/V-4 and M/V-5,
respectively. The total treatment time is unknown.
The estimated treatment cost was about $130,000,
which includes equipment rental, reagent,
mobilization, and labor costs (Geo-Cleanse, 2000).
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ISOTEC™ Fenton System
An ISOTEC™ in situ Fenton treatment system was
used to treat petroleum hydrocarbon-contaminated
soil beneath a private residence in New Jersey,
U.S.A. The type of fuel that contaminated the soil is
unknown. The petroleum hydrocarbons in the soil
are assumed to have been weathered, and most of
the volatile contamination is assumed to have been
lost. As a result, the petroleum hydrocarbons that
remained are assumed to have been SVOCs. The
initial petroleum hydrocarbon concentration in soil
was 14,000 mg/kg. The geology of the contaminated
area consisted of fractured.bedrock. The subsurface
pH is unknown. i
The system had 10 injection wells positioned
vertically and horizontally within the contaminated
area, which was 9.1 m wide, 9.1 m long, and 2.1 m
deep. Reagent application was conducted for a total
of 12 days over a period of 2 months. A total of
4,300 L of 8.75 percent H2O2 and 8,600 L of
proprietary catalyst (organometallic [iron] complex
solution) were injected into the subsurface. I
i
Within 30 days after the final round of injection
activities had been completed, the petroleurh
hydrocarbon concentration had been reduced by
96 percent. The treatment goal of <1,000 mg/kg was
achieved, and the site was issued a no further action
letter by the New Jersey Department of
Environmental Protection 4 months after treatment
activities began. The estimated treatment cost was
about $73,000, which includes equipment, reagent,
mobilization and demobilization, and labor costs
(ISOTEC, 2000).
H&H SSCO™ Fenton System
An H&H SSCO™ ex situ Fenton treatment system
was used to treat SVOC-contaminated soil
excavated from a soil remediation site in Minnesota,
U.S.A. H&H treated 990 m3 of soil, which was
predominantly sand and silt with a high level of peat
moss. The primary contaminants in the soil included
carcinogenic PAHs (680 mg/kg) and PCP
(160 mg/kg). The soil pH before treatment was 5.5;
Fe(0) was applied to soil windrows during the first
pass by the Microenfractionator™; the amount of
Fe(0) added corresponded to 1 percent of the total
weight of the soil to be treated (10,000 mg/kg). After
a 24-hour stabilization period, a second pass was
made by the Microenfractionator™, during which
50 percent H2O2 was applied to the soil; the amount
of H2O2 added corresponded to 1 percent of the total.
weight of the soil to be treated (10,000 mg/kg). Two
additional H2O2 treatments were performed 3 and
6 weeks after the initial H2O2 treatment.
Within 60 days after the initial H2O2 treatment, the
system had achieved >85 percent removal of
carcinogenic PAHs, and the treatment goal of
<100 mg/kg was achieved. PCP concentrations
were reduced to nondetectable levels, and the
treatment goal of <100 mg/kg was achieved.
According to H&H, the soil pH was unchanged after
treatment. The estimated treatment cost was about
$74/m3 of soil treated, which includes capital,
reagent, mobilization and demobilization, and labor
costs (H&H, 2000).
In another application, an H&H SSCO™ ex situ
Fenton treatment system was used to treat
petroleum hydrocarbon-contaminated soil excavated
from a service station tank farm in Alberta, Canada.
The type of fuel that contaminated the soil is
unknown. The petroleum hydrocarbons in the soil
are assumed to have been weathered, and most of
the volatile contamination is assumed to have been
lost. As a result, the petroleum hydrocarbons that
remained are assumed to have been SVOCs. H&H
treated 1,500 m3 of soil, which was predominantly
coarse-grained sand. Petroleum hydrocarbons in
the soil (at 8,000 mg/kg) included hydrocarbons in
the gasoline and diesel ranges; however, on
average, most of the hydrocarbons were in the diesel
range. The soil pH before treatment was 6.2.
Fe(0) was applied to soil windrows during the first
pass by the Microenfractionator™; the amount of
Fe(0) added corresponded to 1 percent of the total
weight of the soil to be treated (10,000 mg/kg). A
second pass was immediately made by the
Microenfractionator™, during which 50 percent H2O2
was applied to the soil; the amount of H2O2 added
corresponded to 0.5 percent of the total weight of the
soil to be treated (5,000 mg/kg).
Within 36 hours after the H2O2 treatment, the system
had achieved >88 percent removal of petroleum
hydrocarbons, and the treatment goal of
<1,000 mg/kg was achieved. According to H&H, the
soil pH was unchanged after treatment. The
estimated treatment cost was about $43/m3 of soil
treated, which includes capital, reagent, mobilization
and demobilization, and labor costs (H&H, 2000).
An H&H SSCO™ ex situ Fenton treatment system
was also used to treat petroleum
hydrocarbon-contaminated soil excavated from an oil
field in Alberta, Canada. H&H treated 400 m3 of soil,
which was predominantly clay. Petroleum
hydrocarbons in the soil (at 8,500 mg/kg) included
5-4
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hydrocarbons in the diesel to crude range. The soil
pH before treatment was 6.2.
Fe(0) was applied to soil windrows during the first
pass by the Microenfractionator™; the amount of
Fe(0) added corresponded to-1 percent of the total
weight of the soil to be treated (10,000 mg/kg). After
a 24-hour stabilization period, a second pass was
made by the Microenfractionator™, during which
50 percent H2O2 was applied to the soil; the amount
of H2O2 added corresponded to 0.5 percent of the
total weight of the soil to be treated (5,000 mg/kg).
Two additional H2O2 treatments were performed 4-
and 8 days after the initial H2O2 treatment.
Within 12 days after the initial H2O2 treatment, the
system had achieved >88 percent removal of
petroleum hydrocarbons, and the treatment goal of
<1,000 mg/kg was achieved. According to H&H, the
soil pH was unchanged after treatment. The
estimated treatment cost was about $73/m3 of soil
treated, which includes capital, reagent, mobilization
and demobilization, and labor costs (H&H, 2000).
Pilot-Scale Applications
This section presents pilot-scale evaluation results
for removal of the following SVOCs from soil using
the Fenton process.
ANPO Process
Fenton
SVOCs Removed
PAHs; petroleum
hydrocarbons
A pilot-scale bioslurry system was
field-demonstrated by International Technology
Corporation under the U.S. EPA's Superfund
Innovative Technology Evaluation Program in 1994
(U.S. EPA, 1997; International Technology
Corporation, 1995). The system was used" to
remove PAHs and carcinogenic PAHs from soil and
combined biological treatment with the Fenton
process. The system consisted of two 60-L, TEKNO
Associates, aerobic bioslurry reactors and a 10-L
reactor operated in semicontinuous, plug flow mode.
The first 60-L reactor (R1) was designed to remove
easily biodegradable carbon and to increase
biological activity against the more resistant PAHs
(those with three or more rings). Slurry from the first
reactor was fed to the 10-L reactor (R2), where
Fenton's reagent was added to accelerate oxidation
of four-to six-ring PAHs. The second 60-L reactor
(R3) was used as a polishing reactor to remove any
partially oxidcad contaminants remaining in the
system. Slurry was removed from R3 and clarified
using gravity settling techniques. The soil used for
the demonstration was collected from a wood
treating facility and contained sand (30 percent) and
clay (70 percent).
Under optimal operating conditions, the total PAH
and carcinogenic PAH concentrations in the soil
slurry were about 320 and 65 mg/kg, respectively.
Forthe demonstration, PAH-contaminated slurry was
wet-sieved through a 30-mesh screen and blended
so as to contain 40 percent solids (30 percent
contaminated soil and 10 percent uncontaminated
clay). The flow rates in the reactors were 6 L/day
(R1 and R2) and 8 L/day (R3). The hydraulic
retention-times in the reactors were 20 days (R1),
2 days (R2), and 15 days (R3); the total hydraulic
retention time in the system was 37 days. The
temperature in all three reactors was maintained at
about 25 °C. Fenton's reagent (a 1:1 volumetric ratio
of 30 percent H2O2 and 0.0084 M ferrous sulfate
heptahydrate) was added at a rate of about 2 L/day.
The pH in R2, where the Fenton process occurred,
was maintained at 2, whereas the pH in R1 and R3
was maintained at about 7.
The system achieved an approximately 95 percent
reduction in the total PAH concentration and an
approximately 84 percent reduction in the
carcinogenic PAH concentration over a total
hydraulic retention time of 37 days. Individually, R1
achieved 87 ± 1 percent removal of PAHs and 65 ±
4 percent removal of carcinogenic PAHs. R2
achieved 45 ± 13 percent removal of PAHs and 49 ±
12 percent removal of carcinogenic PAHs. R3
achieved 3.9 ± 6.0 percent removal of PAHs and
0.42 ± 1.0 percent removal of carcinogenic PAHs.
Also at the pilot scale, Watts (1992) has studied the
effectiveness of treating petroleum
hydrocarbon-contaminated soil using the Fenton
process with no iron addition. The study was
performed ex situ using 0.96-m3 batches of soil in
55-gallon drums. The soil used was an arid soil with
low organic carbon content, low manganese oxide
content, and an unspecified amount of naturally
occurring iron minerals. During the study, 17 m3 of
soil was treated. Initial petroleum hydrocarbon
concentrations ranged from nondetectable levels to
6,900 mg/kg. The petroleum hydrocarbon was
primarily composed of high-molecular-weight
molecules. The study was performed using a 2 or
7 percent dose of 50 percent H2O2, with no iron
addition, and at a pH of 3. Over a period of 1 to
3 days, the system achieved the treatment goal
(100 mg/kg) with two additions of both the 2 and
7 percent H2O2 doses. No percent removals have
been reported. The estimated treatment cost based
5-5
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on the 2 percent H2O2 dose was about $44/m3 of soil
treated, which includes only the cost of the H2O2.
Bench-Scale Studies
This section discusses the effectiveness of the
Fenton process in treating soil contaminated with the
following SVOCs at the bench-scale level. ;
ANPO Process
Fenton
SVOCs Removed
Acenaphthene;
acenaphthylene;
anthracene; benzo(a)-
anthracene; benzo(b)-
fluoranthene; chrysene;
fluoranthene; fluorene;
hexadecane; naphthalene;
PCP; phenanthrene;
pyrene
Kawahara and others (1995) have evaluated the
effectiveness of using the Fenton process'to treat
soil slurries contaminated with various PAHs; the
slurries were collected from a wood treating facility.
The soil composition by weight was as follows:
gravel, 24 percent; sand, 22 percent; silt, 37 percent;
and clay, 18 percent. The PAHs in the soil included
acenaphthene (750 ug/kg), acenaphthylene
(10 ug/kg), anthracene (1,500 ug/kg),
benzo(a)anthracene (350 ug/kg),
benzo(b)fluoranthene (200 ug/kg), chrysene
(350 ug/kg)- fluoranthene (1,700 ug/kg), fluorene
(600 ug/kg), naphthalene (20 ug/kg), phenanthrene
(1,700 ug/kg), and pyrene (1,200 ug/kg). Soil
slurries consisting of 10 grams of contaminated soil
and 30 mL of water were treated with 40 mL of
Fenton's reagent (30 percent H2O2 and 0.009 M
FeSO4 mixed in equal proportions). The untreated
soil pH was 7. PAH removals ranged frorp
72 percent (naphthalene) to 93 percent
(acenaphthene). '.
Ravikumar and Gurol (1992) have evaluated the
effect of external FeSO4 addition when
PGP-contaminated soil was treated using the Fenton
process in a soil column study. The soil used for the
study was coarse-grained sand with a low organic
content (400 mg/kg) and an iron content of
800 mg/kg. The column was about 4 cm in diameter
and 85 cm long. The volume of the soil in the
column was about 1 L. With an initial PCP
concentration of 0.53 millimole in the soil, 73 percent
removal of PCP was achieved using 2.65 millimoles
H2O2 alone, whereas use of 1.76 millimoles H2O2
and 0.1 millimole FeSO4 resulted in 77 percent
removal of PCP. The pH of the system has not been
reported. On a molar basis, H2O2 treatment
generated 0.59 mole chloride per mole of H;,O2, and
Fenton's reagent generated 1.06 moles chloride per
mole of H2O2. The higher chloride liberation resulting
from treatment with Fenton's reagent is a strong
indication that Fenton's reagent is more effective in
oxidizing PCP in soil than H2O2 alone.
Watts and others (1990) have evaluated the
effectiveness of removing PCP in soil slurries using
the Fenton process. Three types of soils (silica sand
and two natural soils) were used. The organic
carbon content (0.05 versus 0.58 percent) was the
major difference between the natural soils. For the
study, 2.5 grams of soil was spiked with 12.5 mL of
65 percent H2O2 and 1 mL of FeSO4 such that the
concentration of Fe(ll) in the soil slurry was
480 mg/L. The initial pH of the slurry was adjusted
to between 2 and 8 using H2SO4. The highest
removal of PCP (>99 percent) was achieved for the
silica sand with a reaction time of 24 hours and a pH
of 3. The optimum pH for all three soil types was in
the range of 2 to 3. The study also evaluated the
effect of iron additions. Under optimal pH conditions,
the PCP treatment efficiency (that is, the ratio of the
contaminant degradation rate to the H2O2
consumption rate) was greater in systems that
(1) had a lower organic carbon content and (2) did
not receive iron amendments.
Tyre and others (1991) have studied the
effectiveness of removing PCP and hexadecane
from soil using the Fenton process. The soil used
was a gravelly, loamy, coarse sand with organic
carbon, concentrations ranging from 2,000 to
16,000 mg/kg; iron oxide concentrations ranging
from 7,800 to 8,800 mg/kg; and a manganese oxide
concentration of 200 mg/kg. For the study,
2.5 grams of soil was spiked with 200 mg/kg each of
PCP and hexadecane. The soil was treated with
12.5 mL of 13 percent H2O2 and 1 mL of FeSO4 or
deionized water such that the Fe(ll) concentration in
the soil was 0, 200, or 400 mg/L. Soil slurry
treatment was conducted at a pH of 3. PCP removal
rates decreased as a function of the soil organic
carbon content, whereas no such effect on
hexadecane removal rates was observed. The
removal efficiency ratios (that is, the ratios of the
contaminant removal rate constant to the H2O2
degradation rate constant) were highest for
treatment with no iron addition. According to the
researchers, these results suggest that iron minerals
in soil and H2O2 cause Fenton process-like oxidation
to occur. This hypothesis was verified by
demonstrating PCP removal in silica sand spiked
'5-6
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with iron minerals (goethite, hemotite, and
magnetite).
5.3 PCB-Contaminated Soil
No evaluations of commercial- or pilot-scale ANPO
processes for removing PCBs from soil have been
reported. However, PCB-contaminated soil has
been treated using the Fenton and gamma-ray
processes at the bench-scale level. This section
summarizes the results of bench-scale studies of the
effectiveness of the Fenton and gamma-ray
processes in removing the following PCBs from soil.
ANPO Process
Fenton
Gamma-ray
PCB Removed
• Aroclor 1248
• Aroclor 1260
Fenton
The effectiveness of the Fenton process in removing
Aroclor 1248 from a clayey humic soil from the
Texas Eastern Gas Pipeline site in Danville,
Kentucky, U.S.A., was studied by the International
Technology Corporation under the U.S. EPA
Superfund Innovative. Technology Evaluation
Program (U.S. EPA, 1995; International Technology
Corporation, 1995). At initial PCB concentrations
ranging from 5,000 to 10,000 ug/kg, minimal PCB
removal (0 to 55 percent) was achieved. The
highest PCB removal (55 percent) was achieved
when the percent moisture was 8.4; the pH was
maintained at 2.5; the Fe(ll) dose was 100 mg/kg;
the H2O2 dose was 18,000 mg/kg; and the retention
time was 118 hours. PCB removal was more
effective for less chlorinated congeners
(dichlorinated and trichlorinated biphenyls) than for
more chlorinated congeners.
Gamma-Ray
Curry and others (1998) have evaluated the
effectiveness of the gamma-ray process in treating
soil spiked with Aroclor 1260 at the bench-scale
level. For initial PCB concentrations of 58 and
200 mg/kg, high doses (100 Mrads) were required to
achieve >81 and >68 percent removals, respectively.
The researchers speculate that the soil surface may
have acted as an electron scavenger or may not
have allowed diffusion of aqueous electrons.
5.4 Pesticide- and Herbicide-
Contaminated Soil
No evaluations of commercial-scale ANPO
processes for removing pesticides or herbicides from
soil have been reported. However, the Fenton
process has been evaluated in terms of pesticide
and herbicide removal at the pilot- and bench-scale
levels. The results of these evaluations are
summarized below.
Pilot-Scale Application
A pilot-scale Fenton system was field-tested using
soil contaminated with organophosphorous
compounds (disulfoton and thiometon) and oxadixyl
(Egli and others, 1992). The system influent, which
was generated by an on-site soil washing system,
contained about 3.5 and 0.3 mg/L of total
organophosphorous compounds and oxadixyl,
respectively. During the field test, the system was
operated using an H2O2 dose of 1,000 mg/L; a
FeSO4 dose of 50 mg/L; a flow rate of 30 m3/hr
(corresponding to a retention time of 40 to
46 minutes); a pH of 3.0; and a temperature of
50 °C. The system achieved removals of about
94 percent for total organophosphorous compounds
in 2 hours. Over the same period, the oxadixyl
concentration was reduced to nondetectable levels,
representing >80 percent removal.
Bench-Scale Studies
This section summarizes the results of bench-scale
studies of the effectiveness of the Fenton process in
removing the following pesticides and herbicides
from synthetic soil matrices.
ANPO Process '
Fenton
Pesticide or
Herbicide Removed
Dieldrin; pendimethalin;
trifluralin
Miller and others (1996) found that pendimethalin
could be removed from soil slurries using the Fenton
process. The highest pendimethalin removal
(99 percent) was achieved when the soil to water
ratio was 0.2 on a weight basis; the H2O2 dose was
360,000 mg/kg of soil; the Fe(ll) dose was
2,000 mg/kg of soil; and the pH was maintained
between 2 and 3. Moreover, the study showed that
the Fenton process can remove the inhibitory effect
that pendimethalin has on biodegrading
microorganisms. Also, the Fenton process was
5-7
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shown to release BOD, COD, TOG, and nitrate ions
into solution. The organic matter released into
solution was biodegradable and served as a
substrate for subsequent microbial growth.
Tyre and others (1991) have studied the
effectiveness of removing dieldrin and trifluralin from
soil using the Fenton process. The soil used wasja
gravelly, loamy, coarse sand with organic carbon
concentrations ranging from 2,000 to 16,000 mg/kg;
iron oxide concentrations ranging from 7,800 to
8,800 mg/kg; and a manganese oxide concentration
of 200 mg/kg. For the study, 2.5 grams of soil was
spiked with 200 mg/kg each of dieldrin and trifluralin.
The soil was treated with 12.5 mL of 13 percent H2O2
and 1 mL of FeSO4 or deionized water such that the
Fe(ll) concentration in the soil was 0, 200, or
400 mg/L. Soil slurry treatment was conducted at £
pH of 3. Trifluralin removal rates decreased as a
function of the soil organic carbon content, whereas
no such effect on dieldrin removal rates was
observed. The removal efficiency ratios (that is, the
ratios of the 'contaminant removal rate constant to
the H2O2 degradation rate constant) were highest for
treatment with no iron addition. According to the
researchers, these results suggest that iron minerals
in soil and H2O2 cause Fenton process-like oxidation
to occur.
5.5 Dioxin-Contaminated Soil
No evaluations of commercial-scale ANPO
processes for removing dioxins from soil have been
reported. However, the Fenton process has been
evaluated in terms of dioxin removal at the pilot- and
bench-scale levels. The results of these evaluations
are summarized below. I
Pilot-Scale Application
A pilot-scale in situ Fenton system has been
field-tested using 2,3,7,8-tetrachlorodibenzo-p-dioxiri.
(TCDD)-contaminated soil (Hu and others, 2000).
The system employed was a proprietary injection
method designed to inject reagents under pressure.,
The soil in the test area was predominantly sand; the:
pH of the soil has not been reported. The TCDD
concentrations in the test area (which was 0.61 m
wide, 0.61 m long, and 1.2 m deep) decreased with!
depth, ranging from 240 ug/kg between 0 and 5.1 cm
bgs to 2.4 ug/kg between 117 and 122 cm bgs. For
the field test, Fenton's reagent was added in five
steps over2 consecutive days: (1) 230 L of 5 percent!
H2O2 at a flow rate of 60 L/hr; (2) 230 L of
2,000 mg/L Fe(ll) at a flow rate of 450 L/hr; (3) 110 L
of 35 percent H2O2 at a flow rate of 450 L/hr
(4) 230 L of 1,000 mg/L Fe(ll) at a flow rate of
450 L/hr; and (5) 1,900 L of 35 percent H2O2 at a
flow rate of 470 L/hr. The interval between
consecutive steps was about 8 hours. The system
achieved TCDD removals ranging from about 73 to
>99; TCDD concentrations were reduced to
nondetectable levels (<0.3 ug/kg) in most depth
intervals.
Bench-Scale Study
Watts and others (1991) have evaluated the
effectiveness of removing
octachlorodibenzo-p-dioxin (OCDD) in soil using the
Fenton process at the bench-scale level. For the
study, 1 gram of soil was spiked with OCDD at
200 ug/kg. The soil was treated with 2 mL of either
3.5 or 35 percent H2O2 and 0.1 millimole of FeSO4.
Removals of OCDD ranged from about 75 to
96 percent; the removal efficiency was inversely
proportional to the organic carbon content of the soil.
The highest removal (96 percent) was achieved
when 35 percent H2O2 was used and the organic
carbon content was 2 percent. To achieve
96 percent removal, the reaction proceeded until the
H2O2 was depleted, and then the reaction was
repeated two more times. OCDD removal generally
increased with increasing temperature and H2O2
concentration. Temperatures of 20, 40, 60, and
80 °C were used. At 20 °C, OCDD removals of
about 10 and 20 percent were achieved with use of
3.5 and 35 percent H2O2, respectively. However, at
80 °C, OCDD removals of about 80 and 85 percent
were achieved with use of 3.5 and 35 percent H2O2,
respectively.
5.6 Explosive- and Degradation
Product-Contaminated Soil
No evaluations of commercial-scale ANPO
processes for removing explosives and their
degradation products from soil have been reported.
However, the Fenton process has been evaluated in
terms of explosive and degradation product removal
at the pilot- and bench-scale levels. The results of
these evaluations are summarized below.
Pilot-Scale Application
A pilot-scale Fenton system was field-tested using
TNT-contaminated soil from the former Nebraska
Ordnance Plant in Mead, Nebraska, U.S.A. (Arienzo
and others, 1998). The system consisted.of three
units: (1) the soil sample preparation unit, (2) the soil
slurry treatment unit, and (3) the soil dewatering unit.
The soil used for the field test was a Charpsburg silty
loam (20 percent sand, 44 percent silt, and
5-8
-------
36 percent clay) that had a pH of 7.5 and an organic
carbon content of 1.9 percent.
TNT was the primary contaminant of concern and
was present in the soil at 400 mg/kg. For the field
test, TNT-contaminated soil was sieved to a particle
size of 0.1 cm in the sample preparation unit. In the
soil slurry treatment unit, 12 kg of soil from the
sample preparation unit was combined with 60 L of
water to produce a soil slurry with 20 percent solids.
O2 was supplied by dissolving compressed air in the
slurry using bubble diffusers. The pH of the slurry
was adjusted to 3.0 using 60 mL of H2SO4. Fenton's
reagent was added in one, four, or eight additions.
When Fenton's reagent was added in one batch,
H2O2 and Fe(ll) were added such that their
concentrations were 1 percent and 640 mg/L in the
reaction mixture, respectively. During the alternate
modes of operation, the total amount of Fenton's
reagent was divided into four or eight equal amounts
that were supplied to the treatment unit at
predetermined time intervals. The temperature in
the slurry treatment unit was maintained in the range
of 34 to 38 °C. Following the addition of Fenton's
reagent, the soil slurry was dewatered in the
dewatering unit using a porous, stainless-steel filter.
Soil in the form of a soft sludge-cake was removed
from the filter, air-dried, sieved to a particle size of
0.2 cm, and subsampled for TNT analysis. Leachate
from the filter was also subsampled for TNT analysis.
The highest removal (98 percent) of TNT achieved
by the system occurred when Fenton's reagent was
supplied in eight additions (one every 4 hours) and
the. total treatment time was 36 hours. The soil
treatment goal of 17 mg/kg was achieved, and the
concentration of TNT in the leachate from the soil
dewatering unit filter was minimal (<2 mg/L).
. Bench-Scale Study
Bier and others (1999) have evaluated the
effectiveness of removing RDX in soil slurry using
the Fenton process at the bench-scale level.
Contaminated soil was collected for the study from a
former drainage ditch adjacent to a former munitions
production building at the Nebraska Ordnance Plant
in Mood, Nebraska, U.S.A. The concentration of
RDX in soil ranged from 900 to 1,500 mg/kg. The
soil slurry was prepared by mixing soil and H2O at a
proportion of 1 gram of soil and 5 mL of H2O. The
H2O2 concentration in the slurry was either 0.59 or
1.18 M, and the Fe'(ll) concentration was 0.003 M.
The temperature was maintained at either 25 or
45 °C. At an initial RDX concentration of
1,300 mg/kg, >99 percent removal was achieved
through stepwise addition of Fenton's reagent over
48 hours at a temperature of 45 °C. CO2, formic
acid, ammonium, and nitrate ions were identified as
reaction by-products.
5-9
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5-14
-------
5.7 References
Arienzo, M., S.D. Comfort, M.Zerkoune, Z.M. Li, and
P.J. Shea. 1998. "Pilot-Scale Devices for
Remediation of Munitions Contaminated Soils."
Journal of Environmental Science and Health,
Part A: Toxic-Hazardous Substances &
Environmental Engineering. Volume 33,
Number 8. November. Pages 1515 through
1531.
Bier, Eleanor L, Jasbir Singh, Zhengming Li, Steve
D. Comfort, and Patrick J. Shea. 1999.
"Remediating Hexahydro-1,3,5-Trinitro-1,2,5-
Trazine[s/c]-Contaminated Water and Soil by
Fenton Oxidation." Environmental Toxicology
and Chemistry. Volume 18, Number 6. June.
Pages 1078 through 1084.
Curry, R.D., T. Clevenger, O. Stancu-Ciolac, W.H.
Miller, J. Farmer, B.J. Mincher, and S. Kapila.
1998. "Decontamination of Soil Contaminated
with Aroclor 1260 Using a Solvent Extraction
Process and v-Ray Radiolysis." Journal of
Advanced Oxidation Technologies. Volume 3,
Number 1. Pages 55 through 62.
Egli, S., S. Lomanto, R. Galli, R. Fitzi, and C. Munz.
1992. "Oxidative Treatment of Process Water in
a Soil Decontamination Plant: II. Pilot Plant and
Large Scale Experiences." Chemical Oxidation:
Technologies for the Nineties. Volume 2.
Edited by W.W. Eckenfelder, A.R. Bowers, and
J.A. Roth. Technomic Publishing Co., Inc.
Lancaster, Pennsylvania. Pages 264 through
277.
Geo-Cleanse International, Inc. (Geo-Cleanse).
2000. Correspondence Regarding Case Studies
on Geo-Cleanse® Fenton Process. From Matt
Dingens, Vice President. To Suzette Tay,
Environmental Scientist, Tetra Tech EM Inc.
H&H Eco Systems, incorporated (H&H). 2000.
Correspondence Regarding Case Studies on
ISOTEC Fenton Process. From Terry Horn,
President. To Suzette Tay, Environmental
Scientist, Tetra Tech EM Inc.
Hu, H., G.D. Hitchens, D. Hodko, T.D. Rogers, M.L.
Madigan, B.K. Carnley, and R.N. Reeves. 2000.
"In-Situ Remediation of Tetrachloro-
dibenzo-p-Dioxin (TCDD) Contaminated Soils
Using Fenton's Reagent." Unpublished.
In-Situ Oxidative Technologies, Inc. (ISOTEC).
2000. Correspondence Regarding Case Studies
on ISOTEC™ Fenton Process. From Dave
Zervas, President. To Suzette Tay,
Environmental Scientist, Tetra Tech EM Inc.
International Technology Corporation. 1995. "Site
Emerging Technologies Program, EO6 Bioslurry
Treatment, Final Report." Prepared for U.S.
Environmental Protection Agency National Risk
Management Research Laboratory.
Kawahara, Fred K., Brunilda Davila, Souhail R.
AI-Abed, Stephen J. Vesper, John C. Ireland,
and Steve Rock. 1995. "Polynuclear Aromatic
Hydrocarbon (PAH) Release from Soil During
Treatment with Fenton's Reagent."
Chemosphere. Volume 31, Number 9.
Pages 4131 through 4142.
Leung, Solomon W., Richard J. Watts, and Glenn C.
Miller. 1992. "Degradation of Perchloroethylene
by Fenton's Reagent: Speciation and Pathway."
Journal of Environmental Quality. Volume 21.
Pages 377 through 381.
Miller, Christopher M., Richard L. Valentine, Marc E.
Roehl, and Pedro J.J.Alvarez. 1996. "Chemical
and Microbiological Assessment of
Pendimethalin-contaminated Soil after
Treatment with Fenton's Reagent." Water
Research. Volume 30, Number 11. Pages 2579
through 2586.
Ravikumar, J.X., and M.D. Gurol: 1992. "Fenton's
Reagent as a Chemical Oxidant for Soil
Contaminants." Chemical Oxidation:
Technologies for the Nineties. Volume 2.
Edited by W.W. Eckenfelder, A.R. Bowers, and
J.A. Roth. Technomic Publishing Co., Inc.
Lancaster, Pennsylvania. Pages 206 through
216.
Tyre, Bryan W., Richard J. Watts, and Glenn C.
Miller. 1991. "Treatment of Four Biorefractory
Contaminants in Soils Using Catalyzed
Hydrogen Peroxide." Journal of Environmental
Quality. Volume 20. October-December.
Pages 832 through 838.
U.S. Environmental Protection Agency (U.S. EPA).
1995. "Superfund Innovative Technology
Evaluation, Emerging Technology Summary:
Bench-Scale Testing of Photolysis, Chemical
Oxidation, and Biodegradation of PCB
Contaminated Soils, and Photolysis of TCDD
Contaminated Soils." Office of Research and
Development Superfund Innovative Technology
5-15
-------
Evaluation (SITE) Program. Washington, DC.
EPA/540/SR-94/531. March.
U.S. EPA. 1997. "Superfund Innovative Technology
Evaluation, Emerging Technology Summary:
Innovative Methods for Bioslurry Treatment."
Office of Research and Development. SITE
Program. Washington, DC. EPA/540/S
R-96/505. August. ;
Wang, X., and M.L. Brussea. 1998. "Effect of
Pyrophosphate on the Dechlorination of
Tetrachloroethene by the Fenton Reaction,"
Environmental Toxicology and Chemistry.
Volume 17, Number 9. Pages 1689 through
1694.
Watts, R.J., M.D. Udell, P.A. Rauch, and S.W.
Leung. 1990. "Treatment of Pentachloro-
phenol-Contaminated Soils Using Fenton's
Reagent." Hazardous Waste & Hazardous
Materials. Volume 7, Number 4. Pages 335
through 345.
Watts, Richard J., B. Randy Smith, and Glenn C.
Miller. 1991. "Catalyzed Hydrogen Peroxide
Treatment of Octachlorodibenzo-p-Dioxin
(OCDD) in Surface Soils." Chemosphere.
Volume 24, Number 7. Pages 949 through 955.
Watts, R.J. 1992. "Hydrogen Peroxide for Physico-
chemically Degrading Petroleum-Contaminated
Soils." Remediation. Autumn. Pages 413
through 425.
5-16
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APPENDIX
TECHNOLOGY VENDOR CONTACT INFORMATION
Applied Process Technology, Inc.
Mr. Charles Borg
233 Sansome Street, Suite 1108
San Francisco, CA94104
Telephone: (415) 675-7280
Internet: www.aptwater.com
Chematur Engineering AB
Mr. Lars Stenmark
Box 430
S-691 27 Karlskoga
Sweden
Telephone: 46-586-641-00
Internet: www.chematur.se
General Atomics
Mr. Michael H. Spritzer
3550 General Atomics Court
San Diego, CA 92121
Telephone: (858) 455-3000
Internet: www.aat.com
Geo-Cleanse International, Inc.
Mr. Matthew M. Dingens
4 Mark Road, Suite C
Kenilworth, NJ 07033
Telephone: (908) 206-1250
Internet: www.geoclea nse .com
H&H Eco Systems, Inc.
Mr. Terry D. Horn
P.O. Box 38
505 Evergreen Drive
North Bonneville, WA 98639
Telephone: (509) 427-7353
Internet: www.hheco.com
High Voltage Environmental Applications, Inc.
Dr. William J. Cooper
601 South College Road
Wilmington, NC 28403
Telephone: (910) 962-2387
Internet: Under construction
In-Situ Oxidative Technologies, Inc.
Mr. Luis Moreno
3858 Benner Road
Miamisburg, OH 45342
Telephone: (800) 448-9760
Internet: www.isotec.com
Mantech Environmental Corporation
Dr. Timothy A. Hall
6300 West Loop South, Suite 500
Houston, TX 77401
Telephone: (713) 585-7000
Internet: www.mantech.com
Oxidation Systems, Inc.
Dr. Joseph A. Pisani
250 West Colorado Boulevard, Suite 190
Arcadia, CA 91007
Telephone: (626) 446-1482
Internet: Under construction
A-1
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