&EPA
United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/600/R-98/128
September 1998
Technical Protocol for
Evaluating Natural
Attenuation of Chlorinated
Solvents in Ground Water
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TECHNICAL PROTOCOL FOR EVALUATING NATURAL
ATTENUATION OF CHLORINATED SOLVENTS IN
GROUND WATER
by
Todd H. Wiedemeier
Parsons Engineering Science, Inc.
Pasadena, California
Matthew A. Swanson, David E. Moutoux, and E. Kinzie Gordon
Parsons Engineering Science, Inc.
Denver, Colorado
John T. Wilson, Barbara H. Wilson, and Donald H. Kampbell
United States Environmental Protection Agency
National Risk Management Research Laboratory
Subsurface Protection and Remediation Division
Ada, Oklahoma
Patrick E. Haas, Ross N. Miller and Jerry E. Hansen
Air Force Center for Environmental Excellence
Technology Transfer Division
Brooks Air Force Base, Texas
Francis H. Chapelle
United States Geological Survey
Columbia, South Carolina
IAG#RW57936164
Project Officer
John T. Wilson
National Risk Management Research Laboratory
Subsurface Protection and Remediation Division
Ada, Oklahoma
NATIONAL RISK MANAGEMENT RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U. S. ENVIRONMENTAL PROTECTION AGENCY
CINCINNATI, OHIO 45268
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NOTICE
The information in this document was developed through a collaboration between the U.S.
EPA (Subsurface Protection and Remediation Division, National Risk Management Research
Laboratory, Robert S. Kerr Environmental Research Center, Ada, Oklahoma [SPRD]) and the U.S.
Air Force (U.S. Air Force Center for Environmental Excellence, Brooks Air Force Base, Texas
[AFCEE]). EPA staff were primarily responsible for development of the conceptual framework for
the approach presented in this document; staff of the U.S. Air Force and their contractors also
provided substantive input. The U.S. Air Force was primarily responsible for field testing the
approach presented in this document. Through a contract with Parsons Engineering Science, Inc.,
the U.S. Air Force applied the approach at chlorinated solvent plumes at a number of U.S. Air
Force Bases. EPA staff conducted field sampling and analysis with support from ManTech
Environmental Research Services Corp., the in-house analytical support contractor for SPRD.
All data generated by EPA staffer by ManTech Environmental Research Services Corp. were
collected following procedures described in the field sampling Quality Assurance Plan for an in-
house research proj ect on natural attenuation, and the analytical Quality Assurance Plan for ManTech
Environmental Research Services Corp.
This protocol has undergone extensive external and internal peer and administrative review by
the U.S. EPA and the U.S. Air Force. This EPA Report provides technical recommendations, not
policy guidance. It is not issued as an EPA Directive, and the recommendations of this EPA Report
are not binding on enforcement actions carried out by the U.S. EPA or by the individual States of
the United States of America. Neither the United States Government (U. S. EPA or U. S. Air Force),
Parsons Engineering Science, Inc., or any of the authors or reviewers accept any liability or
responsibility resulting from the use of this document. Implementation of the recommendations of
the document, and the interpretation of the results provided through that implementation, are the
sole responsibility of the user.
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, and 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 these mandates, 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 reducing risks from threats to human health and
the environment. The focus of the Laboratory's research program is on methods for the prevention
and control of pollution to air, land, water, and subsurface resources; protection of water quality in
public water systems; remediation of contaminated sites and ground water; and prevention and
control of indoor air pollution. The goal of this research effort is to catalyze development and
implementation of innovative, cost-effective environmental technologies; develop scientific and
engineering information needed by EPA to support regulatory and policy decisions; and provide
technical support and information transfer to ensure effective implementation of environmental
regulations and strategies.
The site characterization processes applied in the past are frequently inadequate to allow an
objective and robust evaluation of natural attenuation. Before natural attenuation can be used in the
remedy for contamination of ground water by chlorinated solvents, additional information is required
on the three-dimensional flow field of contaminated ground water in the aquifer, and on the physical,
chemical and biological processes that attenuate concentrations of the contaminants of concern.
This document identifies parameters that are useful in the evaluation of natural attenuation of
chlorinated solvents, and provides recommendations to analyze and interpret the data collected
from the site characterization process. It will also allow ground-water remediation managers to
incorporate natural attenuation into an integrated approach to remediation that includes an active
remedy, as appropriate, as well as natural attenuation.
Clinton W. Hall, Director
Subsurface Protection and Remediation Division
National Risk Management Research Laboratory
in
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IV
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OF
Notice ii
Foreword iii
Acknowledgments viii
List of Acronyms and Abbreviations ix
Definitions xii
SECTION! INTRODUCTION 1
1.1 APPROPRIATE APPLICATION ON NATURAL ATTENUATION 2
1.2 ADVANTAGES AND DISADVANTAGES 4
1.3 LINES OF EVIDENCE 6
1.4 SITE CHARACTERIZATION 7
1.5 MONITORING 9
SECTION 2 PROTOCOL FOR EVALUATING NATURAL ATTENUATION 11
2.1 REVIEW AVAILABLE SITE DATA AND DEVELOP PRELIMINARY
CONCEPTUAL MODEL 13
2.2 INITIAL SITE SCREENING 15
2.2.1 Overview of Chlorinated Aliphatic Hydrocarbon Biodegradation 15
2.2.1.1 Mechanisms of Chlorinated Aliphatic Hydrocarbon Biodegradation 23
2.2.1.1.1 Electron Acceptor Reactions (Reductive Dehalogenation) 23
2.2.1.1.2 Electron Donor Reactions 25
2.2.1.1.3 Cometabolism 25
2.2.1.2 Behavior of Chlorinated Solvent Plumes 26
2.2.1.2.1 Type 1 Behavior 26
2.2.1.2.2 Type 2 Behavior 26
2.2.1.2.3 Type 3 Behavior 26
2.2.1.2.4 Mixed Behavior 27
2.2.2 Bioattenuation Screening Process 27
2.3 COLLECT ADDITIONAL SITE CHARACTERIZATION DATA IN
SUPPORT OF NATURAL ATTENUATION AS REQUIRED 34
2.3.1 Characterization of Soils and Aquifer Matrix Materials 37
2.3.2 Ground-water Characterization 38
2.3.2.1 Volatile and Semivolatile Organic Compounds 38
2.3.2.2 Dissolved Oxygen 38
2.3.2.3 Nitrate 39
2.3.2.4 Iron (II) 39
2.3.2.5 Sulfate 39
2.3.2.6 Methane 39
2.3.2.7 Alkalinity 39
2.3.2.8 Oxidation-Reduction Potential 40
2.3.2.9 Dissolved Hydrogen 40
2.3.2.10 pH, Temperature, and Conductivity 41
2.3.2.11 Chloride 42
2.3.3 Aquifer Parameter Estimation 42
2.3.3.1 Hydraulic Conductivity 42
2.3.3.1.1 Pumping Tests in Wells 43
2.3.3.1.2 Slug Tests in Wells 43
2.3.3.1.3 DownholeFlowmeter 43
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2.3.3.2 Hydraulic Gradient 44
2.3.3.3 Processes Causing an Apparent Reduction in
Total Contaminant Mass 44
2.3.4 Optional Confirmation of Biological Activity 45
2.4 REFINE CONCEPTUAL MODEL, COMPLETE PRE-MODELING
CALCULATIONS, AND DOCUMENT INDICATORS OF NATURAL
ATTENUATION 45
2.4.1 Conceptual Model Refinement 46
2.4.1.1 Geol ogi cLogs 46
2.4.1.2 Cone Penetrometer Logs 46
2.4.1.3 Hydrogeologic Sections 46
2.4.1.4 Potentiometric Surface or Water Table Map(s) 47
2.4.1.5 Contaminant and Daughter Product Contour Maps 47
2.4.1.6 Electron Acceptor, Metabolic By-product, and
Alkalinity Contour Maps 47
2.4.2 Pre-Modeling Calculations 48
2.4.2.1 Analysis of Contaminant, Daughter Product, Electron Acceptor,
Metabolic By-product, and Total Alkalinity Data 48
2.4.2.2 Sorption and Retardation Calculations 49
2.4.2.3 NAPL/Water Partitioning Calculations 49
2.4.2.4 Ground-water Flow Velocity Calculations 49
2.4.2.5 Biodegradation Rate-Constant Calculations 49
2.5 SIMULATE NATURAL ATTENUATION USING SOLUTE FATE AND
TRANSPORT MODELS 49
2.6 CONDUCT A RECEPTOR EXPOSURE PATHWAYS ANALYSIS 50
2.7 EVALUATE SUPPLEMENTAL SOURCE REMOVAL OPTIONS 50
2.8 PREPARE LONG-TERM MONITORING PLAN 50
2.9 PRESENT FINDINGS 52
SECTION 3 53
APPENDIX A Al-1
APPENDIXB Bl-1
APPENDIX C CM
VI
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FIGURES
No. Title
2.1 Natural attenuation of chlorinated solvents flow chart 12
2.2 Reductive dehalogenation of chlorinated ethenes 24
2.3 Initial screening process flow chart 28
2.4 General areas for collection of screening data 31
2.5 A cross section through a hypothetical release 36
2.6 A stacked plan representation of the plumes that may develop from the
hypothetical release 36
2.7 Hypothetical long-term monitoring strategy 51
TABLES
No. Title
i. Contaminants with Federal Regulatory Standards xiv
2.1 Soil, Soil Gas, and Ground-water Analytical Protocol 16
2.2 Objectives for Sensitivity and Precision to
Implement the Natural Attenuation Protocol 21
2.3 Analytical Parameters and Weighting for Preliminary Screening for
Anaerobic Biodegradation Processes 29
2.4 Interpretation of Points Awarded During Screening Step 1 32
2.5 Range of Hydrogen Concentrations for a Given Terminal
Electron-Accepting Process 41
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ACKNOWLEDGMENTS
The authors would like to thank Dr. Robert Hinchee, Doug Downey, and Dr. Guy Sewell for their
contributions and their extensive and helpful reviews of this manuscript. Thanks also to Leigh
Alvarado Benson, R. Todd Herrington, Robert Nagel, Cindy Merrill, Peter Guest, Mark Vesseley,
John Hicks, and Saskia Hoffer for their contributions to this project.
Vlll
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LIST OF
AAR
AFB
AFCEE
ASTM
bgs
BRA
BRAC
BTEX
CAP
CERCLA
cfm
CFR
COPC
CPT
CSM
DAF
DERP
DNAPL
DO
DOD
DQO
EE/CA
FS
gpd
HOPE
HSSM
HSWA
ID
IDW
IRP
L
LEL
LNAPL
LUFT
MAP
MCL
American Association of Railroads
Air Force Base
Air Force Center for Environmental Excellence
American Society for Testing and Materials
below ground surface
baseline risk assessment
Base Realignment and Closure
benzene, toluene, ethylbenzene, xylenes
corrective action plan
Comprehensive Environmental Response, Compensation and Liability
Act
cubic feet per minute
Code of Federal Regulations
chemical of potential concern
cone penetrometer testing
conceptual site model
dilution/attenuation factor
Defense Environmental Restoration Program
Dense Nonaqueous Phase Liquid
dissolved oxygen
Department of Defense
data quality objective
engineering evaluation/cost analysis
feasibility study
gallons per day
standard (Gibbs) free energy
high-density polyethylene
Hydrocarbon Spill Screening Model
Hazardous and Solid Waste Amendments of 1984
inside-diameter
investigation derived waste
Installation Restoration Program
liter
lower explosive limit
light nonaqueous-phase liquid
leaking underground fuel tank
management action plan
maximum contaminant level
IX
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MDL
MS
Hg/L
mg
nig/kg
mg/L
mg/m3
mm Hg
MOC
MOGAS
NAPL
NCP
NFRAP
NOAA
NOEL
NPL
OD
ORP
OSHA
OSWER
PAH
PEL
POA
POC
POL
ppmv
psi
PVC
QA
QC
RAP
RBCA
RBSL
redox
RFI
RI
RME
RPM
SAP
SARA
scfm
SPCC
method detection limit
microgram
microgram per kilogram
microgram per liter
milligram
milligrams per kilogram
milligrams per liter
milligrams per cubic meter
millimeters of mercury
method of characteristics
motor gasoline
nonaqueous-phasc liquid
National Contingency Plan
no further response action plan
National Oceanographic and Atmospheric Administration
no-observed-effect level
National Priorities List
outside-diameter
oxidation-reduction potential
Occupational Safety and Health Administration
Office of Solid Waste and Emergency Response
polycyclic aromatic hydrocarbon
permissible exposure limit
point-of-action
point-of-compliance
petroleum, oil, and lubricant
parts per million per volume
pounds per square inch
polyvinyl chloride
quality assurance
quality control
remedial action plan
risk-based corrective action
risk-based screening level
reduction/oxi dation
RCRA facility investigation
remedial investigation
reasonable maximum exposure
remedial project manager
sampling and analysis plan
Superfund Amendments and Reauthorization Act
standard cubic feet per minute
spill prevention, control, and countermeasures
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SSL soil screening level
SSTL site-specific target level
SVE soil vapor extraction
SVOC semivolatile organic compound
TC toxicity characteristic
TCLP toxicity-characteristic leaching procedure
TI technical impracticability
TMB trimethylbenzene
TOC total organic carbon
TPH total petroleum hydrocarbons
TRPH total recoverable petroleum hydrocarbons
TVH total volatile hydrocarbons
TVPH total volatile petroleum hydrocarbons
TWA time-weighted-average
UCL upper confidence limit
US United States
USGS US Geological Survey
UST underground storage tank
VOCs volatile organic compounds
XI
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DEFINITIONS
Aerohe: bacteria that use oxygen as an electron acceptor.
Anabolism: The process whereby energy is used to build organic compounds such as enzymes and
nucleic acids that are necessary for life functions. In essence, energy is derived from catabolism,
stored in high-energy intermediate compounds such as adenosine triphosphate (ATP), guanosine
triphosphatc (GTP) and acctyl-cocnzymc A, and used in anabolic reactions that allow a cell to
grow.
Anaerobe: Organisms that do not require oxygen to live.
Area of Attainment: The area over which cleanup levels will be achieved in the ground water. It
encompasses the area outside the boundary of any waste remaining in place and up to the boundary
of the contaminant plume. Usually, the boundary of the waste is defined by the source control
remedy. Note: this area is independent of property boundaries or potential receptors - it is the
plume area which the ground water must be returned to beneficial use during the implementation of
a remedy.
Anthropogenic: Man-made.
Autotrophs: Microorganisms that synthesize organic materials from carbon dioxide.
Catabolism: The process whereby energy is extracted from organic compounds by breaking them down
into their component parts.
Coefficient of Variation: Sample standard deviation divided by the mean.
Cofactor. A small molecule required for the function of an enzyme.
Cometabolism: The process in which a compound is fortuitously degraded by an enzyme or cofactor
produced during microbial metabolism of another compound.
Daughter Product: A compound that results directly from the biodegradation of another. For example
cw-l,2-dichloroethene (c/s-l,2-DCE)is commonly a daughter product of trichloroethene (TCE).
Dehydrohalogenation: Elimination of a hydrogen ion and a halide ion resulting in the formation of an
alkene.
Diffusion: The process whereby molecules move from a region of higher concentration to a region of
lower concentration as a result of Brownian motion.
Dihaloelimination: Reductive elimination of two halide substituents resulting in formation of an alkene.
Dispersivity: A property that quantifies mechanical dispersion in a medium.
Effective Porosity: The percentage of void volume that contributes to percolation: roughly equivalent to
the specific yield.
Electron Acceptor: A compound capable of accepting electrons during oxidation-reduction reactions.
Microorganisms obtain energy by transferring electrons from electron donors such as organic
compounds (or sometimes reduced inorganic compounds such as sulfide) to an electron acceptor.
Electron acceptors are compounds that are relatively oxidized and include oxygen, nitrate,
iron (III), manganese (IV), sulfate. carbon dioxide, or in some cases the chlorinated aliphatic
hydrocarbons such as pcrchlorocthcnc (PCE), TCE, DCE, and vinyl chloride.
Electron Donor: A compound capable of supplying (giving up) electrons during oxidation-reduction
reactions. Microorganisms obtain energy by transferring electrons from electron donors such as
organic compounds (or sometimes reduced inorganic compounds such as sulfide) to an electron
acceptor. Electron donors are compounds that are relatively reduced and include fuel
hydrocarbons and native organic carbon.
Electrophile: A reactive species that accepts an electron pair.
Elimination: Reaction where two groups such as chlorine and hydrogen are lost from adjacent carbon
atoms and a double bond is formed in their place.
Epoxidation: A reaction wherein an oxygen molecule is inserted in a carbon-carbon double bond and an
epoxide is formed.
xn
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Facultative Anaerobes: microorganisms that use (and prefer) oxygen when it is available, but can also use
alternate electron acceptors such as nitrate under anaerobic conditions when necessary.
Fermentation: Microbial metabolism in which a particular compound is used both as an electron donor
and an electron acceptor resulting in the production of oxidized and reduced daughter products.
Heterotroph: Organism that uses organic carbon as an external energy source and as a carbon source.
Hydraulic Conductivity: The relative ability of a unit cube of soil, sediment, or rock to transmit water.
Hydraulic Head: The height above a datum plane of the surface of a column of water. In the
ground-water environment, it is composed dominantly of elevation head and pressure head.
Hydraulic Gradient: The maximum change in head per unit distance.
Hydrogenolysis: A reductive reaction in which a carbon-halogen bond is broken, and hydrogen replaces
the halogen substituent.
Hydroxylation: Addition of a hydroxyl group to a chlorinated aliphatic hydrocarbon.
Lithotroph: Organism that uses inorganic carbon such as carbon dioxide or bicarbonate as a carbon
source and an external source of energy.
Mechanical Dispersion: A physical process of mixing along a flow path in an aquifer resulting from
differences in path length and flow velocity. This is in contrast to mixing due to diffusion.
Metabolic Byproduct: A product of the reaction between an electron donor and an electron acceptor.
Metabolic byproducts include volatile fatty acids, daughter products of chlorinated aliphatic
hydrocarbons, methane, and chloride.
Monooxygenase: A microbial enzyme that catalyzes reactions in which one atom of the oxygen molecule
is incorporated into a product and the other atom appears in water.
Nucleophile: A chemical reagent that reacts by forming covalent bonds with electronegative atoms and
compounds.
Obligate Aerobe: Microorganisms that can use only oxygen as an electron acceptor. Thus, the presence
of molecular oxygen is a requirement for these microbes.
Obligate Anaerobes: Microorganisms that grow7 only in the absence of oxygen; the presence of molecular
oxygen either inhibits growth or kills the organism. For example, methanogens are very sensitive
to oxygen and can live only under strictly anaerobic conditions. Sulfate reducers, on the other
hand, can tolerate exposure to oxygen, but cannot grow in its presence (Chapclle, 1993).
Performance Evaluation Well: A ground-water monitoring well placed to monitor the effectiveness of
the chosen remedial action.
Porosity: The ratio of void volume to total volume of a rock or sediment.
Respiration: The process of coupling oxidation of organic compounds with the reduction of inorganic
compounds, such as oxygen, nitrate, iron (III), manganese (TV), and sulfate.
Solvolysis: A reaction in which the solvent serves as the nucleophile.
Xlll
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Table i: Contaminants with Federal Regulatory Standards Considered in this Document
Abbreviation
PCE
TCE
1,1 -DCE
trans- 1,2-DCE
cis-l,2-DCE
VC
1,1,1-TCA
1,1,2-TCA
1,1 -DCA
1,2-DCA
CA
CF
CT
Methylene Chloride
CB
1,2-DCB
1,3-DCB
1,4-DCB
1,2,3-TCB
1,2,4-TCB
1,3,5-TCB
1,2,3,5-TECB
1,2,4,5-TECB
HCB
EDB
Chemical Abstracts Service
(CAS) Name
tetrachloroethene
trichloroethene
1 , 1 -dichloroethene
(E)-l,2-dichloroethene
chloroethene
1,1,1 -trichloroethane
1 , 1 ,2-trichloroethane
1 , 1 -dichloroethane
1 ,2-dichloroethane
chloroethane
trichloromethane
tetrachloromethane
dichloromethane
chlorobenzene
1 ,2-dichlorobenzene
1 ,3 -dichlorobenzene
1 ,4-dichlorobenzene
1,2,3-trichlorobenzene
1 ,2,4-trichlorobenzene
1,3,5 -trichlorobenzene
1 ,2,3 ,5 -tetrachlorobenzene
1, 2,4,5 -tetrachlorobenzene
hexachlorobenzene
1 ,2-dibromoethane
CAS
Number
127-18-4
79-01-6
75-35-4
156-60-5
156-59-2
75-01-4
71-55-6
79-00-5
75-34-3
107-06-02
75-00-3
67-66-3
56-23-5
75-09-2
108-90-7
95-50-1
541-73-1
106-46-7
87-61-6
120-82-1
108-70-3
634-90-2
95-94-3
118-74-1
106-93-4
Other Names
perchloroethylene ; tetrachloroethylene
trichloroethylene
1,1-dichloroethylene; vinylidine chloride
trans- 1 ,2-dichloroethene;trans- 1 ,2- dichloroethylene
cis- 1 ,2-dichloroethene; cis- 1 ,2 -dichloroethylene
vinyl chloride; chloroethylene
chloroform
carbon tetrachloride
methylene dichloride
o-dichlorobenzene
m-dichlorobenzene
p-dichlorobenzene
1,2,3,5-TCB
ethylene dibromide; dibromoethane
Molecular
Formula
C2C14
C2HC13
C2H2C12
C2H2C12
C2H2C12
C2H3C1
C2H3C13
C2H3C13
C2H4C12
C2H4C12
C2H5C1
CHC13
ecu
CH2C12
C6H5C1
CeFi4Cl2
CeFi4Cl2
C6H4C12
CeH3Cl3
C6H3C13
CeH3Cl3
C6H2C14
C6H2C14
\^6\^L()
C2ri4j:>r2
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SECTION 1
INTRODUCTION
Natural attenuation processes (biodegradation, dispersion, sorption, volatilization) affect the
fate and transport of chlorinated solvents in all hydrologic systems. When these processes are
shown to be capable of attaining site-specific remediation objectives in a time period that is reasonable
compared to other alternatives, they may be selected alone or in combination with other more
active remedies as the preferred remedial alternative. Monitored Natural Attenuation (MNA) is a
term that refers specifically to the use of natural attenuation processes as part of overall site
remediation. The United States Environmental Protection Agency (U.S. EPA) defines monitored
natural attenuation as (OSWER Directive 9200.4-17, 1997):
The term "monitored natural attenuation, " as used in this Directive, refers
to the reliance on natural, attenuation processes (within the context of a. carefully
controlled and monitored clean-up approach) to achieve site-specific remedial
objectives within a time frame that is reasonable compared to other 'methods. The
"natural attenuation processes " that are at work in such a remediation approach
include a. variety of physical, chemical, or biological processes that, under favorable
conditions, act without human intervention to reduce the mass, toxicity, mobility,
volume, or concentration of contaminants in soil and ground water. These in-situ
processes include, biodegradation, dispersion, dilution, sorption, volatilization, and
chemical or biological stabilization, transformation, or destruction of contaminants.
Monitored natural attenuation is appropriate as a remedial approach only
when it can be demonstrated capable of achieving a site's remedial objectives within
a time frame that is reasonable compared to that offered by other methods and
where it meets the applicable remedy selection program for a particular OSWER
program. EPA, therefore, expects that monitored natural attenution typically will
be used in conjunction with active remediation measures (e.g., source control), or
as a follow-up to active remediation measures that have already been implemented.
The intent of this document is to present a technical protocol for data collection and analysis
to evaluate monitored natural attenuation through biological processes for remediating ground
water contaminated with mixtures of fuels and chlorinated aliphatic hydrocarbons. This document
focuses on technical issues and is not intended to address policy considerations or specific regulatory
or statutory requirements. In addition, this document does not provide comprehensive guidance on
overall site characterization or long-term monitoring of MNA remedies. Users of this protocol
should realize that different Federal and State remedial programs may have somewhat different
remedial objectives. For example, the CERCLA and RCRA Corrective Action programs generally
require that remedial actions: 1) prevent exposure to contaminated ground water, above acceptable
risk levels; 2) minimize further migration of the plume; 3) minimize further migration of
contaminants from source materials; and 4) restore the plume to cleanup levels appropriate for
current or future beneficial uses, to the extent practicable. Achieving such objectives could often
require that MNA be used in conjunction with other "active" remedial methods. For other cleanup
programs, remedial objectives may be focused on preventing exposures above acceptable levels.
Therefore, it is imperative that users of this document be aware of and understand the Federal and
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State statutory and regulatory requirements, as well as policy considerations that apply to a specific
site for which this protocol will be used to evaluate MNA as a remedial option. As a general
practice (i.e., not just pertaining to this protocol), individuals responsible for evaluating remedial
alternatives should interact with the overseeing regulatory agency to identify likely characterization
and cleanup objectives for a particular site prior to investing significant resources. The policy
framework within which MNA should be considered for Federal cleanup programs is described in
the November 1997 EPA Directive titled, "Use of Monitored Natural Attenuation at Superfund,
RCRA Corrective Action and Underground Storage Tank Sites" (Directive No. 9200.4-17).
This protocol is designed to evaluate the fate in ground water of chlorinated aliphatic
hydrocarbons and/or fuel hydrocarbons. Because documentation of natural attenuation requires
detailed site characterization, the data collected under this protocol can be used to compare the
relative effectiveness of other remedial options and natural attenuation. This protocol should be
used to evaluate whether MNA by itself or in conjunction with other remedial technologies is
sufficient to achieve site-specific remedial objectives. In evaluating the appropriateness of MNA,
the user of this protocol should consider both existing exposure pathways, as well as exposure
pathways arising from potential future uses of the ground water.
This protocol is aimed at improving the characterization process for sites at which a remedy
involving monitored natural attenuation is being considered. It contains methods and recommended
strategies for completing the remedial investigation process. Emphasis is placed on developing a
more complete understanding of the site through the conceptual site model process, early pathways
analysis, and evaluation of remedial processes to include MNA. Understanding the contaminant
flow field in the subsurface is essential for a technically justified evaluation of an MNA remedial
option; therefore, use of this protocol is not appropriate for evaluating MNA at sites where the
contaminant flow field cannot be determined with an acceptable degree of certainty (e.g., complex
fractured bedrock, karst aquifers).
In practice, natural attenuation also is referred to by several other names, such as intrinsic
remediation, intrinsic bioremediation, natural restoration, or passive bioremediation. The goal of
any site characterization effort is to understand the fate and transport of the contaminants of concern
over time in order to assess any current or potential threat to human health or the environment.
Natural attenuation processes, such as biodegradation, can often be dominant factors in the fate and
transport of contaminants. Thus, consideration and quantification of natural attenuation is essential
to a more thorough understanding of contaminant fate and transport.
1.1 APPROPRIATE APPLICATION ON NATURAL ATTENUATION
The intended audience for this document includes Project Managers and their contractors,
scientists, consultants, regulatory personnel, and others charged with remediating ground water
contaminated with chlorinated aliphatic hydrocarbons or mixtures of fuel hydrocarbons and
chlorinated aliphatic hydrocarbons. This protocol is intended to be used within the established
regulatory framework appropriate for selection of a remedy at a particular hazardous waste site
(e.g., the nine-criteria analysis used to evaluate remedial alternatives in the CERCLA remedy
selection process). It is not the intent of this document to replace existing U.S. EPA or state-
specific guidance on conducting remedial investigations.
The EPA does not consider monitored natural attenuation to be a default or presumptive
remedy at any contaminated site (OSWER Directive 9200.4-17, 1997), as its applicability is highly
variable from site to site. In order for MNA to be selected as a remedy, site-specific determinations
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will always have to be made to ensure that natural attenuation is sufficiently protective of human
health and the environment.
Natural attenuation in ground-water systems results from the integration of several subsurface
attenuation mechanisms that are classified as either destructive or nondestructive. Biodegradation
is the most important destructive attenuation mechanism, although abiotic destruction of some
compounds does occur. Nondestructive attenuation mechanisms include sorption, dispersion, dilution
from recharge, and volatilization. The natural attenuation of fuel hydrocarbons is described in the
Technical Protocol for Implementing Intrinsic Remediation with Long- Term. Monitoring for Natural
Attenuation of Fuel Contamination Dissolved in Groundwater, published by the Air Force Center
for Environmental Excellence (AFCEE) (Wiedemeier et al, 1995d). This document differs from
the technical protocol for intrinsic remediation of fuel hydrocarbons because it focuses on the
individual processes of chlorinated aliphatic hydrocarbon biodegradation which are fundamentally
different from the processes involved in the biodegradation of fuel hydrocarbons.
For example, biodegradation of fuel hydrocarbons, especially benzene, toluene, ethylbenzene,
and xylenes (BTEX), is mainly limited by electron acceptor availability, and generally will proceed
until all of the contaminants biochemically accessible to the microbes are destroyed. In the experience
of the authors, there appears to be an adequate supply of electron acceptors in most, if not all,
hydrogeologic environments. On the other hand, the more highly chlorinated solvents such as
perchloroethene (PCE) and trichloroethene (TCE) typically are biodegraded under natural conditions
via reductive dechlorination, a process that requires both electron acceptors (the chlorinated aliphatic
hydrocarbons) and an adequate supply of electron donors. Electron donors include fuel hydrocarbons
or other types of anthropogenic carbon (e.g., landfill leachate) or natural organic carbon. If the
subsurface environment is depleted of electron donors before the chlorinated aliphatic hydrocarbons
are removed, biological reductive dechlorination will cease, and natural attenuation may no longer
be protective of human health and the environment. This is the most significant difference between
the processes of fuel hydrocarbon and chlorinated aliphatic hydrocarbon biodegradation.
For this reason, it is more difficult to predict the long-term behavior of chlorinated aliphatic
hydrocarbon plumes than fuel hydrocarbon plumes. Thus, it is important to have a good
understanding of the important natural attenuation mechanisms. Data collection should include all
pertinent parameters to evaluate the efficacy of natural attenuation. In addition to having a better
understanding of the processes of advection, dispersion, dilution from recharge, and sorption, it is
necessary to better quantify biodegradation. This requires an understanding of the interactions
between chlorinated aliphatic hydrocarbons, anthropogenic or natural carbon, and inorganic electron
acceptors at the site. Detailed site characterization is required to adequately document and understand
these processes. The long-term monitoring strategy should consider the possibility that the behavior
of a plume may change over time and monitor for the continued availability of a carbon source to
support reductive dechlorination.
An understanding of the attenuation mechanisms is also important to characterizing exposure
pathways. After ground water plumes come to steady state, sorption can no longer be an important
attenuation mechanism. The most important mechanisms will be biotransformation, discharge
through advective flow, and volatilization. As an example, Martin and Imbrigiotta (1994) calibrated
a detailed transport and fate model to a release of pure TCE at Picatinny Arsenal, in New Jersey.
The plume was at steady state or declining. Ten years after surface spills ceased, leaching of
contaminants from subsurface DNAPLs and desorption from fine-grained layers were the only
processes identified that continued to contribute TCE to ground water. Desorption of TCE occurred
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at a rate of 15 to 85 mg/second. Anaerobic biotransformation consumed TCE at a rate of up to 30
mg/second, advective flow and discharge of TCE to surface water accounted for up to 2 mg/
second, and volatilization of TCE accounted for 0.1 mg/second. In this case, recharge of
uncontaminated water drove the plume below the water table, which minimized the opportunity for
volatization to the unsaturated zone. As a result, discharge to surface water was the only important
exposure pathway. Volatilization will be more important at sites that do not have significant
recharge to the water table aquifer, or that have NAPLs at the water table that contain chlorinated
organic compounds.
Chlorinated solvents are released into the subsurface as either aqueous-phase or nonaqueous
phase liquids. Typical solvent releases include nonaqueous phase relatively pure solvents that are
more dense than water and aqueous rinseates. Additionally, a release may occur as a mixture of
fuel hydrocarbons or sludges and chlorinated aliphatic hydrocarbons which, depending on the
relative proportion of each compound group, may be more or less dense than water. If the NAPL
is more dense than water, the material is referred to as a "dense nonaqueous-phase liquid," or
DNAPL. If the NAPL is less dense than water the material is referred to as a "light nonaqueous-
phase liquid," or LNAPL. Contaminant sources generally consist of chlorinated solvents present
as mobile NAPL (NAPL occurring at sufficiently high saturations to drain under the influence of
gravity into a well) and residual NAPL (NAPL occurring at immobile, residual saturations that are
unable to drain into a well by gravity). In general, the greatest mass of contaminant is associated
with these NAPL source areas, not with the aqueous phase.
When released at the surface, NAPLs move downward under the force of gravity and tend to
follow preferential pathways such as along the surface of sloping fine-grained layers or through
fractures in soil or rock. Large NAPL releases can extend laterally much farther from the release
point than would otherwise be expected, and large DNAPL releases can sink to greater depths than
expected by following preferential flow paths. Thus, the relative volume of the release and potential
migration pathways should be considered when developing the conceptual model for the distribution
of NAPL in the subsurface.
As water moves through NAPL areas (recharge in the vadose zone or ground water flow in an
aquifer), the more soluble constituents partition into the water to generate a plume of dissolved
contamination and the more volatile contaminants partition to the vapor phase. After surface
releases have stopped, NAPLs remaining in the subsurface tend to "weather" over time as volatile
and soluble components are depleted from NAPL surfaces. Even considering this "weathering"
effect, subsurface NAPLS continue to be a source of contaminants to ground water for a very long
time. For this reason, identification and delineation of subsurface zones containing residual or
free-phase NAPL is an important aspect of the site conceptual model to be developed for evaluating
MNA or other remediation methods.
Removal, treatment or containment of NAPLs may be necessary for MNA. to be a viable
remedial option or to decrease the time needed for natural processes to attain site-specific remediation
objectives. In cases where removal of mobile NAPL is feasible, it is desirable to remove this
source material and decrease the time required to reach cleanup objectives. Where removal or
treatment of NAPL is not practical, source containment may be practicable and necessary for MNA
to be a viable remedial option.
1.2 ADVANTAGES AND DISADVANTAGES
In comparison to engineered remediation technologies, remedies relying on monitored natural
attenuation have the following advantages and disadvantages, as identified in OSWER Directive
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9200.4-17, dated November 1997. (Note that this an iterim, not a final, Directive which was released
by EPA for use. Readers are cautioned to consult the final version of this Directive when it becomes
available.)
The advantages of monitored natural attenuation (MNA) remedies are:
• As with any in situ process, generation of lesser volume of remediation wastes reduced potential
for cross-media transfer of contaminants commonly associated with ex situ treatment, and
reduced risk of human exposure to contaminated media;
* Less intrusion as few surface structures are required;
* Potential for application to all or part of a given site, depending on site conditions and cleanup
objectives;
* Use in conjunction with, or as a follow-up to, other (active) remedial measures; and
* Lower overall remediation costs than those associated, with active remediation.
The potential disadvantages of monitored natural attenuation (MNA) include:
• Longer time frames may be required to achieve remediation objectives, compared to active
remediation;
* Site characterization may be more complex and costly;
* Toxicity of transformation products may exceed that of the parent compound;
* Long-term monitoring will generally be necessary;
* Institutional controls may be necessary to ensure long-term protectiveness;
* Potential exists for continued contamination migration, and/or cross-media, transfer of
contaminants;
* Hydrologic and geochemical conditions amenable to natural attenuation are likely to change
over time and could re suit in renewed mobility of previously stabilized contaminants, adversely
impacting remedial effectiveness; and
* More extensive education and outreach efforts may be required in order to gain public
acceptance of monitored natural attenuation.
At some sites the same geochemical conditions and processes that lead to biodegradation of
chlorinated solvents and petroleum hydrocarbons can chemically transform naturally occurring
manganese, arsenic and other metals in the aquifer matrix, producing forms of these metals that
are more mobile and/or more toxic than the original materials. A comprehensive assessment of
risk at a hazardous waste site should include sampling and analysis for these metals.
This document describes (1) those processes that bring about natural attenuation, (2) the site
characterization activities that may be performed to conduct a full-scale evaluation of natural
attenuation, (3) mathematical modeling of natural attenuation using analytical or numerical solute
fate and transport models, and (4) the post-modeling activities that should be completed to ensure
successful evaluation and verification of remediation by natural attenuation. The objective is to
quantify and provide defensible data to evaluate natural attenuation at sites where naturally occurring
subsurface attenuation processes are capable of reducing dissolved chlorinated aliphatic hydrocarbon
and/or fuel hydrocarbon concentrations to acceptable levels. A comment made by a member of the
regulatory community summarizes what is required to successfully implement natural attenuation:
A regulator looks for the data, necessary to determine that a. proposed.
treatment technology, if properly installed and operated, will reduce the contaminant
concentrations in the soil and water to legally mandated, limits. In this sense, the
use of biological treatment systems calls for the same level of investigation,
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demonstration of effectiveness, and monitoring as any conventional [remediation]
system (National Research Council, 1993),
When the rate of natural attenuation of site contaminants is sufficient to attain site-specific
remediation objectives in a time period that is reasonable compared to other alternatives, MNA
may be an appropriate remedy for the site. This document presents a technical course of action that
allows converging lines of evidence to be used to scientifically document the occurrence of natural
attenuation and quantify the rate at which it is occurring. Such a "weight-of-evidence" approach
will greatly increase the likelihood of successfully implementing natural attenuation at sites where
natural processes are restoring the environmental quality of ground water.
1.3 LINES OF EVIDENCE
The OSWER Directive 9200.4-17 (1997) identifies three lines of evidence that can be used to
estimate natural attenuation of chlorinated aliphatic hydrocarbons, including:
(1) Historical ground-water and/or soil chemistry data that demonstrate a clear and meaningful
trend of decreasing contaminant mass and/or concentration over time at appropriate
monitoring or sampling points. (In the case of a ground water plume, decreasing
concentrations should not be solely the result of plume migration. In the case of inorganic
contaminants, the primary attenuating mechanism should also be understood.)
(2) Hydrogeologic andgeochemical data thai can he used, to demonstrate indirectly the lype(s)
of natural attenuation processes active at the site, and the rate at which such processes
will reduce contaminant concentrations to required levels. For example, characterization
data may be used to quantify the rates of contaminant sorplion, dilution, or volatilization,
or to demonstrate and quantify the rates of biological degradation processes occurring at
the site.
(3) Data from field or microcosm studies (conducted in or with actual contaminated site
media) which directly demonstrate the occurrence of a particular natural attenuation
process at the site and its ability to degrade the contaminants of concern (typically used to
demonstrate biological degradation processes only).
The OSWER Directive provides the following guidance on interpreting the lines of evidence:
Unless EPA or the implementing state agency determines that historical
data (Number 1 above) are of sufficient quality and duration to support a decision
to use monitored natural attenuation, EPA expects that data characterizing the
nature and rates of natural attenuation processes at the site (Number 2 above)
should be provided. Wltere the latter are also inadequate or inconclusive, data
from microcosm studies (Number 3 above) may also be necessary. In general,
more supporting information may be required to demonstrate the efficacy of
monitored natural attenuation at those sites with contaminants which do not readily
degrade through biological processes (e.g., most non-petroleum compounds,
inorganics), at sites with contaminants that transform into more toxic and/or mobile
forms than the parent contaminant, or at sites where monitoring has been performed
for a relatively short period of time. The amount and type of information needed for
such a demonstration will depend upon a number of site-specific factors, such as the
size and nature of the contamination problem, the proximity of receptors and the
potential risk to those receptors, and other physical characteristics of the
environmental setting (e.g., hydrogeology, ground cover, or climatic conditions).
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The first line of evidence does not prove that contaminants are being destroyed. Reduction in
contaminant concentration could be the result of advection, dispersion, dilution from recharge,
sorption, and volatilization (i.e., the majority of apparent contaminant loss could be due to dilution).
However, this line of evidence is critical for determining if any exposure pathways exist for current
or potential future receptors.
In order to evaluate remediation by natural attenuation at most sites, the investigator will
have to determine whether contaminant mass is being destroyed. This is done using either, or
both, of the second or third lines of evidence. The second line of evidence relies on chemical and
physical data to show that contaminant mass is being destroyed, not just being diluted or sorbed to
the aquifer matrix. For many contaminants, biodegradation is the most important process, but for
certain contaminants nonbiological reactions are also important. The second line of evidence is
divided into two components:
* Using chemical analytical data in mass balance calculations to show that decreases in
contaminant and electron acceptor/donor concentrations can be directly correlated to
increases in metabolic end products/daughter compounds. This evidence can be used to
show that electron acceptor/donor concentrations in ground water are sufficient to facilitate
degradation of dissolved contaminants. Solute fate and transport models can be used to
aid mass balance calculations and to collate and present information on degradation.
* Using measured concentrations of contaminants and/or biologically recalcitrant tracers in
conjunction with aquifer hydrogeologic parameters such as seepage velocity and dilution
to show that a reduction in contaminant mass is occurring at the site and to calculate
biodegradation rate constants.
The biodegradation rate constants are used in conjunction with the other fate and transport
parameters to predict contaminant concentrations and to assess risk at downgradient performance
evaluation wells and within the area of the dissolved plume.
Microcosm studies may be necessary to physically demonstrate that natural attenuation is
occurring. Microcosm studies can also be used to show that indigenous biota are capable of degrading
site contaminants at a particular rate. Microcosm studies for the purpose of developing rate
constants should only be undertaken when they are the only means available to obtain biodegradation
rate estimates. There are two important categories of sites where it is difficult or impossible to
extract rate constants from concentrations of contaminants in monitoring wells in the field. In
some sites, important segments of the flow path to receptors are not accessible to monitoring because
of landscape features (such as lakes or rivers) or property boundaries that preclude access to a site
for monitoring. In other sites that are influenced by tides, or the stage of major rivers, or ground
water extraction wells, the ground water plume trajectory changes so rapidly that it must be described
in a statistical manner. A "snapshot" round of sampling cannot be used to infer the plume velocity
in calculations of the rate of attenuation.
1.4 CHARACTERIZATION
The OSWER Directive 9200.4-17 (1997) describes EPA requirements for adequate site
characterization.
Decisions to employ monitored natural attenuation as a remedy or remedy
component should he thoroughly and adequately supported with site-specific
characterization data and analysis. In general, the level of site characterization
necessary to support a comprehensive evaluation of natural attenuation is more
detailed than that needed to support active remediation. Site characterizations for
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natural attenuation generally warrant a quantitative understanding of source mass;
ground water flow; contaminant phase distribution and partitioning between soil,
ground water, and soil gas; rates of biological and non-biological transformation;
and an understanding of how all of these factors are likely to vary with time. This
information is generally necessary since contaminant behavior is governed by
dynamic processes which must be well understood, before natural attenuation can
be appropriately applied at a site. Demonstrating the efficacy of this remediation
approach likely will require analytical or numerical simulation of complex
attenuation processes. Such analyses, which are critical to demonstrate natural
attenuation \s ability to meet remedial action objectives, generally require a detailed
conceptual site model as a foundation.
A conceptual site model is a three-dimensional representation that conveys
what is known or suspected about contamination sources, release mechanisms, and
the transport and fate of those contaminants. The conceptual m.odel provides the
basis for assessing potential remedial technologies at the site. "Conceptual site
model" in not synonymous with "computer model; " however, a computer model
may be helpful for understanding and visualizing current site conditions or for
predictive simulations of potential future conditions. Computer models, which
simulate site processes mathematically, should in turn be based upon sound
conceptual site models to provide meaningful information. Computer models typically
require a lot of data, and the quality of the output from computer models is directly
related to the quality of the input data. Because of the complexity of natural systems,
models necessarily rely on simplifying assumptions that may or may not accurately
represent the dynamics of the natural system..
Site characterization should include collecting data to define (in three spatial
dimensions over time) the nature and distribution of contamination, sources as well
as the extent of the ground water plume and its potential impacts on receptors.
However, where monitored, natural attenuation will be considered as a remedial
approach, certain aspects of site characterization may require more detail or
additional elements. For example, to assess the contributions ofsorption, dilution,
and dispersion to natural attenuation of contaminated groundwater, a very detailed
understanding of aquifer hydraulics, recharge and discharge areas and. volumes,
and chemical properties is required. Where biodegradation will be assessed,
characterization also should include evaluation of the nutrients and. electron, donors
and acceptors present in the ground water, the concentrations of co-metabolites
and metabolic by-products, and perhaps specific analyses to identify the microbial
populations present. The findings of these, and any other analyses pertinent to
characterizing natural attenuation processes, should be incorporated into the
conceptual model of contaminant fate and transport developed for the site.
Development of an adequate database during the iterative site characterization process is an
important step in the documentation of natural attenuation. Site characterization should provide
data on the location, nature, phase distribution, and extent of contaminant sources. Site
characterization also should provide information on the location, extent, and concentrations of
dissolved contamination; ground water geochemical data; geologic information on the type and
distribution of subsurface materials; and hydrogeologic parameters such as hydraulic conductivity,
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hydraulic gradients, and potential contaminant migration pathways to human or ecological receptor
exposure points.
The data collected during site characterization can be used to simulate the fate and transport of
contaminants in the subsurface. Such simulation allows prediction of the future extent and
concentrations of the dissolved contaminant plume. Several types of models can be used to simulate
dissolved contaminant transport and attenuation.
The natural attenuation modeling effort has five primary objectives:
• To evaluate whether M.NA will be likely to attain site-specific remediation objectives in a
time period that is reasonable compared to other alternatives;
* To predict the future extent and concentration of a dissolved contaminant plume by
simulating the combined effects of contaminant loading, advection, dispersion, sorption,
and biodegradation;
• To predict the most useful locations for ground-water monitoring;
• To assess the potential for downgradient receptors to be exposed to contaminant
concentrations that exceed regulatory or risk-based levels intended to be protective of
human health and the environment; and
To provide technical support for remedial options using MNA during screening and detailed
evaluation of remedial alternatives in a CERCLA Feasibility Study or RCRA Corrective
Measures Study.
Upon completion of the fate and transport modeling effort, model predictions can be used to
evaluate whether MNA is a viable remedial alternative for a given site. If the transport and fate
models predict that natural attenuation is sufficient to attain site-specific remediation objectives
and will be protective of human health and the environment, natural attenuation may be an
appropriate remedy for the site. Model assumptions and results should be verified by data obtained
from site characterization. If model assumptions and results are not verified by site data, MNA is
not likely to be a viable option and should not be proposed as the remedy.
1.5 MONITORING
The Monitoring Program OSWER Directive on Monitored Natural Attenuation (9200.4-17)
describes EPA expectations for performance monitoring.
Performance monitoring to evaluate remedy effectiveness and to ensure
protection of human health and the environment is a. critical, element of all response
actions. Performance monitor ing is ofe ven greater importance for monitored natural
attenuation than for other types of remedies due to the longer remediation time
frames, potential for ongoing contaminant migration, and, other uncertainties
associated with using monitored natural attenuation. This emphasis is underscored,
by EPA 's reference to "monitored natural attenuation".
The monitoring program developed for each site should specify the location,
frequency, and type of samples and measurements necessary to evaluate remedy
performance as we II as define the anticipated performance objectives of the remedy.
In addition, all monitoring programs should he designed, to accomplish the follow ing:
• Demonstrate that natural, attenuation is occurring according to expectations;
* Identify any potentially toxic transformation products resulting from,
biodegradation;
* Determine if a plume is expanding (either downgradient, laterally or vertically) ;
• Ensure no impact to downgradient receptors;
* Detect new releases of contaminants to the environment that could impact the
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effectiveness of the natural attenuation remedy;
* Demonstrate the efficacy of institutional controls that were put in place to
protect potential receptors;
* Detect changes in environmental conditions (e.g., hydrogeologic, geochemical,
microbiological, or other changes) that may reduce the efficacy of any of the
natural attenuation processes; and
* Verify attainment of cleanup objectives.
Detection of changes will depend on the proper siting and construction of
monitoring wells/points. Although the siting of monitoring wells is a concern for
any remediation technology, it is of even greater concern with monitored natural
attenuation because of the lack of engineering controls to control contaminant
migration.
Performance monitoring should continue as long as contamination
remains above required cleanup levels. Typically, monitoring is continued for a
specified period (e.g., one to three years) after cleanup levels have been achieved to
ensure that concentration levels are stable and remain below target levels. The
institutional and financial mechanisms for maintaining the monitoring program
should be clearly established, in the remedy decision or other site documents, as
appropriate.
Natural attenuation is achieved when naturally occurring attenuation mechanisms, such as
biodegradation, bring about a reduction in the total mass, toxicity, mobility, volume, or concentration
of a contaminant dissolved in ground water. In some cases, natural attenuation processes will be
capable of attaining site-specific remediation objectives in a time period that is reasonable compared
to other alternatives. However, at this time, the authors are not aware of any sites where natural
attenuation alone has succeeded in restoring ground water contaminated with chlorinated aliphatic
hydrocarbons to drinking water quality over the entire plume.
The material presented here was prepared through the joint effort between the Bioremediation
Research Team at the Subsurface Protection and Remediation Division of U.S. EPA's National
Risk Management Research Laboratory (NRMRL) in Ada, Oklahoma, and the U.S. Air Force
Center for Environmental Excellence, Technology Transfer Division, Brooks Air Force Base,
Texas, and Parsons Engineering Science, Inc. (Parsons ES). It is designed to facilitate proper
evaluation of remedial alternatives including natural attenuation at large chlorinated aliphatic
hydrocarbon-contaminated sites.
This information is the most current available at the time of this writing. The scientific
knowledge and experience with natural attenuation of chlorinated solvents is growing rapidly and
the authors expect that the process for evaluating natural attenuation of chlorinated solvents will
continue to evolve.
This document contains three sections, including this introduction. Section 2 presents the
protocol to be used to obtain scientific data to evaluate the natural attenuation option. Section 3
presents the references used in preparing this document. Appendix A describes the collection of
site characterization data necessary to evaluate natural attenuation, and provides soil and ground-
water sampling procedures and analytical protocols. Appendix B provides an in-depth discussion
of the destructive and nondestructive mechanisms of natural attenuation. Appendix C covers data
interpretation and pre-modeling calculations.
10
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2
EVALUATING NATURAL ATTENUATION
The primary objective of the natural attenuation investigation is to determine whether natural
processes will be capable of attaining site-specific remediation objectives in a time period that is
reasonable compared to other alternatives. Further, natural attenuation should be evaluated to
determine if it can meet all appropriate Federal and State remediation objectives for a given site.
This requires that projections of the potential extent of the contaminant plume in time and space be
made. These projections should be based on historic variations in contaminant concentration, and
the current extent and concentrations of contaminants in the plume in conjunction with measured
rates of contaminant attenuation. Because of the inherent uncertainty associated with such predictions,
it is the responsibility of the proponent of monitored natural attenuation to provide sufficient evidence
to demonstrate that the mechanisms of natural attenuation will meet the remediation objectives
appropriate for the site. This can be facilitated by using conservative parameters in solute fate and
transport models and numerous sensitivity analyses in order to better evaluate plausible contaminant
migration scenarios. When possible, both historical data and modeling should be used to provide
information that collectively and consistently confirms the natural reduction and removal of the
dissolved contaminant plume.
Figure 2.1 outlines the steps involved in a natural attenuation demonstration and shows the
important regulatory decision points for implementing natural attenuation. For example, a Superfund
Feasibility Study is a two-step process that involves initial screening of potential remedial alternatives
followed by more detailed evaluation of alternatives that pass the screening step. A similar process
is followed in a RCRA Corrective Measures Study and for sites regulated by State remediation
programs. The key steps for evaluating natural attenuation are outlined in Figure 2.1 and include:
1) Review available site data and develop a preliminary conceptual model. Determine if
receptor pathways have already been completed. Respond as appropriate.
2) If sufficient existing data of appropriate quality exist, apply the screening process de-
scribed in Section 2.2 to assess the potential for natural attenuation.
3) If preliminary site data suggest natural attenuation is potentially appropriate, perform
additional site characterization to further evaluate natural attenuation. If all the recom-
mended screening parameters listed in Section 2.2 have been collected and the screening
processes suggest that natural attenuation is not appropriate based on the potential for
natural attenuation, evaluate whether other processes can meet the cleanup objectives for
the site (e.g., abiotic degradation or transformation, volatilization, or sorption) or select a
remedial option other than MNA.
4) Refine conceptual model based on site characterization data, complete pre-modeling
calculations, and document indicators of natural attenuation.
5) Simulate, if necessary, natural attenuation using analytical or numerical solute fate and
transport models that allow incorporation of a biodegradation term.
6) Identify potential receptors and exposure points and conduct an exposure pathways analy-
sis.
7) Evaluate the need for supplemental source control measures. Additional source control
may allow MNA to be a viable remedial option or decrease the time needed for natural
processes to attain remedial objectives.
11
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Review Available Site Data
If Site Data are Adequate
Develop Preliminary Conceptual Model
•e-
Gather any Additional Data
Necessary to Complete
the Screening of the Site
Screen the Site using the Procedure
Presented in Figure 2,3
Are
Sufficient
Available to Properly
Screen the Site?
Are
Screening Criteria
Met?
Does it
Appear That
Natural Attenuation Alone
Will Meet Regulat
Criteria?
Engineered Remediation Required,
Implement Other Protocols
Perform Site Characterization
to Evaluate Natural Attenuation
Refine Conceptual Model and
Complete Pre-Modeling
Calculations
Simulate Natural Attenuation
Using Solute Fate and
Transport Models
•4-
3l ASSU
Verify yodel Assumptions and
Results with Site
Characterization Data
Use Results of Modeling and
Site-Specific Information in an
Exposure Pathways Analysis
Will Remediation
Objectives Be Met
Without Posing Unacceptable^
RisksTo Potential
Evaluate Use of
Selected Additional
Remedial Options
Including Source
Removal or Source
Control Along with
Natural Attenuation
Perform Site Characterization
to Support Remedy Decision Making
NO
Potential For
Natural Attenuation
With Remediation
System Installed
Refine Conceptual Model and
Complete Pre-Modeling
Calculations
Simulate Natural Attenuation
Combined with Remedial
Option Selected Above
Using Solute Transport Models
Verily Model Assumptions and
Results with Site
Characterization Data
Use Results of Modeling and
Site-Specific Information in
an Exposure Assessment
Does
Revised Remediation**
'Strategy Meet Remediation^
Objectives Without Posing
Unacceptable Risks
To Potential
hugr^ %l
YES '
Develop Draft Plan for
Performance Evaluation
Monitoring Wells and
Long-Term Monitoring
f
%
Determine Remedial
Measures to be •
Combined with MNA
Ł
^^irf
YES
Present Findings
and Proposed
Remedy in
Feasibility Study
Figure 2,1 Natural attenuation of chlorinated solvents flow chart.
12
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8) Prepare a long-term monitoring and verification plan for the selected alternative. In some
cases, this includes monitored natural attenuation alone, or in other cases in concert with
supplemental remediation systems.
9) Present findings of natural attenuation studies in an appropriate remedy selection docu-
ment, such as a CERCLA Feasibility or RCRA Corrective Measures Study. The appropri-
ate regulatory agencies should be consulted early in the remedy selection process to clarify
the remedial objectives that are appropriate for the site and any other requirements that the
remedy will be expected to meet. However, it should be noted that remedy requirements
are not finalized until a decision is signed, such as a CERCLA Record of Decision or a
RCRA Statement of Basis.
The following sections describe each of these steps in more detail.
2.1 REVIEW AVAILABLE DATA AND DEVELOP PRELIMINARY
CONCEPTUAL
The first step in the natural attenuation investigation is to review available site-specific data.
Once this is done, it is possible to use the initial site screening processes presented in Section 2.2 to
determine if natural attenuation is a viable remedial option. A thorough review of these data also
allows development of a preliminary conceptual model. The preliminary conceptual model will
help identify any shortcomings in the data and will facilitate placement of additional data collection
points in the most scientifically advantageous and cost-effective manner possible.
The following site information should be obtained during the review of available data.
Information that is not available for this initial review should be collected during subsequent site
investigations when refining the site conceptual model, as described in Section 2.3.
« Nature, extent, and magnitude of contamination:
- Nature and history of the contaminant release:
—Catastrophic or gradual release of NAPL ?
—More than one source area possible or present ?
—Divergent or coalescing plumes ?
- Three-dimensional distribution of dissolved contaminants and mobile and residual
NAPLs. Often high concentrations of chlorinated solvents in ground water are the result
of landfill leachates, rinse waters, or ruptures of water conveyance pipes. For LNAPLs
the distribution of mobile and residual NAPL will be used to define the dissolved plume
source area. For DNAPLs the distribution of the dissolved plume concentrations, in addition
to any DNAPL will be used to define the plume source area.
- Ground water and soil chemical data.
- Historical water quality data showing variations in contaminant concentrations both
vertically and horizontally.
- Chemical and physical characteristics of the contaminants.
- Potential forbiodegradation of the contaminants.
- Potential for natural attenuation to increase toxity and/or mobility of natural occurring
metals.
• Geologic and hydrogeologic data in three dimensions (If these data are not available, they
should be collected for the natural attenuation demonstration and for any other remedial
investigation or feasibility study):
- Lithology and stratigraphic relationships.
- Grain-size distribution (gravels vs. sand vs. silt vs. clay).
13
-------
Aquifer hydraulic conductivity (vertical and horizontal, effectiveness of aquitards,
calculation of vertical gradients).
Ground-water flow gradients and potentiometric or water table surface maps (over
several seasons, if possible).
- Preferential flow paths.
- Interactions between ground water and surface water and rates of infiltration/recharge.
» Locations of potential receptor exposure points:
- Ground water production and supply wells, and areas that can be deemed a potential source
of drinking water.
Downgradient and crossgradient discharge points including any discharges to surface waters
or other ecosystems.
Vapor discharge to basements and other confined spaces.
In some cases, site-specific data are limited. If this is the case, all future site characterization
activities should include collecting the data necessary to screen the site for the use of monitored
natural attenuation as a potential site remedy. Much of the data required to evaluate natural attenuation
can be used to design and evaluate other remedial measures.
Available site characterization data should be used to develop a conceptual model for the site.
This conceptual model is a three-dimensional representation of the source area as a NAPL or
region of highly contaminated ground water, of the surrounding uncontaminated area, of ground
water flow properties, and of the solute transport system based on available geological, biological,
geochemical, hydrological, climatological, and analytical data for the site. Data on the contaminant
levels and aquifer characteristics should be obtained from wells and boreholes which will provide
a clear three-dimensional picture of the hydrologic and geochemical characteristics of the site.
High concentrations of dissolved contaminants can be the result of leachates, rinse waters and
rupture of water conveyance lines, and are not necessarily associated with NAPLs.
This type of conceptual model differs from the conceptual site models commonly used by risk
assessors that qualitatively consider the location of contaminant sources, release mechanisms,
transport pathways, exposure points, and receptors. However, the conceptual model of the ground
water system facilitates identification of these risk-assessment elements for the exposure pathways
analysis. After development, the conceptual model can be used to help determine optimal placement
of additional data collection points, as necessary, to aid in the natural attenuation investigation and
to develop the solute fate and transport model. Contracting and management controls must be
flexible enough to allow for the potential for revisions to the conceptual model and thus the data
collection effort.
Successful conceptual model development involves:
» Definition of the problem to be solved (generally the three dimensional nature, magnitude,
and extent of existing and future contamination).
« Identification of the core or cores of the plume in three dimensions. The core or cores contain
the highest concentration of contaminants.
» Integration and presentation of available data, including:
- Local geologic and topographic maps,
- Geologic data,
- Hydraulic data,
- Biological data,
- Geochemical data, and
- Contaminant concentration and distribution data.
14
-------
• Determination of additional data requirements, including:
- Vertical profiling locations, boring locations and monitoring well spacing in three dimensions,
- A sampling and analysis plan (SAP), and
- Any data requirements listed in Section 2.1 that have not been adequately addressed.
Table 2.1 contains the recommended soil and ground water analytical methods for evaluating
the potential for natural attenuation of chlorinated aliphatic hydrocarbons and/or fuel hydrocarbons.
Any plan to collect additional ground water and soil quality data should include the analytes listed
in this table. Table 2.2 lists the availability of these analyses and the recommended data quality
requirements. Since required procedures for field sampling, analytical methods and data quality
objectives vary somewhat among regulatory programs, the methods to be used at a particular site
should be developed in collaboration with the appropriate regulatory agencies. There are many
documents which may aid in developing data quality objectives (e.g.,U.S. EPA Order 5360.1 and
U.S. EPA QA/G-4 Guidance for the Data Quality Objectives Process).
2.2 INITIAL
After reviewing available site data and developing a preliminary conceptual model, an
assessment of the potential for natural attenuation must be made. As stated previously, existing
data can be useful to determine if natural attenuation is capable of attaining site-specific remediation
objectives in a time period that is reasonable compared to other alternatives. This is achieved by
first determining whether the plume is currently stable or migrating and the future extent of the
plume based on (1) contaminant properties, including volatility, sorptive properties, and
biodegradability; (2) aquifer properties, including hydraulic gradient, hydraulic conductivity, porosity
and concentrations of native organic material in the sediment (TOC), and (3) the location of the
plume and contaminant source relative to potential receptor exposure points (i.e., the distance between
the leading edge of the plume and the potential receptor exposure points). These parameters
(estimated or actual) are used in this section to make a preliminary assessment of the effectiveness
of natural attenuation in reducing contaminant concentrations.
If, after completing the steps outlined in this section, it appears that natural attenuation will be
a significant factor in contaminant removal and a viable remedial alternative, detailed site
characterization activities that will allow evaluation of this remedial option should be performed.
If exposure pathways have already been completed and contaminant concentrations exceed protective
levels, or if such completion is likely, an engineered remedy is needed to prevent such exposures
and should be implemented as an early action. For this case, MNA may still be appropriate to attain
long-term remediation objectives for the site. Even so, the collection of data to evaluate natural
attenuation can be integrated into a comprehensive remedial strategy and may help reduce the cost
and duration of engineered remedial measures such as intensive source removal operations or pump-
and-treat technologies.
2.2.1 Overview of Chlorinated Aliphatic Hydrocarbon Blodegradation
Because biodegradation is usually the most important destructive process acting to reduce
contaminant concentrations in ground water, an accurate estimate of the potential for natural
biodegradation is important to consider when determining whether ground water contamination
presents a substantial threat to human health and the environment. This information also will be
useful when selecting the remedial alternative that will be most cost effective at eliminating or
abating these threats should natural attenuation alone not prove to be sufficient.
15
-------
Table 2.1 Soil. Soil Gas. and Ground-water Analytical Methods to Evaluate the Potential for Natural Attenuation of Chlorinated Solvents or Fuel
Hydrocarbons in Ground Water. Analyses other than those listed in this table may be required for regulatory compliance.
Matrix
Soil
Soil
Soil
Soil Gas
Soil Gas
Analysis
Aromatic and
Chlorinated
hydrocarbons
(benzene,
toluene,
ethylbenzene, and
xylene [BTEX];
Chlorinated
Compounds
Biologically
Available Iron
(III)
Total organic
carbon (TOC)
Fuel and
Chlorinated
VOCs
Methane,
Oxygen, Carbon
dioxide
Method/Reference
SW8260A
Under development
SW9060 modified for
soil samples
EPA Method TO- 14
Field Soil Gas
Analyzer
Comments
HCI extraction
followed by
quantification of
released iron (III)
Procedure must
be accurate over
the range of 0.1
to 5 percent TOC
Data Use
Data are used to
determine the extent of
soil contamination, the
contamination mass
present, and the
potential for source
removal.
Optional method that
should be used when
fuel hydrocarbons or
vinyl chloride are
present in the ground
water to predict the
possible extent of
removal of fuel
hydrocarbons and
vinyl chloride via iron
reduction.
The rate of migration
of petroleum
contaminants in
ground water is
dependent upon the
amount of TOC in the
aquifer matrix.
Useful for determining
chlorinated and BTEX
compounds in soil
Useful for determining
bioactivity in vadose
zone.
Recommended
Frequency of
Analysis
Each soil sampling
round
One round of
sampling in five
borings, five cores
from each boring
At initial sampling
At initial sampling
At initial sampling
and respiration
testing
Sample Volume, Sample
Container, Sample
Preservation
Sample volume
approximately 100 ml;
subsample and extract in
the field using methanol
or appropriate solvent;
cool to 4°C.
Minimum 1 inch
diameter core samples
collected into plastic
liner. Cap and prevent
aeration.
Collect 100 g of soil in a
glass container with
Teflon-lined cap; cool to
4°C.
1 -liter Summa Canister
3-liters in a Tedlar bag,
bags are reusable for
analysis of methane,
oxygen, or carbon
dioxide.
Field or
Fixed-Base
Laboratory
Fixed-base
Laboratory
Fixed-base
Fixed-base
Field
-------
Table 2.1 (Continued)
Matrix
Water
Water
Water
Water
Analysis
Alkalinity
Aromatic and
chlorinated
hydrocarbons
(BTEX,
trimethylbenzene
isomers,
chlorinated
compounds)
Arsenic
Chloride
(optional, see
data use)
Method/Reference
Hach Alkalinity test kit
model AL AP MG-L
SW8260A
EPA 200.7 or EPA
200.9
Hach Chloride test kit
model 8-P
Comments
Phenolphthalein
method
Analysis may be
extended to higher
molecular weight
alkyl benzenes
Silver nitrate
titration
Data Use
General water quality
parameter used (1) as a
marker to verify that all site
samples are obtained from
the same ground-water
system and (2) to measure
the buffering capacity of
ground water.
Method of analysis for
BTEX and chlorinated
solvents/byproducts, which
are the primary target
analytes for monitoring
natural attenuation; method
can be extended to higher
molecular weight alkyl
benzenes; trimethylben-
zenes are used to monitor
plume dilution if
degradation is primarily
anaerobic.
To determine if anaerobic
biological activity is
solubilizing arsenic from
the aquifer matrix material.
As above, and to guide
selection of additional data
points in real time while in
the field.
Recommended
Frequency of
Analysis
Each sampling
round
Each sampling
round
One round of
sampling
Each sampling
round
Sample Volume,
Sample Container,
Sample Preservation
Collect 100 mL of
water in glass container.
Collect water samples
in a 40 mL VGA vial;
cool to 4°C; add
hydrochloric acid to
pH2.
Collect 100 ml in a
glass or plastic
container that is rinsed
in the field with the
ground water to be
sampled. Unfiltered
samples obtained using
low flow sampling
methods are preferred
for analysis of dissolved
metals. Adjust pH to 2
with nitric acid. Do not
insert pH paper or an
electrode into the
sample.
Collect 100 mL of
water in a glass
container.
Field or
Fixed-Base
Laboratory
Field
Fixed-base
Laboratory
Field
-------
Table 2.1 (Continued)
Matrix
Water
Water
Water
Water
Water
Water
Analysis
Chloride
Chloride
(optional, see
data use)
Conductivity
Iron (II) (Fe+2)
Hydrogen (H2)
Manganese
Method/Reference
Mercuric nitrate
titration A4500-Cr C
Hach Chloride test kit
model 8-P
E120.1/SW9050, direct
reading meter
Colorimetric
Hach Method #8 146
Equilibration with gas
in the field.
Determined with a
reducing gas detector.
EPA 200.7 or EPA
200.9
Comments
Ion chromatography
(1C) method E300
or method SW9050
may also be used
Silver nitrate
titration
Filter if turbid.
Optional
specialized analysis
Data Use
General water quality
parameter used as a marker
to verify that site samples
are obtained from the same
ground-water system. Final
product of chlorinated
solvent reduction.
As above, and to guide
selection of additional data
points in real time while in
the field.
General water quality
parameter used as a marker
to verify that site samples
are obtained from the same
ground-water system.
May indicate an anaerobic
degradation process due to
depletion of oxygen,
nitrate, and manganese.
Determined terminal
electron accepting process.
Predicts the possiblity for
reductive dechlorination.
To determine if anaerobic
biological activity is
solubilizing manganese
from the aquifer matrix
material.
Recommended
Frequency of
Analysis
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
One round of
sampling on
selected wells.
One round of
sampling
Sample Volume,
Sample Container,
Sample Preservation
Collect 250 mL of
water in a glass
container.
Collect 100 mL of
water in a glass
container.
Collect 100 to 250 mL
of water in a glass or
plastic container.
Collect from a flow-
through or over-flow
cell / analyze at the well
head.
Sampled at well head
requires the production
of 300 mL per minute
of water for 30 minutes.
Collect 100 ml in a
glass or plastic
container that is rinsed
in the field with the
ground water to be
sampled. Unfiltered
samples obtained using
low flow sampling
methods are preferred
for analysis of dissolved
metals. Adjust pH to 2
with nitric acid. Do not
insert pH paper or an
electrode into the
sample.
Field or
Fixed-Base
Laboratory
Fixed-base
Field
Field
Field
Field
Laboratory
oo
-------
Table 2.1 (Continued)
Matrix
Water
Water
Water
Water
Water
Analysis
Methane, ethane,
and ethene
Nitrate
Oxidation-
reduction
potential
Oxygen
PH
Method/Reference
Kampbelle?a/.,1989
andl998orSW3810
Modified
1C method E300
A2580B
Dissolved oxygen meter
calibrated between each
well according to the
supplier's specifications
Field probe with direct
reading meter calibrated
in the field according to
the supplier's
specifications.
Comments
Method published
by researchers at the
U.S. Environmental
Protection Agency.
Limited to few
commercial labs.
Measurements made
with electrodes;
results are displayed
on a meter; protect
samples from
exposure to oxygen.
Report results
against a
silver/silver chloride
reference electrode.
(Eh) is calculated by
adding a correction
factor specific to the
electrode used.
Refer to
method A4500
for a comparable
laboratory
procedure.
Field
Data Use
The presence of CH4
suggests BTEX degradation
via methanogenesis.
Ethane and ethene data are
used where chlorinated
solvents are suspected of
undergoing biological
transformation.
Substrate for microbial
respiration if oxygen is
depleted.
The ORP of ground water
influences and is influenced
by the nature of the
biologically mediated
degradation of
contaminants; the ORP
(expressed as Eh) of
ground water may range
from more than 800 mV to
less than -400 mV.
The oxygen concentration
is a data input to the
Bioplume model;
concentrations less than
1 mg/L generally indicate
an anaerobic pathway.
Aerobic and anaerobic
biological processes are
pH- sensitive.
Recommended
Frequency of
Analysis
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
Sample Volume,
Sample Container,
Sample Preservation
Collect water samples
in 50 mL glass serum
bottles with gray butyl
/Teflon-faced septa and
crimp caps; add H2SO4
to pH less than 2, cool
to 4°C.
Collect up to 40 mL of
water in a glass or
plastic container; add
H2SO4 to pH less than
2, cool to 4°C.
Measure in a flow
through cell or an over-
flowing container filled
from the bottom to
prevent exposure of the
ground water to the
atmosphere.
Measure dissolved
oxygen on site using a
flow-through cell or
over-flow cell.
Measure dissolved
oxygen on site using a
flow-through cell or
over-flow cell.
Field or
Fixed-Base
Laboratory
Fixed-base
Fixed-base
Field
Field
Field
-------
Table 2.1 (Continued)
Matrix
Water
Water
Water
Water
Analysis
Sulfate (SO4-2)
Sulfate (SO4-2)
Temperature
Total Organic
Carbon also
called DOC
Method/Reference
1C method E300
Hach method #8051
Field probe with direct
reading meter.
SW9060
Comments
If this method is
used for sulfate
analysis, do not use
the field method.
Colorimetric, if this
method is used for
sulfate analysis, do
not use the fixed-
base laboratory
method.
Field only
Laboratory
Data Use
Substrate for anaerobic
microbial respiration.
Same as above.
To determine if a well is
adequately purged for
sampling.
Used to classify plume and
to determine if reductive
dechlorination is possible
in the absence of
anthropogenic carbon.
Recommended
Frequency of
Analysis
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
Sample Volume,
Sample Container,
Sample Preservation
Collect up to 40 mL of
water in a glass or
plastic container; cool
to 4°C.
Collect up to 40 mL of
water in a glass or
plastic container; cool
to 4°C.
Read from oxygen
meter.
Measure using a flow-
through cell or over-
flow cell.
Field or
Fixed-Base
Laboratory
Fixed-base
Field
Field
Laboratory
O
NOTES:
1. "Hach" refers to the Hach Company catalog, 1990.
2. "A" refers to Standard Methods for the Examination of Water and Wastewater, 18th edition, 1992.
3. "E" refers to Methods for Chemical Analysis of Water and Wastes, U.S. EPA, 1983.
4. "SW" refers to the Test Methods for Evaluating Solid Waste, Physical, and Chemical Methods, SW-846, U.S. EPA, 3rd edition, 1986.
-------
Table 2.2 Objectives for Sensitivity and Precision to Implement the Natural Attenuation Protocol.
required for regulatory compliance.
Analyses other than those listed in this table may be
Matrix
Soil
Soil
Soil
Soil Gas
Soil Gas
Water
Water
Water
Water
Water
Analysis
Aromatic and
chlorinated
hydrocarbons
(benzene,
toluene,
ethylbenzene, and
xylene [BTEX];
chlorinated
compounds)
Biologically
Available Iron
(HI)
Total organic
carbon (TOC)
Fuel and
Chlorinated
VOCs
Methane, O2, CO2
Alkalinity
Aromatic and
chlorinated
hydrocarbons
(BTEX,
trimethylbenzene
isomers,
chlorinated
compounds)
Chloride
Chloride
(optional, see
data use)
Conductivity
Method/Reference
SW8260A
Under development
SW9060 modified for
soil samples
EPA Method TO- 14
Field Soil Gas Analyzer
Hach alkalinity test kit
model AL AP MG-L
SW8260A
1C method E300
Hach Chloride test kit
model 8-P
E120.1/SW9050, direct
reading meter
Minimum Limit of
Quantification
1 mg/Kg
50 mg/Kg
0.1 percent
1 ppm
(volume/volume)
1 percent
(volume/volume)
50mg/L
MCLs
Img/L
1 mg/L
50 laS/cm2
Precision
Coefficient of Variation of
20 percent.
Coefficient of Variation of
40 percent.
Coefficient of Variation of
20 percent.
Coefficient of Variation of
20 percent.
Coefficient of Variation of
20 percent.
Standard deviation of 20
mg/L.
Coefficient of Variation of
10 percent.
Coefficient of Variation of
20 percent.
Coefficient of Variation of
20 percent.
Standard deviation of 50
pS/cm2.
Availability
Common laboratory
analysis.
Specialized laboratory
analysis.
Common laboratory
analysis.
Common laboratory
analysis.
Readily available field
instrument.
Common field analysis.
Common laboratory
analysis.
Common laboratory
analysis.
Common field analysis.
Common field probe.
Potential Data Quality
Problems
Volatiles lost during shipment
to laboratory; prefer extraction
in the field.
Sample must not be allowed
to oxidize.
Samples must be collected
from contaminant-
transporting (i.e.,
transmissive) intervals.
Potential for atmospheric
dilution during sampling.
Instrument must be properly
calibrated.
Analyze sample within 1 hour
of collection.
Volatilization during shipment
and biodegradation due to
improper preservation.
—
Possible interference from
turbidity.
Improperly calibrated
instrument.
-------
Table 2.2 (Continued)
Matrix
Water
Water
Water
Water
Water
Water
Water
Water
Water
Water
Water
Water
Analysis
Hydrogen (H2)a/
Iron (II) (Fe2+)
XX
Major Cations
Methane, ethane,
and ethene
Nitrate
Oxidation-
reduction
potential (ORP)
Oxygen
Sulfate (SO42-)
Sulfate (SO42-)
XX
pH
Temperature
Total Organic
Carbon
Method/Reference
See Appendix A
Colorimetric
Hach Method #8 146
SW6010
Kampbelle/a/., 1989 or
SW3810 Modified
1C method E300
A2580B
Dissolved oxygen meter
1C method E300
Hach method # 8051
Field probe with direct
reading meter.
Field probe with direct
reading meter.
SW9060
Minimum Limit of
Quantification
0.1 nM
0.5 mg/L
1 mg/L
lUg/L
0.1 mg/L
plus or minus
300 mV
0.2 mg/L
5 mg/L
5 mg/L
0.1 standard units
0 degrees Celsius
0.1 mg/L
Precision
Standard deviation of
O.lnM.
Coefficient of Variation of
20 percent.
Coefficient of Variation of
20 percent.
Coefficient of Variation of
20 percent.
Standard deviation of 0.1
mg/L
plus or minus 50 mV.
Standard deviation of 0.2
mg/L.
Coefficient of Variation of
20 percent.
Coefficient of Variation of
20 percent.
0.1 standard units.
Standard deviation of 1
degrees Celsius.
Coefficient of Variation of
20 percent.
Availability
Specialized field
analysis.
Common field analysis.
Common laboratory
analysis.
Specialized laboratory
analysis.
Common laboratory
analysis.
Common field probe.
Common field
instrument.
Common laboratory.
Common field analysis.
Common field meter.
Common field probe.
Common laboratory
analysis.
Potential Data Quality
Problems
Numerous, see Appendix A.
Possible interference from
turbidity (must filter if turbid).
Keep out of sunlight and
analyze within minutes of
collection.
Possible colloidal
interferences.
Sample must be preserved
against biodegradation and
collected without headspace
(to minimize volatilization).
Must be preserved.
Improperly calibrated
electrodes or introduction of
atmospheric oxygen during
sampling.
Improperly calibrated
electrodes or bubbles behind
the membrane or a fouled
membrane or introduction of
atmospheric oxygen during
sampling.
Fixed-base.
Possible interference from
turbidity (must filter if turbid).
Keep sample cool.
Improperly calibrated
instrument; time sensitive.
Improperly calibrated
instrument; time sensitive.
Notes:
** Filter if turbidity gives a response from the photometer before addition of the reagents that is as large or larger than the specified minimum quantification limit.
-------
Over the past two decades, numerous laboratory and field studies have demonstrated that
subsurface microorganisms can degrade a variety of chlorinated solvents (e.g., Bouwer era/., 1981;
Miller and Guengerich, 1982; Wilson and Wilson, 1985; Nelson etal., 1986; Bouwer and Wright,
1988; Lee, 1988; Little et a/., 1988; Mayer et a/., 1988; Arciero et al., 1989; Cline and Delfino,
1989; Freedman and Gossett, 1989; Folsom et al., 1990; Marker and Kim, 1990; Alvarez-Cohen
and McCarty, 1991 a, 1991b; DeStefanoera/., 1991; Henry, 1991; McCarty era/., 1992; Hartmans
and de Bont, 1992; McCarty and Semprini, 1994; Vogel, 1994). Whereas fuel hydrocarbons are
biodegraded through use as a primary substrate (electron donor), chlorinated aliphatic hydrocarbons
may undergo biodegradation under three different circumstances: intentional use as an electron
acceptor; intentional use as an electron donor; or, through cometabolism where degradation of the
chlorinated organic is fortuitous and there is no benefit to the microorganism. At a given site, one
or all of these circumstances may pertain, although at many sites the use of chlorinated aliphatic
hydrocarbons as electron acceptors appears to be most important under natural conditions. In this
case, biodegradation of chlorinated aliphatic hydrocarbons will be an electron-donor-limited process.
Conversely, biodegradation of fuel hydrocarbons is an electron-acceptor-limited process.
In an uncontaminated aquifer, native organic carbon is used as an electron donor, and dissolved
oxygen (DO) is used first as the prime electron acceptor. Where anthropogenic carbon (e.g., as fuel
hydrocarbons) is present, it also will be used as an electron donor. After the DO is consumed,
anaerobic microorganisms typically use additional electron acceptors (as available) in the following
order of preference: nitrate, ferric iron oxyhydroxide, sulfate, and finally carbon dioxide. Evaluation
of the distribution of these electron acceptors can provide evidence of where and how chlorinated
aliphatic hydrocarbon biodegradation is occurring. In addition, because chlorinated aliphatic
hydrocarbons may be used as electron acceptors or electron donors (in competition with other
acceptors or donors), isopleth maps showing the distribution of these compounds and their daughter
products can provide evidence of the mechanisms of biodegradation working at a site. As with
BTEX, the driving force behind oxidation-reduction reactions resulting in chlorinated aliphatic
hydrocarbon degradation is electron transfer. Although thermodynamically favorable, most of the
reactions involved in chlorinated aliphatic hydrocarbon reduction and oxidation do not proceed
abiotically. Microorganisms are capable of carrying out the reactions, but they will facilitate only
those oxidation-reduction reactions that have a net yield of energy.
2.2.1.1 Mechanisms of Chlorinated Aliphatic Hydrocarbon Biodegradation
The following sections describe the biodegradation of those compounds that are most prevalent
and whose behavior is best understood.
2.2.1.1.1 Electron Acceptor Reactions (Reductive Dehalogenation)
The most important process for the natural biodegradation of the more highly chlorinated
solvents is reductive dechlorination. During this process, the chlorinated hydrocarbon is used as an
electron acceptor, not as a source of carbon, and a chlorine atom is removed and replaced with a
hydrogen atom. Figure 2.2 illustrates the transformation of chlorinated ethenes via reductive
dechlorination. In general, reductive dechlorination occurs by sequential dechlorination from PCE
to TCE to DCE to VC to ethene. Depending upon environmental conditions, this sequence may be
interrupted, with other processes then acting upon the products. During reductive dechlorination,
all three isomers of DCE can theoretically be produced. However, Bouwer (1994) reports that
under the influence of biodegradation, cis-l ,2-DCE is a more common intermediate than trans-1,2-
DCE, and that 1,1-DCE is the least prevalent of the three DCE isomers when they are present as
daughter products. Reductive dechlorination of chlorinated solvent compounds is associated with
23
-------
1,1-DCE
TCE
cis -1,2,
.c\) Chlorine Atom
—x
cj Carbon Atom
—x
Hj Hydrogen Atom
Chemical
Bond
Double Chemical
Bond
trans-1,2- DCE
Complete
H) ( H) Co) (o*
Figure 2.2 Reductive dehalogencttion of chlorinated ethenes.
24
-------
the accumulation of daughter products and an increase in the concentration of chloride ions. Reductive
dechlorination affects each of the chlorinated ethenes differently. Of these compounds, PCE is the
most susceptible to reductive dechlorination because it is the most oxidized. Conversely, VC is the
least susceptible to reductive dechlorination because it is the least oxidized of these compounds.
As a result, the rate of reductive dechlorination decreases as the degree of chlorination decreases
(Vogel and McCarty, 1985; Bouwer, 1994). Murray and Richardson (1993) have postulated that
this rate decrease may explain the accumulation of VC in PCE and TCE plumes that are undergoing
reductive dechlorination. Reductive dechlorination has been demonstrated under nitrate- and iron-
reducing conditions, but the most rapid biodegradation rates, affecting the widest range of chlorinated
aliphatic hydrocarbons, occur under sulfate-reducing and methanogenic conditions (Bouwer, 1994).
Because chlorinated aliphatic hydrocarbon compounds are used as electron acceptors during reductive
dechlorination, there must be an appropriate source of carbon for microbial growth in order for this
process to occur (Bouwer, 1994). Potential carbon sources include natural organic matter, fuel
hydrocarbons, or other anthropogenic organic compounds such as those found in landfill leachate.
2.2,1.1.2 Electron Donor Reactions
Murray and Richardson (1993) write that microorganisms are generally believed to be incapable
of growth using PCE and TCE as a primary substrate (i.e., electron donor). However, under aerobic
and some anaerobic conditions, the less oxidized chlorinated aliphatic hydrocarbons (e.g., VC) can
be used as the primary substrate in biologically mediated oxidation-reduction reactions (McCarty
and Semprini, 1994). In this type of reaction, the facilitating microorganism obtains energy and
organic carbon from the degraded chlorinated aliphatic hydrocarbon. In contrast to reactions in
which the chlorinated aliphatic hydrocarbon is used as an electron acceptor, only the least oxidized
chlorinated aliphatic hydrocarbons can be used as electron donors in biologically mediated oxidation-
reduction reactions. McCarty and Semprini (1994) describe investigations in which VC and 1,2-
dichloroethane (DCA) were shown to serve as primary substrates under aerobic conditions. These
authors also document that dichloromethane has the potential to function as a primary substrate
under either aerobic or anaerobic environments. In addition, Bradley and Chapelle (1996) show
evidence of mineralization of VC under iron-reducing conditions so long as there is sufficient
bioavailable iron (III). Aerobic metabolism of VC may be characterized by a loss of VC mass and
a decreasing molar ratio of VC to other chlorinated aliphatic hydrocarbon compounds. In addition,
Klier et al ( 1998) and Bradley and Chapelle (1997) show mineralization of DCE to carbon dioxide
under aerobic, Fe(III) reducing, and methanogenic conditions, respectively.
2.2.1.1.3 Cometabolism
When a chlorinated aliphatic hydrocarbon is biodegraded via cometabolism, the degradation
is catalyzed by an enzyme or cofactor that is fortuitously produced by the organisms for other
purposes. The organism receives no known benefit from the degradation of the chlorinated aliphatic
hydrocarbon. Rather, the cometabolic degradation of the chlorinated aliphatic hydrocarbon may in
fact be harmful to the microorganism responsible for the production of the enzyme or cofactor
(McCarty and Semprini, 1994). Cometabolism is best documented in aerobic environments, although
it potentially could occur under anaerobic conditions. It has been reported that under aerobic
conditions chlorinated ethenes, with the exception of PCE, are susceptible to cometabolic degradation
(Murray and Richardson, 1993; Vogel, 1994; McCarty and Semprini, 1994). Vogel (1994) further
elaborates that the rate of cometabolism increases as the degree of dechlorination decreases. During
cometabolism, the chlorinated alkene is indirectly transformed by bacteria as they use BTEX or
25
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another substrate to meet their energy requirements. Therefore, the chlorinated alkene does not
enhance the degradation of BTEX or other carbon sources, nor will its cometabolism interfere with
the use of electron acceptors involved in the oxidation of those carbon sources.
2.2.1.2 Behavior of Chlorinated Solvent Plumes
Chlorinated solvent plumes can exhibit three types of behavior depending on the amount of
solvent, the amount of biologically available organic carbon in the aquifer, the distribution and
concentration of natural electron acceptors, and the types of electron acceptors being used. Individual
plumes may exhibit all three types of behavior in different portions of the plume. The different
types of plume behavior are summarized below.
2.2.1.2.1 Type 1 Behavior
Type 1 behavior occurs where the primary substrate is anthropogenic carbon (e.g., BTEX or
landfill leachate), and microbial degradation of this anthropogenic carbon drives reductive
dechlorination. When evaluating natural attenuation of a plume exhibiting Type 1 behavior, the
following questions must be answered:
1) Is the electron donor supply adequate to allow microbial reduction of the chlorinated
organic compounds? In other words, will the microorganisms "strangle" before they
"starve" (i.e., will they run out of chlorinated aliphatic hydrocarbons used as electron
acceptors before they run out of anthropogenic carbon used as the primary substrate)?
2) What is the role of competing electron acceptors (e.g., dissolved oxygen, nitrate, iron (III)
and sulfate)?
3) Is VC oxidized, or is it reduced?
Appendices B and C discuss what these questions mean and how they are answered. Type 1
behavior results in the rapid and extensive degradation of the more highly-chlorinated solvents
such as PCE, TCE, and DCE.
2.2,1.2.2 Type 2 Behavior
Type 2 behavior dominates in areas that are characterized by relatively high concentrations of
biologically available native organic carbon. Microbial utilization of this natural carbon source
drives reductive dechlorination (i.e., it is the primary substrate for microorganism growth). When
evaluating natural attenuation of a Type 2 chlorinated solvent plume, the same questions as those
posed in the description of Type 1 behavior must be answered. Type 2 behavior generally results in
slower biodegradation of the highly chlorinated solvents than Type 1 behavior, but under the right
conditions (e.g., areas with high natural organic carbon contents), this type of behavior also can
result in rapid degradation of these compounds.
2.2.1.2.3 Type 3 Behavior-
Type 3 behavior dominates in areas that are characterized by inadequate concentrations of
native and/or anthropogenic carbon, and concentrations of dissolved oxygen that are greater than
l.Omg/L. Under these aerobic conditions, reductive dechlorination will not occur. The most
significant natural attenuation mechanisms for PCE, TCE, and DCE will be advection, dispersion,
and sorption. However, VC can be rapidly oxidized under these conditions. Type 3 behavior also
occurs in ground water that does not contain microbes capable of biodegradation of chlorinated
solvents.
26
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2.2.1.2.4 Mixed Behavior
As mentioned above, a single chlorinated solvent plume can exhibit all three types of behavior
in different portions of the plume. This can be beneficial for natural biodegradation of chlorinated
aliphatic hydrocarbon plumes. For example, Wiedemeier et al. (1996a) describe a plume at
Plattsburgh AFB, New York, that exhibits Type 1 behavior in the source area and Type 3 behavior
downgradient from the source. The most fortuitous scenario involves a plume in which PCE, TCE,
and DCE are reductively dechlorinated with accumulation of VC near the source area (Type 1 or
Type 2 behavior), then VC is oxidized (Type 3 behavior), either aerobically or via iron reduction
further downgradient. Vinyl chloride is oxidized to carbon dioxide in this type of plume and does
not accumulate. The following sequence of reactions occurs in a plume that exhibits this type of
mixed behavior.
PCE->TCE-»DCE->VC->Carbon Dioxide
In general, TCE, DCE, and VC may attenuate at approximately the same rate, and thus these
reactions may be confused with simple dilution. Note that no ethene is produced during this reaction.
Vinyl chloride is removed from the system much faster under these conditions than it is under VC-
reducing conditions.
A less desirable scenario, but one in which all contaminants may be entirely biodegraded,
involves a plume in which all chlorinated aliphatic hydrocarbons are reductively dechlorinated via
Type 1 or Type 2 behavior. Vinyl chloride is reduced to ethene, which may be further reduced to
ethane or methane. The following sequence of reactions occurs in this type of plume.
PCE^ TCE-» DCE->VC->Ethene->Ethane
This sequence has been investigated by Freedman and Gossett (1989). In this type of plume,
VC degrades more slowly than TCE, and thus tends to accumulate.
2.2.2 Bioattenuation Screening Process
An accurate assessment of the potential for natural biodegradation of chlorinated compounds
should be made before investing in a detailed study of natural attenuation. The screening process
presented in this section is outlined in Figure 2.3. This approach should allow the investigator to
determine if natural bioattenuation of PCE, TCE, DCE, TCA, and chlorobenzenes is likely to be a
viable remedial alternative before additional time and money are expended. If the site is regulated
under CERCLA, much of the data required to make the preliminary assessment of natural attenuation
will be used to evaluate alternative engineered remedial solutions as required by the NCR Table 2.3
presents the analytical screening criteria.
For most of the chlorinated solvents, the initial biotransformation in the environment is a
reductive dechlorination. The initial screening process is designed to recognize geochemical
environments where reductive dechlorination is plausible. It is recognized, however, that
bioodegradation of certain halogenated compounds can also proceed via oxidative pathways.
Examples include DCE, VC, the dichloroethanes, chloroethane, dichlorobenzenes,
monochlorobenzene, methylene chloride, and ethylene dibromide.
The following information is required for the screening process:
« The chemical and geochemical data presented in Table 2.3 for background and target
areas of the plume as depicted in Figure 2.4. Figure 2.4 shows the schematic locations of
these data collection points. Note: If other contaminants are suspected, then data on the
concentrations and distribution of these compounds also should be obtained.
• Locations of source(s) and potential points of exposure. If subsurface NAPLs are
sources, estimate extent of residual and free-phase NAPL.
• An estimate of the transport velocity and direction of ground-water flow.
27
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Analyze Available Site Data
Along Core of Plume
to Determine if Biodegradation
is Occurring
Collect More Screening Data
Are
Sufficient Data
Available ?
Biodegradation
Occurring?
Engineered
Remediation Required,
Implement Other
Protocols
Locate source(s)and
potential points of
exposure. Estimate
extent of NPAL,
residual and free-phase
Determine Groundwater Flow and
Solute Transport Parameters Along
Core of Plume using
Site-Specific Data; Porosity and
Dispersivity May be Estimated
Estimate Biodegradation
Rate Constant
Compare the Rate of Transport
to the Rate of Attenuation using
Analytical Solute Transport Model
Does it
Appear that Natural
Attenuation Alone will Meet
Regulatory Criteria?
Evaluate use of Selected
Additional Remedial Options
along with Natural Attenuation
Proceed to
Figure 2.1
Perform Site Characterization
to Evaluate Natural Attenuation
>
Proceed to
Figure 2.1
f
Figure 2.3 Initial screening process flow chart.
28
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Table 2.3 Analytical Parameters and Weighting for Preliminary Screening for Anaerobic
Biodegradation Processes'1''
Analysis
Oxygen*
Oxygen*
Nitrate*
Iron II*
Sulfate*
Sulfide*
Methane*
Oxidation Reduction
Potential* (ORP)
against Ag/AgCI
electrode
pH*
roc
Temperature*
Carbon Dioxide
Alkalinity
Chloride*
Hydrogen
Hydrogen
Volatile Fatty Acids
BTEX*
Fetrachloroethene
Trichloroethene*
DCE*
VC*
1 , 1 , 1-Trichloroethane*
DCA
Carbon Tetrachloride
Chloroethane*
Ethene/Ethane
Chloroform
Dichloromethane
Concentration in
Most Contaminated
Zone
<0.5 mg/L
>5 mg/L
<1 mg/L
>1 mg/L
<20 mg/L
>1 mg/L
<0.5 mg/L
>0.5 mg/L
<50 millivolts (mV)
<-100mV
5 < pH < 9
5 > pH >9
> 20 mg/L
>20°C
>2x background
>2x background
>2x background
>1 nM
<1 nM
> 0.1 mg/L
> 0.1 mg/L
>0.01mg/L
>0.1 mg/L
Interpretation
Tolerated, suppresses the reductive pathway at higher
concentrations
Not tolerated; however, VC may be oxidized aerobically
At higher concentrations may compete with reductive pathway
Reductive pathway possible; VC may be oxidized under Fe(lll)-
reducing conditions
At higher concentrations may compete with reductive pathway
Reductive pathway possible
VC oxidizes
Ultimate reductive daughter product, VC Accumulates
Reductive pathway possible
Reductive pathway likely
Optimal range for reductive pathway
Outside optimal range for reductive pathway
Carbon and energy source; drives dechlorination; can be
natural or anthropogenic
At T >20°C biochemical process is accelerated
Ultimate oxidative daughter product
Results from interaction between CO2 and aquifer minerals
Daughter product of organic chlorine
Reductive pathway possible, VC may accumulate
VC oxidized
ntermediates resulting from biodegradation of more complex
compounds; carbon and energy source
Carbon and energy source; drives dechlorination
Material released
Material released
Daughter product of PCE
Material released
Daughter product of TCE
If cis is > 80% of total DCE it is likely a daughter product
1,1 -DCE can be chemical reaction product of TCA
Material released
Daughter product of DCE
Material released
Daughter product of TCA under reducing conditions
Material released
Daughter product of DCA or VC under reducing conditions
Daughter product of VC/ethene
Material released
Daughter product of Carbon Tetrachloride
Material released
Daughter product of Chloroform
Value
3
-3
2
3
2
3
0
3
1
2
0
-2
2
1
1
1
2
3
0
2
2
0
0
^a/
0
2a/
0
^a/
0
2
0
2
2
3
0
2
0
2
* Required analysis, a/ Points awarded only if it can be shown that the compound is a daughter product (i.e., not a constituent of the source
NAPL).
29
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Once these data have been collected, the screening process can be undertaken. The following
steps summarize the screening processes:
1) Determine if biodegradation is occurring using geochemical data. If biodegradation is
occurring, proceed to step 2. If it is not, assess the amount and types of data available. If
data are insufficient to determine if biodegradation is occurring, collect supplemental data.
If all the recommended screening parameters listed in section 2.2 have been collected and
the screening processes suggest that natural attenuation is not appropriate, the screening
processes are finished. Perform site characterization to evaluate other remediation alterna-
tives.
2) Determine ground-water flow and solute transport parameters from representative field
data. Dispersivity and porosity may be estimated from literature but the hydraulic conduc-
tivity and the ground-water gradient and flow direction must be determined from field
data. The investigator should use the highest valid hydraulic conductivity measured at the
site during the preliminary screening because solute plumes tend to follow the path of
least resistance (i.e., highest hydraulic conductivity). This will give the "worst-case"
estimate of the solute migration distance over a given period of time. Compare this
"worst-case" estimate with the rate of plume migration determined from site characteriza-
tion data. Determine what degree of plume migration is accepable or unacceptable with
respect to site-specific remediation objectives.
3) Locate source(s) and potential points of exposure. If subsurface NAPLs are sources,
estimate extent of residual and free-phase NAPL.
4) Estimate the biodegradation rate constant. Biodegradation rate constants can be estimated
using a conservative tracer found commingled with the contaminant plume, as described
in Appendix C and by Wiedemeier et al. (1996b). When dealing with a plume that con-
tains chlorinated solvents, this procedure can be modified to use chloride as a tracer. Rate
constants derived from microcosm studies can also be used when site specific field data
are inadequate or inconclusive. If it is not possible to estimate the biodegradation rate
using these procedures, then use a range of accepted literature values for biodegradation of
the contaminants of concern. Appendix C presents a range of biodegradation rate con-
stants for various compounds. Although literature values may be used to estimate
biogradation rates in the bioattenuation screening process described in Section 2.2, litera-
ture values should not be used in the later more detailed analysis of natural attenuation,
described in Section 2.3.
5) Compare the rate of transport to the rate of attenuation.
Use analytical solutions or a screening model such as BIO SCREEN.
6) Determine if screening criteria are met.
Step 1: Determine if Biodegradation is Occurring
The first step in the screening process is to sample or use existing data for the areas represented
in Figure 2.4 and analyze them for the parameters listed in Table 2.3 (see also Section 2.3.2). These
areas should include (1) the most contaminated portion of the aquifer (generally in the "source"
area with NAPL or high concentrations of contaminants in ground water ; (2) downgradient from
the source area but still in the dissolved contaminant plume; (3) downgradient from the dissolved
contaminant plume; and (4) upgradient and lateral locations that are not impacted by the plume.
Although this figure is a simplified two-dimensional representation of the features of a contaminant
plume, real plumes are three-dimensional objects. The sampling should be conducted in accordance
with Appendix A.
30
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Dissolved Contaminant Plume
Source Area
\
O
O
Direction of Plume Migration
>-
O Representative Sampling Location
Figure 2.4 Target areas for collecting screening data. Note that the number and location of monitoring
wells will vary with the three dimensional complexity of the plume (s).
The sample collected in the NAPL source area provides information as to the predominant
terminal electron-accepting process at the source area. In conjunction with the sample collected in
the NAPL source zone, samples collected in the dissolved plume downgradient from the NAPL
source zone allow the investigator (1) to determine if the plume is degrading with distance along
the flow path and (2) to determine the distribution of electron acceptors and donors and metabolic
by-products along the flow path. The sample collected downgradient from the dissolved plume
aids in plume delineation and allows the investigator to determine if metabolic byproducts are
present in an area of ground water that has been remediated. The upgradient and lateral samples
allow delineation of the plume and determination of background concentrations of the electron
acceptors and donors.
After these samples have been analyzed for the parameters listed in Table 2.3, the investigator
should analyze the data to determine if biodegradation is occurring. The right-hand column of
Table 2.3 contains scoring values that can be used as a test to assess the likelihood that biodegradation
is occurring. This method relies on the fact that biodegradation will cause predictable changes in
ground water chemistry. For example, if the dissolved oxygen concentration in the area of the
plume with the highest contaminant concentration is less than 0.5 milligrams per liter (mg/L), 3
points are awarded. Table 2.4 summarizes the range of possible scores and gives an interpretation
for each score. If the score totals 15 or more points, it is likely that biodegradation is occurring, and
the investigator should proceed to Step 2.
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Table 2.4 Interpretation of Points A\v circled During Screening Step 1
Score Interpretation
0 to 5 Inadequate evidence for anaerobic biodegradation* of chlorinated organics
6 to 14 Limited evidence for anaerobic biodegradation* of chlorinated organics
15 to 20 Adequate evidence for anaerobic biodegradation* of chlorinated organics
> 20 Strong evidence for anaerobic biodegradation* of chlorinated organics
*reduclive dechlorinalion
The following two examples illustrate how Step 1 of the screening process is implemented.
The site used in the first example is a former fire training area contaminated with chlorinated
solvents mixed with fuel hydrocarbons. The presence of the fuel hydrocarbons appears to reduce
the ORP of the ground water to the extent that reductive dechlorination is favorable. The second
example contains data from a dry cleaning site contaminated only with chlorinated solvents. This
site was contaminated with spent cleaning solvents that were dumped into a shallow dry well situated
just above a well-oxygenated, unconfmed aquifer with low organic carbon concentrations of dissolved
organic carbon.
Example 1: Strong Evidence for Anaerobic Biodegradation (Reductive Dechlorination) of
Chlorinated Organics
Analyte Concentration in Most Contaminated Zone Points Awarded
Dissolved Oxygen
Nitrate
Iron (11)
Sulfate
Methane
ORP
Chloride
PCE (released)
TCE (none released)
cis-DCE (none released)
VC (none released)
O.lmg/L
0.3 mg/L
lOmg/L
2 mg/L
5 mg/L
-190mV
3 times background
1,000 ng/L
1,200 ng/L
500 ug/L
50 ug/L
Total Points Awarded
3
2
3
2
3
2
2
0
2
2
2
23 Points
In this example, the investigator can infer that biodegradation is likely occurring at the time of
sampling and may proceed to Step 2.
Example 2: Anaerobic Biodegradation (Reductive Dechlorination) Unlikely
Analyte Concentration in Most Contaminated Zone Points Awarded
Dissolved Oxygen
Nitrate
Iron (II)
Sulfate
Methane
ORP
Chloride
TCE (released)
cis-DCE (none released)
VC (none released)
3 mg/L
0.3 mg/L
Not Detected (ND)
10 mg/L
ND
+ lOOmV
background
1,200 ug/L
ND
ND
Total Points Awarded
-3
2
0
2
0
0
0
0
0
0
1 Point
32
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In this example, the investigator can infer that biodegradation is probably not occurring or is
occurring too slowly to contribute to natural attenuation at the time of the sampling. In this case,
the investigator should evaluate whether other natural attenuation processes can meet the cleanup
objectives for the site (e.g., abiotic degradation or transformation, volatilization or sorption) or
select a remedial option other than MNA.
Step 2: Determine Ground-water Flow Solute Transport Parameters
After it has been shown that biodegradation is occurring, it is important to quantify ground-
water flow and solute transport parameters. This will make it possible to use a solute transport
model to quantitatively estimate the concentration of the plume and its direction and rate of travel.
To use an analytical model, it is necessary to know the hydraulic gradient and hydraulic conductivity
for the site and to have estimates of porosity and dispersivity. It also is helpful to know the coefficient
of retardation. Quantification of these parameters is discussed in detail in Appendix B.
In order to make the modeling as accurate as possible, the investigator must have site-specific
hydraulic gradient and hydraulic conductivity data. To determine the ground-water flow and solute
transport direction, it is necessary to have at least three accurately surveyed wells in each
hydrogeologic unit of interest at the site. The porosity and dispersivity are generally estimated
using accepted literature values for the aquifer matrix materials containing the plume at the site. If
the investigator has total organic carbon data for soil, it is possible to estimate the coefficient of
retardation; otherwise, it is conservative to assume that the solute transport and ground-water
velocities are the same. Techniques to collect these data are discussed in the appendices.
Step 3: Locate Sources Receptor Exposure Points
To determine the length of flow for the predictive modeling to be conducted in Step 5, it is
important to know the distance between the source of contamination, the leading edge along the
core of the dissolved plume, and any potential downgradient or cross-gradient receptor exposure
points.
Step 4: Estimate the Biodegradation Mate
Biodegradation is the most important process that degrades contaminants in the subsurface;
therefore, the biodegradation rate is one of the most important model input parameters.
Biodegradation of chlorinated aliphatic hydrocarbons can be represented as a first-order rate constant.
Whenever possible, use site-specific biodegradation rates estimated from field data collected along
the core of the plume. Calculation of site-specific biodegradation rates is discussed in Appendix C.
If it is not possible to determine site-specific biodegradation rates, then literature values may be
used in a sensitivity analysis (Table C.3.5). A useful approach is to start with average values, and
then to vary the model input to predict "best-case" and "worst-case" scenarios. Estimated
biodegradation rates can be used only after it has been shown that biodegradation is occurring (see
Step 1). Although literature values may be used to estimate biodegradation rates in the bioattenuation
screening process described in Section 2.2, additional site information should be collected to
determine biodegradation rates for the site when refining the site conceptual model, as described in
Section 2.3. Literature values should not be used during the more detailed analysis.
Step 5: Compare the Rate of Transport to the Rate of Attenuation
At this early stage in the natural attenuation demonstration, comparison of the rate of solute
transport to the rate of attenuation is best accomplished using an analytical model. Several models
are available. It is suggested that the decay option be first order for use in any of the models.
The primary purpose of comparing the rate of transport to the rate of natural attenuation is to
determine if natural attenuation processes will be capable of attaining site-specific remediation
objectives in a time period that is reasonable compared to other alternatives (i.e., to quantitatively
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estimate if site contaminants are attenuating at a rate fast enough to prevent further plume migration
and restore the plume to appropriate cleanup levels). The analytical model BIOSCREEN can be
used to determine whether natural attenuation processes will be capable of meeting site-specific
remediation objectives at some distance downgradiant of a source. The numerical model BIOPLUME
III can be used to estimate whether site contaminants are attenuating at a rate fast enough to restore
the plume to appropriate cleanup levels It is important to perform a sensitivity analysis to help
evaluate the confidence in the preliminary screening modeling effort. For the purposes of the
screening effort, if modeling shows that the screening criteria are met, the investigator can proceed
with the natural attenuation evaluation.
Step 6: Determine If Screening Criteria are Met
Before proceeding with the full-scale natural attenuation evaluation, the investigator should
ensure that the answers to both of the following questions are "yes":
» Has the plume moved a shorter distance than would be expected based on the known (or
estimated) time since the contaminant release and the contaminant velocity in ground
water, as calculated from site-specific measurements of hydraulic conductivity and
hydraulic gradient, and estimates of effective porosity and contaminant retardation?
• Is it likely that site contaminants are attenuating at rates sufficient to meet remediation
objectives for the site in a time period that is reasonable compared to other alternatives?
If the answers to these questions are "yes," then the investigator is encouraged to proceed with
the full-scale natural attenuation demonstration.
2.3 COLLECT ADDITIONAL TO EVALUATE
NATURAL ATTENUATION AS
It is the responsibility of the proponent to "make the case" for natural attenuation. Thus, a
credible and thorough site assessment is necessary to document the potential for natural attenuation
to meet cleanup objectives. As discussed in Section 2.1, review of existing site characterization
data is particularly useful before initiating site characterization activities. Such review should
allow identification of data gaps and guide the most effective placement of additional data collection
points.
There are two goals during the site characterization phase of a natural attenuation investigation.
The first is to collect the data needed to determine if natural mechanisms of contaminant attenuation
are occurring at rates sufficient to attain site-specific remediation objectives in a time period that is
reasonable compared to other alternatives. The second is to provide sufficient site-specific data to
allow prediction of the future extent and concentrations of a contaminant plume through solute fate
and transport modeling. Thus, detailed site characterization is required to achieve these goals and
to support this remedial option. Adequate site characterization in support of natural attenuation
requires that the following site-specific parameters be determined:
• Location, nature, and extent of contaminant source area(s) (i.e., areas containing mobile
or residual NAPL or highly contaminated ground water).
• Chemical properties (e.g., composition, solubility, volatility, etc.) of contaminant source
materials.
• The potential for a continuing source due to sewers, leaking tanks, or pipelines, or other
site activity.
» Extent and types of soil and ground-water contamination.
« Aquifer geochemical parameters (Table 2.1).
-------
» Regional hydrogeology, including:
- Drinking water aquifers, and
- Regional confining units.
• Local and site-specific hydrogeology, including:
- Local drinking water aquifers;
- Location of industrial, agricultural, and domestic water wells;
- Patterns of aquifer use (current and future);
-Lithology;
- Site stratigraphy, including identification of transmissive and nontransmissive units;
- Potential pathways for NAPL migration (e.g., surface topography and dip of confining
layers);
- Grain-size distribution (sand vs. silt vs. clay);
- Aquifer hydraulic conductivity;
- Ground water hydraulic information;
- Preferential flow paths;
- Locations and types of surface water bodies; and
- Areas of local ground water recharge and discharge.
« Identification of current and future potential exposure pathways, receptors, and exposure
points.
Many chlorinated solvent plumes have enough three-dimensional expression to make it
impossible for a single well to adequately describe the plume at a particular location on a map of
the site.
Figure 2.5 depicts a cross section of a hypothetical site with three-dimensional expression of
the plume. A documented source exists in the capillary fringe just above the water table. Such
sources are usually found by recovering, extracting, and analyzing core material. This material can
be (1) a release of LNAPL containing chlorinated solvents; (2) a release of pure chlorinated solvents
that has been entrapped by capillary interactions in the capillary fringe; or (3) material that has
experienced high concentrations of solvents in solution in ground water, has sorbed the solvents,
and now is slowly desorbing the chlorinated solvents. Recharge of precipitation through this source
produces a plume that appears to dive into the aquifer as it moves away from the source. This effect
can be caused by recharge of clean ground water above the plume as it moves downgradient of the
source, by collection of the plume into more hydraulically conductive material at the bottom of
aquifer, or by density differences between the plume and the unimpacted ground water.
Below the first hydrologic unit there is a second unit that has fine-textured material at the top
and coarse-textured material at the bottom of the unit. In the hypothetical site, the fine-textured
material at the top of the second unit has inhibited downward migration of a DNAPL, causing it to
spread laterally at the bottom of the first unit and form a second source of ground-water contamination
in the first unit. Because DNAPL below the water table tends to exi st as diffuse and widely extended
ganglia rather than of pools filling all the pore space, it is statistically improbable that the material
sampled by conventional core sampling will contain DNAPL. Because these sources are so difficult
to sample, these sources are cryptic to conventional sampling techniques.
At the hypothetical site, DNAPL has found a pathway past the fine-textured material and has
formed a second cryptic source area at the bottom of the second hydrologic unit. Compare Figure 2.6.
The second hydrological unit at the hypothetical site has a different hydraulic gradient than the first
unit. As a result, the plume in the second unit is moving in a different direction than the plume in
the first unit. Biological processes occurring in one hydrological unit may not occur in another; a
plume may show Type 2 behavior in one unit and Type 3 behavior in another.
35
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Documented NAPL
Cryptic
NAPL
Figure 2.5 A. cross section through a hypothetical release, illustrating the three-dimensional character
of the plumes that may develop from a release of chlorinated solvents.
Documented NAPL
Cryptic NAPL
Figure 2.6 A stacked plan representation of the plumes that may develop from the hypothetical release
depicted in Figure 2.5. Each plan representation depicts a separate plume that can
originate from discrete source areas produced from the same release of chlorinated solvents.
36
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As a consequence, it is critical to sample and evaluate the three-dimensional character of the
site with respect to (1) interaction of contaminant releases with the aquifer matrix material, (2)
local hydological features that control development and migration of plumes, and (3) the geochemical
interactions that favor bioattenuation of chlorinated solvents.
The following sections describe the methodologies that should be implemented to allow
successful site characterization in support of natural attenuation.
2.3.1 Characterization of Soils Aquifer Matrix Materials
In order to adequately define the subsurface hydrogeologic system and to determine the three-
dimensional distribution of mobile and residual NAPL that can act as a continuing source of ground-
water contamination, credible and thorough soil characterization must be completed. As appropriate,
soil gas data may be collected and analyzed to better characterize soil contamination in the vadose
zone. Depending on the status of the site, this work may have been completed during previous
remedial investigation work. The results of soils characterization will be used as input into a solute
fate and transport model to help define a contaminant source term and to support the natural
attenuation investigation.
The purpose of sampling soil and aquifer matrix material is to determine the subsurface
distribution of hydrostratigraphic units and the distribution of mobile and residual NAPL, as well
as pore water that contains high concentrations of the contaminants in the dissolved phase. These
objectives can be achieved through the use of conventional soil borings or direct-push methods
(e.g., Geoprobe® or cone penetrometer testing), and through collection of soil gas samples. All
samples should be collected, described, analyzed, and disposed of in accordance with local, State,
and Federal guidance. Appendix A contains suggested procedures for sample collection. These
procedures may require modification to comply with local, State, and Federal regulations or to
accommodate site-specific conditions.
The analytical methods to be used for soil, aquifer matrix material, and soil gas sample analyses
is presented in Table 2.1. This table includes all of the parameters necessary to document natural
attenuation, including the effects of sorption, volatilization, and biodegradation. Each analyte is
discussed separately below.
• Volatile Organic Compounds: Knowledge of the location, distribution, concentration,
and total mass of contaminants sorbed to soils or present as mobile or immobile NAPL is
required to calculate contaminant partitioning from NAPL into ground water. This
information is useful to predict the long-term persistence of source areas. Knowledge of
the diffusive flux of volatile organic compounds from NAPLs or ground water to the
atmosphere or other identified receptor for vapors is required to estimate exposure of the
human population or ecological receptors to contaminant vapors. If the flux of vapors
can be compared to the discharge of the contaminants in ground water, the contribution of
volatilization to natural attenuation of contamination in ground water can be documented.
• Total Organic Carbon: Knowledge of the TOC content of the aquifer matrix is
important for sorption and solute-retardation calculations. TOC samples should be
collected from a background location in the stratigraphic horizon(s) where most
contaminant transport is expected to occur.
» Oxygen and Carbon Dioxide: Oxygen and carbon dioxide soil gas measurements can be
used to identify areas in the unsaturated zone where biodegradation is occurring. This
can be a useful and relatively inexpensive way to identify NAPL source areas, particularly
when solvents are codisposed with fuels or greases (AFCEE, 1994).
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» Fuel and Chlorinated Volatile Organic Compounds: Knowledge of the distribution of
contaminants in soil gas can be used as a cost-effective way to estimate the extent of soil
contamination.
2.3.2 Ground-water Characterkation
To adequately determine the amount and three-dimensional distribution of dissolved
contamination and to document the occurrence of natural attenuation, ground-water samples must
be collected and analyzed. Biodegradation of organic compounds, whether natural or anthropogenic,
brings about measurable changes in the chemistry of ground water in the affected area. By measuring
these changes, it is possible to document and quantitatively evaluate the importance of natural
attenuation at a site.
Ground-water sampling is conducted to determine the concentrations and distribution of
contaminants, daughter products, and ground-water geochemical parameters. Ground-water samples
may be obtained from monitoring wells or with point-source sampling devices such as a Geoprobe®,
Hydropunch®, or cone penetrometer. All ground-water samples should be collected, handled, and
disposed of in accordance with local, State, and Federal guidelines. Appendix A contains suggested
procedures for ground-water sample collection. These procedures may need to be modified to
comply with local, State, and Federal regulations or to accommodate site-specific conditions.
The analytical protocol for ground-water sample analysis is presented in Table 2.1. This
analytical protocol includes all of the parameters necessary to delineate dissolved contamination
and to document natural attenuation, including the effects of sorption and biodegradation. Data
obtained from the analysis of ground water for these analytes is used to scientifically document
natural attenuation and can be used as input into a solute fate and transport model. The following
paragraphs describe each ground-water analytical parameter and the use of each analyte in the
natural attenuation demonstration.
2.3.2.1 Volatile and Semivolatile Organic Compounds
These analytes are used to determine the type, concentration, and distribution of contaminants
and daughter products in the aquifer. In many cases, chlorinated solvents are found commingled
with fuels or other hydrocarbons. At a minimum, the volatile organic compound (VOC) analysis
(Method SW8260A) should be used, with the addition of the trimethylbenzene isomers if fuel
hydrocarbons are present or suspected. The combined dissolved concentrations of BTEX and
trimethylbenzenes should not be greater than about 30 mg/L for a JP-4 spill (Smith et a/., 1981) or
about 135 mg/L for a gasoline spill (Cline et al, 1991; American Petroleum Institute, 1985). If
these compounds are found in higher concentrations, sampling errors such as emulsification of
LNAPL in the ground-water sample likely have occurred and should be investigated.
Maximum concentrations of chlorinated solvents dissolved in ground water from neat solvents
should not exceed their solubilities in water. Appendix B contains solubilities for common
contaminants. If contaminants are found in concentrations greater than their solubilities, then
sampling errors such as emul si fi cation of NAPL in the ground-water sample have likely occurred
and should be investigated.
2.3.2.2 Dissolved Oxygen
Dissolved oxygen is the most thermodynamically favored electron acceptor used by microbes
for the biodegradation of organic carbon, whether natural or anthropogenic. Anaerobic bacteria
generally cannot function at dissolved oxygen concentrations greater than about 0.5 mg/L and,
hence, reductive dechlorination will not occur. This is why it is important to have a source of
carbon in the aquifer that can be used by aerobic microorganisms as a primary substrate. During
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aerobic respiration, dissolved oxygen concentrations decrease. After depletion of dissolved oxygen,
anaerobic microbes will use nitrate as an electron acceptor, followed by iron (III), then sulfate, and
finally carbon dioxide (methanogenesis). Each sequential reaction drives the ORP of the ground
water downward into the range within which reductive dechlorination can occur. Reductive
dechlorination is most effective in the ORP range corresponding to sulfate reduction and
methanogenesis, but dechlorination of PCE and TCE also may occur in the ORP range associated
with denitrification or iron (III) reduction. Dehalogenation of DCE and VC generally are restricted
to sulfate reducing and methanogenic conditions.
Dissolved oxygen measurements should be taken during well purging and immediately before
and after sample acquisition using a direct-reading meter. Because most well purging techniques
can allow aeration of collected ground-water samples, it is important to minimize the potential for
aeration as described in Appendix A.
2.3.2.3 Nitrate
After dissolved oxygen has been depleted in the microbiological treatment zone, nitrate may
be used as an electron acceptor for anaerobic biodegradation of organic carbon via denitrification.
In order for reductive dechlorination to occur, nitrate concentrations in the contaminated portion of
the aquifer must be less than 1.0 mg/L.
2.3.2.4 Iron (II)
In some cases, iron (III) is used as an electron acceptor during anaerobic biodegradation of
organic carbon. During this process, iron (III) is reduced to iron (II), which may be soluble in water.
Iron (II) concentrations can thus be used as an indicator of anaerobic degradation of fuel compounds,
and vinyl chloride (see Section 2.2.1.1.2). Native organic matter may also support reduction of iron
(II). Care must be taken when interpreting iron (II) concentrations because they may be biased low
by reprecipitation as sulfides or carbonates.
2.3.2.5 Sulfate
After dissolved oxygen and nitrate have been depleted in the microbiological treatment zone,
sulfate may be used as an electron acceptor for anaerobic biodegradation. This process is termed
"sulfate reduction" and results in the production of sulfide. Concentrations of sulfate greater than
20 mg/L may cause competitive exclusion of dechlorination. However, in many plumes with high
concentrations of sulfate, reductive dechlorination still occurs.
2.3.2.6 Methane
During methanogenesis acetate is split to form carbon dioxide and methane, or carbon dioxide
is used as an electron acceptor, and is reduced to methane. Methanogenesis generally occurs after
oxygen, nitrate, and sulfate have been depleted in the treatment zone. The presence of methane in
ground water is indicative of strongly reducing conditions. Because methane is not present in fuel,
the presence of methane above background concentrations in ground water in contact with fuels is
indicative of microbial degradation of hydrocarbons. Methane also is associated with spills of pure
chlorinated solvents (Weaver et al, 1996). It is not known if the methane comes from chlorinated
solvent carbon or from native dissolved organic carbon.
2.3.2.7 Alkalinity
There is a positive correlation between zones of microbial activity and increased alkalinity.
Increases in alkalinity result from the dissolution of rock driven by the production of carbon dioxide
produced by the metabolism of microorganisms. Alkalinity is important in the maintenance of
ground-water pH because it buffers the ground water system against acids generated during both
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aerobic and anaerobic biodegradation. In the experience of the authors, biodegradation of organic
compounds rarely, if ever, generates enough acid to impact the pH of the ground water.
2.3.2.8 Oxidation-Reduction Potential
The ORP of ground water is a measure of electron activity and is an indicator of the relative
tendency of a solution to accept or transfer electrons. Oxidation-reduction reactions in ground
water containing organic compounds (natural or anthropogenic) are usually biologically mediated,
and, therefore, the ORP of a ground water system depends upon and influences rates of
biodegradation. Knowledge of the ORP of ground water also is important because some biological
processes operate only within a prescribed range of ORP conditions.
ORP measurements can be used to provide real-time data on the location of the contaminant
plume, especially in areas undergoing anaerobic biodegradation. Mapping the ORP of the ground
water while in the field helps the field scientist to determine the approximate location of the
contaminant plume. To map the ORP of the ground water while in the field, it is important to have
at least one ORP measurement (preferably more) from a well located upgradient from the plume.
ORP measurements should be taken during well purging and immediately before and after sample
acquisition using a direct-reading meter. Because most well purging techniques can allow aeration
of collected ground-water samples (which can affect ORP measurements), it is important to minimize
potential aeration by using a flow-through cell as outlined in Appendix A.
Most discussion of oxidation reduction potential expresses the potential as if it were measured
against the standard hydrogen electrode. Most electrodes and meters to measure oxidation-reduction
potential use the silver/silver chloride electrode (Ag/AgCl) as the reference electrode. This protocol
uses the potential against the Ag/AgCl electrode as the screening potential, not Eh as would be
measured against the standard hydrogen electrode.
2.3.2.9 Dissolved Hydrogen
In some ground waters, PCE and TCE appear to attenuate, although significant concentrations
of DCE and VC do not accumulate. In this situation, it is difficult to distinguish between Type 3
behavior where the daughter products are not produced, and Type 1 or Type 2 behavior where the
daughter products are removed very rapidly. In cases like this, the concentration of hydrogen can
be used to identify ground waters where reductive dechlorination is occurring. If hydrogen
concentrations are very low, reductive dechlorination is not efficient and Type 3 behavior is indicated.
If hydrogen concentrations are greater than approximately 1 nM, rates of reductive dechlorination
should have environmental significance and Type 1 or Type 2 behavior would be expected.
Concentrations of dissolved hydrogen have been used to evaluate redox processes, and thus
the efficiency of reductive dechlorination, in ground-water systems (Lovley and Goodwin, 1988;
Lovley etal., 1994; Chapelle etal., 1995). Dissolved hydrogen is continuously produced in anoxic
ground-water systems by fermentative microorganisms that decompose natural and anthropogenic
organic matter. This H2 is then consumed by respiratory microorganisms that use nitrate, Fe(III),
sulfate, or CO2 as terminal electron acceptors. This continuous cycling of H, is called interspecies
hydrogen transfer. Significantly, nitrate-, Fe(III)-, sulfate- and CO2-reducing (methanogenic)
microorganisms exhibit different efficiencies in utilizing the H2that is being continually produced.
Nitrate reducers are highly efficient H2 utilizers and maintain very low steady-state H2 concentrations.
Fe(III) reducers are slightly less efficient and thus maintain somewhat higher H9 concentrations.
Sulfate reducers and methanogenic bacteria are progressively less efficient and maintain even higher
H9 concentrations. Because each terminal electron accepting process has a characteristic H2
concentration associated with it, EL, concentrations can be an indicator of predominant redox
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processes. These characteristic ranges are given in Table 2.5. An analytical protocol for quantifying
H2 concentrations in ground water is given in Appendix A.
Table 2.5 Range of Hydrogen Concentrations for a Given Terminal Electron-Accepting Process
Terminal Electron Hydrogen (H2)
Accepting Process Concentration (nanomoles per liter)
Denitrification <().!
Iron (III) Reduction 0.2 to 0.8
Sulfate Reduction 1 to 4
Reductive Dechlorination >1
Methanogenesis 5-20
Oxidation-reduction potential (ORP) measurements are based on the concept of thermodynamic
equilibrium and, within the constraints of that assumption, can be used to evaluate redox processes
in ground water systems. The H2 method is based on the ecological concept of interspecies hydrogen
transfer by microorganisms and, within the constraints of that assumption, can also be used to
evaluate redox processes. These methods, therefore, are fundamentally different. A direct comparison
of these methods (Chapelle et al., 1996) has shown that ORP measurements were effective in
delineating oxic from anoxic ground water, but that ORP measurements could not distinguish between
nitrate-reducing, Fe(III)-reducing, sulfate-reducing, or methanogenic zones in an aquifer. In contrast,
the El, method could readily distinguish between different anaerobic zones. For those sites where
distinguishing between different anaerobic processes is important, H2 measurements are an available
technology for making such distinctions. At sites where concentrations of redox sensitive parameters
such as dissolved oxygen, iron (II), sulfide, and methane are sufficient to identify operative redox
processes, EL concentrations are not always required to identify redox zonation and predict
contaminant behavior.
In practice, it is preferable to interpret EL, concentrations in the context of electron acceptor
availability and the presence of the final products of microbial metabolism (Chapelle etal, 1995).
For example, if sulfate concentrations in ground water are less than 0.5 mg/L, methane concentrations
are greater than 0.5 mg/L, and H9 concentrations are in the 5 to 20 nM range, it can be concluded
with a high degree of certainty that methanogenesis is the predominant redox process in the aquifer.
Similar logic can be applied to identifying denitrification (presence of nitrate, EL<0.1 nM), Fe(III)
reduction (production of Fe(II), H2 concentrations ranging from 0.2 to 0.8 nM), and sulfate reduction
(presence of sulfate, production of sulfide, EL, concentrations ranging from 1 to 4 nM). Reductive
dechlorination in the field has been documented at hydrogen concentrations that support sulfate
reduction or methanogenesi s. If hydrogen concentrations are high enough to support sulfate reduction
or methanogenesis, then reductive dechlori nation is probably occurring, even if other geochemical
indicators as scored in Table 2.3 do not indicate that reductive dechlorination is possible.
2.3.2.10 pH, Temperature, and Conductivity
Because the pH, temperature, and conductivity of a ground-water sample can change
significantly within a short time following sample acquisition, these parameters must be measured
in the field in unfiltered, unpreserved, "fresh" water collected by the same technique as the samples
taken for dissolved oxygen and ORP analyses. The measurements should be made in a clean
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container separate from those intended for laboratory analysis, and the measured values should be
recorded in the ground-water sampling record.
The pH of ground water has an effect on the presence and activity of microbial populations in
ground water. This is especially true for methanogens. Microbes capable of degrading chlorinated
aliphatic hydrocarbons and petroleum hydrocarbon compounds generally prefer pH values varying
from 6 to 8 standard units.
Ground-water temperature directly affects the solubility of dissolved gasses and other
geochemical species. Ground-water temperature also affects the metabolic activity of bacteria.
Conductivity is a measure of the ability of a solution to conduct electricity. The conductivity
of ground water is directly related to the concentration of ions in solution; conductivity increases as
ion concentration increases.
2.3.2.11 Chloride
Chlorine is the most abundant of the halogens. Although chlorine can occur in oxidation
states ranging from Cl" to Cl+7, the chloride form (Cl~) is the only form of major significance in
natural waters (Hem, 1985). Chloride forms ion pairs or complex ions with some of the cations
present in natural waters, but these complexes are not strong enough to be of significance in the
chemistry of fresh water (Hem, 1985). Chloride ions generally do not enter into oxidation-reduction
reactions, form no important solute complexes with other ions unless the chloride concentration is
extremely high, do not form salts of low solubility, are not significantly adsorbed on mineral surfaces,
and play few vital biochemical roles (Hem, 1985). Thus, physical processes control the migration
of chloride ions in the subsurface. Kaufman and Orlob (1956) conducted tracer experiments in
ground water, and found that chloride moved through most of the soils tested more conservatively
(i.e., with less retardation and loss) than any of the other tracers tested.
During biodegradation of chlorinated hydrocarbons dissolved in ground water, chloride is
released into the ground water. This results in chloride concentrations in ground water in the
contaminant plume that are elevated relative to background concentrations. Because of the neutral
chemical behavior of chloride, it can be used as a conservative tracer to estimate biodegradation
rates, as discussed in Appendix C.
2.3.3 Aquifer Parameter Estimation
Estimates of aquifer parameters are necessary to accurately evaluate contaminant fate and
transport.
2.3.3.1 Hydraulic Conductivity
Hydraulic conductivity is a measure of an aquifer's ability to transmit water, and is perhaps the
most important aquifer parameter governing fluid flow in the subsurface. The velocity of ground
water and dissolved contamination is directly related to the hydraulic conductivity of the saturated
zone. In addition, subsurface variations in hydraulic conductivity directly influence contaminant
fate and transport by providing preferential paths for contaminant migration. Estimates of hydraulic
conductivity are used to determine residence times for contaminants and tracers, and to determine
the seepage velocity of ground water.
The most common methods used to quantify hydraulic conductivity are aquifer pumping tests
and slug tests (Appendix A). Another method that may be used to determine hydraulic conductivity
is the borehole dilution test. One drawback to these methods is that they average hydraulic properties
over the screened interval. To help alleviate this potential problem, the screened interval of the test
wells should be selected after consideration is given to subsurface stratigraphy.
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Information about subsurface stratigraphy should come from geologic logs of continuous cores
or from cone penetrometer tests. The rate of filling of a Hydropunch® can be used to obtain a
rough estimate of the local hydraulic conductivity at the same time the water sample is collected.
The results of pressure dissipation data from cone penetrometer tests can be used to supplement the
results obtained from pumping tests and slug tests. It is important that the location of the aquifer
tests be designed to collect information to delineate the range of hydraulic conductivity both vertically
and horizontally at the site.
2.3.3.1.1 Pumping Tests in Wells
Pumping tests done in wells provide information on the average hydraulic conductivity of the
screened interval, but not the most transmissive horizon included in the screened interval. In
contaminated areas, the extracted ground water generally must be collected and treated, increasing
the difficulty of such testing. In addition, a minimum 4-inch-diameter well is typically required to
complete pumping tests in highly transmissive aquifers because the 2-inch submersible pumps
available today are not capable of producing a flow rate high enough for meaningful pumping tests.
In areas with fairly uniform aquifer materials, pumping tests can be completed in uncontaminated
areas, and the results can be used to estimate hydraulic conductivity in the contaminated area.
Pumping tests should be conducted in wells that are screened in the most transmissive zones in the
aquifer. If pumping tests are conducted in wells with more than fifteen feet of screen, a down-hole
flowmeter test can be used to determine the interval actually contributing to flow.
2.3.3.1.2 Slug Tests in Wells
Slug tests are a commonly used alternative to pumping tests. One commonly cited drawback to
slug testing is that this method generally gives hydraulic conductivity information only for the area
immediately surrounding the monitoring well. Slug tests do, however, have two distinct advantages
over pumping tests: they can be conducted in 2-inch monitoring wells, and they produce no water.
If slug tests are going to be relied upon to provide information on the three-dimensional distribution
of hydraulic conductivity in an aquifer, multiple slug tests must be performed. It is not advisable to
rely on data from one slug test in one monitoring well. Because of this, slug tests should be
conducted at several zones across the site, including a test in at least two wells which are narrowly
screened in the most transmissive zone. There should also be tests in the less transmissive zones to
provide an estimate of the range of values present on the site.
2.3.3.1.3 Downhole Flow meter
Borehole flowmeter tests are conducted to investigate the relative vertical distribution of
horizontal hydraulic conductivity in the screened interval of a well or the uncased portion of a
borehole. These tests can be done to identify any preferential flow pathways within the portion of
an aquifer intersecting the test well screen or the open borehole. The work of Molz and Young
(1993), Molz etal. (1994) , Young and Pearson (1995), and Young (1995) describes the means by
which these tests may be conducted and interpreted.
In general, measurements of ambient ground-water flow rates are collected at several regularly
spaced locations along the screened interval of a well. Next, the well is pumped at a steady rate,
and the measurements are repeated. The test data may be analyzed using the methods described by
Molz and Young (1993) and M^olz et al. (1994) to define the relative distribution of horizontal
hydraulic conductivity within the screened interval of the test well. Estimates of bulk hydraulic
conductivity from previous aquifer tests can be used to estimate the absolute hydraulic conductivity
distribution at the test well.
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Using flowmeter test data, one may be able to more thoroughly quantify the three-dimensional
hydraulic conductivity distribution at a site. This is important for defining contaminant migration
pathways and understanding solute transport at sites with heterogeneous aquifers. Even at sites
where the hydrogeology appears relatively homogeneous, such data may point out previously
undetected zones or layers of higher hydraulic conductivity that control contaminant migration. In
addition, ground-water velocities calculated from hydraulic head, porosity, and hydraulic conductivity
data may be used to evaluate site data or for simple transport calculations. In these cases, it is also
important to have the best estimate possible of hydraulic conductivity for those units in which the
contaminants are migrating.
2.3.3.2 Hydraulic Gradient
The horizontal hydraulic gradient is the change in hydraulic head (feet of water) divided by the
distance of ground-water flow between head measurement points. To accurately determine the
hydraulic gradient, it is necessary to measure ground-water levels in all monitoring wells and
piezometers at a site. Because hydraulic gradients can change over a short distance within an
aquifer, it is essential to have as much site-specific ground-water elevation information as possible
so that accurate hydraulic gradient calculations can be made. In addition, seasonal variations in
ground-water flow direction can have a profound influence on contaminant transport. Sites in
upland areas are less likely to be affected by seasonal variations in ground-water flow direction than
low-elevation sites situated near surface water bodies such as rivers and lakes.
To determine the effect of seasonal variations in ground-water flow direction on contaminant
transport, quarterly ground-water level measurements should be taken over a period of at least one
year. For many sites, these data may already exist. If hydraulic gradient data over a one-year period
are not available, natural attenuation can still be implemented, pending an analysis of seasonal
variation in ground-water flow direction.
2.3.3.3 Processes Causing an Apparent Reduction in Total Contaminant Mass
Several processes cause reductions in contaminant concentrations and apparent reductions in
the total mass of contaminant in a system. Processes causing apparent reductions in contaminant
mass include dilution, sorption, and hydrodynamic dispersion. In order to determine the mass of
contaminant removed from the system, it is necessary to correct observed concentrations for the
effects of these processes. This is done by incorporating independent assessments of these processes
into the comprehensive solute transport model. The following sections give a brief overview of the
processes that result in apparent contaminant reduction. Appendix B describes these processes in
detail.
Dilution results in a reduction in contaminant concentrations and an apparent reduction in the
total mass of contaminant in a system due to the introduction of additional water to the system. The
two most common causes of dilution (real or apparent) are infiltration and sampling from monitoring
wells screened over large vertical intervals. Infiltration can cause an apparent reduction in
contaminant mass by mixing unaffected waters with the contaminant plume, thereby causing dilution.
Monitoring wells screened over large vertical distances may dilute ground-water samples by mixing
water from clean aquifer zones with contaminated water during sampling. To avoid potential dilution
during sampling, monitoring wells should be screened over relatively small vertical intervals (e.g.
5 feet). Nested wells should be used to define the vertical extent of contamination in the saturated
zone. Appendix C contains example calculations showing how to correct for the effects of dilution.
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The retardation of organic solutes caused by sorption is an important consideration when
simulating the effects of natural attenuation over time. Sorption of a contaminant to the aquifer
matrix results in an apparent decrease in contaminant mass because dissolved contamination is
removed from the aqueous phase. The processes of contaminant sorption and retardation are
discussed in Appendix B.
The dispersion of organic solutes in an aquifer is another important consideration when
simulating natural attenuation. The dispersion of a contaminant into relatively pristine portions of
the aquifer allows the solute plume to mix with uncontaminated ground water containing higher
concentrations of electron acceptors. Dispersion occurs vertically as well as parallel and perpendicular
to the direction of ground-water flow.
To accurately determine the mass of contaminant transformed to innocuous by-products, it is
important to correct measured contaminant concentrations for those processes that cause an apparent
reduction in contaminant mass. This is accomplished by normalizing the measured concentration
of each of the contaminants to the concentration of a tracer that is biologically recalcitrant. Because
chloride is produced during the biodegradation of chlorinated solvents, this analyte can be used as
a tracer. For chlorinated solvents undergoing reductive dechlorination, it is also possible to use the
organic carbon in the original chlorinated solvent and daughter products as a tracer. Trimethylbenzene
and tetramethylbenzene are two chemicals found in fuel hydrocarbon plumes that also may be
useful as tracers. These compounds are difficult to biologically degrade under anaerobic conditions,
and frequently persist in ground water longer than BTEX. Depending on the composition of the
fuel that was released, other tracers may be used.
2.3.4 Optional Confirmation of Biological Activity
Extensive evidence can be found in the literature showing that biodegradation of chlorinated
solvents and fuel hydrocarbons frequently occurs under natural conditions. Many references from
the large body of literature in support of natural attenuation are listed in Section 3 and discussed in
Appendix B. The most common technique used to show explicitly that microorganisms capable of
degrading contaminants are present at a site is the microcosm study.
If additional evidence (beyond contaminant and geochemical data and supporting calculations)
supporting natural attenuation is required, a microcosm study using site-specific aquifer materials
and contaminants can be undertaken.
If properly designed, implemented, and interpreted, microcosm studies can provide very
convincing documentation of the occurrence of biodegradation. Results of such studies are strongly
influenced by the nature of the geological material submitted for study, the physical properties of
the microcosm, the sampling strategy, and the duration of the study. Because microcosm studies
are time-consuming and expensive, they should be undertaken only at sites where there is considerable
uncertainty concerning the biodegradation of contaminants.
Biodegradation rate constants determined by microcosm studies often are higher than rates
achieved in the field. The collection of material for the microcosm study, the procedures used to set
up and analyze the microcosm, and the interpretation of the results of the microcosm study are
presented in Appendix C.
2.4 CONCEPTUAL CALCULA-
TIONS, AND OF NATURAL ATTENUATION
Site investigation data should first be used to refine the conceptual model and quantify ground-
water flow, sorption, dilution, and biodegradation. The results of these calculations are used to
scientifically document the occurrence and rates of natural attenuation and to help simulate natural
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attenuation over time. It is the responsibility of the proponent to "make the case" for natural
attenuation. This being the case, all available data must be integrated in such a way that the evidence
is sufficient to support the conclusion that natural attenuation is occurring.
2.4.1 Conceptual Model Refinement
Conceptual model refinement involves integrating newly gathered site characterization data to
refine the preliminary conceptual model that was developed on the basis of previously collected
site-specific data. During conceptual model refinement, all available site-specific data should be
integrated to develop an accurate three-dimensional representation of the hydrogeologic and
contaminant transport system. This refined conceptual model can then be used for contaminant
fate and transport modeling. Conceptual model refinement consists of several steps, including
preparation of geologic logs, hydrogeologic sections, potentiometric surface/water table maps,
contaminant and daughter product contour (isopleth) maps, and electron acceptor and metabolic
by-product contour (isopleth) maps.
2.4.1.1 Geologic Logs
Geologic logs of all subsurface materials encountered during the soil boring phase of the field
work should be constructed. Descriptions of the aquifer matrix should include relative density,
color, major and minor minerals, porosity, relative moisture content, plasticity of fines, cohesiveness,
grain size, structure or stratification, relative permeability, and any other significant observations
such as visible contaminants or contaminant odor. It is also important to correlate the results of
VOC screening using soil sample headspace vapor analysis with depth intervals of geologic materials.
The depth of lithologic contacts and/or significant textural changes should be recorded to the nearest
0.1 foot. This resolution is necessary because preferential flow and contaminant transport paths
may be limited to thin stratigraphic units.
2.4.1.2 Cone Penetrometer Logs
Cone Penetrometer Logs provide a valuable tool for the rapid collection of large amounts of
stratigraphic information. When combined with the necessary corroborative physical soil samples
from each stratigraphic unit occurring on the site, they can provide a three-dimensional model of
subsurface stratigraphy.
Cone penetrometer logs express stratigraphic information as the ratio of sleeve friction to tip
pressure. Cone penetrometer logs also may contain fluid resistivity data and estimates of aquifer
hydraulic conductivity. To provide meaningful data, the cone penetrometer must be capable of
providing stratigraphic resolution on the order of 3 inches. To provide accurate stratigraphic
information, cone penetrometer logs must be correlated with continuous subsurface cores. At a
minimum, there must be one correlation for every hydrostratigraphic unit found at the site. Cone
penetrometer logs, along with geologic boring logs, can be used to complete the hydrogeologic
sections discussed in Section 2.4.1.3.
2.4.1.3 Hydrogeologic Sections
Hydrogeologic sections should be prepared from boring logs and/or CPT data. A minimum of
two hydrogeologic sections are required; one parallel to the direction of ground-water flow and one
perpendicular to the direction of ground water flow. More complex sites may require more
hydrogeologic sections. Hydraulic head data including potentiometric surface and/or water table
elevation data should be plotted on the hydrogeologic section. These sections are useful in identifying
potential pathways of contaminant migration, including preferential pathways of NAPL migration
(e.g., surface topography and dip of confining layers) and of aqueous contaminants (e.g., highly
46
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transmissive layers). The potential distribution NAPL sources as well as preferential pathways for
solute transport should be considered when simulating contaminant transport using fate and transport
models.
2.4.1A Potentiometric Surface or Water Table Map(s)
A potentiometric surface or water table map is a two-dimensional graphic representation of
equipotential lines shown in plan view. These maps should be prepared from water level
measurements and surveyor's data. Because ground water flows from areas of higher hydraulic
head to areas of lower hydraulic head, such maps are used to estimate the probable direction of
plume migration and to calculate hydraulic gradients. These maps should be prepared using water
levels measured in wells screened in the same relative position within the same hydrogeologic unit.
To determine vertical hydraulic gradients, separate potentiometric maps should be developed for
different horizons in the aquifer to document vertical variations in ground-water flow. Flow nets
should also be constructed to document vertical variations in ground-water flow. To document
seasonal variations in ground-water flow, separate potentiometric surface or water table maps
should be prepared for quarterly water level measurements taken over a period of at least one year.
In areas with mobile LNAPL, a correction must be made for the water table deflection caused by
accumlation of the LNAPL in the well. This correction and potentiometric surface map preparation
are discussed in Appendix C.
2.4.1.5 Contaminant and Daughter Product Contour Maps
Contaminant and daughter product contour maps should be prepared for all contaminants
present at the site for each discrete sampling event. Such maps allow interpretation of data on the
distribution and the relative transport and degradation rates of contaminants in the subsurface. In
addition, contaminant contour maps are necessary so that contaminant concentrations can be gridded
and used for input into a numerical model. Detection of daughter products not present in the
released NAPL (e.g., c/s-l,2-DCE, VC, or ethene) provides evidence of reductive dechlorination.
Preparation of contaminant isopleth maps is discussed in Appendix C.
If mobile and residual NAPLs are present at the site, a contour map showing the thickness and
vertical and horizontal distribution of each should be prepared. These maps will allow interpretation
of the distribution and the relative transport rate of NAPLs in the subsurface. In addition, these
maps will aid in partitioning calculations and solute fate and transport model development. It is
important to note that, because of the differences between the magnitude of capillary suction in the
aquifer matrix and the different surface tension properties of NAPL and water, NAPL thickness
observations made at monitoring points may not provide an accurate estimate of the actual volume
of mobile and residual NAPL in the aquifer. To accurately determine the distribution of NAPLs, it
is necessary to take continuous soil cores or, if confident that chlorinated solvents present as NAPL
are commingled with fuels, to use cone penetrometer testing coupled with laser-induced fluorescence.
Appendix C discusses the relationship between actual and apparent NAPL thickness.
2.4.1.6 Electron Acceptor, Metabolic By-product, and Alkalinity Contour Maps
Contour maps should be prepared for electron acceptors consumed (dissolved oxygen, nitrate,
and sulfate) and metabolic by-products produced [iron (II), chloride, and methane] during
biodegradation. In addition, a contour map should be prepared for alkalinity and ORP. The electron
acceptor, metabolic by-product, alkalinity, and ORP contour maps provide evidence of the occurrence
of biodegradation at a site. If hydrogen data are available, they also should be contoured.
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During aerobic biodegradation, dissolved oxygen concentrations will decrease to levels below
background concentrations. Similarly, during anaerobic degradation, the concentrations of nitrate
and sulfate will be seen to decrease to levels below background. The electron acceptor contour
maps allow interpretation of data on the distribution of the electron acceptors and the relative transport
and degradation rates of contaminants in the subsurface. Thus, electron acceptor contour maps
provide visual evidence of biodegradation and a visual indication of the relationship between the
contaminant plume and the various electron acceptors.
Contour maps should be prepared for iron (II), chloride, and methane. During anaerobic
degradation, the concentrations of these parameters will be seen to increase to levels above
background. These maps allow interpretation of data on the distribution of metabolic by-products
resulting from the microbial degradation of fuel hydrocarbons and the relative transport and
degradation rates of contaminants in the subsurface. Thus, metabolic by-product contour maps
provide visual evidence of biodegradation and a visual indication of the relationship between the
contaminant plume and the various metabolic by-products.
A contour map should be prepared for total alkalinity (as CaCO3). Respiration of dissolved
oxygen, nitrate, iron (III), and sulfate tends to increase the total alkalinity of ground water. Thus,
the total alkalinity inside the contaminant plume generally increases to levels above background.
This map will allow visual interpretation of alkalinity data by showing the relationship between the
contaminant plume and elevated alkalinity.
2.4.2 Pre-Modeling Calculations
Several calculations must be made prior to implementation of the solute fate and transport
model. These calculations include sorption and retardation calculations, NAPL/water partitioning
calculations, ground-water flow velocity calculations, and biodegradation rate-constant calculations.
Each of these calculations is discussed in the following sections. The specifics of each calculation
are presented in the appendices referenced below.
2.4.2.1 Analysis of Contaminant, Daughter Product, Electron Acceptor, Metabolic By-product,
and Total Alkalinity Data
The extent and distribution (vertical and horizontal) of contamination, daughter product, and
electron acceptor and metabolic by-product concentrations are of paramount importance in
documenting the occurrence of biodegradation and in solute fate and transport model implementation.
Comparison of contaminant, electron acceptor, electron donor, and metabolic by-product
distributions can help identify significant trends in site biodegradation. Dissolved oxygen
concentrations below background in an area with organic contamination are indicative of aerobic
biodegradation of organic carbon. Similarly, nitrate and sulfate concentrations below background
in an area with contamination are indicative of anaerobic biodegradation of organic carbon. Likewise,
elevated concentrations of the metabolic by-products iron (II), chloride, and methane in areas with
contamination are indicative of biodegradation of organic carbon. In addition, elevated concentrations
of total alkalinity (as CaCO,) in areas with contamination are indicative of biodegradation of organic
compounds via aerobic respiration, denitrification, iron (III) reduction, and sulfate reduction. If
these trends can be documented, it is possible to quantify the relative importance of each
biodegradation mechanism, as described in Appendices B and C. The contour maps described in
Section 2.4.1 can be used to provide graphical evidence of these relationships.
Detection of daughter products not present in the released NAPL (e.g., cis-l,2-DCE, VC, or
ethene) provides evidence of reductive dechlorination. The contour maps described in Section 2.4.1
in conjunction with NAPL analyses can be used to show that reductive dechlorination is occurring.
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2.4.2.2 Sorption and Retardation Calculations
Contaminant sorption and retardation calculations should be made based on the TOC content
of the aquifer matrix and the organic carbon partitioning coefficient (Koc) for each contaminant.
The average TOC concentration from the most transmissive zone in the aquifer should be used for
retardation calculations. A sensitivity analysis should also be performed during modeling using a
range of TOC concentrations, including the lowest TOC concentration measured at the site. Sorption
and retardation calculations should be completed for all contaminants and any tracers. Sorption
and retardation calculations are described in Appendix C.
2.4.2.3 NAPL/Water Partitioning Calculations
If NAPL remains at the site, partitioning calculations should be made to account for the
partitioning from this phase into ground water. Several models for NAPL/water partitioning have
been proposed in recent years, including those by Hunt et al. (1988), Bruce etal. (1991), Cline et al.
(1991), and Johnson and Pankow (1992). Because the models presented by Cline et al. (1991) and
Bruce et al. (1991) represent equilibrium partitioning, they are the most conservative models.
Equilibrium partitioning is conservative because it predicts the maximum dissolved concentration
when NAPL in contact with water is allowed to reach equilibrium. The results of these equilibrium
partitioning calculations can be used in a solute fate and transport model to simulate a continuing
source of contamination. The theory behind fuel/water partitioning calculations is presented in
Appendix B, and example calculations are presented in Appendix C.
2.4.2.4 Ground-water Flow Velocity Calculations
The average linear ground-water flow velocity of the most transmissive aquifer zone containing
contamination should be calculated to check the accuracy of the solute fate and transport model and
to allow calculation of first-order biodegradation rate constants. An example of a ground-water
flow velocity calculation is given in Appendix C.
2.4.2.5 Apparent Biodegradation Rate-Constant Calculations
Biodegradation rate constants are necessary to accurately simulate the fate and transport of
contaminants dissolved in ground water. In many cases, biodegradation of contaminants can be
approximated using first-order kinetics. In order to calculate first-order biodegradation rate constants,
the apparent degradation rate must be normalized for the effects of dilution, sorption, and
volatilization. Two methods for determining first-order rate constants are described in Appendix C.
One method involves the use of a biologically recalcitrant compound found in the dissolved
contaminant plume that can be used as a conservative tracer. The other method, proposed by Buscheck
and Alcantar (1995) is based on the one-dimensional steady-state analytical solution to the advection-
dispersion equation presented by Bear (1979). It is appropriate for plumes where contaminant
concentrations are in dynamic equilibrium between plume formation at the source and plume
attenuation downgradient. Because of the complexity of estimating biodegradation rates with these
methods, the results are more accurately referred to as "apparent" biodegradation rate constants.
Apparent degradation rates reflect the difference between contaminant degradation and production
which is important for some daughter products (e.g., TCE, DCE, and VC).
2.5 NATURAL ATTENUATION USING SOLUTE TRANS-
Simulating natural attenuation allows prediction of the migration and attenuation of the
contaminant plume through time. Natural attenuation modeling is a tool that allows site-specific
data to be used to predict the fate and transport of solutes under governing physical, chemical, and
49
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biological processes. Hence, the results of the modeling effort are not in themselves sufficient
proof that natural attenuation is occurring at a given site. The results of the modeling effort are only
as good as the original data input into the model; therefore, an investment in thorough site
characterization will improve the validity of the modeling results. In some cases, straightforward
analytical models of solute transport are adequate to simulate natural attenuation.
Several well-documented and widely accepted solute fate and transport models are available
for simulating the fate and transport of contaminants under the influence of advection, dispersion,
sorption, and biodegradation.
2.6 CONDUCT A ANALYSIS
After the rates of natural attenuation have been documented, and predictions from appropriate
fate and transport models indicate that MNA is a viable remedy, the proponent of natural attenuation
should combine all available data and information to provide support for this remedial option.
Supporting the natural attenuation option generally will involve performing a receptor exposure
pathways analysis. This analysis includes identifying potential human and ecological receptors and
points of exposure under current and future land and ground-water use scenarios. The results of
solute fate and transport modeling are central to the exposure pathways analysis. If conservative
model input parameters are used, the solute fate and transport model should give conservative
estimates of contaminant plume migration. From this information, the potential for impacts on
human health and the environment from contamination present at the site can be assessed.
2.7 EVALUATE SUPPLEMENTAL REMOVAL OPTIONS
Additional source removal, treatment, or containment measures, beyond those previously
implemented, may be necessary for MNA to be a viable remedial option or to decrease the time
needed for natural processes to attain site-specific remedial objectives. Several technologies suitable
for source reduction or removal are listed on Figure 2.1. Other technologies may be used as dictated
by site conditions and regulatory requirements. If a solute fate and transport model has been prepared
for a site, the impact of source removal can readily be evaluated by modifying the contaminant
source term; this will allow for a reevaluation of the exposure pathways analysis.
In some cases (particularly if the site is regulated under CERCLA), the removal, treatment, or
containment of the source may be required to restore the aquifer as a source of drinking water, or to
prevent discharge of contaminants to ecologically sensitive areas. If a solute fate and transport
model has been prepared, it can also be used to forecast the benefits of source control by predicting
the time required to restore the aquifer to drinking water quality, and the reduction in contaminant
loadings to sensitive ecosystems.
2.8 LONG-TERM PLAN
This plan is used to monitor the plume over time and to verify that natural attenuation is
occurring at rates sufficient to attain site-specific remediation objectives and within the time frame
predicted at the time of remedy selection. In addition, the long-term monitoring plan should be
designed to evaluate long-term behavior of the plume, verify that exposure to contaminants does
not occur, verify that natural attenuation breakdown products do not pose additional risks, determine
actual (rather than predicted) attenuation rates for refining predictions of remediation time frame,
and to document when site-specific remediation objectives have been attained.
50
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The long-term monitoring plan should be developed based on site characterization data, analysis
of potential exposure pathways, and the results of solute fate and transport modeling. EPA is
developing additional guidance on long-term monitoring of MNA remedies, which should be
consulted when available.
The long-term monitoring plan includes two types of monitoring wells. Long-term monitoring
wells are intended to determine if the behavior of the plume is changing. Performance evaluation
wells are intended to confirm that contaminant concentrations meet regulatory acceptance levels,
and to trigger an action to manage potential expansion of the plume. Figure 2.7 depicts a schematic
that describes the various categories of wells in a comprehensive monitoring plan. Figure 2.7 is
intended to depict categories of wells, and does not depict monitoring well placement at a real site.
Included in the schematic representation are: 1) wells in the source area; 2) wells in unimpacted
ground water; 3) wells downgradient of the source area in a zone of natural attenuation; 4) wells
located downgradient from the plume where contaminant concentrations are below regulatory
acceptance levels but geochemical indicators are altered and soluble electron acceptors are depleted
with respect to unimpacted ground water; and 5) performance evaluation wells.
The final number and placement of long-term monitoring wells and performance evaluation
wells will vary from site to site, based on the behavior of the plume as revealed during the site
characterization and on the site-specific remediation objectives. In orderto provide a valid monitoring
system, all monitoring wells must be screened in the same hydrogeologic unit as the contaminant
plume being monitored. This generally requires detailed stratigraphic correlation. To facilitate
accurate stratigraphic correlation, detailed visual descriptions of all sub surf ace materials encountered
during borehole drilling or cone penetrometer testing should be prepared prior to monitoring well
installation.
Dissolved Contaminant Plume
Source Area \ Plume of Geochemical Indicators
\
Direction of Plume Migration
>
O Long Term Monitoring Wells
A Performance Evaluation Wells
Figure 2.7 Hypothetical long-term monitoring strategy. Note that number and location of monitoring
wells will vary with the three-dimensional complexity of the plume (s) and site-specific
remediation objectives.
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Although the final number and placement of long-term monitoring wells and performance
evaluation wells should be determined through regulatory negotiation, the locations of long-term
monitoring wells should be based on the behavior of the plume as revealed during the site
characterization and on regulatory considerations. The final number and location of performance
evaluation wells will also depend on regulatory considerations.
A ground-water sampling and analysis plan should be prepared in conjunction with a plan for
placement of performance evaluation wells and long-term monitoring wells. For purposes of
monitoring natural attenuation of chlorinated solvents, ground water from the long-term monitoring
wells should be analyzed for the contaminants of concern, dissolved oxygen, nitrate, iron (II), sulfate,
and methane. For performance evaluation wells, ground-water analyses should be limited to
contaminants of concern. Any additional specific analytical requirements, such as sampling for
contaminants that are metals, should be addressed in the sampling and analysis plan to ensure that
all data required for regulatory decision making are collected. Water level and NAPL thickness
measurements should be made during each sampling event.
Except at sites with very low hydraulic conductivity and gradients, quarterly sampling of both
long-term monitoring wells and performance evaluation wells is recommended during the first year
to help determine whether the plume is stable or migrating, the direction of plume migration and to
establish abaseline for behavior of the plume. After the first year, an appropriate sampling frequency
should be established which considers seasonal variations in water table elevations, ground-water
flow direction and flow velocity at the site. If the hydraulic conductivity or hydraulic gradient are
low, the time required for ground water to move from upgradient monitoring wells to downgradient
monitoring wells should also be considered in determining the appropriate monitoring frequency.
Monitoring of long-term performance of an MNA remedy should continue as long as contamination
remains above required cleanup levels.
2.9
Results of natural attenuation studies should be presented in the remedy selection document
appropriate for the site, such as CERCLA Feasibility Study or RCRA Corrective Measures Study.
This will provide scientific documentation that allows an objective evaluation of whether MNA is
the most appropriate remedial option for a given site.
All available site-specific data and information developed during the site characterization,
conceptual model development, pre-modeling calculations, biodegradation rate calculation, ground-
water modeling, model documentation, and long-term monitoring plan preparation phases of the
natural attenuation investigation should be presented in a consistent and complementary manner in
the feasibility study or similar document. Of particular interest to the site decision makers will be
evidence that natural attenuation is occurring at rates sufficient to attain site-specific remediation
objectives in a time period that is reasonable compared to other alternatives, and that human health
and the environment will be protected over time. Since a weight-of-evidence argument will be
presented to support an MNA remedy, all model assuptions should be conservative and all available
evidence in support of MNA should be presented.
52
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SECTION 3
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A
INVESTIGATION
-------
OF CONTENTS - A
A-l INTRODUCTION Al-4
A-2 SUB SURFACE INVESTIGATION METHODOLOGIES A2-5
A.2.1 TRADITIONAL DRILLING TECHNIQUES A2-5
A.2.2 CONE PENETROMETER TESTING A2-6
A.2.3 HYDRAULIC PERCUSSION SYSTEMS A2-7
A-3 SOIL CHARACTERIZATION METHODOLOGIES A3-8
A.3.1 SAMPLE ACQUISITION A3-8
A.3.2 PHYSICAL DESCRIPTION A3-8
A.3.3 FIXED-BASE LABORATORY ANALYSES A3-9
A-4 GROUND-WATER CHARACTERIZATION METHODOLOGIES A4-10
A.4.1 GROUND-WATER MONITORING LOCATIONS, DEPTHS, AND SCREENED
INTERVALS A4-10
A.4.2 TYPES OF GROUND-WATER SAMPLING LOCATIONS A4-10
A.4.2.1 Monitoring Wells A4-11
A.4.2.2 Monitoring Points A4-11
A.4.2.3 Grab Sampling A4-12
A.4.3 MEASUREMENT OF STATIC FLUID LEVELS A4-12
A.4.3.1 Water Level and Total Depth Measurements A4-12
A.4.3.2 Mobile LNAPL Thickness Measurements A4-13
A.4.3.4 Mobile DNAPL Thickness Measurements A4-13
A.4.3 GROUND-WATER EXTRACTION A4-13
A.4.3.1 Methods A4-13
A.4.3.2 Development A4-14
A.4.3.3 Purging A4-15
A.4.3.4 Sampling A4-16
A.4.4 GROUND-WATER ANALYTICAL PROCEDURES A4-17
A.4.4.1 Standard Well Head Analyses A4-18
A.4.4.2 Dissolved Hydrogen Analysis A4-18
A.4.4.2.1 Sampling Method A4-18
A.4.4.2.2 Analytical Method A4-19
A.4.4.3 Field Analytical Laboratory Analyses A4-20
A.4.4.4 Fixed-Base Laboratory Analyses A4-22
A-5 SURFACE WATER AND SEDIMENT CHARACTERIZATION
METHODOLOGIES A5-23
A.5.1 Surface Water Sample Collection A5-23
A.5.2 Sediment Sample Collection A5-23
A-6 SAMPLE HANDLING A6-24
A.6.1 SAMPLE PRESERVATION, CONTAINERS, AND LABELS A6-24
A.6.2 SAMPLE SHIPMENT A6-24
A.6.3 CHAIN-OF-CUSTODY CONTROL A6-24
A.6.4 SAMPLING RECORDS A6-25
Ai-2
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A-7 AQUIFER CHARACTERIZATION METHODOLOGIES A7-26
A.7.1 HYDRAULIC CONDUCTIVITY A7-26
A.7.1.1 Pump Tests A7-26
A.7.1.1.1 Pumping Test Design A7-26
A.7.1.1.2 Preparation for Testing A7-27
A.7.1.1.3 Conducting the Pumping Test A7-28
A.7.1.2 Slug Tests A7-29
A.7.1.3 Downhole Flow Meter Measurements A7-30
A.7.2 HYDRAULIC GRADIENT A7-31
A.7.3 DIRECT MEASUREMENT OF GROUND-WATER VELOCITY A7-31
FIGURES
No. Title Page
A.4.1 Overflow Cell to Prevent Alteration of Geochemical
Properties of Ground Water by Exposure to the Atmosphere A4-17
A.4.2 Flowthrough Cell to Prevent Alteration of Geochemical
Properties of Ground Water by Exposure to the Atmosphere A4-17
A.4.3 Schematic Showing the "Bubble Strip" Method for Measuring
Dissolved Hydrogen Concentrations in Ground water A4-19
A1-3
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SECTION A-l
INTRODUCTION
Detailed site characterization is an important aspect of the remediation by monitored natural
attenuation. Typically, it is necessary to collect additional site-specific data in order to successfully
complete the demonstration. This appendix presents an overview of field techniques that can be
used to collect the data used to evaluate monitored natural attenuation. These techniques are most
appropriate for aquifers in unconsolidated sediments. They are less appropriate for fractured rock,
and karst hydrogeologic settings. Selection of locations for field investigation activities and analyti-
cal protocols used for soil and water samples are discussed in Section 2 of the protocol document.
During all field investigation activities, special care should be taken to prevent contamination of
the sampled matrices. The primary way that sample contamination can occur is through contact with
improperly cleaned equipment. To prevent such contamination, proper equipment decontamination
procedures must be developed and followed. Procedures will vary according to site contaminants,
equipment type, field activity, sample matrix, rinseate handling requirements, and regulatory require-
ments. All equipment requires decontamination prior to initiation of site activities and between
sampling locations. New, disposable equipment does not require decontamination if factory-sealed
and found acceptable according to the appropriate data quality objectives and the site specific Qual-
ity Assurance Plan. In addition to the use of properly cleaned equipment, new, clean, disposable
gloves (of a material appropriate to the activity and contaminant type/concentration) should be worn
at each new sampling location.
Basic health and safety precautions are required for every piece of equipment and every meth-
odology discussed in this section. It is the responsibility of the investigator to be aware of and to
communicate all health and safety issues to the field team; therefore, a site specific health and safety
plan must be developed prior to initiating investigation activities. At a minimum this plan must
contain:
« A safety and health risk analysis for chemical, physical, and biological hazards associated
with the site conditions, anticipated contaminants, equipment, field activities, and climate;
« An emergency response plan with applicable emergency response numbers; and
* Precautionary measures to be implemented to insure the safety of site workers.
This appendix consists of seven sections, including this introduction. Section A-2 discusses
subsurface investigation methodologies. Section A-3 discusses soil characterization methodologies.
Section A-4 discusses groundwater characterization methodologies. Section A-5 discusses surface
water and sediment characterization methodologies. Section A-6 discusses sample handling proce-
dures. Section A-7 discusses aquifer characterization methodologies.
Ai-4
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SECTION A-2
The ideal technologies for an investigation of monitored natural attenuation are those which can
rapidly provide a large amount of information in a very short period of time while producing low
quantities of waste. The following subsections briefly introduce several alternatives that are avail-
able for performing subsurface investigations to evaluate remediation by monitored natural attenua-
tion. Although some of these alternatives more closely achieve the objectives of remediation by
monitored natural attenuation investigation than others, considerations such as site geology, site
hydrogeology, future well use, or regulatory concerns may dictate the selection of the subsurface
investigation method for any given site. It is crucial to the evaluation of monitored natural attenua-
tion to consider all of these issues prior to selecting a technology appropriate for their site. If during
the investigation it becomes necessary to change methodologies, the same concerns must be re-
addressed.
Prior to initiating any intrusive subsurface activities, proposed drilling locations must be
cleared. It is particularly useful if all utility lines in the investigation area are marked should changes
to the investigation become necessary. In addition, in order to expedite the investigation, all neces-
sary digging, coring, drilling, and ground-water monitoring point installation permits should be
obtained prior to mobilizing to the field. Care should be taken not to cross-contaminate deeper
aquifers by drilling through an aquitard underlying a DNAPL.
At the conclusion of subsurface investigations, each sampling location that is not used to install
a ground-water monitoring point or well should be restored as closely to its original condition as
possible. Where possible, holes should be sealed with bentonite chips, pellets, or grout to eliminate
any creation or enhancement of contaminant migration pathways to the ground water.
A.2.1 TRADITIONAL
Traditional drilling techniques include those methods that traditionally have been used to install
drinking water supply wells. Examples of traditional drilling techniques include hollow stem auger,
rotary, air percussion, and cable tool or chain tool. They have in common the advantage of being
capable of installing wells of varying diameters to drinking water well specifications. Each of these
techniques also allows for visual description of the materials and can allow for easy strati graphic
correlation. In general, the equipment required by each of these techniques is readily available.
Disadvantages of traditional drilling techniques include their expense, time requirements, and waste
generation. Not only do these techniques produce soil/fluids from the drilling process, frequently, in
order to properly develop wells by these techniques, a large volume of ground water must be ex-
tracted during a lengthy development. Although the advantages and disadvantages listed above are
common to most traditional drilling techniques, they are applicable to varying degrees. Furthermore,
drilling depth and subsurface stratigraphy are important considerations when evaluating the efficacy
of each of these techniques.
Hollow-stem auger has been the most widely used traditional drilling technique in environmen-
tal investigations, because it is very effective in the most commonly investigated geologic setting
encountered during environmental investigations: unconsolidated deposits at shallow depths. Al-
though less common, a chain tool can also be effective under similar geologic conditions. When
installing wells, a chain tool may require a little more time, but may prove to be less disruptive to the
formation in the vicinity of the well screen. Both techniques are well suited to collecting continuous
soil samples using a split-barrel continuous sampling device. This capability is extremely important
because detailed knowledge of the subsurface can be critical to the successful demonstration of
remediation by monitored natural attenation.
A2-5
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At greater depths and in more competent formations, rotary and air hammer techniques are
frequently used. Rotary techniques are also suited to penetration of cobbly units that may prove
difficult or impenetrable to a hollow-stem auger or chain tool. With rotary rigs, the fastest drilling
rates are usually achieved by using drilling fluids such as mud or water; however, these fluids may
require handling as IDW and may clog the pore space in the vicinity of the well screen. As long as
air circulation can be maintained in the borehole, an air hammer can be particularly useful in compe-
tent bedrock formations without introducing drilling fluids.
A.2.2 CONE
CPT is increasingly being used for successful site characterization. CPT is accomplished using
a cone penetrometer truck, which consists of an instrumented probe that is forced into the ground
using a hydraulic load frame mounted on a heavy truck, with the weight of the truck providing the
necessary force. Penetration force is typically supplied by a pair of large hydraulic cylinders bolted
to the truck frame. In tight soils, push capacity is more often limited by the structural bending
capacity of the push rods than by the weight of the truck. Cone penetrometers operate well in most
unconsolidated deposits; however, they may not be able to penetrate and may be damaged by
cobbles, gravel layers, very stiff clays, and cemented units.
The penetrometer probe generally consists of a 60-degree conical tip attached to a friction
sleeve. Inside the probe, two load cells independently measure the vertical resistance against the
conical tip and the side friction along the sleeve. Each load cell is a cylinder of uniform cross sec-
tion inside the probe which is instrumented with four strain gauges in a full-bridge circuit. Forces
are sensed by the load cells, and the data are transmitted from the probe assembly via a cable running
through the push tubes. The analog data are digitized, recorded, and plotted by computer in the
penetrometer truck. Penetration, dissipation, and resistivity data are used to determine site strati-
graphy.
The cone penetrometer can be a very effective tool for collecting large quantities of subsurface
information in a short period of time with virtually no waste generation. A cone penetrometer also
can be used for installation of ground-water monitoring points, and specially equipped penetrometers
can be used to screen for mobile and residual fuel hydrocarbon contamination using laser induced
fluorescence (LIF). Although the equipment is fairly expensive, the overall efficiency can make this
option relatively inexpensive.
Most of the disadvantages of CPT are linked to the advantages. For instance, the speed and
minimal waste associated with CPT are directly related to the process of determining lithology in
situ; however, this does not allow for visual description of subsurface materials. Isolated soil
samples can be retrieved for visual description to calibrate the cone penetrometry log, but the proce-
dure cannot be performed frequently (nor continuously) without impairing the efficiency of the
penetrometer. And while CPT can be very effective at precisely determining changes in lithology on
the basis of grain size, the lack of a visual description prevents stratigraphic correlation on the basis
of other parameters, such as color. The U.S. DoD supports a technology development program for
site characterization using cone penetrometers (the SCAPS program). SCAPS has developed a
down-hole CCD camera and light source that can visualize subsurface sediments.
Monitoring points installed using a cone penetrometer illustrate another advantage that comes
with disadvantages. CPT allows for rapid placement of discreet ground-water sampling points at a
precise depth selected on the basis of real-time, detailed, stratigraphic logs. The most effective
emplacement technique allows for installation of monitoring points of not greater than approximately
0.5 inch ID. While these points may not require much development or purging, ground-water extrac-
tion for development, purging, and sampling becomes extremely inefficient if the depth to ground
A2-6
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water is greater than approximately 25 feet. In addition, the monitoring point emplacement tech-
nique typically does not allow for installation of a sand pack, bentonite seal, and grout slurry as may
be required by regulations.
A.2.3 HYDRAULIC PERCUSSION SYSTEMS
A variety of sampling tools can be advanced through unconsolidated soils using relatively
inexpensive hydraulically powered percussion/probing machines (e.g., Geoprobe®). These sorts of
systems are frequently mounted on pickup trucks or all-terrain vehicles and, as a result of their small
size and versatility, can access many locations that larger equipment cannot.
Hydraulic percussion systems provide for the rapid collection of soil, soil gas, and ground-water
samples at shallow depths while minimizing the generation of investigation-derived waste materials.
Specifically undisturbed, continuous soils samples can rapidly be collected for visual observation,
field analysis, and/or laboratory analysis. In addition, ground-water samples can be collected
through the probe rods, or ground-water monitoring points can be installed for later sample collec-
tion. Although monitoring points installed by hydraulic percussion systems can vary considerably in
design and can include sandpacks and seals, monitoring points are typically narrow in diameter. As a
result, it can be difficult to sample points where the ground-water elevation is greater than 25 feet
bgs. Furthermore, the narrow diameter may not comply with regulatory standards or future use
needs.
A2-7
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SECTION A-3
As part of an evaluation of monitored natural attenuation for contaminants in ground water, soil
characterization factors into development of a site conceptual model, estimation of continuing source
strength, and modeling of fate and transport. The following sections describe soil sample acquisi-
tion, description, field screening, and laboratory analysis procedures. Samples should be collected in
accordance with local, State, and Federal requirements.
A.3.1 SAMPLE
Soil samples can be collected using a variety of methods, depending upon the method used to
advance boreholes. In all cases, the goal is to collect samples to allow lithologic logging and to
provide useable samples for field screening and for submission to an analytical laboratory. The
samples should meet the appropriate data quality objectives as identified in the site-specific Quality
Assurance Plan.
When using hollow-stem auger or chain tool methods, relatively undisturbed continuous soil
samples can be collected with split-barrel samplers that are either advanced using a hydraulic ham-
mer or are driven along with the advancing auger. These are well-tested methods that are useful in
most types of soils except for saturated sands, in which samples tend to liquify and slide out of the
barrel. Collection of continuous samples allows a more thorough description of site geology, with
only a slight increase in the time required for drilling. These methods also can be used to collect
samples in various types of liners, such as acetate or brass sleeves. These sleeves can be cut, capped,
and shipped with a minimum of effort. When using sleeves, the samples are disturbed less, but
description of the soils may be hindered if the liners are not clear. Other traditional drilling methods
(i.e., rotary) do not produce samples that can be used for chemical analysis, and will also make
geologic interpretation more difficult due to the disturbed nature of the material.
If CPT or hydraulic percussion methods are used, soil sampled can be collected using a hydrau-
lically driven sampler. When soil samples are collected using a probe-drive sampler, the probe-drive
sampler serves as both the driving point and the sample collection device and is attached to the
leading end of the driving rods. To collect a soil sample, the sampler is pushed or driven to the
desired sampling depth, the drive point is retracted to open the sampling barrel, and the sampler is
subsequently pushed into the undisturbed soils. The soil cores are retained within brass, stainless
steel, or clear acetate liners inside the sampling barrel. The probe rods are then retracted, bringing
the sampling device to the surface. The soil sample can then be extruded from the liners for litho-
logic logging, or the liners can be capped and undisturbed samples submitted to the analytical labora-
tory for testing.
If a hand auger is used, samples will be slightly disturbed, but still useful for logging purposes.
Removing soil from the auger bucket may prove difficult where soils are clayey. Below the water
table, it may be impossible to retain sandy soils in the bucket. Hand driven samplers are similar to
probe-drive samplers, except that all pushing power is provided manually,
Following sample acquisition, the coordinates and elevation of all soil sampling locations
should be surveyed. Horizontal coordinates should be measured to the nearest 0.1 foot relative to an
established coordinate system, such as state planar. The elevation of the ground surface also should
be measured to the nearest 0.1 foot relative to USGS mean sea level (msl) data.
A.3.2 PHYSICAL
Physical characterization of soils should be performed at all sampling locations and a descrip-
tive log prepared for the materials encountered. If using CPT, the descriptive logs should consist of
continuous computer-generated interpretations supplemented by periodic sensory confirmation and
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description. Otherwise, continuous sampling with interpretation and description is recommended in
order to precisely identify and isolate changes in lithology. The descriptive log should contain:
« Sample interval (top and bottom depth);
* Sample recovery;
« Presence or absence of contamination;
* Lithologic description, including relative density, color, major textural constituents, minor
constituents, porosity, relative moisture content, plasticity of fines, cohesiveness, grain
size, structure or stratification, relative permeability, and any other significant observa-
tions; and
* Depths of lithologic contacts and/or significant textural changes measured and recorded to
the nearest 0.1 foot.
In addition, representative samples should be photographed, labeled, and stored. Additional site
characterization features are frequently being added to the list of desirable parameters. Static pore
pressure and transient pore pressures during penetration with a cone penetrometer are examples.
A.3.3 FIXED-BASE ANALYSES
Portions of selected samples should be sent to the fixed-base laboratory for analysis. It is
desirable to sample and submit a relatively undisturbed sample, if possible. Undisturbed samples are
typically collected in brass, stainless steel, or clear acetate liners inside of a sampling barrel. Upon
removal from the barrel, liners are cut to length (if desired) and capped. If the selected drilling
technique, site conditions, or project requirements do not permit collection of undisturbed soils,
samples for analysis of volatile constituents should be transferred immediately to an appropriate
container in such a way as to minimize volatilization during the transfer and headspace in the sample
container. The analytical protocol to be used for soil sample analysis is presented in Table 2.1. This
analytical protocol includes the parameters necessary to document the effects of sorption and to
estimate the magnitude of the continuing source. The protocol document describes each soil analyti-
cal parameter and the use of each analyte in the demonstration of remediation by monitored natural
attenuation.
Each laboratory soil sample will be placed in an analyte-appropriate sample container and
delivered as soon as possible to the analytical laboratory for analysis of total hydrocarbons, aromatic
hydrocarbons, VOCs, and moisture content using the procedures presented in Table 2.1. In addition,
at least two samples from locations upgradient, crossgradient, or far downgradient of the contami-
nant source will be analyzed for TOC, and the chemical and geochemical parameters necessary to
characterize the processes and rates of reaction occurring within the plume.
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A-4
GROUND-WATER CHARACTERIZATION METHODOLOGIES
This section describes the scope of work required to collect ground-water quality samples and
to perform field analyses to evaluate the demonstration of remediation by monitored natural attenua-
tion. Ground-water sampling should be conducted only by qualified scientists and technicians
trained in the conduct of well sampling, sampling documentation, and chain-of-custody procedures.
In addition, sampling personnel should thoroughly review this protocol document and the site-
specific work plan and quality assurance plan prior to sample acquisition and have a copy of the
work plan and quality assurance plan available onsite for reference. Samples should be collected in
accordance with local, State, and Federal requirements.
A.4.1 GROUND-WATER LOCATIONS,
INTERVALS
Ground-water monitoring locations should be selected on the basis of the preliminary concep-
tual site model and information on the three-dimensional distribution of contaminants. At a mini-
mum, one monitoring location should be placed upgradient from the contaminant plume, one loca-
tion should be placed in the suspected source area, two locations should be placed within the plume,
and three locations should be placed various distances downgradient and crossgradient from the
plume. The actual number of monitoring locations could be considerably higher and should be
related to site conditions and the size of the source.
It is necessary to collect samples that document the vertical extent of contamination at several
or at all of the ground-water monitoring locations. This decision is based on the presence of confin-
ing units, the thickness of the aquifer, the type and source of contamination, and suspected variations
in subsurface transmissivity. The position of well screens should be selected by the field scientist
after consideration is given to the geometry and hydraulic characteristics of the stratum in which the
well will be screened. Wells should be screened so that the vertical distribution of contaminants and
hydraulic gradients can be delineated. Typically the shallowest ground-water monitoring depth is
chosen to intersect the water table. This allows for the monitoring of LNAPL and seasonal water
level fluctuations, as well as dissolved contaminant concentrations in the portion of the aquifer
closest to the typical source. Deeper locations are selected on the basis of contaminant distribution,
typically above or below suspected confining units or in zones believed to possess higher transmis-
sivity. To ensure well integrity, clustered monitoring wells/monitoring points generally should be
completed in separate boreholes.
Screen lengths of not more than 5 feet are recommended to help mitigate the dilution of water
samples from potential vertical mixing of contaminated and uncontaminated ground water. Screen-
ing a larger area of the saturated zone will result in averaging of contaminant concentrations and
hydraulic properties. In addition, short screened intervals used in nested pairs give important infor-
mation on the nature of vertical hydraulic gradients in the area.
A.4.2 TYPES OF LOCATIONS
Ground-water samples for the demonstration of remediation by monitored natural attenuation
can be collected from monitoring wells, monitoring points, or grab sampling locations. Monitoring
points and grab locations provide rapid and inexpensive access to shallow ground-water, and yield
ground-water samples that are appropriate for site characterization and plume definition. Conven-
tional monitoring wells are required for sites with ground-water elevations more than approximately
25 feet below ground surface. They also are recommended for long-term monitoring (LTM) and
performance evaluation ground-water sampling, and may be required for regulatory compliance.
Following installation, the location and elevation of all ground-water monitoring locations
should be surveyed. Horizontal coordinates should be measured to the nearest 0.1 foot relative to an
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established coordinate system, such as state planar. The elevation of the ground surface also should
be measured to the nearest 0.1 foot relative to USGS mean sea level (msl) data. Other elevations,
including the measuring point, should be measured to the nearest 0.01 foot.
A.4.2.1 Monitoring Wells
Monitoring wells are commonly installed to evaluate remediation by monitored natural attenua-
tion. As used in this document, monitoring wells are assumed to have, at a minimum, a sand pack, a
bentonite seal, an annular seal, a surface seal, and an inside diameter of at least 2 inches. Monitoring
wells are extremely versatile and can be used for ground-water sampling, aquifer testing, product
recovery systems, long-term monitoring, and performance evaluation monitoring. Although versa-
tile, monitoring wells are relatively expensive to install and create relatively large quantities of waste
during installation, development, and sampling. Detailed well installation procedures are described
in the following paragraphs. Of course, local protocols, regulations, type of drill rig, site conditions
and site-specific data uses should dictate actual well completion details.
The monitoring well should be installed in a bore hole with a diameter at least 4 inches larger
than the outside diameter of the well. At a minimum, blank well casing and screen should be con-
structed of Schedule 40 polyvinyl chloride (PVC) with an inside diameter (ID) of 2 inches. Fre-
quently, this diameter must be increased if the well may be used for a pumping test or certain types
of product or ground-water recovery. The screens should be factory slotted with appropriately sized
openings (typically 0.010-inch). All well sections should be flush-threaded; glued joints should not
be used. The casing at each well should be fitted with a threaded bottom plug and a top cap con-
structed of the same type of material as the well casing. The top should be vented to maintain
ambient atmospheric pressure within the well casing. It is possible that PVC will not be suitable for
use in wells intended to monitor high concentrations of volatile organic constituents.
Once the well is in place, sand, bentonite, and grout are used to fill the remaining borehole
annulus. Appropriately-sized sand must be packed along the entire length of the screen; however, it
is desirable to limit the vertical distance that the sand pack extends to either side of the screen (i.e., at
least 6 inches but less than 2 feet) because the added sand pack can increase the portion of the
aquifer that is effectively screened. A bentonite seal is placed on top of the sand pack. If conditions
permit, this seal should have a minimum thickness of 2 feet. A cement-bentonite grout is used to fill
the remainder of the annular space between the bentonite seal and the surface completion. Depend-
ing on site conditions and facility preferences, either flush-mount or stick-up surface completions
can be used. Site conditions and local, State, and Federal requirements should ultimately dictate
materials selection and construction details.
The field scientist should verify and record the boring depth, the lengths of all casing and screen
sections, and the depth to the top of all well completion materials placed in the annulus between the
casing and borehole wall. All lengths and depths should be measured to the nearest 0.1 foot.
A.4.2.2 Monitoring Points
Where site conditions and the regulatory environment permit, monitoring points are ideal tools
for rapidly and cost-effectively obtaining site data to evaluate a remediation by monitored natural
attenuation. Monitoring points can be installed and sampled rapidly while generating a minimal
volume of waste. Furthermore, some monitoring points cannot be used for ground-water or free
product level measurements. It is always useful when a site has a reasonable and adequate number
of monitoring wells. Detailed monitoring point installation procedures are described in the following
paragraphs. Of course, local protocols, regulations, available equipment, and site conditions should
dictate actual well completion details.
In this document, monitoring points are considered temporary or permanent ground-water
sampling locations that do not meet the specifications of monitoring wells. Typically monitoring
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points are installed in small diameter boreholes using CPT, hydraulic percussion, or manually-
powered equipment. As a result, monitoring points usually have an ID of less than 2 inches. In
addition, because of the extremely small to nonexistent annular space between the borehole wall and
the monitoring point materials, they seldom have a sand pack, bentonite seal, and grout seal, particu-
larly with an annulus of 2 inches. Because these components are missing, ground-water monitoring
points should be installed only in shallow aquifers where installation of such devices will not result
in the cross-contamination of adjacent water-bearing strata.
Like monitoring wells, monitoring points are typically constructed of Schedule 40 PVC casing
and screen; however, monitoring points also can be constructed from Teflon®-lined tubing attached
to a stainless steel, wire mesh screen. Because the screens are often installed without a sand pack, a
slot size of 0.010 inch or smaller should be used. All monitoring point casing and screen sections
should be flush-threaded; glued joints should not be used. The casing at each monitoring point
should be fitted with a bottom cap and a top cap constructed of PVC. The top cap should be vented
to maintain ambient atmospheric pressure within the monitoring point casing. Site conditions and
local, State, and Federal requirements should ultimately dictate materials selection and construction
details.
The field hydrogeologist should verify and record the total depth of the monitoring point, the
lengths of all casing and screen sections, and the depth to the top of all monitoring point completion
materials. All lengths and depths should be measured to the nearest 0.1 foot.
A.4.2.3 Grab Sampling
Ground-water grab samples are temporally and spatially discrete samples collected from bore-
holes that are abandoned upon completion of sampling. In highly transmissive aquifers, the collec-
tion of grab samples can provide a rapid, cost-effective alternative to the use of monitoring points.
Like monitoring points, collection of grab samples generates minimal waste; however, they are not
appropriate for aquifer testing, remediation systems, or long-term monitoring. Furthermore, because
the locations are abandoned upon completion of sampling, analytical results cannot be confirmed,
and ground-water levels at all locations cannot be collected over the space of a few hours for use in
the development of ground-water flow maps. In addition, if the aquifer is not particularly transmis-
sive, sample collection can require hours resulting in inefficient equipment utilization. For these
reasons, installation and sampling of monitoring points typically is recommended where feasible.
Several of the more common instruments used to collect ground-water grab samples include the
HydroPunch®, Geoprobe®, cone penetrometer, or hand-driven points. An optimal site characteriza-
tion approach often involves use of grab samples acquired by push technologies such as the
HydroPunch®, Geoprobe®, cone penetrometer, or hand-driven points for a rapid, three-dimensional
characterization of the site, then using that information to select locations and screened intervals for
permanent monitoring points.
A.4.3 MEASUREMENT OF LEVELS
A.4.3.1 Water Level and Total Depth Measurements
Prior to purging or developing any water from a ground-water sampling location, the static
water level should be measured. At all locations of sufficient diameter, an electric water level probe
should be used to measure the depth to ground water below the datum to the nearest 0.01 foot. Small
diameter probes are commercially available for measurement of water levels in monitoring points
and through Geoprobe®, HydroPunch®, and CPT pushrods. After measuring the static water level,
the water level probe should be slowly lowered to the bottom of the well, and the total well depth
should be measured to the nearest 0.01 foot. If measuring from the ground surface, an accuracy
better than 0.1 foot is probably not practical. Based on these measurements the volume of water to
be developed or purged from the location can be calculated. If mobile LNAPL is encountered, the
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LNAPL thickness should be determined, and attempts should be made to sample both the ground
water below the LNAPL layer as well as the LNAPL.
If a sufficiently narrow water level probe is unavailable, hollow, high-density polyethylene
(HOPE) tubing connected to a manometer can be used to determine depth to ground water. The
manometer will indicate when ground water is reached as the HDPE tubing is inserted into the
monitoring location. The HDPE attached to the manometer will then be marked at the level of the
ground surface and removed. The depth to water will be determined by placing a tape measure next
to the HDPE tubing and measuring the length from the base of the tubing to the ground level mark to
the nearest 0.01 foot, if possible.
A.4.3.2 Mobile LNAPL Thickness Measurements
At sites where phase-separated hydrocarbons are present in the ground-water system, it is
important to accurately measure the thickness of floating hydrocarbons. Accurate measurement of
hydrocarbon thickness allows for estimation of the amount and distribution of the hydrocarbon and
correction of measured ground-water elevations. There are three methods that can be used to deter-
mine the thickness of mobile LNAPL in a well, including use of an interface probe, a bailer, or tape
and paste. Interface probes generally operate on either light refraction sensors or density float
switches to detect hydrocarbons and the hydrocarbon/water interface. The depth to mobile LNAPL
and depth to water should be measured to the nearest 0.01 foot. The thickness of phase-separated
hydrocarbons should also be measured to the nearest 0.01 foot. Three consecutive measurements
should be made to ensure the accuracy of the measuring instrument. A clear bailer can be slowly
lowered into the well until it intersects the fluid but is not totally immersed. The bailer is then
retrieved, and the floating LNAPL can be visually observed and measured with an engineer's tape.
The third method for measurement of floating hydrocarbon thickness is hydrocarbon paste and an
engineer's tape. The paste, when applied to the tape, changes color when it intersects the hydrocar-
bon and the hydrocarbon/water interface. Measurements of the mobile LNAPL thickness can be
made directly from the engineer's tape. It is extremely important to remember to thoroughly decon-
taminate all equipment between well measurement events to prevent cross-contamination of wells.
Equipment blanks, part of the Quality Assurance Program, will confirm the suitability of the decon-
tamination activities.
Measurements of mobile LNAPL thickness made in monitoring wells provide only an estimate
of the actual thickness of NAPL at that location. Actual mobile and residual LNAPL thicknesses can
only be obtained from continuous soil cores. Correcting apparent mobile LNAPL thickness as
measured in monitoring wells to true thickness is discussed in Appendix C.
A.4.3.3 Mobile DNAPL Thickness Measurements
DNAPL thickness in wells cannot be used to estimate actual DNAPL quantities on a site.
A.4.3 GROUND-WATER EXTRACTION
Varied equipment and methods are available for the extraction of ground water. The approach
is determined on the basis of application (development, purging, or sampling), hydrogeologic condi-
tions, monitoring location dimensions, and regulatory requirements.
Ground water produced during extraction activities must be handled in a manner consistent with
the investigation-derived waste (IDW) plan for the site. The method of handling and disposal will
depend on location and type of source, site contaminants, degree of contamination (e.g., free product,
odor, air monitoring measurements), and applicable local, State, and Federal regulations.
A.4.3.1 Methods
Portable ground-water extraction devices from three generic classifications are commonly used
for investigations of monitored natural attenuation: grab, suction lift, and positive displacement.
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The selection of the type of device(s) for the investigation is based on type of activity, well/point
dimensions, and hydrogeologic conditions.
Bailers are common grab sampling devices. Disposable bailers can be used to avoid decontami-
nation expenses and potential cross-contamination problems. Drawbacks for bailers include agita-
tion/aeration of the ground water and the inability to maintain a steady, non-turbulent flow required
to establish a true flow-through cell. Aeration also can be an issue during transfer of the sample
from the bailer to the sample container. As a result of aeration, and because a true flow-through cell
cannot be established, accurate dissolved oxygen and ORP measurements can be difficult to obtain.
The suction lift technology is best represented in environmental investigations by the peristaltic
pump. A peristaltic pump extracts water using a vacuum created by cyclically advancing a sealed
compression along flexible tubing. This pumping technique means that extracted water contacts
nothing other than tubing that can be easily replaced between sampling locations. This reduces the
possibility of cross-contamination. Furthermore, peristaltic pumps can be used to extract minimally-
disturbed ground water from any size monitoring location at variable low-flow rates. Because of
these features, representative samples are simple to collect, and reliable flow-through cells are
simple to establish. The biggest drawback with a peristaltic pump is the maximum achievable
pumping depth which is equivalent to the height of water column that can be supported by a perfect
vacuum. This effectively limits the use of a peristaltic pump to monitoring locations with ground-
water depths of less than approximately 25 feet. Also, off-gasing can occur in the tubing as a result
of the reduced pressures and high-rate of cyclical loading. If bubbles are observed in the tubing
during purging or sampling, the flow rate of the peristaltic pump must be slowed. If bubbles are still
apparent, the tubing should be checked for holes and replaced. The final potential disadvantage with
a peristaltic pump is the low flow rate. Although advantageous for sampling, this can be inappropri-
ate during purging or development at locations with large extraction volumes. Puls and Barcelona
(1996) show that the use of peristaltic pumps does not compromise sample integrity as long as no
bubbles form during sampling. If the ground water is saturated with methane or carbon dioxide, it is
practically impossible to collect samples without a gas headspace. Pankow (1986) gives advice on
how to correct for this problem.
Positive displacement pumps, also called submersible pumps, include, for example, bladder
pumps, Keck®, Grundfos Redi-Flo II®, Bennett® and Enviro-Tech Purger ES® pumps. Each of these
pumps operates downhole at depths of up to a few hundred feet and rates of up to several gallons per
minute. Therefore, submersible pumps are particularly useful for applications requiring the extrac-
tion of large volumes of water or for the extraction of ground water from depths in excess of 25 feet.
Because the pumps operate downhole, they require appropriately-sized wells. At a minimum, an
inside well diameter of at least 1.5 inches typically is required; however, much larger well diameters
can be required depending on the selected pump type, extraction depth, and extraction rate. Because
typical submersible pump design results in contact between the ground water and internal as well as
external surfaces of the pump, rigorous decontamination and quality assurance procedures must be
implemented to avoid cross-contamination if a pump that is not dedicated to the well is used for
sampling.
A.4.3.2 Development
Monitoring wells and points should be developed prior to sampling to remove fine sediments
from the portion of the formation adjacent to the screen. Development is not required for grab
sampling locations. Because development is intended to enhance ground-water production and
quality through the removal of fine sediments in the immediate vicinity of the screen, high flow rates
and downhole turbulence are beneficial. This is particularly true for monitoring wells because of the
formation disturbance usually associated with installation. Development can be accomplished using
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any of the methods discussed in Section A.4.3.1 with selection dependent on well/point dimensions,
well/point installation procedures, and hydrogeologic conditions.
Development is accomplished through the removal of water from the well/point in combination
with screen/sand pack cleansing through agitation of the downhole ground water. The "agitation" is
typically provided by pumping at a high flow rate; surging with the pump, a surge block, or a bailer;
and/or pumping along the entire length of the screen. As a rule, the more "agitation" that can be
provided, the "better" the development. Typically during development, ground water is extracted
until dissolved oxygen, pH, temperature, specific conductivity, and water clarity (turbidity) stabilize.
Monitoring well/point development should occur a minimum of 24 hours prior to sampling. Devel-
opment water must be handled in accordance with the site IDW plan.
It is important to maintain a record of development for each location. The development record
should include the following information, at a minimum:
• Monitoring point/well number;
• Date and time of development;
• Development method;
• Monitoring point/well depth;
• Volume of water produced;
• Description of water produced;
• Post-development water level and monitoring point/well depth; and
• Field analytical measurements, including pH, temperature, and specific conductivity.
A.4.3.3 Purging
Purging consists of the evacuation of water from the monitoring location prior to sampling, so
that "fresh" formation water will enter the monitoring location and be available for sampling. Be-
cause sampling can occur immediately upon completion of purging, it is best to limit ground-water
agitation, and consequently, aeration of the ground water and volatilization of contaminants. Two
sources for agitation include the purging device and the cascading of water down the screen as the
water level in the well drops. To avoid agitation, a low-disturbance device such as a peristaltic pump
or bladder pump is recommended for purging, while equipment such as bailers should be avoided.
To avoid aeration, wells or points that were initially screened below the water table should be
pumped at a rate which prevents lowering of the water table to below the top of the screen, and if
practical, wells or points screened across the water table should be pumped at a rate that lowers the
total height of the water column no more than 10 percent of the screened interval. Purging should
follow the recommendations of Puls and Barcelona (1996).
Typically, the volume of water contained within the monitoring well/point casing is used to
estimate the amount of ground water that should be removed during the purge. As a general rule,
three times the calculated volume should be removed from the well/monitoring point; however, this
can be reduced to between 1 and 3 volumes for low-producing wells and wells with a very large
water column, but a very short screened interval. Purging should continue until parameters such as
pH, temperature, specific conductance, dissolved oxygen, and ORP stabilize. Sampling should occur
as soon after purging as practical, and definitely within 24 hours. Purge waters must be handled in
accordance with the site IDW plan.
If a monitoring well/monitoring point is evacuated to a dry state during purging, the monitoring
well/monitoring point should be allowed to recharge, and the sample should be collected as soon as
sufficient water is present in the monitoring well or monitoring point to obtain the necessary sample
quantity. Sample compositing or sampling over a lengthy period by accumulating small volumes of
water at different times to obtain a sample of sufficient volume should be avoided.
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It is important to record purge information as a part of the sampling record for each location. At
a minimum, the following information pertaining to the purge should be recorded:
« Monitoring point/well number;
* Date and time of purge;
« Purge method;
* Monitoring point/well depth;
« Volume of water produced;
* Description of water produced;
« Post-purge water level; and
* Field analytical measurements, including pH, temperature, specific conductivity, dissolved
oxygen concentration, and ORP;
* Thickness of LNAPL, if present, in the point/well prior to purging;
« Volume of LNAPL removed during purging.
A.4.3.4 Sampling
Sampling should occur immediately after purging. If well yield is less than 1/10 of a liter per
minute, sample according to the guidance provided by Puls and Barcelona (1996). The object of
sampling is the collection of representative ground-water samples. This means that impact to the
sample as a result of turbulence, contact with equipment, or a change in conditions must be mini-
mized. The use of a peristaltic pump with dedicated HOPE tubing is recommended for monitoring
locations where the depth to water is less than 25 feet because the peristaltic pump is capable of
providing a steady, low-flow, stream of ground water which has contacted only dedicated tubing. In
addition, conditions are relatively unchanged, so long as care is taken to ensure that the pumping
suction does not cause the ground water to boil as a result of the reduced pressure. Where the depth
to ground water is greater than 25 feet, a dedicated positive displacement pump, when available, is
best. Because of the decontamination difficulties and the resulting potential for cross-contamination
associated with most positive displacement pumps, sampling through these pumps is not recom-
mended unless the pumps are dedicated. A bailer should be used only if it is the only means of
obtaining a sample.
An overflow cell, such as the one pictured on Figure A.4.1, or a flow-through cell as pictured in
Figure A.4.2, should be used for the measurement of well-head parameters, including pH, tempera-
ture, specific conductance, dissolved oxygen, and ORP. When using a pump to purge or sample, the
pump intake tubing should be positioned near the bottom of the cell. If using a bailer, the water
should be drained from the bottom of the bailer through tubing into the cell. In either case, the
tubing should be immersed alongside the dissolved oxygen probe beneath the water level in the cell.
This will minimize aeration and keep water flowing past the dissolved oxygen probe's sampling
membrane. The probes for the other parameters are less sensitive to positioning within the flow-
through cell.
Samples should be collected directly from the pump discharge tube or bailer into a sample
container of appropriate size, style, and preservation for the desired analysis. Water should be
directed down the inner walls of the sample bottle to minimize aeration of the sample. All samples
to be analyzed for volatile constituents (e.g., SW8010, SW8020, SW8240, SW8260, and TPH-g) or
dissolved gases (e.g., methane, ethane, and ethene) must be filled and sealed so that no air space
remains in the container. Sample handling procedures are further described in Section A.6.
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Tubing from Pump
or Bailer
Dissolved Cwygen
or ORP Probe
Erlenmeyer Flask
or Flow-Through Cell
Figure A. 4.1 Overflow cell to prevent alteration of geochemical properties of ground water by exposure to
the atmosphere.
Figure A. 4.2. Mow-through cell to prevent alteration of geochemical properties of ground water by
exposure to the atmosphere.
A.4.4 GROUND-WATER ANALYTICAL
In order to demonstrate the efficacy of monitored natural attenuation, field and laboratory
analyses should be performed on all ground-water samples using the analytical procedures listed in
Table 2.1. As a result of analyte properties and available detection equipment, analyses can be
performed at the sampling location, a portable field laboratory, or a fixed-base laboratory. The
dissolved hydrogen analysis is unique in that it requires a combination of well-head and field labora-
tory procedures that are somewhat different from other field methods; therefore, it is presented in a
separate subsection. Several of the analytes or parameters can be measured in more than one man-
ner; consequently, the methods provided in this section should not be considered absolute. Rather,
these methods have been proven to provide reliable information. The site-specific data quality needs
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of each project will be determined during the Data Quality Objective Process and documented in the
Quality Assurance Plan.
In order to obtain accurate and defensible data, it is critical that quality assurance procedures are
followed for all analyses. These procedures generally fall into the following categories:
« Collection and handling of samples;
* Calibration of direct read meters, chromatographs, colorimeters, and field instruments per
manufacturer's instructions;
* Decontamination of equipment and containers; and
« Confirmation of results through analysis of blanks, duplicates, and other quality control
samples.
Actual procedures are equipment and analysis specific, and must be developed accordingly.
A.4.4.1 Standard Weil-Head Analyses
Standard well-head analyses include pH, conductivity, temperature, dissolved oxygen, and ORP
because these parameters can be measured with a direct-reading meter. This allows all of these
parameters to be used as indicators for ground-water stability during development and purging
activities. In addition, dissolved oxygen and ORP can be used to provide real time data on the
location of the contaminant plume, especially in areas undergoing anaerobic biodegradation. Tem-
perature, dissolved oxygen, and ORP must be measured at the well head in unfiltered, unpreserved,
"fresh" water because these parameters can change significantly within a short time following
sample acquisition. Section 2.3.2 of the protocol document describes each analysis and its use in the
demonstration of monitored natural attenuation.
It is critical that samples collected for well-head analyses are disturbed and aerated as little as
possible; therefore, the use of a flow-through cell, as described in Section A.4.3 and illustrated on
Figure A.4.1, is recommended. Where this is not possible, measurements can be made in a clean
glass container separate from those intended for laboratory analysis. Where ground-water extraction
disturbs the sample, downhole probes can be used for dissolved oxygen analyses, but such probes
must be thoroughly decontaminated between wells. In some cases, decontamination procedures can
be harmful to the dissolved oxygen probe, and inadequate decontamination can create potential
cross-contamination problems if performed prior to sample collection for the other analytes. After
sample acquisition, the downhole ground water may be too disturbed to collect an accurate downhole
DO measurement.
A.4.4.2 Dissolved Hydrogen Analysis
As described in Section 2.3.2.9, dissolved hydrogen (H,,) concentrations can be an indicator of
microbially mediated redox processes in ground-water systems. Determination of H, concentrations
is a two-step process in the field: sampling at the well head and analysis with a reducing gas detector.
Hydrogen is highly volatile, and this chemical property can be used to measure H2 concentra-
tions in ground water. The principle is to continuously pump ground water through a gas-sampling
bulb containing a nitrogen or air "bubble" so that the H, can partition between the gas and liquid
phases until the concentration of H9 in the bubble comes into equilibrium with concentration of H, in
the ground water. The bubble is then analyzed for H, and the concentration of H2 in the ground
water is calculated using the Ideal Gas Law and Henry's Law. This method is referred to as the
"bubble strip" method (Chapelle et al., 1995,1997), because the bubble "strips" H, out of the water.
A.4.4.2.1 Sampling Method
The following procedures are recommended for the collection of a sample for analysis by the
"bubble strip" method:
1. Place the intake hose of a peristaltic pump, a Bennett positive displacement pump, or a blad-
der pump into the sampling well at the depth of the screened interval.
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Do not sample for H2 with electrical submersible pumps because they may produce hydrogen.
Do not sample for Ft, from wells with metal screens or casings because they may produce hydrogen
and interfere with measurements.
2. Attach a glass, 250-ml gas-sampling bulb (Figure A.4.3) to the outflow end of the tube.
3. Turn on the pump and adjust the flow rate to between 400 and 700 mL/min.
4. Briefly hold the outlet end of the sampling bulb in the upright position to remove any gas
bubbles from the bulb.
5. Place the bulb in a horizontal position and inject 20 mL of hydrogen-free N0 gas through the
septum (Figure A.4.3).
6. Allow the N2 bubble to come into equilibrium with the flowing ground water for 30 minutes.
This equilibration process takes approximately 20 minutes.
7. Remove 3-5 mL of the gas bubble using a 10 mL glass syringe with attached mini-inert valve.
8. Close the valve to seal the sample.
9. Wait an additional 5 minutes and repeat steps 7 and 8.
10. Analyze both samples on the hydrogen detector, as described in Section A.4.4.2.2.
Resample the well if the H concentrations of the duplicate samples do not agree within 10 percent.
Syringe
Septum
Nitrogen
Bubble
Water Flow
From Pump
250 ml Gas-Sampling Bulb
Water
Discharge
Figure A.4.3 Schematic showing the "bubble strip" method for measuring dissolved hydrogen
concentrations in ground water.
A.4.4.2.2 Analytical Method
Concentrations of H, in the nitrogen bubble are determined by gas chromatography (GC) with
reduction gas detection (Trace Analytical, Menlo Park, CA). To perform this analysis, a gaseous
sample is injected into the stream of a carrier gas such as N0. The sample is transported by the
carrier through a separation column where the components of the sample are separated on the basis
of variations in their transport efficiency through the column matrix. The column is packed with
CarboSieve II which separates chemical species primarily on the basis of molecular size. The sepa-
rated components elute from the column and pass through a heated bed of HgO where the reduced
gases (primarily H9 and CO) are oxidized and Hg vapor is released. The concentration of Hg vapor
released is directly proportional to the concentration of reduced gases present in the sample and is
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detected by means of an ultraviolet photometer. Because chlorinated solvents can destroy the HgO
bed, the column is backflushed immediately after the H2 peak is quantified.
The concentration of H2 dissolved in the ground water can be calculated from the equilibrated
concentration in the nitrogen gas bubble as follows:
1) Prepare a calibration curve for H9 using a 100 ppm Scotty II standard gas mixture. The cali-
bration curve should range from 0.1 to 10.0 [iL/L (ppm).
2) Analyze the gas sample taken from the gas-sampling bulb, obtaining results (CB) in units of
|iL/L (ppm) in the gas phase.
3) Calculate the aqueous concentration of H, (Cw in nanomoles per liter (nM)) in equilibrium
with the equilibrated bubble gas (CB, |JL/L (ppm)) sample using the conversion factor:
CW=Q.S1CB eq. A.4.1
This conversion factor is derived from the Ideal Gas Law and Henry's Law as follows:
PV= nRT (Ideal Gas Law) eq. A.4.2
Rearrange to give:
n P
- = — eq.A.4.3
Where:
n = the quantity of gas in moles
V = the volume the gas occupies in Liters
P = the partial pressure of the gas in atm
T = the temperature in °K
R = the gas constant (R = 0.08205 atm L mole'1 °K-1)
Thus the concentration of a pure gas at atmospheric pressure and room temperature is
40.9mmoles/L.
For a 1.0 ppm calibration standard (i.e., 1.0 |JL/L), the H2 concentration in molar units would be:
(40.9mmoles/ LH )(W~6 LH I Lgas)(W6 nmoles I mmoles) = 40.9nmoles I Lgas eq. A.4.4
The dissolved H, concentration in the aqueous phase is given by Henry's Law:
C,
C =
H eq. A.4.5
„ - (4Q.9nmolesL lppm l)
Conversion factor = - — - = 0.81 eq A 4 6
50.4 4' ' '
Where:
C = the dissolved H, concentration in nmoles/L
w ~
Ch = the equilibrated bubble H2 concentration in nmoles/L
Hlp = the dimensionless Henry's Law coefficient for the distribution of H, between the
gaseous and dissolved phases (Hm = 50.4).
4) Identify the predominant terminal electron accepting process for the water sample using the
characteristic ranges presented in Table 2.5.
A.4.4.3 Field Analytical Laboratory Analyses
The field analytical laboratory analyses to be used for ground-water samples are presented in
Table 2.1. These analyses include parameters that are time-sensitive or can be performed accurately,
easily, and inexpensively on site. In addition, results obtained from field laboratory analyses provide
real-time data on the location of the contaminant plume, especially in areas undergoing anaerobic
biodegradation. This real-time data can be used to guide the investigation of monitored natural
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attenuation at sites with limited or ambiguous hydrogeologic and plume information. Section 2.3.2
of the protocol document describes each analysis and its use in the demonstration of monitored
natural attenuation.
In preparation for field laboratory analysis, all glassware or plasticware used in the analyses
must be cleaned thoroughly by washing with a solution of laboratory-grade, phosphate-free detergent
(such as Alconox®) and water, and rinsing with deionized water and ethanol to prevent interference
or cross-contamination between measurements. If concentrations of an analyte are above the range
detectable by the titrimetric method, the analysis should be repeated by diluting the ground-water
sample with double-distilled water until the analyte concentration falls to a level within the range of
the method. All rinseate and sample reagents accumulated during ground-water analysis must be
handled appropriately, including collection, labeling, storage, and disposal.
Carbon dioxide (CO2) is a byproduct of naturally occurring aerobic and anaerobic biodegrada-
tion processes that occur in ground water. Carbon dioxide concentrations in ground water can be
measured in the field by titrimetric analysis using CHEMetrics® Method 4500 (0 to 250 mg/L as
CO2), or similar.
An increase in the alkalinity of ground water above background may be produced when carbon
dioxide produced by biological activity reacts with carbonate minerals in the aquifer matrix material.
Alkalinity of the ground-water sample will be measured in the field by titrimetric analysis using
U.S. EPA-approved Hach® Method 8221 (0 to 5,000 mg/L as calcium carbonate), or similar.
Nitrate-nitrogen concentrations are of interest because nitrate can act as an electron acceptor
during hydrocarbon biodegradation under anaerobic soil or ground-water conditions. Nitrate-nitro-
gen is also a potential nitrogen source for hydrocarbon-degrading bacteria biomass formation. Ni-
trite-nitrogen is an intermediate byproduct in both ammonia nitrification and in nitrate reduction in
anaerobic environments. Nitrate- and nitrite-nitrogen concentrations in ground water can be mea-
sured in the field by colorimetric analysis using a portable colorimeter (such as the Hach® DR/700).
Nitrate concentrations in ground-water samples can be analyzed after preparation with Hach®
Method 8039 (0 to 30.0 mg/L nitrate), or similar. Nitrite concentrations in ground-water samples can
be analyzed after preparation with U.S. EPA-approved Hach® Method 8507 (0 to 0.35 mg/L nitrite),
or similar.
Sulfate in ground water is a potential electron acceptor for fuel-hydrocarbon biodegradation in
anaerobic environments, and sulfide is produced by biological sulfate reduction. Sulfate and sulfide
concentrations can be measured by colorimetric analysis with a portable colorimeter (such as the
Hach® DR/700) after appropriate sample preparation. U.S. EPA-approved Hach® Methods 8051 (0
to 70.0 mg/L sulfate) and 8131 (0.60 mg/L sulfide) (or similar) can be used to prepare samples and
analyze sulfate and sulfide concentrations, respectively.
Iron III is an electron acceptor for biological metabolism under anaerobic conditions. Iron III is
the substrate for biological iron reduction; Iron II is the product. Iron concentrations can be mea-
sured in the field by colorimetric analysis with a portable colorimeter (such as a Hach® DPJ700) after
appropriate sample preparation. Hach® Method 8008 for total soluble iron (0 to 3.0 mg/L ferric +
ferrous iron) and Hach® Method 8146 for ferrous iron (0 to 3.0 mg/L) (or similar) can be used to
prepare and quantitate the samples. Ferric iron is quantitated by subtracting ferrous iron levels from
total iron levels.
Manganese is a potential electron acceptor under anaerobic environments. Manganese concen-
trations can be quantitated in the field using colorimetric analysis with a portable colorimeter (such
as a Hach® DR/700). U.S. EPA-approved Hach® Method 8034 (0 to 20.0 mg/L), or similar, can be
used to prepare the samples for quantitation of manganese concentrations.
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A.4.4.4 Fixed-Base Laboratory Analyses
The fixed-base laboratory analyses to be used for ground-water samples are presented in
Table 2.1. These analyses include the parameters that cannot be easily or accurately performed in the
field, but are necessary to document monitored natural attenuation of fuel hydrocarbons and chlori-
nated solvents in ground water. Section 2.3.2 of the protocol document describes each analysis and
its use in the demonstration of monitored natural attenution.
Prior to sampling, arrangements should be made with the analytical laboratory (or other sup-
plier) to provide a sufficient number of appropriate sample containers for the samples (including
quality control samples) to be collected. All containers, preservatives, and shipping requirements
should be consistent with the analytical protocol. For samples requiring chemical preservation,
preservatives are best added to containers by the laboratory (or other supplier) prior to shipping.
Sample handling is discussed in Section A.6.
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SECTION A-5
METHODOLOGIES
At sites where surface water bodies are affected (or potentially affected) by contamination,
surface water and sediment sample collection and analysis may be required as a component of the
remediation by monitored natural attenuation demonstration.
A.5.1 SURFACE COLLECTION
Surface water can be collected with a peristaltic pump using exactly the same equipment and
procedures to collect water from a well. The sampling tube can be introduced into the water from a
barge or boat, or from a dock. The depth to the sediment should be sounded, then the tube intro-
duced to a level a very few inches above the sediment layer. A weight can be used to keep the tube
straight. Alternately, 1A inch PVC pipe can be inserted to the correct depth, then sampled with a tube
just as if it were a well.
Many plumes discharge at some distance away from the shoreline of lakes or large rivers.
Samples should be taken at locations where the elevation of the sediment-to-water interface corre-
sponds to the elevation of the contaminant plume in the aquifer. Many plumes are driven down into
aquifers by recharge. Conversely, the flow path bends sharply up underneath a gaining stream at the
point of discharge. Water just above the sediment in the center of a stream or small river should be
sampled. If possible, the stage of a stream or river at a gauging station near the point of sampling
should be determined to estimate the discharge of the stream or river at the time of sampling. Losing
streams or rivers should not be sampled at high stage when they are losing water because groundwa-
ter plumes would be pushed away from the sediment interface. To ensure that the stream is not
losing, the elevation of standing water in monitoring wells near the river should be higher than the
stage of the river or stream at the time of sampling. The same considerations apply to tidal environ-
ments or areas with wind seiches on large bodies of water. Surface water should be sampled when
the tide is out, or the wind is blowing off-shore. Additionally, contaminant plumes may be deflected
strongly downstream by flow occurring within the saturated material surrounding the surface water
channel. This is particularly true when the hydraulic conductivity of the stream sediments is much
greater than the hydraulic conductivity of the surrounding material that supplies ground water to the
stream. A great deal of thought as to when and where to sample is necessary to yield meaningful
results.
A.5.2 SEDIMENT COLLECTION
Sediment samples below the water surface can be collected using a core barrel. The core barrel
can be hand driven to the desired depth from a boat, then pulled back up using a mechanical jack
after sampling is finished. An alternative technique is to place open-end, two-inch diameter PVC
tubing to a desired depth, then insert flexible tubing and collect the sediment as a slurry into a suc-
tion flask connected to a peristaltic pump.
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SECTION A-6
This section describes the handling of soil and ground-water samples from the time of sampling
until the samples arrive at the laboratory.
A.6.1 SAMPLE AND
Sample containers and appropriate container lids must be purchased or provided by the analyti-
cal laboratory. Any required chemical preservatives can be added to the sample containers by the
analytical laboratory prior to shipping the containers to the site or alternatively, at the time of sam-
pling. The sample containers should be filled and tightly sealed in accordance with accepted proce-
dures for the sample matrix and the type of analysis to be conducted. The sample label should be
firmly attached to the container side, and the following information legibly and indelibly written on
the label:
« Facility name;
* Sample identification;
« Sample type (groundwater, surface water, etc.);
* Sampling date;
« Sampling time;
* Preservatives added; and
« Sample collector's initials.
A.6.2 SAMPLE SHIPMENT
After the samples are sealed and labeled, they should be packaged for transport to the analytical
laboratory. The packaged samples should be delivered to the analytical laboratory shortly after
sample acquisition using an overnight delivery service. The following packaging and labeling
procedures are to be followed:
* Abide by all U.S. Department of Transportation (DOT) shipping regulations;
« Package samples so that they will not leak, spill, or vaporize from their containers;
* Place samples in a cooler containing ice to maintain a shipping temperature of approxi-
mately 4 degrees centigrade (°C), if required by the requested analyses;
* Include a properly completed chain-of-custody form, as described in the following subsec-
tion; and
* Label shipping container with
- Sample collector's name, address, and telephone number;
- Laboratory's name, address, and telephone number;
- Description of sample;
- Quantity of sample; and
- Date of shipment.
A.6.3 CHAIN-OF-CUSTODY CONTROL
After the samples are collected, chain-of-custody procedures must be followed to establish a
written record of sample handling and movement between the sampling site and the analytical
laboratory. Each shipping container should include a chain-of-custody form completed in triplicate
by the sampling personnel. One copy of this form should be kept by the sampling contractor after
sample delivery to the analytical laboratory; the other two copies should be retained at the labora-
tory. One of the laboratory copies will become a part of the permanent record for the sample and
will be returned with the sample analytical results. The chain-of-custody form should contain the
following information:
A6-24
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« Unique sample identification number;
* Sample collector's printed name and signature;
« Date and time of collection;
* Sample location;
« Sample matrix;
* Sample size and container;
« Chemical preservatives added;
* Analyses requested;
« Signatures of individuals involved in the chain of possession; and
* Inclusive dates of possession.
The chain-of-custody documentation should be placed inside the shipping container so that it
will be immediately apparent to the laboratory personnel receiving the container, but cannot be
damaged or lost during transport. The shipping container is to be sealed so that it will be obvious if
the seal has been tampered with or broken.
A.6.4 SAMPLING RECORDS
In order to provide complete documentation of the sampling event, detailed records must be
maintained by the field scientist. At a minimum, these records must include the following informa-
tion:
* Sample location (facility name);
« Sample identification;
* Sample location map or detailed sketch;
« Date and time of sampling;
* Sampling method;
« Field observations of
- Sample appearance,
- Sample odor;
* Weather conditions;
« Water level prior to purging (ground-water samples);
* Total well depth (ground-water samples);
« Purge volume (ground-water samples);
* Water level after purging (ground-water samples);
« Well condition (ground-water samples);
* Sample depth;
« Sampler's identification;
* Field measurements such as pH, temperature, specific conductivity, dissolved oxygen
concentration, and redox potential (ground-water samples); and
* Any other relevant information.
A6-25
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SECTION A-7
Adequate characterization of the ground-water flow and contaminant transport system is an
important component of the monitored natural attenuation demonstration. The following sections
describe methodologies that are recommended to characterize the hydrogeologic system.
A.7.1 HYDRAULIC CONDUCTIVITY
Hydraulic conductivity is a measure of an aquifer's capacity to transmit water and governs
ground-water flow and contaminant transport in the subsurface. Methods for determining hydraulic
conductivity in the field can include slug tests, pumping tests, and downhole flowmeter measure-
ments. Hydraulic conductivity can also be measured during penetration with a cone penetrometer by
measuring the transient pressure excursions in the pore water in front of the cone using a cone
equipped with a pressure transducer in contact with the pore water. The method selected for a given
site will depend on the dimensions, locations, and screened intervals of site wells and monitoring
points; site stratigraphy; equipment availability; budget; and waste handling requirements.
A.7.1.1 Pump Tests
A pumping test involves pumping one well at a constant rate for a specified length of time and
collecting periodic water level measurements in both the pumped well and nearby observation wells
in order to determine aquifer hydraulic characteristics representative of a large area. As a rule,
pumping tests provide more representative measurements of hydraulic parameters; however, they
require a greater commitment of resources (time, money, and equipment) that cannot be afforded by
all projects. In addition, for pumping test results to be representative, site hydrogeologic conditions
should not change appreciably over short distances. This section outlines methods that can be used
for conducting pump tests in both confined and unconfined aquifers. For a more detailed discussion
of how to conduct a pumping test, the reader is referred to the work of Dawson and Istok (1991),
Kruseman and de Ridder (1991), and Driscoll (1986).
The interpretation of aquifer pumping test data is not unique. Similar sets of data can be ob-
tained from various combinations of geologic conditions. The interpretation of pumping test data is
discussed in Appendix C of this protocol document.
A.7.1.1.1 Pumping Test Design
Prior to performing an aquifer pumping test, all available site and regional hydrogeologic
information should be assembled and evaluated. Such data should include ground-water flow direc-
tion, hydraulic gradients, other geohydraulic properties, site stratigraphy, well construction details,
regional water level trends, and the performance of other pumping wells in the vicinity of the test
area. This information is used to select test duration, proposed pumping rates, and pumping well and
equipment dimensions.
The precise location of an aquifer test is chosen to be representative of the area under study. In
addition, the location is selected on the basis of numerous other criteria, including:
* Size of the investigation area;
« Uniformity and homogeneity of the aquifer;
* Distribution of contaminant sources and dissolved contaminant plumes;
« Location of known or suspected recharge or barrier boundary conditions;
* Availability of pumping and/or observation wells of appropriate dimension and screened at
the desired depth; and
* Requirements for handling discharge.
The dimensions and screened interval of the pumping well must be appropriate for the tested
aquifer. For example, the diameter of the well must be sufficient to accommodate pumping equip-
A7-26
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ment capable of sustaining the desired flow rate at the given water depth. In addition, if testing a
confined aquifer that is relatively thin, the pumping well should be screened for the entire thickness
of the aquifer. For an unconfined aquifer, the wells should be screened in the bottom one-third or
two-thirds of the saturated zone.
Any number of observation wells may be used. The number chosen is contingent upon both
cost and the need to obtain the maximum amount of accurate and reliable data. If three or more
observation wells are to be installed, and there is a known boundary condition, the wells should be
placed along a radial line extending from the pumping well toward the boundary, with one well
placed perpendicular to the line of observation wells to determine whether radial anisotropy exists
within the aquifer. If two observation wells are to be installed, they should be placed in a triangular
pattern, non-equidistant from the pumping well. Observation wells should be located at distances
and depths appropriate for the planned method for analysis of the aquifer test data. Observation well
spacing should be determined based upon expected drawdown conditions that are the result of the
studies of geohydraulic properties, proposed pumping test duration, and proposed pumping rate.
Preliminary pumping results should also be used (if available). Not all projects can afford the luxury
of preliminary testing.
The equipment needed to perform aquifer pumping tests includes:
• Pumps • Conductivity meter, pH meter, and thermometer
* Gate valve « Barometer
• Electrical generator * Semi-log and log-log graph paper
* Flow meter with totalizer « Portable computer
• Water level indicators « Field printer for data
* Pressure gauge « Type matching curves
• Field logbook/forms * Meter and stopwatch for discharge measurement
* Pressure transducers and data recorder « Hose or pipe for transfer of water
• Engineer's tape calibrated to 0.01 ft * Adequately sized tank for storing contaminated
* 5-gallon pail water
Pumping equipment should conform to the size of the well and be capable of delivering the
estimated range of pumping rates. The selection of flow meter, gate valve, and water transfer lines
should be based on anticipated rates of water discharge. Both the discharge rate and test duration
should be considered when selecting a tank for storing discharge water if the water cannot be re-
leased directly to the ground, sanitary sewer, storm sewer, or nearby water treatment facility.
In areas of severe winter climates, where the frost line may extend to depths of several feet,
pumping tests should be avoided during cold weather months where the water table is less than 12
feet from the surface. Under certain conditions, the frozen soil acts as a confining stratum, and
combined with leaky aquifer and delayed storage characteristics, test results may be unreliable.
A.7.1.1.2 Preparation for Testing
Barometric changes may affect water levels in wells, particularly in semiconfined and confined
aquifers. A change in barometric pressure may cause a change in the water level. Therefore, for at
least 24 hours prior to performing a pumping test, barometric pressure and water levels in the test
well, observation wells, and a well beyond the influence of the pumping well should be measured
hourly to establish trends in ground-water level fluctuation. If a trend is apparent, the barometric
pressure should be used to develop curves depicting the change in water level versus time. These
curves should be used to correct the water levels observed during the pumping test. Ground-water
levels in the background well as well as barometric pressures should continue to be recorded
throughout the duration of the test.
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Test wells should undergo preliminary pumping or step drawdown tests prior to the actual test.
This will enable fines to be flushed from the adjacent formation near the well and a steady flow rate
to be established. The preliminary pumping should determine the maximum drawdown in the well
and the proper pumping rate should be determined by step drawdown testing. The aquifer should
then be given time to recover before the actual pumping test begins (as a rule-of-thumb, one day).
A record should be maintained in the field logbook of the times of pumping and discharge of
other wells in the area, and if their radii of influence intersect the cone of depression of the test well.
All measurements and observations should be recorded in a field notebook or on an Aquifer Test
Data Form. If data loggers with transducers are used, field measurements should be performed in
case of data logger malfunction.
A.7.1.1.3 Conducting the Pumping Test
Immediately prior to starting the pump, the water levels should be measured and recorded for all
wells to determine the static water levels upon which all drawdowns will be based. Data loggers
should be reset for each well to a starting water level of 0.0 foot.
Water pumped from an unconfined aquifer during a pumping test should be disposed of in such
a manner as not to allow the aquifer to be recharged by infiltration during the test. This means that
the water must be piped away from the well and associated observation wells. Recharge could
adversely affect the results. Also, if contaminated water is pumped during the test, the water must be
stored and treated or disposed of according to the project work plan for the study. The discharge
water may be temporarily stored in drums, a lined, bermed area, or tanks. If necessary, it should be
transported and staged in a designated secure area.
The discharge rate should be measured frequently throughout the test and controlled to maintain
it as constant as possible, after the initial excess discharge has been stabilized. This can be achieved
by using a control valve.
The pitch or rhythm of the pump or generators provides a check on performance. If there is a
sudden change in pitch, the discharge should be checked immediately and proper adjustments to the
control valve or the engine speed should be made, if necessary. Do not allow the pump to break
suction during the test. Allow for maximum drawdown of the well during the step drawdown test. If
done properly, the flow control valve can be pre-set for the test and will not have to be adjusted
during pumping. If the pump does shut down during the test, make necessary adjustments and restart
the test after the well has stabilized. For a confined aquifer, the water level in the pumping well
should not be allowed, if possible, to fall below the bottom of the upper confining stratum during a
pumping test.
At least 10 measurements of drawdown for each log cycle of time should be made both in the
test well and the observation wells. Data loggers can be set to record in log time, which is very
useful for data analysis. A suggested schedule for recording water level measurements made by
hand is as follows:
• Oto 10 minutes-0.5, 1.0,2.5,2.0,2.5, 3.0,4.5, 6.5, 8, and 10 minutes. It is important in
the early part of the test to record with maximum accuracy the time at which readings are
taken.
• 10 to 100 minutes - 10, 15, 20, 25, 30, 40, 50, 65, 80, and 100 minutes.
« Then, at 1-hour intervals from 120 minutes to 1,440 minutes (one day) and every 2 hours
after 1 complete day.
Initially, there should be sufficient field personnel to station one person at each well used in the
pumping test (unless an automatic water-level recording system has been installed). After the first
two hours of pumping, two people are usually sufficient to complete the test. A third person may be
needed when treatment of the pumped water is required prior to discharge. It is advisable for at least
A7-28
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one field member to have experience in the performance of pump tests, and for all field personnel to
have a basic familiarity with conducting the test and gathering data.
Field personnel should be aware that electronic equipment sometimes fails in the field. Some
field crews have experienced complete loss of data due to failure of a logger or transducer. It is a
good idea to record data in the field logbook or on a manual form as the data are produced. That
way, the data are not lost should the equipment fail.
The discharge or pumping rate should be measured with a flow meter that also has a totalizer.
When the pumping is complete, the total gallons pumped are divided by the time of pumping to
obtain the average discharge rate for the test. Periodic checking and recording of the pumping rate
during the test also should be performed.
The total pumping time for a test depends on the type of aquifer and degree of accuracy desired.
Economizing on the duration of pumping is not recommended. More reliable results are obtained if
pumping continues until the cone of depression achieves a stabilized condition. The cone of depres-
sion will continue to expand at an ever-decreasing rate until recharge of the aquifer equals the pump-
ing rate, and a steady-state condition is established. The time required for steady-state flow to occur
may vary from a few hours to years.
Under normal conditions, it is a good practice to continue a pumping test in a confined aquifer
for at least 24 hours, and in an unconfined aquifer for a minimum of 72 hours. A longer duration of
pumping may reveal the presence of boundary conditions or delayed yield. Use of portable comput-
ers allows time/drawdown plots to be made in the field. If data loggers are used to monitor water
levels, hard copies of the data printed on field printers should be obtained before transporting the
logger back to the office for downloading.
A.7.1.2 Slug Tests
A slug test is a single-well hydraulic test used to determine the hydraulic conductivity of an
aquifer in the immediate vicinity of the well. Because hydraulic conductivity varies spatially within
and between aquifers and because slug test results reflect aquifer conditions only in the immediate
vicinity of the tested well, slug tests should be conducted in as many wells as possible at a site. Slug
tests can be used for both confined and unconfined aquifers that have transmissivities of less than
approximately 7,000 square feet per day (ft2/day). Slug tests are accomplished by removing a solid
slug (rising head) or introducing a solid slug (falling head), and then allowing the water level to
stabilize while taking water level measurements at closely spaced time intervals. The method pre-
sented herein discusses the use of falling head and rising head slug tests in sequence. The analysis of
slug test data is discussed in Appendix C.
Slug testing should not proceed until water level measurements show that static water level
equilibrium has been achieved. Unvented wells should be uncapped at least 24 hours prior to initiat-
ing the test in order to allow the static water level to come to equilibrium. The protective casing
should remain locked during this time to prevent vandalism. During the slug test, the water level
change should be influenced only by the introduction or removal of the slug volume. Other factors,
such as inadequate well development or extended pumping, may lead to inaccurate results. It is the
field scientist's responsibility to decide when static equilibrium has been reached in the well.
The following equipment is needed to conduct a slug test:
• Teflon®, PVC, or metal slug
• Nylon or polypropylene rope
• Electric water level indicator
• Pressure transducer/sensor
* Field logbook/forms
• Automatic data recorder (such as the Hermit Environmental Data Logger®, In-Situ, Inc.
Model SE1000B, or equal)
A7-29
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The falling head test is the first step in the two-step slug-testing procedure. The following steps
describe the recommended falling head slug test procedure:
I. Decontaminate all downhole equipment.
2. Record pre-test information including: well number, personnel, climatic data, ground surface
elevation, measuring point elevation, equipment identifications, and date.
3. Measure and record the static water level in the well to the nearest 0.01 foot.
4. Lower the decontaminated pressure transducer into the well and allow the displaced water to
return to within 0.01 foot of the original static level.
5. Lower the decontaminated slug into the well to just above the water surface in the well.
6. Start the data logger and quickly lower the slug below the water table being careful not to
disturb the pressure transducer. Follow the owner's manual for proper operation of the data
logger.
oo
7. Terminate data recording when the water level has recovered at least 80 percent from the
initial slug displacement.
Immediately following completion of the falling head test, the rising head test is performed.
The following steps describe the rising head slug test procedure:
1. Measure the static water level in the well to the nearest 0.01 foot to ensure that it has returned
to the static water level.
2. Initiate data recording and quickly withdraw the slug from the well. Follow the owner's
manual for proper operation of the data logger.
3. Terminate data recording when the water level has recovered at least 80 percent from the
initial slug displacement.
It is advisable to produce hard copies or backup electronic copies of the data logger output (draw-
down vs. time) daily and before transporting the logger from the field site.
A.7.1.3 Downhole Flow Meter Measurements
Downhole flow meter measurements are used to investigate the relative vertical distribution of
horizontal hydraulic conductivity in an open borehole or the screened portion of a well. These
measurements are useful for identifying zones of elevated hydraulic conductivity that may contribute
to preferential flow pathways and affect contaminant migration. Methodologies for interpreting data
from borehole surveys are described by Molz et al. (1994).
Flowmeter measurements should be performed at 1- to 3-foot intervals in test wells during both
ambient conditions and induced flow conditions. Test data may be analyzed using the methods
described by Molz el al. (1994) to define the relative distribution of horizontal hydraulic conductiv-
ity within the screened interval of each well. Final results should be presented in tabular and graphi-
cal forms and accompanied by appropriate interpretation and discussion. Estimates of bulk hydraulic
conductivity from previous aquifer tests or results of single-well tests conducted in conjunction with
the flow meter survey can be used to estimate the absolute hydraulic conductivity distribution at each
well.
Borehole flowmeters should be calibrated prior to testing. Generally, 0.5-inch-ID and 1.0-inch-
ID probes will be calibrated using a range of volumetric flowrates potentially applicable to most sites
[e.g., approximately 0.04 liters per minute (L/min) to 10 L/min]. The following nine steps outline
general procedures that can be used to conduct a downhole flow meter survey at a given location.
* Measure the water level, organic liquid (NAPL) interfaces (if present), and total depth
(TD) prior to initiating the test.
* Calibrate the flow meter for the range of anticipated flow velocities before introducing the
flow meter into the well or borehole.
* Lower the flow meter to the bottom of the well/borehole.
A7-30
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« Slowly withdraw the flow meter, pausing to obtain measurements at intervals of approxi-
mately 1 to 3 feet, depending on site conditions. This will provide a baseline under static
(ambient) conditions.
* Conduct a short-term, single-well pumping test in the test well to stress the aquifer.
« Record drawdown using an electronic data logger with a pressure transducer.
* Monitor and adjust the ground-water extraction rate, as necessary, to maintain constant
flow.
* Obtain the profile of the vertical flow at the same elevations occupied during the ambient
profile upon stabilization of the flow rate.
* Analyze the data collected during the tests to estimate relative distribution of flow into the
tested wells and the relative hydraulic conductivity distribution at each location (Molz et
al., 1994).
A.7.2 HYDRAULIC
Hydraulic gradient, defined as the change in ground-water elevation with distance, is a key
parameter governing the direction and rate of ground-water flow and contaminant migration. Be-
cause ground water can flow in both the horizontal and vertical planes, both horizontal and vertical
gradients are required for a successful demonstration of monitored natural attenuation. Hydraulic
gradients are generally calculated on the basis of ground-water elevations measured in site monitor-
ing wells or monitoring points using an electric water level indicator. Therefore, for the most com-
plete representation of site hydrogeology, it is important to measure ground-water elevations from as
many depths and locations as available. Interpretation of ground-water elevations and the subse-
quent calculations for hydraulic gradient are discussed in Appendix C.
A.7.3 DIRECT OF VELOCITY
Ground-water velocity is directly related to contaminant velocity; therefore, a determination of
groundwater velocity is critical to the fate and transport portion of a demonstration of monitored
natural attenuation. Typically, ground-water velocity is estimated from the hydraulic conductivity,
hydraulic gradient, and effective porosity as described in Appendix C; however, direct measurement
of ground-water velocity can be obtained from dye tracer studies.
A7-31
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APPENDIX B
AFFECTING AND OF
IN THE
-------
OF CONTENTS - B
B-l INTRODUCTION Bl-6
B.I.I FATE AND TRANSPORT MECHANISMS Bl-6
B. 1.2 MATHEMATICAL DESCRIPTION OF SOLUTE FATE
AND TRANSPORT Bl-6
B-2 NONDESTRUCTIVE ATTENUATION MECHANISMS B2-9
B.2.1 ADVECTION B2-9
B.2.2 HYDRODYNAMIC DISPERSION B2-9
B.2.2.1 Mechanical Dispersion B2-11
B.2.2.2 Molecular Diffusion B2-13
B.2.2.3 Equation of Hydrodynamic Dispersion B2-13
B.2.2.4 One-Dimensional Advection-Dispersion Equation B2-14
B.2.3 SORPTION B2-15
B.2.3.1 Mechanisms of Sorption B2-16
B.2.3.2 Sorption Models and Isotherms B2-17
B.2.3.2.1 Langmuir Sorption Model B2-17
B.2.3.2.2 Freundlich Sorption Model B2-18
B.2.3.3 Distribution Coefficient B2-19
B.2.3.4 Coefficient of Retardation B2-20
B.2.3.4.1 Determining the Coefficient of Retardation using K B2-20
° ° oc
B.2.3.4.2 Determining the Coefficient of Retardation using
Laboratory Tests B2-24
B.2.3.5 One-Dimensional Advection-Dispersion Equation with Retardation B2-25
B.2.4 VOLATILIZATION B2-26
B.2.5 RECHARGE B2-26
B-3 DESTRUCTIVE ATTENUATION MECHANISMS - BIOLOGICAL B3-29
B.3.1 OVERVIEW OF B1ODEGRADATION B3-30
B.3.2 BIODEGRADATION OF ORGANIC COMPOUNDS VIA USE
AS A PRIMARY GROWTH SUBSTRATE B3-33
B.3.2.1 Aerobic Biodegradation of Primary Substrates B3-33
B.3.2.1.1 Aerobic Oxidation of Petroleum Hydrocarbons B3-35
B.3.2.1.2 Aerobic Oxidation of Chlorinated Ethenes B3-35
B.3.2.1.3 Aerobic Oxidation of Chlorinated Ethanes B3-35
B.3.2.1.4 Aerobic Oxidation of Chlorobenzenes B3-36
B.3.2.2 Anaerobic Biodegradation of Primary Substrates B3-36
B.3.2.2.1 Anaerobic Oxidation of Petroleum Hydrocarbons B3-36
B.3.2.2.2 Anaerobic Oxidation of Chlorinated Ethenes B3-36
B.3.2.2.3 Anaerobic Oxidation of Chlorinated Ethanes B3-36
B.3.2.2.4 Anaerobic Oxidation of Chlorobenzenes B3-37
B.3.3 BIODEGRADATION OF ORGANIC COMPOUNDS VIA USE AS AN
ELECTRON ACCEPTOR (REDUCTIVE DECHLORINATION) B 3 -3 7
B.3.3.1 Reductive Dechlorination of Chlorinated Ethenes B3-38
Bl-2
-------
B.3.3.2 Reductive Dechlorination of Chlorinated Ethanes B3-40
B.3.3.3 Reductive Dechlorination of Chlorobenzenes B3-40
B.3.4 BIODEGRADATION OF ORGANIC COMPOUNDS VIA
COMETABOLISM B3-40
B.3.5 THERMODYNAMIC CONSIDERATIONS B3-41
B.3.6 ONE-DIMENSIONAL ADVECTION-DISPERSION EQUATION WITH
RETARDATION AND BIODEGRADATION B3-59
B-4 DESTRUCTIVE ATTENUATION MECHANISMS - ABIOTIC B4-60
B.4.1 HYDROLYSIS AND DEHYDROHALOGENATION B4-60
B.4.1.1 Hydrolysis B4-60
B.4.1.2 Dehydrohalogenation B4-61
B.4.2 REDUCTION REACTIONS B4-63
Bl-3
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FIGURES
No. Title Page
B.2.1 Breakthrough curve in one dimension showing plug flow with
continuous source resulting from advection only B2-10
B.2.2 Breakthrough curve in one dimension showing plug flow with
instantaneous source resulting from advection only B2-10
B.2.3 Plume migration in two dimensions (plan view) showing plume
migration resulting from advective flow only with continuous and
instantaneous source B2-10
B.2.4 Physical processes causing mechanical dispersion at the microscopic scale B2-11
B.2.5 Breakthrough curve in one dimension showing plug flow with
instantaneous source resulting from advection only and the combined
processes of advection and hydrodynamic dispersion B2-12
B.2.6 Breakthrough curve in one dimension showing plug flow with
instantaneous source resulting from advection only and the combined
processes of advection and hydrodynamic dispersion B2-12
B.2.7 Relationship between dispersivity and scale B2-15
B.2.8 Breakthrough curve in one dimension showing plug flow with
continuous source resulting from advection only; the combined
processes of advection and hydrodynamic dispersion; and the combined
processes of advection, hydrodynamic dispersion, and sorption B2-16
B.2.9 Breakthrough curve in one dimension showing plug flow with
instantaneous source resulting from advection only; the combined
processes of advection and hydrodynamic dispersion; and the combined
processes of advection, hydrodynamic dispersion, and sorption B2-16
B.2.10 Characteristic adsorption isotherm shapes B2-18
B.2.11 Plot of sorbed concentration versus equilibrium concentration B2-25
B.3.1 Breakthrough curve in one dimension showing plug flow with
continuous source resulting from advection only; the combined
processes of advection and hydrodynamic dispersion; the combined
processes of advection, hydrodynamic dispersion, and sorption; and
the combined processes of advection, hydrodynamic dispersion,
sorption, and biodegradation B3-30
B.3.2 Breakthrough curve in one dimension showing plug flow with
instantaneous source resulting from advection only; the combined
processes of advection and hydrodynamic dispersion; the combined
processes of advection, hydrodynamic dispersion, and sorption; and
the combined processes of advection, hydrodynamic dispersion,
sorption, and biodegradation B3-30
B.3.3 Oxidation-reduction potentials for various oxidation-reduction reactions B3-34
B.3.4 Expected sequence of microbially-mediated redox reactions and
Gibbs free energy of the reaction B3-42
Bl-4
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TABLES
No. Title Page
B.I.I Summary of Important Processes Affecting Solute Fate and Transport Bl-7
B.2.1 Values of Aqueous Solubility and Koc for Selected Chlorinated Compounds B2-22
B.2.2 Values of Aqueous Solubility and K „ for BTEX and Trimethylbenzene Isomers B2-23
B.2.3 Data from Hypothetical Batch Test Experiment B2-25
B.2.4 Henry's Law Constants and Vapor Pressures for Common Fuel Hydrocarbons
and Chlorinated Solvents B2-27
B.3.1 Biologic and Abiotic Degradation Mechanisms for Various
Anthropogenic Organic Compounds B3-29
B.3.2 Some Microorganisms Capable of Degrading Organic Compounds B3-31
B.3.3 Trends in Contaminant, Electron Acceptor, Metabolic By-product, and Total
Alkalinity Concentrations During Biodegradation B3-34
B.3.4 Sources, Donors, Acceptors, and Products of Reported Reductive
Dechlorinating Laboratory Systems B3-39
B.3.5 Electron Donor and Electron Acceptor Half Cell Reactions B3-43-- B3-44
B.3.6 Gibbs Free Energy of Formation for Species used in Half Cell Reactions
and Coupled Oxidation-Reduction Reactions B3-45—B3-46
B.3.7 Coupled Oxidation-Reduction Reactions B3-47-B3-58
B.4.1 Approximate Half-Lives of Abiotic Hydrolysis and Dehydrohalogenation
Reactions Involving Chlorinated Solvents B4-62
Bl-5
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SECTION B-l
INTRODUCTION
B.1.1 FATE
This appendix presents an overview of the important processes affecting the fate and transport
of chlorinated solvents and fuel hydrocarbons dissolved in ground water. The environmental fate
and transport of a contaminant is controlled by the compound's physical and chemical properties and
the nature of the subsurface media through which the compound is migrating. Several processes are
known to cause a reduction in the concentration and/or mass of a contaminant dissolved in ground
water. Those processes that result only in the reduction of a contaminant's concentration but not of
the total contaminant mass in the system are termed "nondestructive." Those processes that result in
degradation of contaminants are referred to as "destructive." Nondestructive processes include
advection, hydrodynamic dispersion (mechanical dispersion and diffusion), sorption, dilution, and
volatilization. Destructive processes include biodegradation and abiotic degradation mechanisms.
Biodegradation may be the dominant destructive attenuation mechanism acting on a contaminant,
depending upon the type of contaminant and the availability of electron donors or carbon sources.
Abiotic degradation processes are also known to degrade chlorinated solvents; where biodegradation
is not occurring, these may be the only destructive processes operating. However, the rates of abiotic
processes are generally slow relative to biodegradation rates.
Remediation by monitored natural attenuation results from the integration of all the subsurface
attenuation mechanisms (both nondestructive and destructive) operating at a given site. Table B.1.1
summarizes the processes that affect fate and transport of chlorinated solvents and fuel hydrocarbons
dissolved in ground water. Important factors to consider include:
« The compound's soil/water distribution coefficient (Kd);
* The compound's organic carbon/water partition coefficient (Koc);
« The compound's octanol/water partition coefficient (Kow);
* The compound's water solubility;
« The compound's vapor pressure;
* The compound's Henry's Law constant (air/water partition coefficient, H);
« Indigenous bacterial population;
* Hydraulic conductivity of aquifer materials;
« Porosity of aquifer materials;
* Total organic carbon content of aquifer materials;
« Bulk density of aquifer materials;
* Aquifer heterogeneity; and
« Ambient ground-water geochemistry.
Nondestructive attenuation mechanisms are discussed in Section B-2. Biodegradation is dis-
cussed in Section B-3. Abiotic degradation mechanisms are discussed in Section B-4. It is impor-
tant to separate nondestructive from destructive attenuation mechanisms during the natural attenua-
tion demonstration. The methods for correcting apparent attenuation caused by nondestructive
attenuation mechanisms are discussed in Appendix C.
B.1.2 MATHEMATICAL OF AND
The partial differential equation describing contaminant migration and attenuation in the
saturated zone includes terms for advection, dispersion, sorption, and degradation. In one dimen-
sion, the partial differential equation describing solute transport in the saturated zone is:
D -N 2 eq.
R dx R dx '
Bl-6
-------
Table B. 1.1 Summary of Important Processes Affecting Solute Fate and Transport
Process
Description
Dependencies
Effect
Advection
Movement of solute by bulk
ground-water movement.
Dependent on aquifer properties,
mainly hydraulic conductivity and
effective porosity, and hydraulic
gradient. Independent of contaminant
properties.
Main mechanism driving
contaminant movement in the
subsurface.
Dispersion
Fluid mixing due to ground-
water movement and aquifer
heterogeneities.
Dependent on aquifer properties and
scale of observation. Independent of
contaminant properties.
Causes longitudinal, transverse,
and vertical spreading of the
plume. Reduces solute
concentration.
Diffusion
Spreading and dilution of
contaminant due to molecular
diffusion.
Dependent on contaminant properties
and concentration gradients.
Described by Pick's Laws.
Diffusion of contaminant from
areas of relatively high
concentration to areas of relatively
low concentration. Generally
unimportant relative to dispersion
at most ground-water flow
velocities.
Sorption
Reaction between aquifer matrix
and solute whereby relatively
hydrophobic organic compounds
become sorbed to organic
carbon or clay minerals.
Dependent on aquifer matrix
properties (organic carbon and clay
mineral content, bulk density, specific
surface area, and porosity) and
contaminant properties (solubility,
hydrophobicity, octanol-water
partitioning coefficient).
Tends to reduce apparent solute
transport velocity and remove
solutes from the ground water via
sorption to the aquifer matrix.
Recharge
(Simple Dilution)
Movement of water across the
water table into the saturated
zone.
Dependent on aquifer matrix
properties, depth to ground water,
surface water interactions, and
climate.
Causes dilution of the contaminant
plume and may replenish electron
acceptor concentrations, especially
dissolved oxygen.
Volatilization
Volatilization of contaminants
dissolved in ground water into
the vapor phase (soil gas).
Dependent on the chemical's vapor
pressure and Henry's Law constant.
Removes contaminants from
ground water and transfers them to
soil gas.
Biodegradation
Microbially mediated oxidation-
reduction reactions that degrade
contaminants.
Dependent on ground-water
geochemistry, microbial population
and contaminant properties.
Biodegradation can occur under
aerobic and/or anaerobic conditions.
May ultimately result in complete
degradation of contaminants.
Typically the most important
process acting to truly reduce
contaminant mass.
Abiotic Degradation
Chemical transformations that
degrade contaminants without
microbial facilitation; only
halogenated compounds are
subject to these mechanisms in
the ground-water environment.
Dependent on contaminant properties
and ground-water geochemistry.
Can result in partial or complete
degradation of contaminants.
Rates typically much slower than
for biodegradation.
Partitioning from
NAPL
Partitioning from NAPL into
ground water. NAPL plumes,
whether mobile or residual, tend
to act as a continuing source of
ground-water contamination.
Dependent on aquifer matrix and
contaminant properties, as well as
ground-water mass flux through or
past NAPL plume.
Dissolution of contaminants from
NAPL represents the primary
source of dissolved contamination
in ground water.
Bl-7
-------
Where:
C = solute concentration [M]
t = time [T]
Dx = hydrodynamic dispersion [L2/T]
R = coefficient of retardation [dimensionless]
jc = distance along flow path [L]
vx = transport velocity in x direction [L/T]
<9v = general source or sink term for reactions involving the
production or loss of solute (e.g., biodegradation) [M/L3/T]
The degradation of organic contaminants commonly can be approximated using first-order
kinetics. In one dimension, the partial differential equation describing solute transport with first-
order decay in the saturated zone is given by:
ac Dxd2c vxdc ._
37 = ~H"3~2 D3 AC eq. B.1.2
at R ax R ax n
Where:
C = concentration [M/L3]
t = time [T]
Dx = hydrodynamic dispersion [L2/T]
x = distance along flow path [L]
R = coefficient of retardation [dimensionless]
vv = transport velocity in x direction [L/T]
X = first-order decay rate [T1]
These equations serve to illustrate how the processes of advection, dispersion, sorption, and
biotic and abiotic degradation are integrated to describe the fate and transport of solutes in the
saturated zone. These relationships were derived using the continuity (conservation of mass) equa-
tion, which states that the rate of change of contaminant mass within a unit volume of porous media
is equal to the flux of contaminant into the unit volume minus the flux out of the unit volume (Freeze
and Cherry, 1979). Processes governing flux into the unit volume include advection and hydrody-
namic dispersion (including mechanical dispersion and diffusion). Processes governing flux out of
the unit volume include advection, hydrodynamic dispersion, dilution, sorption, and chemical reac-
tions (most notably biodegradation). The change in solute concentration may, therefore, be stated
mathematically as:
Change in Solute Concentration = Flux In - Flux Out ± Reactions
The following sections describe the most significant reactions affecting this mass balance (and
therefore the fate and transport) of organic contaminants in the subsurface. Methods for evaluating
the flux through the system will be discussed in Appendix C.
Bl-8
-------
SECTION B-2
NONDESTRUCTIVE ATTENUATION
B.2.1 ADVECTION
Advective transport is the transport of solutes by the bulk movement of ground water. Advec-
tion is the most important process driving dissolved contaminant migration in the subsurface. The
linear groundwater velocity in the direction parallel to ground-water flow caused by advection is
given by:
KdH
V*=~^ eq-B-2J
Where:
vx = average linear velocity [L/T]
K = hydraulic conductivity [L/T]
ne = effective porosity [L3/L3]
dH/dL = hydraulic gradient [L/L]
Solute transport by advection alone yields a sharp solute concentration front. Immediately
ahead of the front, the solute concentration is equal to the background concentration (generally zero).
At and behind the advancing solute front, the concentration is equal to the initial contaminant con-
centration at the point of release. This is referred to as plug flow and is illustrated in Figures B.2.1,
B.2.2, and B.2.3. In reality, the advancing front spreads out due to the processes of dispersion and
diffusion, as discussed in Section B-3, and is retarded by sorption and biodegradation, as discussed
in Sections B-4 and B-5, respectively.
The one-dimensional advective transport component of the advection-dispersion equation is
given by:
dC dC
37 = ~v*3~~ eq. B.2.2
dt dx l
Where:
vv = average linear velocity [L/T]
C = contaminant concentration [M/L3]
t = time [T]
x = distance along flow path [L]
Equation B.2.2 considers only advective transport of the solute. In some cases this may be a
fair approximation for simulating solute migration because advective transport is the main force
behind contaminant migration. However, because of dispersion, diffusion, sorption, and biodegrada-
tion, this equation generally must be combined with the other components of the modified advection-
dispersion equation (equation B.I.I) to obtain an accurate mathematical description of solute trans-
port.
HYDRODYNAMIC
Hydrodynamic dispersion is the process whereby a contaminant plume spreads out in directions
that are longitudinal and transverse to the direction of plume migration. Dispersion of organic
solutes in an aquifer is an important consideration when modeling remediation by natural attenua-
tion. Dispersion of a contaminant dilutes the concentrations of the contaminant, and introduces the
contaminant into relatively pristine portions of the aquifer where it may admix with more electron
acceptors crossgradient to the direction of ground-water flow. Two very different processes cause
B2-9
-------
Contaminant front with
/•"advection only
a C/C. 0.5 -
O
0.0
Figure B.2.1 Breakthrough curve in one dimension showing plug flow with continuous source resulting
from advection only.
«I
8
.3 a
U O
l.On
C/C0 0.5-
^Contaminant front with
advection only
0.0
Time or Distance from Source
Figure B.2.2 Breakthrough curve in one dimension showing plug flow with instantaneous source resulting
from advection only.
Ground-water Flow Direction
Source'
Continuous Source
Instantaneous Source
Figure B.2.3 Plume migration in two dimensions (plan view) showing plume migration resulting from
advective flow only with continuous and instantaneous sources.
B2-10
-------
hydrodynamic dispersion; mechanical dispersion and molecular diffusion. The variable describing
hydrodynamic dispersion, D, is the sum of mechanical dispersion and molecular diffusion. Mechani-
cal dispersion is the dominant mechanism causing hydrodynamic dispersion at normal ground-water
velocities. At extremely low ground-water velocities, molecular diffusion can become the dominant
mechanism of hydrodynamic dispersion. Molecular diffusion is generally ignored for most ground-
water studies. The following sections describe these processes and how they are incorporated into
the modified advection-dispersion equation (Equation B.I.I).
B.2.2.1 Mechanical Dispersion
As defined by Domenico and Schwartz (1990), mechanical dispersion is mixing that occurs as a
result of local variations in velocity around some mean velocity of flow. With time, a given volume
of solute will gradually become more dispersed as different portions of the mass are transported at
the differing velocities. In general, the main cause of variations of both rate and direction of trans-
port velocities is the heterogeneity of the porous aquifer medium. These heterogeneities are present
at scales ranging from microscopic (e.g., pore to pore) to macroscopic (e.g., well to well) to megas-
copic (e.g., a regional aquifer system).
Three processes are responsible for mechanical dispersion on the microscopic scale
(Figure B.2.4). The first process is the variation in flow velocity through pores of various sizes. As
ground water flows through a porous medium, it flows more slowly through large pores than through
smaller pores. The second cause of mechanical dispersion is tortuosity, or flow path length. As
ground water flows through a porous medium, some of the ground water follows less tortuous
(shorter) paths, while some of the ground water takes more tortuous (longer) paths. The longer the
flow path, the slower the average linear velocity of the ground water and the dissolved contaminant.
The final process causing mechanical dispersion is variable friction within an individual pore.
Groundwater traveling close to the center of a pore experiences less friction than ground water
traveling next to a mineral grain, and therefore moves faster. These processes cause some of the
contaminated ground water to move faster than the average linear velocity of the ground water and
some to move slower. This variation in average velocity of the solute causes dispersion of the
contaminant.
Small,
Fast
Pore Size
Tortuosity
Low,
Fast
Friction in
Pore Throat
Figure B.2.4 Physical processes causing mechanical dispersion at the microscopic scale.
Heterogeneity at the macroscopic and megascopic scales also creates variability in ground water
and solute velocities, therefore producing dispersion on a larger scale. Geologic features that con-
B2-11
-------
tribute to dispersion at the macroscopic scale include stratification characteristics such as changing
unit geometry, discontinuous units, and contrasting lithologies, and permeability characteristics such
as nonuniform permeability, directional permeability, and trending permeability (Domenico and
Schwartz, 1990). Even in aquifer material that appears to be homogeneous, relatively small changes
in the fraction of fine sediment can change hydraulic conductivity characteristics enough to produce
significant variations in fluid and solute velocities and thus introduce dispersion. Larger geological
features will introduce dispersion at the megascopic scale. At this scale, structural features such as
faults, dipping strata, folds, or contacts will create inhomogeneity, as will stratigraphic features such
as bedding or other depositional structures.
As a result of dispersion, the solute front travels at a rate that is faster than would be predicted
based solely on the average linear velocity of the ground water. The overall result of dispersion is
spreading and mixing of the contaminant plume with uncontaminated ground water. Figures B.2.5
and B.2.6 illustrate the effects of hydrodynamic dispersion on an advancing solute front. The com-
ponent of hydrodynamic dispersion contributed by mechanical dispersion is given by the relation-
ship:
Mechanical Dispersion = axvx eq. B.2.3
Where:
vx = average linear groundwater velocity [L/T]
ocx = dispersivity [L]
Mechanical dispersion has two components, longitudinal dispersion and transverse (both hori-
zontal and vertical) dispersion. Longitudinal dispersion is the spreading of a solute in a direction
parallel to the direction of ground-water flow. On the microscopic scale, longitudinal dispersion
,- Contaminant front with
// advection only
il
Ł a
c/c
l.U
n s
n n
^
/ Contaminant front
/ with advection and
/ hydrodynamic
y / dispersion
Distance from Source, x ^
Figure B.2.5 Breakthrough curve in one dimension showing plug flow with continuous source resulting from
advection. only and the combined processes of advection and hydrodynamic dispersion.
^Contaminant slug with
advection only
/ \
Initial
contaminant
slug
l.O
i
C/C,, 0.5 -
I Contaminant slug
with advection
and hydrodynamic
dispersion
Time or Distance from Source
Figure B. 2,6 Breakthrough curve in one dimension showing plug flow with instantaneous source resulting
from advection only and the combined processes of advection and hydrodynamic dispersion.
B2-12
-------
occurs because of velocity changes due to variations in pore size, friction in the pore throat, and
tortuosity. Transverse dispersion is the spreading of a solute in directions perpendicular to the
direction of ground-water flow. Transverse dispersion on the microscopic scale is caused by the
tortuosity of the porous medium, which causes flow paths to branch out from the centerline of the
contaminant plume.
B.2.2.2 Molecular
Molecular diffusion occurs when concentration gradients cause solutes to migrate from zones of
higher concentration to zones of lower concentration, even in the absence of ground-water flow.
Molecular diffusion is only important at low ground-water velocities, and therefore can be ignored in
areas with high ground-water velocities (Davis et al., 1993).
The molecular diffusion of a solute in ground water is described by Pick's Laws. Pick's First
Law applies to the diffusive flux of a dissolved contaminant under steady-state conditions and, for
the one-dimensional case, is given by:
Z7 T^C
F = -D— eq.B.2.4
dx
Where:
F = mass flux of solute per unit area of time [M/T]
D = diffusion coefficient (L2/T)
C = solute concentration (M/L3)
dC
~~T~ = concentration gradient (M/L3/L)
For systems where the dissolved contaminant concentrations are changing with time, Pick's Second
Law must be applied. The one-dimensional expression of Pick's Second Law is:
dC ^d2C
— = D—r eq. B.2.5
dt dx
Where:
dC
—— = change in concentration with time [M/T]
The process of diffusion is slower in porous media than in open water because the ions must
follow more tortuous flow paths (Fetter, 1988). To account for this, an effective diffusion coeffi-
cient, D*, is used.
The effective diffusion coefficient is expressed quantitatively as (Fetter, 1988):
D* = wD eq. B.2.6
Where:
w = empirical coefficient determined by laboratory experiments [dimensionless]
The value of w generally ranges from 0.01 to 0.5 (Fetter, 1988).
B.2.2.3 Equation of Hydrodynamic Dispersion
Hydrodynamic dispersion, D, has two components, mechanical dispersion and molecular
diffusion. For one-dimensional flow, hydrodynamic dispersion is represented by the following
equation (Freeze and Cherry, 1979):
Dx=axvx+D* eq. B.2.7
Where:
DX = longitudinal coefficient of hydrodynamic dispersion in the x direction [L2/T]
ocx = longitudinal dispersivity [L]
vx = average linear ground-water velocity [L/T]
D* = effective molecular diffusion [L2/T]
B2-13
-------
Dispersivity is a parameter that is characteristic of the porous medium through which the
contaminant migrates. Dispersivity represents the spreading of a contaminant over a given length of
flow, and therefore has units of length. It is now commonly accepted (on the basis of empirical
evidence) that as the scale of the plume or the system being studied increases, the dispersivity will
also increase. Therefore, dispersivity is scale-dependent, but at a given scale, data compiled by
Gelhar el al. (1985 and 1992) show that dispersivity may vary over three orders of magnitude. The
data of Gelhar elal. (1992) are presented on Figure B.2.7 (with permission from Newell et al., 1996).
Several approaches can be used to estimate longitudinal dispersivity, otx., on the field scale (i.e.,
macroscopic to megascopic scales). One technique involves conducting a tracer test. Although this
is potentially the most reliable method, time and monetary constraints can be prohibitive. Another
method commonly used to estimate dispersivity when implementing a solute transport model is to
start with a longitudinal dispersivity of 0.1 times the plume length (Lallemand-Barres and
Peaudecerf, 1978; Pickens and Grisak, 1981; Spitz and Moreno, 1996). This assumes that
dispersivity varies linearly with scale. However, Xu and Eckstein (1995) evaluated the same data
presented by Gelhar el al. (1992) and, by using a weighted least-squares method, developed the
following relationship for estimating dispersivity:
ax =0.83(LogwLP)2AU eq. B.2.8
Where:
otx = longitudinal dispersivity [L]
L = plume length [L]
Both relationships are shown on Figure B.2.7. In either case, the value derived for dispersivity
will be an estimate at best, given the great variability in dispersivity for a given plume length. How-
ever, for modeling studies, an initial estimate is needed, and these relationships provide good starting
points for a modeling study.
In addition to estimating longitudinal dispersivity, it may be necessary to estimate the transverse
and vertical dispersivities (ar and ocz., respectively) for a given site. Several empirical relationships
between longitudinal dispersivity and transverse and vertical dispersivity have been described.
Commonly, aT is estimated as 0.1 ocx. (based on data from Gelhar etal, 1992), or as 0.33ax. (ASTM,
1995; US EPA, 1986). Vertical dispersivity (az) may be estimated as 0.05ccx. (ASTM, 1995), or as
0.025ax. to 0.1 ax. (US EPA, 1986).
Some solute transport modelers will start with an accepted literature value for the types of
materials found in the aquifer matrix. After selecting initial dispersivity values, the contaminant
transport model is calibrated by adjusting the dispersivities (along with other transport parameters, as
necessary) within the range of accepted literature values until the modeled and observed contaminant
distribution patterns match (Anderson, 1979). This is a two-step process. The first step is to cali-
brate the flow model to the hydraulic conditions present at the site. After the ground-water flow-
model is calibrated to the hydraulics of the system, the contaminant transport model is calibrated by
trial and error using various values for dispersivity. There is no unique solution because several
hydraulic parameters, including hydraulic conductivity, effective porosity, and dispersivity, are
variable within the flow system (Anderson, I979; Davis et al., I993), and other transport parameters
such as retardation and biodegradation may not be well-defined.
B.2.2.4 One-Dimensional Advection-Dispersion Equation
The advection-dispersion equation is obtained by adding hydrodynamic dispersion to
equation B.2.2. In one dimension, the advection-dispersion equation is given by:
ac n a2c ac
^7=D*^^~V*^ eq. B.2.9
H
B2-14
-------
10
10'
E 10'
e 10
0)
Q.
_w
O 10'
to
T3
~D) 10
10"
10"
Longitudinal Dispersivity
= 10% of scale
(Pickens and
Grisak, 1981) -
• Longitudinal Dispersivity =;
= 0.83 [Log,,(scale)]''"' J
(Xu and Eckstein, 1995) ~i
Reliability 5
o Low
o Intermediate
O High
Data Source: Gelhar et a/., 1992 '"'
10" 10' 10' 10' 10' 10' 10* 10'
(m)
Source: Newell etal., 1996
Figure B.2.7 Relationship between dispersivity and scale.
Where:
vv = average linear velocity [L/T]
C = contaminant concentration [M/L3]
D = hydrodynamic dispersion [L2/T]
t = time [T]
x = distance along flow path [L]
This equation considers both advection and hydrodynamic dispersion. Because of sorption and
biodegradation, this equation generally must be combined with the other components of the modified
advection-dispersion equation presented as equation B.I.I to obtain an accurate mathematical de-
scription of solute transport.
Many organic contaminants, including chlorinated solvents and BTEX, are removed from
solution by sorption onto the aquifer matrix. Sorption is the process whereby dissolved contami-
nants partition from the ground water and adhere to the particles comprising the aquifer matrix.
Sorption of dissolved contamination onto the aquifer matrix results in slowing (retardation) of the
contaminant relative to the average advective ground-water flow velocity and a reduction in dis-
solved BTEX concentrations in ground water. Sorption can also influence the relative importance of
volatilization and biodegradation (Lyman et al, 1992). Figures B.2.8 and B.2.9 illustrate the effects
of sorption on an advancing solute front.
Keep in mind that sorption is a reversible reaction and that at a given solute concentration, some
portion of the solute is partitioning to the aquifer matrix and some portion is also desorbing and
reentering solution. As solute concentrations change, the relative amounts of contaminant that are
sorbing and desorbing will change. For example, as solute concentrations decrease (perhaps due to
B2-15
-------
o
••§ s
•3 §
e4 O
1.0
C/C 0.5 -
0.0
Contaminant front with advection,
hydrodynamic dispersion
and sorption
Distance from Source, x
- Contaminant front with
advection only
Contaminant front
with advection and
hydrodynamic
dispersion
Figure B.2.8 Breakthrough curve in one dimension shoving plug flow with continuous source resulting
from advection only; the combined processes of advection and hydrodynamic dispersion; and
the combined processes of advection, hydrodynamic dispersion, and sorption.
Initial
contaminant
slug
1.0-,
§
•43
si
31
4J O
C/C0 0.5-
0.0
Contaminant slug
with advection,
hydrodynamic dispersion,
and sorption
Contaminant slug with
advection only
Contaminant slug
with advection
and hydrodynamic
dispersion
Time or Distance from Source
Figure B.2.9 Breakthrough curve in one dimension showing plug flow with instantaneous source resulting
from advection only; the combined processes of advection and hydrodynamic dispersion; and
the combined processes of advection, hydrodynamic dispersion, and sorption.
plume migration or solute biodegradation and dilution), the amount of contaminant reentering solu-
tion will likely increase. The affinity of a given compound for the aquifer matrix will not be suffi-
cient to permanently isolate it from ground water, although for some compounds, the rates of desorp-
tion may be so slow that the loss of mass may be considered permanent for the time scale of interest.
Sorption, therefore, does not permanently remove solute mass from ground water; it merely retards
migration. It is this slowing of contaminant migration that must be understood in order to effectively
predict the fate of a dissolved contaminant. This section provides information on how retardation
coefficients are determined in the laboratory. It is not the intent of this document to instruct people
in how to perform these experiments; this information is provided for informational purposes only.
Linear isotherms and previously determined soil sorption coefficients (Koc) are generally used to
estimate sorption and retardation.
B.2.3.1 Mechanisms of Sorption
Sorption of dissolved contaminants is a complex phenomenon caused by several mechanisms,
including London-van der Waals forces, Coulomb forces, hydrogen bonding, ligand exchange,
chemisorption (covalent bonding between chemical and aquifer matrix), dipole-dipole forces, dipole-
induced dipole forces, and hydrophobic forces. Because of their nonpolar molecular structure,
hydrocarbons most commonly exhibit sorption through the process of hydrophobic bonding. When
B2-16
-------
the surfaces comprising the aquifer matrix are less polar than the water molecule, as is generally the
case, there is a strong tendency for the nonpolar contaminant molecules to partition from the ground
water and sorb to the aquifer matrix. This phenomenon is referred to as hydrophobic bonding and is
an important factor controlling the fate of many organic pollutants in soils (Devinny el al., 1990).
Two components of an aquifer have the greatest effect on sorption: organic matter and clay minerals.
In most aquifers, the organic fraction tends to control the sorption of organic contaminants.
B.2.3.2 Sorption Models and Isotherms
Regardless of the sorption mechanism, it is possible to determine the amount of sorption to be
expected when a given dissolved contaminant interacts with the materials comprising the aquifer
matrix. Bench-scale experiments are performed by mixing water-contaminant solutions of various
concentrations with aquifer materials containing various amounts of organic carbon and clay miner-
als. The solutions are then sealed with no headspace and left until equilibrium between the various
phases is reached. The amount of contaminant left in solution is then measured.
Both environmental conservative isotherms (ECI) and constant soil to solution isotherms (CSI)
can be generated. The ECI study uses the same water concentration but changes the soil to water
ratio. In CSI isotherm studies, the concentration of contaminant in water is varied while the amount
of water and sediment is constant. In some instances, actual contaminated water from the site is
added. Typically, the samples are continually rotated and concentrations measured with time to
document equilibrium. True equilibrium may require hundreds of hours of incubation but 80 to 90
percent of equilibrium may be achieved in one or two days.
The results are commonly expressed as a plot of the concentration of chemical sorbed (|ig/g)
versus the concentration remaining in solution (jig/L). The relationship between the concentration of
chemical sorbed (Ca) and the concentration remaining in solution (C() at equilibrium is referred to as
the sorption isotherm because the experiments are performed at constant temperature.
Sorption isotherms generally exhibit one of three characteristic shapes depending on the
sorption mechanism. These isotherms are referred to as the Langmuir isotherm, the Freundlich
isotherm, and the linear isotherm (a special case of the Freundlich isotherm). Each of these sorption
isotherms, and related equations, are discussed in the following sections.
B.2.3.2.1 Langmuir Sorption Model
The Langmuir model describes sorption in solute transport systems wherein the sorbed con-
centration increases linearly with increasing solute concentration at low concentrations and ap-
proaches a constant value at high concentrations. The sorbed concentration approaches a constant
value because there are a limited number of sites on the aquifer matrix available for contaminant
sorption. This relationship is illustrated in Figure B.2.10. The Langmuir equation is described
mathematically as (Devinny elal., 1990):
c _K^b_
C--I7]^ eq. B.2.10
Where:
C = sorbed contaminant concentration (mass contaminant/mass soil)
a ^ ^
K = equilibrium constant for the sorption reaction (|ig/g)
Cj = dissolved contaminant concentration (|ig/ml)
b = number of sorption sites (maximum amount of sorbed contaminant)
The Langmuir model is appropriate for highly specific sorption mechanisms where there are a
limited number of sorption sites. This model predicts a rapid increase in the amount of sorbed
contaminant as contaminant concentrations increase in a previously pristine area. As sorption sites
become filled, the amount of sorbed contaminant reaches a maximum level equal to the number of
sorption sites, b.
B2-17
-------
•s?
c
a
Langmulr
Dissolved Concentration C, (ng/ml)
Figure B.2.10 Characteristic adsorption isotherm shapes.
B.2.3.2.2 Freundlich Sorption Model
The Langmuir isotherm model can be modified if the number of sorption sites is large (assumed
infinite) relative to the number of contaminant molecules. This is generally a valid assumption for
dilute solutions (e.g., downgradient from a petroleum hydrocarbon spill in the dissolved BTEX
plume) where the number of unoccupied sorption sites is large relative to contaminant concentra-
tions. The Freundlich model is expressed mathematically as (Devinny et al., 1990):
eq.B.2.11
Where:
c =
<=
n =
distribution coefficient
sorbed contaminant concentration (mass contaminant/mass soil, mg/g)
dissolved concentration (mass contaminant/volume solution, (mg/ml)
chemical-specific coefficient
The value of n in this equation is a chemical-specific quantity that is determined experimentally.
Values of 1/n typically range from 0.7 to 1.1, but may be as low as 0.3 and as high as 1.7 (Lyman et
al. 1992).
The simplest expression of equilibrium sorption is the linear sorption isotherm, a special form
of the Freundlich isotherm that occurs when the value of n is 1. The linear isotherm is valid for a
dissolved species that is present at a concentration less than one half of its solubility (Lyman et al.,
1992). This is a valid assumption for BTEX compounds partitioning from fuel mixtures into ground
water. Dissolved BTEX concentrations resulting from this type of partitioning are significantly less
than the pure compound's solubility in pure water. The linear sorption isotherm is expressed as (Jury
etal, 1991):
Ca=KdC, eq.B.2.12
Where:
Kd= distribution coefficient (slope of the isotherm, ml/g).
Ca = sorbed contaminant concentration (mass contaminant/mass soil, Jlg/g)
Cl = dissolved contaminant concentration (mass contaminant/volume solution, |ig/ml)
The slope of the linear isotherm is the distribution coefficient, Kd.
B2-18
-------
B.2.3.3 Distribution Coefficient
The most commonly used method for expressing the distribution of an organic compound
between the aquifer matrix and the aqueous phase is the distribution coefficient, Kd, which is defined
as the ratio of the sorbed contaminant concentration to the dissolved contaminant concentration:
Kd=~C~ eq. B.2.13
Where:
Kd= distribution coefficient (slope of the sorption isotherm, ml/g)
Ca = sorbed concentration (mass contaminant/mass soil or |ig/g)
(7| = dissolved concentration (mass contaminant/volume solution or Jig/ml)
The transport and partitioning of a contaminant is strongly dependent on the chemical's
soil/water distribution coefficient and water solubility. The distribution coefficient is a measure of
the sorption/desorption potential and characterizes the tendency of an organic compound to be
sorbed to the aquifer matrix. The higher the distribution coefficient, the greater the potential for
sorption to the aquifer matrix. The distribution coefficient is the slope of the sorption isotherm at the
contaminant concentration of interest. The greater the amount of sorption, the greater the value of Kd.
For systems described by a linear isotherm, Kd is a constant. In general terms, the distribution
coefficient is controlled by the hydrophobicity of the contaminant and the total surface area of the
aquifer matrix available for sorption. Thus, the distribution coefficient for a single compound will
vary with the composition of the aquifer matrix. Because of their extremely high specific surface
areas (ratio of surface area to volume), the organic carbon and clay mineral fractions of the aquifer
matrix generally present the majority of sorption sites in an aquifer.
Based on the research efforts of Ciccioli et al. (1980), Karickhoff etal. (1979), and
Schwarzenbach and Westall (1981), it appears that the primary adsorptive surface for organic chemi-
cals is the organic fraction of the aquifer matrix. However, there is a "critical level of organic mat-
ter" below which sorption onto mineral surfaces is the dominant sorption mechanism (McCarty et
al., 1981). The critical level of organic matter, below which sorption appears to be dominated by
mineral-solute interactions, and above which sorption is dominated by organic carbon-solute interac-
tions, is given by (McCarty et al., 1981):
f°°°=lJW~K^ eq.B.2.14
Where:
fOCf = critical level of organic matter (mass fraction)
Ay = surface area of mineralogical component of the aquifer matrix (m2/g)
Kow = octanol-water partitioning coefficient
From this relationship, it is apparent that the total organic carbon content of the aquifer matrix
is less important for solutes with low octanol-water partitioning coefficients (Kow). Also apparent is
the fact that the critical level of organic matter increases as the surface area of the mineralogic
fraction of the aquifer matrix increases. The surface area of the mineralogic component of the
aquifer matrix is most strongly influenced by the amount of clay. For compounds with low Kow
values in materials with a high clay content, sorption to mineral surfaces could be an important factor
causing retardation of the chemical.
Several researchers have found that if the distribution coefficient is normalized relative to the
aquifer matrix total organic carbon content, much of the variation in observed Kd values between
different soils is eliminated (Dragun, 1988). Distribution coefficients normalized to total organic
carbon content are expressed as Koc. The following equation gives the expression relating Kd to KOC:
B2-19
-------
-
Koc - ~r eq. B.2.15
Joe
Where:
Koc = soil sorption coefficient normalized for total organic carbon content
K, = distribution coefficient
d
foc = fraction total organic carbon (mg organic carbon/mg soil)
In areas with high clay concentrations and low total organic carbon concentrations, the clay
minerals become the dominant sorption sites. Under these conditions, the use of Koc to compute Kd
might result in underestimating the importance of sorption in retardation calculations, a source of
error that will make retardation calculations based on the total organic carbon content of the aquifer
matrix more conservative. In fact, aquifers that have a high enough hydraulic conductivity to spread
hydrocarbon contamination generally have low clay content. In these cases, the contribution of
sorption to mineral surfaces is generally trivial.
Earlier investigations reported distribution coefficients normalized to total organic matter
content (Kom). The relationship between fm and foc is nearly constant and, assuming that the organic
matter contains approximately 58 percent carbon (Lyman el al., 1992):
Koc=l.724Kom eq.B.2.16
B.2.3.4 Coefficient of Retardation
As mentioned earlier, sorption tends to slow the transport velocity of contaminants dissolved in
ground water. The coefficient of retardation, R, is used to estimate the retarded contaminant veloc-
ity. The coefficient of retardation for linear sorption is determined from the distribution coefficient
using the relationship:
_n1_
eq.B.2.17
n
Where:
R = coefficient of retardation [dimensionless]
pft = bulk density of aquifer [M/L3]
Kd= distribution coefficient [L3/M]
n = porosity [L3/L3]
The retarded contaminant transport velocity, v , is given by:
eq.B.2.18
Where:
vc = retarded contaminant transport velocity [L/T]
vv = advective ground-water velocity [L/T]
R = coefficient of retardation [dimensionless]
Two methods used to quantify the distribution coefficient and amount of sorption (and thus
retardation) for a given aquifer/contaminant system are presented below. The first method involves
estimating the distribution coefficient by using Koc for the contaminants and the fraction of organic
carbon comprising the aquifer matrix. The second method involves conducting batch or column
tests to determine the distribution coefficient. Because numerous authors have conducted experi-
ments to determine K values for common contaminants, literature values are reliable, and it gener-
OC ' ' O
ally is not necessary to conduct laboratory tests.
B.2.3.4. 1 Determining the Coefficient of Retardation using K
O O Oc
Batch and column tests have been performed for a wide range of contaminant types and concen-
trations and aquifer conditions. Numerous studies have been performed using the results of these
B2-20
-------
tests to determine if relationships exist that are capable of predicting the sorption characteristics of a
chemical based on easily measured parameters. The results of these studies indicate that the amount
of sorption is strongly dependent on the amount of organic carbon present in the aquifer matrix and
the degree of hydrophobicity exhibited by the contaminant (Bailey and White, 1970; Karickhoff el
al., 1979; Kenaga and Goring, 1980; Brown and Flagg, 1981; Schwarzenbach and Westall, 1981;
Hassett el a/., 1983; Chiou el al., 1983). These researchers observed that the distribution coefficient,
Kd, was proportional to the organic carbon fraction of the aquifer times a proportionality constant.
This proportionality constant, Koc, is defined as given by equation B.2.15. In effect, equation B.2.15
normalizes the distribution coefficient to the amount of organic carbon in the aquifer matrix. Be-
cause it is normalized to organic carbon, values of Koc are dependent only on the properties of the
compound (not on the type of soil). Values of Koc have been determined for a wide range of chemi-
cals. Table B.2.1 lists K values for selected chlorinated compounds, and Table B.2.2 lists K
DC l ' OC
values for BTEX and trimethylbenzene.
By knowing the value of Koc for a contaminant and the fraction of organic carbon present in the
aquifer, the distribution coefficient can be determined by using the relationship:
Kd=Kocfoc eq.B.2.19
When using the method presented in this section to predict sorption of the BTEX compounds,
total organic carbon concentrations obtained from the most transmissive aquifer zone should be
averaged and used for predicting sorption. This is because the majority of dissolved contaminant
transport occurs in the most transmissive portions of the aquifer. In addition, because the most
transmissive aquifer zones generally have the lowest total organic carbon concentrations, the use of
this value will give a conservative prediction of contaminant sorption and retardation.
B2-21
-------
Table B. 2.1
Values of Aqueous Solubility and K Jor Selected Chlorinated Compounds
Compound
Tetrachloroethene
Tetrachloroethene
Tetrachloroethene
Trichloroethene
Trichloroethene
Trichloroethene
1 , 1 -Dichloroethene
1 , 1 -Dichloroethene
1 , 1 -Dichloroethene
c/s-l,2-Dichloroethene
c/s-l,2-Dichloroethene
/ra«s-l,2-Dichloroethene
/ra«5-l,2-Dichloroethene
/raw5-l,2-Dichloroethene
Vinyl Chloride
Vinyl Chloride
1,1,1 -Trichloroethane
1,1,2-Trichloroethane
1 , 1 -Dichloroethane
1 ,2-Dichloroethane
Chloroethane
Hexachlorobenzene
1 ,2-Dichlorobenzene
1 ,3 -Dichlorobenzene
1 ,4-Dichlorobenzene
Chlorobenzene
Carbon Tetrachloride
Chloroform
Methylene Chloride
Solubility (mg/L)
150a
1,503C
l,100a
1,100C
2,250a
2,500d
3,500C
6,300a
6,300C
l,100a
2,763d
1,495C
4,420e
5,060d
8,520C
5,710e
0.006'
156C
lllg
74 to 87d
472d
805g
7,950C
13,000C
KOC
(L/Kg)
263a
359b
209 - 238C
107a
137b
87 - 150C
64.6a
80.2b
150d
80.2b
49C
58.9a
80.2b
36C
2.45a
0.4 - 56d
183C
70e
40d
33 to 152C
33 to 143e
—
272 - 1480C
293to31,600g
273 to 1833d
83 to 389d
110g
<34C
48C
From KJIOX etaL, 1993
From Jeng et al., 1992; Temperature = 20°C
From Howard, 1990; Temperature = 25"C
From Howard, 1989; Temperature = 25"C
From Howard, 1989; Temperature = 20°C
ATSDR, 1990; Temperature = 2(fC
From Howard, 1990; Temperature = 20°C
B2-22
-------
Table B.2.2 Values of Aqueous Solubility and K Jbr BTEXand Trimethylbenzene Isomers
Compound
Benzene
Benzene
Benzene
Benzene
Benzene
Benzene*
Benzene
Toluene
Toluene
Toluene
Toluene
Toluene*
Ethylbenzene
Ethylbenzene
Ethylbenzene
Ethylbenzene
Ethylbenzene
Ethylbenzene*
Ethylbenzene
o-xylene
o-xylene
o-xylene*
m-xylene
m-xylene
m-xylene
m-xylene
m-xylene*
p-xylene
p-xylene
p-xylene*
1 ,2,3 -trimethylbenzene*
1 ,2,4-trimethylbenzene
1,2,4-trimethylbenzene*
1 , 3 ,5 -trimethylbenzene *
Solubility (mg/L)
1750a
1780C
1780C
1780h
1780h
1780c'h
515a
537C
537C
537C
152a
167C
167C
140h
140h
167C
152a
152a
158a
162C
162C
162C
198a
198a
75
591
591
72.60g
KOC
(L/Kg)
87. la
83b
190c,d,t
62c'e''
?2h,i
79ly.-
89k
151a
303b
380^'
noc,e,t
190k'*
158.5a
519b
680^'
200c'e''
50 111'1
468hJ
398k
128.8a
519b
422k'*
519b
720c,d,t
210c,e,t
405. 37k'*
204a
519b
357k'*
884b'*
884b
772k'*
676k'*
" From Knox et al., 1993
* From Jeng et al., 1992; Temperature 20°C
c From Lyman et al., 1992; Temperature ::::: 25°C
fl Estimated from Km
" Estimated from solubility
Estimate from solubility generally considered more reliable
From Lyman et al., 1992; Temperature ::::: 20°C
From Fetter, 1993
' Average of 12 equations used to estimate K from K or K
^ *' * oc-J fm am
' Average of 5 equations used to estimate K from Solubility
'r Average using equations from Kenaga and Goring (1980), Means et al. (1980), and Hassett et al. (1983) to
estimate Kocfrom solubility
' From Sutton and Colder (1975)
* Recommended value
B2-23
-------
B.2.3.4.2 Determining the Coefficient of Retardation using Laboratory Tests
The distribution coefficient may be quantified in the laboratory using batch or column tests.
Batch tests are easier to perform than column tests. Although more difficult to perform, column tests
generally produce a more accurate representation of field conditions than batch tests because con-
tinuous flow is involved. Knox et al. (1993) suggest using batch tests as a preliminary screening
tool, followed by column studies to confirm the results of batch testing. The authors of this docu-
ment feel that batch tests, if conducted properly, will yield sufficiently accurate results for fate and
transport modeling purposes provided that sensitivity analyses for retardation are conducted during
the modeling.
Batch testing involves adding uncontaminated aquifer material to a number of vessels, adding
solutions prepared using uncontaminated ground water from the site mixed with various amounts of
contaminants to produce varying solute concentrations, sealing the vessel and shaking it until equi-
librium is reached, analyzing the solute concentration remaining in solution, and calculating the
amount of contaminant sorbed to the aquifer matrix using mass balance calculations. A plot of the
concentration of contaminant sorbed versus dissolved equilibrium concentration is then made using
the data for each reaction vessel. The slope of the line formed by connecting each data point is the
distribution coefficient. The temperature should be held constant during the batch test, and should
approximate that of the aquifer system through which solute transport is taking place.
Table B.2.3 contains data from a hypothetical batch test. These data are plotted (Figure B.2.11)
to obtain an isotherm unique to the aquifer conditions at the site. A regression analysis can then be
performed on these data to determine the distribution coefficient. For linear isotherms, the distribu-
tion coefficient is simply the slope of the isotherm. In this example, Kd = 0.0146 L/g. Batch-testing
procedures are described in detail by Roy et al. (1992).
Column testing involves placing uncontaminated aquifer matrix material in a laboratory column
and passing solutions through the column. Solutions are prepared by mixing uncontaminated ground
water from the site with the contaminants of interest and a conservative tracer. Flow rate and time
are accounted for and samples are periodically taken from the effluent of the column and analyzed to
determine contaminant and tracer concentrations. Breakthrough curves are prepared for the contami-
nants by plotting chemical concentration versus time (or relative concentration versus number of
pore volumes). The simplest way to determine the coefficient of retardation (or the distribution
coefficient) from the breakthrough curves is to determine the time required for the effluent concen-
tration to equal 0.5 of the influent concentration. This value can be used to determine average
velocity of the center of mass of the contaminant. The retardation factor is determined by dividing
the average flow velocity through the column by the velocity of the center of mass of the contami-
nant. The value thus obtained is the retardation factor. The coefficient of retardation also can be
determined by curve fitting using the CXTFIT model of Parker and van Genuchten (1984). Break-
through curves also can be made for the conservative tracer. These curves can be used to determine
the coefficient of dispersion by curve fitting using the model of Parker and van Genuchten (1984).
When using the method presented in this section to predict sorption of the BTEX compounds,
aquifer samples should be obtained from the most transmissive aquifer zone. This is because the
majority of dissolved contaminant transport occurs in the most transmissive portions of the aquifer.
In addition, because the most transmissive aquifer zones generally have the lowest organic carbon
concentrations, the use of these materials will give a conservative prediction of contaminant sorption
and retardation.
B2-24
-------
Table B.2.3 Data from Hypothetical Batch Test Experiment
Initial Concentration
(U2/L)
250
500
1000
1500
2000
3800
6000
9000
Equilibrium Concentration
(U2/L)
77.3
150.57
297.04
510.1
603.05
1198.7
2300.5
3560.7
Weight of Solid
Matrix (g)
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
Sorbed Concentration* (p.g/g)
1.69
3.42
6.89
9.70
13.68
25.48
36.23
53.27
* Adsorbed concentration ((Initial concentration - Equilibrium Concentration) x Volume of Solution) / Weight of
Solid Matrix
60.00
.50.00 "
^40.00
g 30.00
> 20.00 "
10.00 --
0.00
1000 2000 3000 4000
Equilibrium Concentration (|Xg/L)
Figure B. 2.11 Plot of 'sorbed concentration vs. equilibrium concentration.
B.2.3.5 One-Dimensional Advection-Dispersion Equation with Retardation
The advection-dispersion equation is obtained by adding hydrodynamic dispersion to
equation B.2.2. In one dimension, the advection-dispersion equation is given by:
ac
--v.
a*
eq. B.2.20
Where:
v =
X
J >
C =
r~j
x
t =
average linear velocity ground-water velocity [L/T]
coefficient of retardation [dimensionless]
contaminant concentration [M/L3]
hydrodynamic dispersion [L2/T]
time [T]
x = distance along flow path [L]
This equation considers advection, hydrodynamic dispersion, and sorption (retardation). Be-
cause of biodegradation, this equation generally must be combined with the other components of the
modified advection-dispersion equation, presented as equation B. 1.1, to obtain an accurate math-
ematical description of solute transport.
B2-25
-------
B.2.4 VOLATILIZATION
While not a destructive attenuation mechanism, volatilization does remove contaminants from
the ground-water system. In general, factors affecting the volatilization of contaminants from ground
water into soil gas include the contaminant concentration, the change in contaminant concentration
with depth, the Henry's Law constant and diffusion coefficient of the compound, mass transport
coefficients for the contaminant in both water and soil gas, sorption, and the temperature of the water
(Larson and Weber, 1994).
Partitioning of a contaminant between the liquid phase and the gaseous phase is governed by
Henry's Law. Thus, the Henry's Law constant of a chemical determines the tendency of a contami-
nant to volatilize from ground water into the soil gas. Henry's Law states that the concentration of a
contaminant in the gaseous phase is directly proportional to the compound's concentration in the
liquid phase and is a constant characteristic of the compound. Stated mathematically, Henry's Law is
given by (Lyman et aL, 1992):
Ca=HC, eq.B.2.21
Where:
H = Henry's Law Constant (atm m3/mol)
Ca= concentration in air (atm)
Cl = concentration in water (mol/m3)
Henry's Law constants for chlorinated and petroleum hydrocarbons range over several orders of
magnitude. For petroleum hydrocarbons, Henry's Law constants (H) for the saturated aliphatics,
H range from I to 10 atm mVmol @ 25°C; for the unsaturated and cyclo-aliphatics ranges from 0.1 to
1 atm mVmol @ 25°C; and for the light aromatics (e.g., BTEX) H ranges from 0.007 to
0.02 atm mVmol @ 25°C (Lyman etal., 1992). Values of Henry's Law constants for selected chlori-
nated solvents and the BTEX compounds are given in Table B.2.4. As indicated on the table, values
of H for chlorinated compounds also vary over several orders of magnitude, although most are
similar to those for BTEX compounds.
The physiochemical properties of chlorinated solvents and the BTEX compounds give them low
Henry's Law constants, with the exception of vinyl chloride. Because of the small surface area of the
ground-water flow system exposed to soil gas, volatilization of chlorinated solvents and BTEX
compounds from ground water is a relatively slow process that, in the interest of being conservative,
generally can be neglected when modeling biodegradation. Chiang et al. (1989) demonstrated that
less than 5 percent of the mass of dissolved BTEX is lost to volatilization in the saturated ground-
water environment. Moreover, Rivett (1995) observed that for plumes more than about 1 meter
below the air-water interface, little, if any, solvent concentrations will be detectable in soil gas due to
the downward ground-water velocity in the vicinity of the water table. This suggests that for por-
tions of plumes more than 1 meter below the water table, very little, if any, mass will be lost due to
volatilization. In addition, vapor transport across the capillary fringe can be very slow (McCarthy
and Johnson, 1993), thus further limiting mass transfer rates. Because of this, the impact of volatil-
ization on dissolved contaminant reduction can generally be neglected, except possibly in the case of
vinyl chloride. However, Rivett's (1995) findings should be kept in mind even when considering
volatilization as a mechanism for removal of vinyl chloride from ground water.
B.2.5 RECHARGE
Groundwater recharge can be defined as the entry into the saturated zone of water made avail-
able at the water-table surface (Freeze and Cherry, I979). In recharge areas, flow near the water
table is generally downward. Recharge defined in this manner may therefore include not only pre-
cipitation that infiltrates through the vadose zone, but water entering the ground-water system due to
discharge from surface water bodies (i.e., streams and lakes). Where a surface water body is in
B2-26
-------
Table B.2.4 Henry s Law Constants and Vapor Pressures for Common Fuel Hydrocarbons and
Chlorinated Solvents
Compound
Benzene
Ethylbenzene
Toluene
o-Xylene
/w-Xylene
p-Xylene
1,2, 3-Trimethylbenzene
1,2, Ą-Trimethylbenzene
1,3, 5-Trimethylbenzene
1,2,4, 5-Tetramethylbenzene
Tetrachloroethene
Trichloroethene
1 , 1 -Dichloroethene
c/s-l,2-Dichloroethene
/ra«s-l,2-Dichloroethene
Vinyl Chloride
1,1,1 -Trichloroethane
1 , 1 ,2-Trichloroethane
1 , 1 -Dichloroethane
1 ,2-Dichloroethane
Chloroethane
Hexachlorobenzene
1 ,2-Dichlorobenzene
1 ,3 -Dichlorobenzene
1 ,4-Dichlorobenzene
Chlorobenzene
Carbon Tetrachloride
Chloroform
Methylene Chloride
Vapor Pressure (mmHg
@25°C)
95
10
28.4
10
10
10
14
57.8
591
200
265
2,580
123.7
30.3
227
78.7
766
0.0000109
1.47
2.3
1.76
11.9
113.8
246
434.9
Henry's Law Constant
(atm-m^/mol)
0.0054
0.0066
0.0067
0.00527
0.007
0.0071
0.00318
0.007
0.006
0.0249
0.0153
0.0091
0.018
0.0037
0.0072
1.22
0.008
0.0012
0.0059
0.00098
0.0085
0.00068
0.0012
0.0018
0.0015
0.0035
0.0304
0.00435
0.00268
B2-27
-------
contact with or is part of the ground-water system, the definition of recharge above is stretched
slightly. However, such bodies are often referred to as recharging lakes or streams. Recharge of a
water table aquifer has two effects on the natural attenuation of a dissolved contaminant plume.
Additional water entering the system due to infiltration of precipitation or from surface water will
contribute to dilution of the plume, and the influx of relatively fresh, electron-acceptor-charged water
will alter geochemical processes and in some cases facilitate additional biodegradation.
Recharge from infiltrating precipitation is the result of a complex series of processes in the
unsaturated zone. Description of these processes is beyond the scope of this discussion; however, it
is worth noting that the infiltration of precipitation through the vadose zone brings the water into
contact with the soil and thus may allow dissolution of additional electron acceptors and possibly
organic soil matter (a potential source of electron donors). Infiltration, therefore, provides fluxes of
water, inorganic species, and possibly organic species into the ground water. Recharge from surface
water bodies occurs when the hydraulic head of the body is greater than that of the adjacent ground
water. The surface water may be a connected part of the ground-water system, or it may be perched
above the water table. In either case, the water entering the ground-water system will not only aid in
dilution of a contaminant plume but it may also add electron acceptors and possibly electron donors
to the ground water.
An influx of electron acceptors will tend to increase the overall electron-accepting capacity
within the contaminant plume. In addition to the inorganic electron acceptors that may be dissolved
in the recharge (e.g., dissolved oxygen, nitrate, or sulfate), the introduction of water with different
geochemical properties may foster geochemical changes in the aquifer. For example, iron (II) will be
oxidized back to iron (III). Vroblesky and Chapelle (1994) present data from a site where a major
rainfall event introduced sufficient dissolved oxygen into the contaminated zone to cause
reprecipitation of iron (III) onto mineral grains. This reprecipitation made iron (III) available for
reduction by microorganisms, thus resulting in a shift from methanogenesis back to iron (III) reduc-
tion (Vroblesky and Chapelle, 1994). Such a shift may be beneficial for biodegradation of com-
pounds used as electron donors, such as fuel hydrocarbons or vinyl chloride. However, these shifts
can also make conditions less favorable for reductive dehalogenation.
Evaluating the effects of recharge is typically difficult. The effects of dilution might be esti-
mated if one has a detailed water budget for the system in question, but if a plume has a significant
vertical extent, it cannot be known with any certainty what proportion of the plume mass is being
diluted by the recharge. Moreover, because dispersivity, sorption, and biodegradation are often not
well-quantified, separating out the effects of dilution may be very difficult indeed. Where recharge
enters from precipitation, the effects of the addition of electron acceptors may be qualitatively appar-
ent due to elevated electron acceptor concentrations or differing patterns in electron acceptor con-
sumption or byproduct formation in the area of the recharge. However, the effects of short-term
variations in such a system (which are likely due to the intermittent nature of precipitation events in
most climates) may not be easily understood. Where recharge enters from surface water, the influx
of mass and electron acceptors is more steady over time. Quantifying the effects of dilution may be
less uncertain, and the effects of electron acceptor replenishment may be more easily identified
(though not necessarily quantified).
B2-28
-------
SECTION B-3
ATTENUATION -
Many anthropogenic organic compounds, including certain chlorinated solvents, can be de-
graded by both biological and abiotic mechanisms. Biological degradation mechanisms are dis-
cussed in this section; abiotic degradation mechanisms are discussed in Section B.4. Table B.3.1
summarizes the various biotic and abiotic mechanisms that result in the degradation of anthropo-
genic organic compounds. Biological degradation mechanisms tend to dominate in most ground-
water systems, depending on the type of contaminant and the ground-water chemistry.
Table B.3.1 Biologic and Abiotic Degradation Mechanisms for Various Anthropogenic Organic
Compounds
Compound
PCE
TCE
DCE
Vinyl Chloride
TCA
1,2-DCA
Chloroethane
Carbon Tetrachloride
Chloroform
Methylene Chloride
Chlorobenzenes
Benzene
Toluene
Ethylbenzene
Xylenes
1 ,2-Dibromoethane
Degradation Mechanism
Reductive dechlorination
Reductive dechlorination, cometabolism
Reductive dechlorination, direct biological oxidation
Reductive dechlorination, direct biological oxidation
Reductive dechlorination, hydrolysis,
dehydrohal ogenati on
Reductive dechlorination, direct biological oxidation
Hydrolysis
Reductive dechlorination, cometabolism, abiotic
Reductive dechlorination, cometabolism
Direct biological oxidation
Direct biological oxidation, reductive dechlorination,
cometabolism
Direct biological oxidation
Direct biological oxidation
Direct biological oxidation
Direct biological oxidation
Reductive dehal ogenati on, hydrolysis, direct
biological oxidation
Many organic contaminants are biodegraded by microorganisms indigenous to the subsurface
environment. During biodegradation, dissolved contaminants are ultimately transformed into in-
nocuous byproducts such as carbon dioxide, chloride, methane, and water. In some cases, intermedi-
ate products of these transformations may be more hazardous than the original compound; however,
they may also be more easily degraded. Biodegradation of organic compounds dissolved in ground
water results in a reduction in contaminant concentration (and mass) and slowing of the contaminant
front relative to the average advective ground-water flow velocity. Figures B.3.1 and B.3.2 illustrate
the effects of biodegradation on an advancing solute front.
B3-29
-------
8
•s §
1.0
§
C/C0 0.5-
0.0
Contaminant front with advectionj
hydrodynamic dispersion
and.sprptip_n
Contaminant front with advection,
hydrodynamic dispersion,
sorption, and biodegradation
- Contaminant front with
advection only
Contaminant front
with advection and
hydrodynamic
dispersion
0
Distance from Source, x
Figure B.3.1 Breakthrough curve in one dimension show ing plug flow with continuous source resulting
from advection only; the combined processes of advection and hydrodynamic dispersion; the
combined processes of advection, hydrodynamic dispersion, and sorption; and the combined
processes of advection, hydrodynamic dispersion, sorption, and biodegradation.
§
Initial
contaminant
slug
l.C
g | C/CQ 0.5-
• H
•0
43 fl
0.0
Contaminant slug
with advection,
hydrodynamic dispersion,
and sorption
t,,x,
Contaminant slug
with advection,
hydrodynamic dispersion,
sorption, and
biodegradation
Contaminant slug with
advection only
Contaminant slug
with advection
and hydrodynamic
dispersion
Time or Distance from Source
Figure B.3.2 Breakthrough curve in one dimension show ing plug flow with instantaneous source resulting
from, advection only; the combined processes of advection and hydrodynamic dispersion; the
combined processes of advection, hydrodynamic dispersion, and sorption; and the combined
processes of advection, hydrodynamic dispersion, sorption, and biodegradation.
B.3.1 OVERVIEW OF
As recently as 1975 the scientific literature reported the subsurface/aquifer environment as
devoid of significant biological activity. It is now known that soils and shallow sediments contain a
large variety of microorganisms, ranging from simple prokaryotic bacteria and cyanobacteria to more
complex eukaryotic algae, fungi, and protozoa. Over the past two decades, numerous laboratory and
field studies have shown that microorganisms indigenous to the subsurface environment can degrade
a variety of organic compounds, including components of gasoline, kerosene, diesel, jet fuel, chlori-
nated ethenes, chlorinated ethanes, the chlorobenzenes, and many other compounds (e.g., for fuels
see Jamison etal, 1975; Atlas, 1981, 1984, and 1988; Young, 1984; Bartha, 1986; B. H. Wilson et
al, 1986 and 1990; Barker etal, 1987; Baedecker etal, 1988; Lee, 1988; Chiang era/., 1989;
Cozzarelli etal., 1990; Leahy and Colewell, 1990; Alvarez and Vogel, 1991; Evans etal., 1991a and
1991b; Edwards etal., 1992; Edwards and Grbic-Galic, 1992; Thierrin etal., 1992; Mai one etal.,
1993; Davis et al., 1994a and 1994b; and Lovley et al, 1995; and for chlorinated solvents see
Brunner and Leisinger, 1978; Brunner etal, 1980; Rittman and McCarty, 1980; Bouwer etal, 1981;
B3-30
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Table B.3.2 Some Microorganisms Capable of Degrading Organic Compounds(Modified from Riser-
Roberts, 1992)
Contaminant
Benzene
Toluene
Ethylbenzene
Xylenes
Jet Fuels
Kerosene
Chlorinated
Ethenes
Chlorinated
Ethanes
Chlorinated
Methanes
Chlorobenzenes
Microorganisms
Pseudomonas putida, P. rhodochrous, P. aeruginosa,
Acinetobacter sp., Methylosinus trichosporium OB3b, Nocardia
sp., methanogens, anaerobes
Methylosinus trichosporium OB3b, Bacillus sp., Pseudomonas
sp., P. putida, Cunninghamella elegans, P. aeruginosa, P.
mildenberger, P. aeruginosa, Achromobacter sp., methanogens,
anaerobes
Pseudomonas putida
Pseudomonas putida, methanogens, anaerobes
Cladosporium, Hormodendrum
Torulopsis, Candidatropicalis, Corynebacterium
hydrocarboclastus, Candidaparapsilosis, C. guilliermondii, C.
lipolytica, Trichosporon sp., Rhohosporidium toruloides,
Cladosporium resinae
Dehalobacter restrictus, Dehalospirillum multivorans,
Enterobacter agglomerans, Dehalococcus entheogenes strain
\95,Desulfitobacterium sp. strain PCE1, Pseudomonas putida
(multiple strains), P. cepacia G4, P. mendocina,
Desulfobacterium sp., Methanobacterium sp., Methanosarcina
sp. strain DCM, Alcaligenes eutrophus JMP 134, Methylosinus
trichosporium OB3b, Escherichia coli, Nitorsomonas europaea,
Methylocystis parvus OBBP, Mycobacterium sp., Rhodococcus
erythopolis
Desulfobacterium sp., Methanobacterium sp., Pseudomonas
putida, Clostridium sp., C. sp. strain TCAIIB,
Acetobacterium woodii, Desulfobacterium sp.,
Methanobacterium sp., Pseudomonas sp. strain KC, Escherichia
coli K-12, Clostridium sp., Methanosarcina sp.,
Hyphomicrobium sp. strain DM2,
Alcaligenes sp. (multiple strains), Pseudomonas sp. (multiple
strains,) , P. putida, Staphylococcus epidermis
Comments/
Biodegradability
Moderate to High
High
High
High
High
High
Moderate
Moderate
Moderate
Moderate to High
Miller and Guengerich, 1982; Roberts et a/., 1982; Bouwer and McCarty, 1983; Stuck! et al., 1983;
Reineke and Knackmuss, 1984; Wilson and Wilson, 1985; Fogel et al, 1986; Egli et al, 1987; Vogel
and McCarty, 1987; Vogel etal, 1987; Bouwer and Wright, 1988; Little etal, 1988; Freedman and
Gossett, 1989; Sewell and Gibson, 1991; Chapelle, 1993; DeBruin etal, 1992; Ramanand etal,
1993; Vogel, 1994; Suflita and Townsend, 1995; Adriaens and Vogel, 1995; Bradley and Chapelle,
1996; Gossett and Zinder, 1996; Spain, 1996). Table B.3.2 presents a partial list of microorganisms
known to degrade anthropogenic organic compounds.
Although we now recognize that microorganisms are ubiquitous in drinking water aquifers, the
study of the microbial ecology and physiology of the subsurface, below the rhizosphere, is still in its
infancy. However, great progress has been made at least in identifying, if not fully understanding,
B3-31
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the numerous and diverse types of microbially-mediated contaminant transformations that can occur
in the subsurface.
Chemothrophic organisms, such as humans and most microorganisms, obtain energy for growth
and activity from physiologically coupling oxidation and reduction reactions and harvesting the
chemical energy that is available. Under aerobic conditions (in the presence of molecular oxygen)
humans and many bacteria couple the oxidation of organic compounds (food) to the reduction of
oxygen (from the air). However in the absence of oxygen (anaerobic conditions), microorganisms
may use other compounds as electron acceptors. Anaerobic microorganisms can obtain energy from
a variety of electron donors such as natural organic carbon or many forms of anthropogenic carbon
and electron acceptors such as nitrate, iron (III), sulfate, carbon dioxide, as well as many of the
chlorinated solvents.
The introduction of oxidizable soluble organic contaminants into ground water initiates a series
of complex responses by subsurface microorganisms. Field and laboratory research suggests that
distinct communities defined by the dominant electron acceptor develop which are spatially and
temporally separate. These communities are most likely ecologically defined by the flux of biologi-
cally available electron donors and acceptors. The biological processes of these communities are
potentially useful as natural attenuation mechanisms, as the basis of new bioremediation technolo-
gies, and as indicators of the extent and severity of the release. As electron acceptors and nutrients
are depleted by microbial activity during biodegradation of contaminants, the redox potential of
contaminated aquifers decreases. This results in a succession of bacterial types adapted to specific
redox regimes and electron acceptors. Metabolic byproducts of contaminant biodegradation also
exert selective forces, either by presenting different carbon sources or by further modifying the
physical and chemical environment of the aquifer. Like organic and inorganic colloids, microorgan-
isms possess complex surface chemistry, and can themselves serve as mobile and immobile reactive
sites for contaminants.
Under anaerobic conditions, most organic compounds are degraded by groups of interacting
microorganisms referred to as a consortium. In the consortium, individual types of organisms carry
out different specialized reactions which, when combined, can lead to the complete mineralization of
a particular compound. The metabolic interaction between organisms can be complex and may be so
tightly linked under a given set of conditions that stable consortia can be mistakenly identified as a
single species. There seems to be several advantages to the consortial system, including: 1) This
system allows for the creation of microenvironments where certain types of organisms can survive in
otherwise hostile conditions; 2) Reactions that are thermodynamically unfavorable can be driven by
favorable reactions when they are metabolically linked within the consortium; and, 3) This system
takes advantage of the diverse metabolic capabilities of microorganisms by allowing for the forma-
tion and enrichment of associations that can utilize an introduced substrate faster than a single
species could evolve a novel complex enzyme pathway to degrade the same compound.
It appears that subsurface microbial communities contain the metabolic diversity required to
utilize a wide variety of organic contaminants as a primary growth substrate in the presence of
electron acceptors such as oxygen. Some pollutants, especially the highly oxidized chlorinated
hydrocarbons, are not amenable to use as a primary growth substrate. Instead, these compounds are
used as electron acceptors in reactions that rely on another source of carbon as a primary substrate or
are degraded fortuitously via cometabolism. Thus, biodegradation of organic compounds in ground
water occurs via three mechanisms:
« Use of the organic compound as the primary growth substrate;
* Use of the organic compound as an electron acceptor; and
« Cometabolism.
B3-32
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The first two biodegradation mechanisms involve the microbial transfer of electrons from
electron donors (primary growth substrate) to electron acceptors. This process can occur under
aerobic or anaerobic conditions. Electron donors include natural organic material, fuel hydrocar-
bons, chlorobenzenes, and the less oxidized chlorinated ethenes and ethanes. Electron acceptors are
elements or compounds that occur in relatively oxidized states. The most common naturally occur-
ring electron acceptors in ground water include dissolved oxygen, nitrate, manganese (IV), iron (III),
sulfate, and carbon dioxide. In addition, the more oxidized chlorinated solvents such as PCE, TCE,
DCE, TCA, DC A, and polychlorinated benzenes can act as electron acceptors under favorable
conditions. Under aerobic conditions, dissolved oxygen is used as the terminal electron acceptor
during aerobic respiration. Under anaerobic conditions, the electron acceptors listed above are used
during denitrification, manganese (IV) reduction, iron (III) reduction, sulfate reduction,
methanogenesis, or reductive dechlorination. Chapelle (1993) and Atlas (1988) discuss terminal
electron accepting processes in detail.
The third biodegradation mechanism is cometabolism. During cometabolism the compound
being degraded does not benefit the organism. Instead, degradation is brought about by a fortuitous
reaction wherein an enzyme produced during an unrelated reaction degrades the organic compound.
As discussed in sections B.3.2, B.3.3, and B.3.4, biodegradation causes measurable changes in
ground-water chemistry. Table B.3.3 summarizes these trends. During aerobic respiration, oxygen is
reduced to water, and dissolved oxygen concentrations decrease. In anaerobic systems where nitrate
is the electron acceptor, the nitrate is reduced to NO,;, N2O, NO, NH4", or N2 via denitrification or
dissimilatory nitrate reduction, nitrate concentrations decrease. In anaerobic systems where iron (III)
is the electron acceptor, it is reduced to iron (II) via iron (III) reduction, and iron (II) concentrations
increase. In anaerobic systems where sulfate is the electron acceptor, it is reduced to H2S via sulfate
reduction, and sulfate concentrations decrease. During aerobic respiration, denitrification, iron (III)
reduction, and sulfate reduction, total alkalinity will increase. In anaerobic systems where CO,, is
used as an electron acceptor, it is reduced by methanogenic bacteria during methanogenesis, and CH4
is produced. In anaerobic systems where contaminants are being used as electron acceptors, they are
reduced to less chlorinated daughter products; in such a system, parent compound concentrations
will decrease and daughter product concentrations will increase at first and then decrease as the
daughter product is used as an electron acceptor or is oxidized.
As each subsequent electron acceptor is utilized, the ground water becomes more reducing and
the redox potential of the water decreases. Figure B.3.3 shows the typical ORP conditions for
ground water when different electron acceptors are used. The main force driving this change in ORP
is microbially mediated oxidation-reduction reactions. ORP can be used as a crude indicator of
which oxidation-reduction reactions may be operating at a site. The ORP determined in the field
using an electrode is termed Eh. Eh can be expressed as pE, which is the hypothetical measure of the
electron activity associated with a specific Eh. High pE means that the solution or redox couple has
a relatively high oxidizing potential.
OF VIA USE AS A
GROWTH
Many organic compounds including natural organic carbon, fuel hydrocarbons, and the less
oxidized chlorinated compounds such as DCE, 1,2-DCA, chlorobenzene, or vinyl chloride can be
used as primary growth substrates (electron donor) for microbial metabolism. The following sec-
tions describe biodegradation of organic compounds through use as a primary substrate under both
aerobic and anaerobic conditions.
B.3.2.1 Aerobic Biodegradation of Primary Substrates
Biodegradation of organic compounds is often an aerobic process that occurs when indigenous
populations of microorganisms are supplied with the oxygen and nutrients necessary to utilize
B3-33
-------
Table 115. J Trends in Contaminant, Electron Acceptor, Metabolic By-product and Total Alkalinity
Concentrations During Bi ode gradation
Analyte
Fuel Hydrocarbons
Highly Chlorinated Solvents and
Daughter Products
Lightly Chlorinated Solvents
Dissolved Oxygen
Nitrate
Manganese (II)
Iron (II)
Sulfate
Methane
Chloride
ORP
Alkalinity
Terminal Electron Accepting Process
Aerobic Respiration, Denitrification,
Manganese (IV) Reduction, Iron (III) Reduction,
Methanogenesis
Reductive Dechlorination
Aerobic Respiration, Denitrification,
Manganese (IV) Reduction, Iron (III) Reduction
(Direct Oxidation)
Aerobic Respiration
Denitrification
Manganese (IV) Reduction
Iron (III) Reduction
Sulfate Reduction
Methanogenesis
Reductive Dechlorination or Direct Oxidation of
Chlorinated Compound
Aerobic Respiration, Denitrification,
Manganese (IV) Reduction, Iron (III) Reduction,
Methanogenesis
Aerobic Respiration, Denitrification, Iron (III)
Reduction, and Sulfate Reduction
Trend in Analyte Concentration During
Biodegradation
Decreases
Parent Compound Concentration Decreases, Daughter
Products Increase Initially and Then
May Decrease
Compound Concentration Decreases
Decreases
Decreases
Increases
Increases
Decreases
Increases
Increases
Decreases
Increases
Redox Potential (I
In Millivolts @pH=V
1000 -r
Aerobic
Posslbta Rungs
for Reductive—
- 2NO,-+12H*+10er
Anaerobic
500-h
o--
OpUmal Range
tor Reductive
Dechkxtutlon
-500-1-
2H.O
MnCO,(a) + 2H.O
FeOOH(8) + HCO; + 2IT + e
• HS- + 4H.O (Eh" = -220)
• CH. + 2HJS (E,' = -2«)
Modified From Bomrar (1994)
Figure B.3.3 Oxidation-reduction potentials for various oxidation-reduction reactions.
B3-34
-------
organic carbon as an energy source. The biodegradation of fuel hydrocarbons occurs rapidly under
aerobic conditions and is discussed in Wiedemeier et al (1995a). Some pollutants, especially the
highly oxidized chlorinated hydrocarbons (i.e., those containing more chlorine substituents), are
biologically recalcitrant under aerobic conditions. However, some of the less chlorinated ethenes
and ethanes such as DCE, VC, and 1,2-DCA, and many of the chlorinated benzenes can be utilized
as primary substrates and oxidized under aerobic conditions. During aerobic biodegradation (oxida-
tion) of chlorinated solvents, the facilitating microorganism obtains energy and organic carbon from
the degraded solvent.
Of the chlorinated ethenes, vinyl chloride is the most susceptible to aerobic biodegradation, and
PCE the least. Of the chlorinated ethanes, 1,2-DCA is the most susceptible to aerobic biodegrada-
tion (chloroethane is more likely to abiotically hydrolyze to ethanol), while TCA, tetrachloroethane,
and hexachloroethane are less so. Chlorinated benzenes with up to 4 chlorine atoms (i.e., chloroben-
zene, dichlorobenzene, trichlorobenzene, and tetrachlorobenzene) also have been shown to be readily
biodegradable under aerobic conditions (Spain, 1996). Pentachlorobenzene and hexachlorobenzene
are unlikely to be oxidized by microbial activity.
B.3.2.1.1 Aerobic Oxidation of Petroleum Hydrocarbons
Fuel hydrocarbons are rapidly biodegraded when they are utilized as the primary electron donor
for microbial metabolism under aerobic conditions. Biodegradation of fuel hydrocarbons occurs
naturally when sufficient oxygen (or other electron acceptors) and nutrients are available in the
ground water. The rate of natural biodegradation is generally limited by the lack of oxygen or other
electron acceptors rather than by the lack of nutrients such as nitrogen or phosphorus. The rate of
natural aerobic biodegradation in unsaturated soil and shallow aquifers is largely dependent upon the
rate at which oxygen enters the contaminated media. Biodegradation of fuel hydrocarbons is dis-
cussed by Wiedemeier et al. (1995a).
B.3.2.1.2 Aerobic Oxidation of Chlorinated Ethenes
In general, the highly chlorinated ethenes (e.g., PCE and TCE) are not likely to serve as electron
donors or substrates for microbial degradation reactions. This is because the highly chlorinated
compounds tend to be much more oxidized than many compounds present in a natural ground-water
system.. Several microbes or microbial enrichments have been shown to be capable of TCE oxida-
tion (Fogel et al, 1986; Nelson et al., 1986; Little et al., 1988); however, as noted by Vogel (1994),
no strong evidence for the oxidation of highly chlorinated solvents has been derived from actual
hazardous waste sites.
Using microcosms from two different sites with no prior history of exposure to DCE,
Klier et al. (1998) show that all three isomers of DCE (i.e., 1,1-DCE, I-1,2-DCE, and trans-1,2-
DCE) can be biodegraded in aerobic systems. In these experiments, it was observed that cis-1,2-
DCE degraded more rapidly than the other isomers. Hartmans et al. (1985) and Hartmans and de
Bont (1992) show that vinyl chloride can be used as a primary substrate under aerobic conditions,
with vinyl chloride apparently being directly mineralized to carbon dioxide and water. This has also
been reported by Davis and Carpenter (1990). Aerobic biodegradation is rapid relative to other
mechanisms of vinyl chloride degradation, especially reductive dehalogenation.
B.3.2.1.3 Aerobic Oxidation of Chlorinated Ethanes
Of the chlorinated ethanes, only 1,2-dichloroethane has been shown to be aerobically mineral-
ized/oxidized. Stucki el al. (1983) and Janssen el al. (1985) show that 1,2-DCA can be used as a
primary substrate under aerobic conditions. In this case, the bacteria transform 1,2-DCA to
chloroethanol, which is then mineralized to carbon dioxide. Evidence of oxidation of chloroethane
is scant, however, it appears to rapidly degrade via abiotic mechanisms (hydrolysis) and is thus less
likely to undergo biodegradation.
B3-35
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B.3.2.1.4 Aerobic Oxidation of Chlorobenzenes
Chlorobenzene and polychlorinated benzenes (up to and including tetrachlorobenzene) have
been shown to be biodegradable under aerobic conditions. Several studies have shown that bacteria
are able to utilize chlorobenzene (Reineke and Knackmuss, 1984), 1,4-DCB (Reineke and
Knackmuss, 1984; SchraaefaL, 1986; Spain and Nishino, 1987), 1,3-DCB (de Bont etal, 1986),
1,2-DCB (Haigler etal, 1988), 1,2,4-TCB (van derMeeref or/., 1987; Sander eta/., 1991), and
1,2,4,5-TeCB (Sander etal, 1991) as primary growth substrates in aerobic systems. Nishino etal
(1994) note that aerobic bacteria able to grow on chlorobenzene have been detected at a variety of
chlorobenzene-contaminated sites, but not at uncontaminated sites. Spain (1996) notes that this
provides strong evidence that the bacteria are selected for their ability to derive carbon and energy
from chlorobenzene degradation in situ.
The pathways for all of these reactions are similar, and are also similar to that of benzene
(Chapelle, 1993; Spain, 1996). In general, the aerobic biodegradation involves hydroxylation of the
chlorinated benzene to a chlorocatechol, followed by ortho cleavage of the benzene ring. This
produces a muconic acid, which is dechlorinated, and the non-chlorinated intermediates are then
metabolized. The only significant difference between this process and aerobic benzene degradation
is the elimination of chlorine at some point in the pathway (Chapelle, 1993).
B.3.2.2 Anaerobic Biodegradation of Primary Substrates
Rapid depletion of dissolved oxygen caused by microbial respiration results in the establish-
ment of anaerobic conditions in areas with high organic carbon concentrations. Certain requirements
must be met in order for anaerobic (anoxic) bacteria to degrade organic compounds, including:
absence of dissolved oxygen; availability of carbon sources (natural or anthropogenic), electron
acceptors, and essential nutrients; and proper ranges of pH, temperature, salinity, and redox potential.
When oxygen is absent, nitrate, manganese (IV), iron (III), sulfate, and carbon dioxide can serve as
terminal electron acceptors during oxidation of organic carbon. While there is a large body of
evidence for anaerobic mineralization (oxidation) of fuel hydrocarbons, there is very little evidence
of such transformations involving chlorinated compounds.
B.3.2.2.1 Anaerobic Oxidation of Petroleum Hydrocarbons
Biodegradation of fuel hydrocarbons will occur under anaerobic conditions in most, if not all,
ground-water environments via denitrification, manganese (IV) reduction, iron (III) reduction, sulfate
reduction, and methanogenesis. Biodegradation of fuel hydrocarbons is discussed by Wiedemeier et
al. (1995a), and many primary references are cited therein.
B.3.2.2.2 Anaerobic Oxidation of Chlorinated Ethenes
In general, due to the oxidized nature of polychlorinated ethenes, they are unlikely to undergo
oxidation in groundwater systems. However, Bradley and Chapelle (1996) show that vinyl chloride
(with only one chlorine substituent) can be directly oxidized to carbon dioxide and water via
iron (III) reduction. Reduction of vinyl chloride concentrations in microcosms amended with iron
(III)-EDTA closely matched the production of carbon dioxide. Slight mineralization was also noted
in unamended microcosms. The rate of this reaction apparently depends on the bioavailability of the
iron (III). At this time, it is not known if other workers have demonstrated other anaerobic mineral-
ization reactions involving chlorinated ethenes.
B.3.2.2.3 Anaerobic Oxidation of Chlorinated Ethanes
During preparation of this protocol, no evidence of anaerobic oxidation of chlorinated ethanes
was found; this does not necessarily indicate that such reactions have not been described. However,
the lack of discussion of such transformations in surveys of chlorinated hydrocarbon biodegradation
(e.g., Vogel et al., 1987; McCarty and Semprini, 1994; Vogel, 1994, Adriaens and Vogel, 1995;
Spain, 1996) suggests that there has indeed been little, if any, work on this subject.
B3-36
-------
B.3.2.2.4 Anaerobic Oxidation of Chlorobenzenes
While aerobic mineralization of chlorobenzenes is similar to that of benzene, similar activity
under anaerobic conditions has not been documented. As discussed above, there is little, if any,
discussion of this topic in the literature.
BIODEGRADATION OF VIA USE AS AN
ACCEPTOR (REDUCTIVE
Bouwer el al. (1981) were the first to show that halogenated aliphatic hydrocarbons could be
biologically transformed under anaerobic conditions in the subsurface environment. Since that time,
numerous investigators have shown that chlorinated compounds can degrade via reductive dechlori-
nation under anaerobic conditions. Anaerobically, biodegradation of chlorinated solvents most often
proceeds through a process called reductive dechlorination. During this process, the halogenated
hydrocarbon is used as an electron acceptor, not as a source of carbon, and a halogen atom is re-
moved and replaced with a hydrogen atom. As an example, Dehalobacler reslriclus was shown by
Holliger el ai, (1993) to use tetrachloroethene as an electron acceptor during reductive dechlorina-
tion to produce c/'s-/,2-dichloroethene. Because chlorinated compounds are used as electron accep-
tors during reductive dechlorination, there must be an appropriate source of carbon for microbial
growth in order for reductive dehalogenation to occur (Back and Jaffe, 1989; Freedman and Gossett,
1989; Fathepure and Boyd, 1988; Bouwer, 1994). Potential carbon sources can include low molecu-
lar weight organic compounds (lactate, acetate, methanol, glucose, etc.), fuel hydrocarbons,
byproducts of fuel degradation (e.g., volatile fatty acids), or naturally occurring organic matter.
In some situations, reductive dechlorination may be a cometabolic process, in that the reaction
is incidental to normal metabolic functions and the organisms derive no benefit from the reaction.
Such cometabolism typically results in slow, incomplete dechlorination (Gantzer and Wackett, 1991;
Gossett and Zinder, 1996). More important, recent studies are discovering direct dechlorinators
(typically isolated from contaminated subsurface environments or treatment systems) that use chlori-
nated ethenes as electron acceptors in reactions that provide growth and energy (e.g., Holliger el a/,.,
1992; Holliger el al., 1993; Holliger and Schumacher, 1994; Neumann et al., 1994; Krumholz, 1995;
Maymo-Gatell etal., 1995; Sharma and McCarty, 1996; Gerritse elal., 1996). This process has been
termed both halorespiralion and dehalorespiration.
Biotic transformations of chlorinated solvents under anaerobic conditions generally are reduc-
tions that involve either hydrogenolysis or dihaloelimination (McCarty and Semprini, 1994).
Hydrogenolysis occurs when a chlorine atom is replaced with hydrogen. Dihaloelimination occurs
when two adjacent chlorine atoms are removed and a double bond is formed between the respective
carbon atoms. The most important process for the natural biodegradation of the more highly chlori-
nated solvents is reductive dechlorination (hydrogenolysis).
Higher ratios of chlorine to carbon represent higher oxidation levels; highly chlorinated com-
pounds are more oxidized than lesser chlorinated compounds and thus are less susceptible to oxida-
tion. Thus, highly chlorinated compounds such as PCE, TCE, TCA, or HCB are more likely to
undergo reductive reactions than oxidative reactions. During these reductive reactions, electrons are
transferred to the chlorinated compound, and a chlorine atom is replaced with a hydrogen atom. As
an example, consider the reductive dechlorination of PCE to TCE and then TCE to DCE, and finally
DCE to vinyl chloride. Because of the relatively low oxidation state of VC, this compound more
commonly undergoes aerobic biodegradation as a primary substrate than reductive dechlorination.
Reductive dechlorination processes result in the formation of intermediates which are more
reduced than the parent compound. These intermediates are often more susceptible to oxidative
bacterial metabolism than to further reductive anaerobic processes. Actual mechanisms of reductive
dehalogenation are still unclear, and in some cases may be a form of cometabolism (Gantzer and
Wackett, 1991; Adriaens and Vogel, 1995; Wackett, 1995). In addition, other factors that will influ-
B3-37
-------
ence the process include the type of electron donor and the presence of competing electron acceptors
(Adriaens and Vogel, 1995; Suflita and Townsend, 1995), temperature, and substrate availability.
Recent evidence suggests that dechlorination is dependent upon the supply of hydrogen (H2),
which acts as the electron donor in many such reactions (Gossett and Zinder, 1996; Smatlak el al.,
1996). The hydrogen is produced as a result of the microbial degradation of a primary substrate
(e.g., lactate, acetate, butyrate, ethanol, BTEX, or other such compounds). Bacteria that facilitate
dechlorination compete with sulfate-reducers and methanogens for the H, produced in such a system.
When degradation of the original substrate/electron donor rapidly yields high concentrations of H.,,
the sulfate-reducers and methanogens appear to be favored over the dechlorinators. Conversely,
when substrate degradation produces a steady supply of H, at low concentrations, the dechlorinators
are favored (Gossett and Zinder, 1996; Smatlak el al., 1996). Complete dechlorination is thus
apparently favored when a steady, low-concentration supply of H, is produced through microbial
degradation of substrates such as proprionate or benzoate (and, by extension from benzoate, the
BTEX compounds) (Gossett and Zinder, 1996). Therefore, the type of substrate/electron donor can
also play a role in how thoroughly a natural system is able to dechlorinate solvents.
One or more of the following generally is observed at a site where reductive dechlorination of
alkenes is ongoing:
1) Ethene is being produced (even low concentrations are indicative of biodegradation);
2) Methane is being produced;
3) Iron II is being produced;
4) Hydrogen concentrations are between 1-4 nM; and
5) Dissolved oxygen concentrations are low.
B.3.3.1 Reductive Dechlorination of Chlorinated Ethenes
PCE and TCE have been shown to undergo reductive dechlorination in a variety of anaerobic
systems from different environments, with various electron donors/carbon sources (Table B.3.4)
(Wilson, 1988; Sewell et al., 1991; Roberts et al., 1982). This is particularly true if the subsurface
also contains other anthropogenic or native organic compounds that can serve as electron donors and
whose utilization by subsurface bacteria will deplete any available oxygen. In general, reductive
dechlorination of chlorinated ethenes occurs by sequential dechlorination from PCE to TCE to DCE
to VC to ethene. Depending upon environmental conditions, this sequence may be interrupted, with
other processes then acting upon the products. With sufficient quantities or appropriate types of
electron donors (e.g., slow but steady Reproduction), the final end-product of anaerobic reductive
dehalogenation can be ethene (Freedman and Gossett, 1989). Reductive dehalogenation of chlori-
nated solvent compounds is associated with the accumulation of daughter products and an increase
in chloride.
Studies have shown that PCE and TCE can be anaerobically reduced to either 1,1-DCE, c/s-1,2-
DCE, or /ra«s-l,2-DCE, all of which can be further transformed to vinyl chloride (Miller and
Guengerich, 1982; Wilson and Wilson, 1985; Mayer el al, 1988; Nelson, et al, 1986; Henson et al,
1989; Tsien etal, 1989; Henry, 1991; McCarty, 1994; Wilson etal., 1994). During reductive
dehalogenation, all three isomers of DCE can theoretically be produced; however, Bouwer (1994)
reports that cis-J,2-DCE is a more common intermediate than trans-J,2-DCE and that 7,7-DCE is
the least prevalent intermediate of the three DCE isomers. Vinyl chloride produced from
dehalogenation of DCE may be subsequently reduced to innocuous products such as ethane or
carbon dioxide. The removal of vinyl chloride occurs more readily under aerobic conditions, such as
those encountered at the edge of the plume. Vinyl chloride may also be used as a primary substrate
by aerobic organisms, as previously discussed.
B3-38
-------
Table B.3.4
Sources, Donors, Acceptors, and Products of Reductive De chlorinating Laboratory Systems
Reference
Bouwer & McCarty,1983
Vogel & McCarty, 1985
Kleopfere^a/., 1985
3arrio-Lage etal., 1987
"athepure etal., 1987
Baek& Jaffe, 1989
"reedman & Gossett, 1989
Scholz-Muramatsu etal., 1990
Gibson & Sewell, 1990
Sewell & Gibson, 1990
Sewell etal., 1991
Lyonetal., 1995
Source
Digester
3ioreactor
Soil
Swamp Muck
Soil
VIethanosarcina
DCB-1
digester
digester
Bioreactor
Aquifer
Aquifer
Aquifer
Landfill
Aquifer
Donor
Organic Material
Acetate
Soybean Meal
Organic Material
Methanol (?)
Methanol
3CBa,Pyruvate,RFb
Formate
Methanol
Methanol
Glucose
H2
Formate
Acetate
Benzoate
VFAd
Toluene
VFA
VFA
Acceptor-Product
PCE-TCE
PCE-VC, CO2
TCE-DCE
PCE-VC
PCE-VC
PCE-TCE
PCE-TCE
TCE-VC,CAC
TCE-VC,CA
PCE-VC, Ethene
PCE-VC, Ethene
PCE-VC, Ethene
PCE-VC, Ethene
PCE-VC, Ethene
PCE-DCE
PCE-DCE
PCE-DCE
PCE-DCE
PCE-VC
Native Organic Matter [PCE-DCE
a 3-Chlorobenzoate
h Rumen Fluid
c CMoroethane
d Volatile FattvAcid
B3-39
-------
B.3.3.2 Reductive Deehlorinatlon of Chlorinated Ethanes
As with the ethenes, chlorinated ethanes will also undergo reductive dehalogenation in the
subsurface via use as electron acceptors. Dechlorination of TCA has been described by Vogel and
McCarty (1987) and Cox et al. (1995), but this pathway is complicated by the abiotic reactions that
can affect TCA and its byproducts (Vogel, 1994).
B.3.3.3 Reductive Dechlorinatlon of Chlorobenzenes
For the highly chlorinated benzenes (e.g., hexachlorobenzene and pentachlorobenzene, as well
as tetrachlorobenzene, and trichlorobenzene), reductive dechlorination is the most likely biodegrada-
tion mechanism (Holliger et al., 1992; Ramanand et al., 1993; Suflita and Townsend, 1995). As
discussed by Suflita and Townsend (1995), reductive dehalogenation of aromatic compounds has
been observed in a variety of anaerobic habitats, including aquifer materials, marine and freshwater
sediments, sewage sludges, and soil samples; however, isolation of specific microbes capable of
these reactions has been difficult. As with the chlorinated ethenes and ethanes, the chlorobenzenes
are most likely acting as electron acceptors as other sources of carbon and energy are being utilized
by microbes or microbial consortia (Suflita and Townsend, 1995). Evidence has been presented
suggesting that oxidation of hydrogen using halogenated aromatics as electron acceptors may yield
more energy than if more commonly available electron acceptors were used (Dolfing and Harrison,
1992).
As discussed previously, the actual mechanisms of reductive dehalogenation are not well under-
stood. Further, reductive dehalogenation of chlorinated benzenes has not been as well-documented
as for other chlorinated solvents. However, reductive dechlorination of chlorobenzenes has been
documented more frequently in the past several years (e.g., Bosma et al., 1988; Fathepure et al.,
1988; Fathepure and Vogel, 1991; Holliger et al, 1992; Ramanand el al, 1993). As with other
chlorinated solvents, the reductive dehalogenation of chlorobenzenes is affected by the degree of
chlorination of the compound. The more chlorinated aromatic compounds are typically more ame-
nable to this reaction (Suflita and Townsend, 1995; Adriaens and Vogel, 1995), but as they are
dechlorinated, the daughter products will become more resistant to further dehalogenation reactions
(Fathepure et al, 1988; Bosma et al., 1988; Holliger et al., 1992). The reductive dechlorination of
chlorobenzenes is analogous to reactions involving chlorinated ethenes and ethanes in that such
degradation will make them more amenable to aerobic biodegradation (Schraa, etal., 1986; Spain
and Nishino, 1987; Ramanand et al., 1993).
B.3.4 BIODEGRADATION OF VIA
When a chlorinated solvent is biodegraded through cometabolism, it serves as neither an elec-
tron acceptor nor a primary substrate in a biologically mediated redox reaction. Instead, the degrada-
tion of the compound is catalyzed by an enzyme cofactor that is fortuitously produced by organisms
for other purposes. The best-documented cometabolism reactions involve catabolic oxygenases that
catalyze the initial step in oxidation of their respective primary or growth substrate (BTEX or other
organic compounds). These oxygenases are typically nonspecific and, therefore, fortuitously initiate
oxidation of a variety of compounds, including many of the CAHs (McCarty and Semprini, 1994).
The organism receives no known benefit from the degradation of the chlorinated solvent; in some
cases the cometabolic degradation of the solvent may, in fact, be harmful to the microorganism
responsible for the production of the enzyme or cofactor (McCarty and Semprini, 1994). Chlorinated
solvents are usually only partially transformed during cometabolic processes, with additional biotic
or abiotic degradation generally required to complete the transformation (McCarty and Semprini,
1994).
Cometabolism is best documented for CAHs in aerobic environments; evidence of
cometabolism of chlorobenzenes is scant, as is clear evidence of anaerobic cometabolism. In an
B3-40
-------
aerobic environment, many chlorinated organic compounds can only be degraded via cometabolism.
It has been reported that under aerobic conditions chlorinated ethenes, with the exception of PCE,
are susceptible to cometabolic degradation (Murray and Richardson, 1993; Vogel, 1994; McCarty
and Semprini, 1994; Adriaens and Vogel, 1995). Vogel (1994) further elaborates that the oxidation
rate increases as the degree of chlorination decreases. Aerobic cometabolism of ethenes may be
characterized by a loss of contaminant mass, the presence of intermediate degradation products (e.g.,
chlorinated oxides, aldehydes, ethanols, and epoxides), and the presence of other products such as
chloride, carbon dioxide, carbon monoxide, and a variety of organic acids (Miller and Guengerich,
1982; McCarty and Semprini, 1994).
The lack of clear evidence for anaerobic cometabolism does not necessarily imply that such
transformations do not occur; in some cases, reductive dechlorination may be a result of
cometabolism (e.g., Gantzer and Wackett, 1991), depending upon the relationship between the
microbes, substrates, contaminants, and other electron acceptors. However, as with aerobic
cometabolism, anaerobic cometabolism will be slow relative to dehalorespiration and might not be
distinguishable at the field scale (Gossett and Zinder, 1996).
Several groups of aerobic bacteria currently are recognized as being capable of transforming
TCE and other CAHs via cometabolism; these groups include:
* Methane Oxidizers (Methanotrophs) (Fogel el al., 1986; Little el al., 1988, Mayer el al.,
1988; Oldenhuis etal., 1989; Tsien etal., 1989; Henry and Grbic-Galic, 1990; Alvarez-
Cohen and McCarty, 1991a,b; Henry and Grbic-Galic, 1991a,b; Lanzarone and McCarty,
1990; Oldenhuis etal., 1991);
* Propane Oxidizers (Wackett el al., 1989);
« Ethene Oxidizers (Henry, 1991);
• Toluene, Phenol, or Cresol Oxidizers (Nelson el al, 1986, 1987, 1988; Wackett and
Gibson, 1988; Folsom etal., 1990; Barker and Kim, 1990);
* Ammonia Oxidizers (Arciero el al., 1989; Vannelli el al., 1990);
« Isoprene Oxidizers (Ewers etal., 1991); and
* Vinyl Chloride Oxidizers (Hartmans and de Bont, 1992).
These bacteria all have catabolic oxygenases that catalyze the initial step in oxidation of their
respective primary or growth substrates and have the potential for initiating the oxidation of CAHs.
Cometabolism is not nearly as important a degradation mechanism for chlorinated solvents in
the saturated zone as reductive dehalogenation. Due to the need for a substrate that may be present
in limited concentrations, as well as the fortuitous nature of the reactions, rates of cometabolism are
often slow enough that this process may not be detectable unless the system is stimulated with
additional substrate mass. For a discussion of this topic, see McCarty and Semprini (1994) or
Wackett (1995).
Electron transfer results in oxidation of the electron donor and reduction of the electron accep-
tor and the production of usable energy. The energy produced by these reactions is quantified by the
Gibbs free energy of the reaction (Gr) which is given by:
AGr0 = Ł &G°Lproducts - Ł &G°Lreactants eq. B.3.1
Where:
AGr = Gibbs Free Energy of the Reaction at Standard State
f products = Gibbs Free Energy of Formation for Products at Standard State
,reactants = Gibbs Free Energy of Formation for the Reactants at Standard State
B3-41
-------
The Gr defines the maximum useful energy change for a chemical reaction at a constant tem-
perature and pressure. Table B.3.5 presents select electron acceptor and electron donor half-cell
reactions and the calculated Gr values. Table B.3.6 gives the Gibbs free energy of formation (Gf) for
species used in these half-cell reactions. Table B.3.7 presents coupled oxidation-reduction reactions.
In general, those reactions that yield the most energy tend to take precedence over less energy-
yielding reactions. However, the calculated energy yield of processes involving anthropogenic
organic compounds may not be reflected in the true energy yield of the metabolic process.
Figure B.3.4 illustrates the expected sequence of microbially mediated redox reactions based on Gr.
There is sufficient energy in the reaction of fuel hydrocarbons with chlorinated solvents to allow
their use by microorganisms as physiological electron acceptors.
B3-42
-------
Table B. 3.5 Electron Donor and Electron Acceptor Half-Cell Reactions
HALF-CELL REACTIONS
AG°r(kcal/
equiv)*
AG°r(kJ/
equiv)*
E°
(V)
Eh
(V)
pe
Conditions
for Eh and pe §
ELECTRON- ACCEPTOR (REDUCTION) HALF CELL REACTIONS
5e + 6H+ + NO i => 0.5N2 + 3Hf>
Denitrification
4e + 4H+ + O2=> 2H2O
Aerobic Respiration
2e~ + 4H+ + MnO, =>Mn2+ + 2Hfi
Pyrolusite Dissolution/Reduction
CO, + e + H+ + MnOOH =>MnCO3 + H,O
Manganite Carbonation/Reduction
e + H+ + MnO, => MnOOH
Pyrolusite Hydrolysis/Reduction
e + 3H+ + FefOHh^r,!, _=>Fe2+ + 3H2O
Amorphous "Goethite " Dissolution/Reduction
8e + 10H+ + NO' 3 =^NH+4 + 3H2O
Nitrate Reduction
2e + 2H+ + NO' 3 ^NO~2 + H2O
Nitrate Reduction
e + 3H+ + FeOOH =>Fe2+ + 2H,O
"Ferric oxyhydroxide " Dissolution/Reduction
e~ + 3H+ + FeiOH),,,,,, =>Fe2+ + 3H2O
Crystallized "Goethite " Dissolution/Reduction
e + H+ + CO2_g + Fe{QHJ_3:amph. =>FeCO3 + 2H2O
Amorphous "Goethite " Carbonation/Reduction
8e + 9H+ + SO2- 4 =>HS~ + 4H2O
Sulfate Reduction
8e + 10H+ + SO2- 4 =>H2S" + 4H2O
Sulfate Reduction
8e + 8H+ + CO2_g =* CH4,S + 2H2O
Methanogenesis
C2C14 +H+ + 2e => C2HC13 + Ct
PCE Reductive Dechlorination
C2HC13 +H+ + 2e =$ C2H2C12 + Ct
TCE Reductive Dechlorination
C3H2C12 +H+ + 2e =$ C2H3Cl + Ct
c-DCE Reductive Dechlorination
C2H3Cl +H+ + 2e =$ C2H4 + Ct
VC Reductive Dechlorination
C2H2C14 + H+ + 2e =$ C2H3C13 + CT
PCA Reductive Dechlorination
C2H3C13 +H+ + 2e =1 C2H4C12 + Ct
TCA Reductive Dechlorination
C2H4C12 + H+ + 2e =1 C2HSCI + Ct
DCA Reductive Dechlorination
C6C16 + H+ + 2e => CnHCls + CT
Hexachlorobenzene Reductive Dechlorination
CtfiCls +H+ + 2e =$ Cfffk + Ct
Pentachlorobenzene Reductive Dechlorination
dfi2Cl4 + H+ + 2e =$ dfi3Cl, + Ct
Tetrachlorobenzene Reductive Dechlorination
CfL3Cl3 +H+ + 2e =$ CfL4Cl2 + Ct
Trichlorobenzene Reductive Dechlorination
-28.7
-28.3
-28.3
-23.1
-22.1
-21.5
-20.3
-18.9
-15.0
-11.8
-11.0
-5.74
-6.93
-3.91
-14.79
-14.50
-12.12
-13.75
-13.59
-15.26
-14.08
-14.36
-14.64
-13.66
-13.20
-120.
-119.
-119
-96.8
-92.5
-89.9
-84.9
-78.9
-62.9
-49.2
-46.2
-24.0
-28.9
-16.4
-61.8
-60.6
-50.7
-57.5
-56.8
-63.8
-58.9
-60.0
-61.2
-57.1
-55.2
+1.24
+1.23
+1.23
+1.00
+0.959
+0.932
+0.879
+0.819
+0.652
+0.510
+0.479
+0.249
+0.301
+0.169
+0.641
+0.628
+0.525
+0.596
+0.589
+0.661
+0.610
+0.622
+0.634
+0.592
+0.572
+0.708
+0.805
+1.169
+0.408
+0.545
+0.163
+0.362
+0.404
-0.118
-0.259
-0.113
-0.278
-0.143
-0.259
+0.552
+0.539
+0.436
+0.507
+0.500
+0.572
+0.521
+0.533
+0.545
+0.503
+0.483
+12.0
+13.6
+19.8
+6.90
+9.21
+2.75
+6.12
+6.82
-1.99
-4.38
-1.90
-4.70
-2.42
-4.39
+9.33
+9.12
+7.38
+8.57
+8.45
+9.67
+8.81
+9.01
+9.22
+8.50
+8.17
pH = 7
Ł[N]=1Q-3
pH = 7
Po =0.21 atm
pH = 7
S[Mn]=10'5
pH = 8
Pco2=10-2
pH = 7
pH = 6
X[Fe]=10-5
pH = 7
pH = 7
pH = 6
I [Fe]=10'5
pH = 6
I [Fe]=10'5
pH = 8
Pco =10"2atm
pH = 8
pH = 6
pH = 7
Pco2=10-2
PCH4=10°
pH = 7
[Cl-]=10'4
pH = 7
[Cl-]=10'4
pH = 7
[Cl-]=10'4
pH = 7
[Cl-]=10~4
pH = 7
[C1-]=1Q-4
pH = 7
[Cl-]=10'4
pH = 7
[Cl-]=10'4
pH = 7
[Cl-]=10'4
pH = 7
[Cl-]=10'4
pH = 7
[Cl-]=10~4
pH = 7
[C1-]=1Q-4
B3-43
-------
Table R 3.5
Continued.
HALF-CELL REACTIONS
AG°r(kcal/
equiv)*
AG°r(kJ/
equiv)*
E°
(V)
Eh
(V)
pe
Conditions
for Eh and pe §
ELECTRON-DONOR (OXIDATION) HALF CELL REACTIONS
12H20 + C6H6 => 6CO2 + 30H+ + 30e
Benzene Oxidation
14H20 + CtfisCH, =* 7CO2 + 36H+ + 36e
Toluene Oxidation
16H20 + C6H$C2H5 =* SCO 2 + 42H+ + 42e
Ethylbenzene Oxidation
16H20 + C6H4(CH3)2=> SCO 2 + 42H+ + 42e
m-Xylene Oxidation
20H2O + CioHg => 10CO2 + 48H+ + 48e
Naphthalene Oxidation
ISHff + C6H3(CH3)3 => 9CO2 + 48H+ + 48e
1,3,5-Trimethylbenzene Oxidation
ISHff + C6H3(CH3)3 => 9CO2 + 48H+ + 48e
1,2,4-Trimethylbenzene Oxidation
4H2O + C2H2C12 => 2CO2 + 10H+ + 8e + 2CT
DCE Oxidation
4H2O + C2H3Cl => 2CO2 + 11H+ + We + CT
Vinyl Chloride Oxidation
12H2O + C6H2C14 => 6CO2 + 26H+ + 22e + 4CT
Tetrachlorobenzene Oxidation
12H2O + CtHsCls => 6CO2 + 27H+ + 24e + 3Ct
Trichlorobenzene Oxidation
12H2O + C6H4C12 => 6CO2 + 28H+ + 26e + 2Ct
Dichlorobenzene Oxidation
12H2O + CnHffl => 6CO2 + 29H+ + 28e + Ct
Chlorobenzene Oxidation
+2.83
+2.96
+2.96
+3.03
+2.98
+3.07
+3.07
-3.88
-0.55
-0.64
+0.42
+1.40
+2.22
+11.8
+12.4
+12.4
+12.7
+12.5
+12.8
+12.9
-16.2
-2.31
-2.68
+1.77
+5.84
+9.26
-0.122
-0.128
-0.128
-0.132
-0.130a
-0.133"
-0.134"
+0.168
+0.024"
+0.028
-0.018
-0.060
-0.096"
+0.316
+0.309
+0.309
+0.303
+0.309
+0.303
+0.302
-0.131
-0.006
+0. 016
-0.030
-0.071
-0.107
+5.34
+5.22
+5.21
+5.12
+5.22
+5.12
+5.11
-2.21
-0.10
+0.27
-0.50
-1.21
-1.80
pH = 7
Pco2=10-2
pH = 7
Pco2=10-2
pH = 7
Pco2=10'2
pH = 7
Pco2=10-2
pH = 7
Pco2 =10'2
pH = 7
Pco2=10'2
pH = 7
Pco2=10'2
pH = 7
Pco2=10-2
pH = 7
Pco2=10-2
pH = 7
Pco2=10-2
pH = 7
Pco2=10'2
pH = 7
Pco2=10'2
pH = 7
Pco2=10-2
NOTES:
* = AG°r for half-cell reaction as shown divided by the number of electrons involved in reaction.
§ = Conditions assumed for the calculation of Eh and pe (pc = Eh/0.05916). Where two dissolved species arc involved,
other than those mentioned in this column, their activities are taken as equal. Note, this does not affect the free
energy values listed.
a = E° calculated using the following equation; E° = AG°r (J/nF) * 1.0365xl()'5 (VF/J) from Stumm and Morgan, 1981.
B3-44
-------
Table B.3.6 Gibbs Free Energy of Formation for Species used in Half-Cell Reactions and Coupled
Oxidation-Reduction Reactions
Species
e"
H+
O2
H20
State
i
i
R
1
AG°f)298.i5
(kcal/mole)
0
0
0
-56.687
Source
std
std
std
Dean (1972)
Carbon Species
C02
CH2O, formaldehyde
CgHs, benzene
CH4, methane
CgHsCHs, toluene
C6H5C2H5, ethylbenzene
C6H4(CH3)2, o-xylene
C6H4(CH3)2, m-xylene
C6H4(CH3)2, p-xylene
C2C14, PCE
C2HC13, TCE
C2H2C12 1,1-dichloroethene
C2H2C12 cis-l,2-dichloroethene
C2H2C12 trans- 1,2-
dichloroethene
C2H4 Ethene
C2H6 Ethane
HC1 hydrochloric acid
C2H2C14, 1,1,2,2-PCA
C2H3C13, 1,1,2-TCA
C2H4C12, 1,2-DCA
C2H5Cll5 Chloroethane
Ci0H8, naphthalene
C6H3(CH3)3, 1,3,5-TMB
C6H3(CH3)3, 1,2,4-TMB
C2H3C1, Vinyl chloride
C6C16, Hexachlorobenzene
CsHiCls, Pentachlorobenzene
C6H2C14, 1,2,4,5-
Tetrachlorobenzene
C6H3C13, 1,2,4-
Trichlorobenzene
C6H4C12, 1 ,4-Dichlorobenzene
CsHsCl, chlorobenzene
C14H10, phenanthrene
R
aq
1
R
1
1
1
1
1
1
1
1
1
1
R
aq, m=l
R
aq, m=l
aq, m=l
1
R
R
R
1
1
1
R
1
1
1
1
1
1
1
-94.26
-31.02
+29.72
-12.15
+27.19
+28.61
+26.37
+25.73
+26.31
+1.1
+2.9
+5.85
5.27
+6.52
+16.28
+19.43
-7.68
-4.09
-31.372
-22.73
-18.54
-17.68
-14.47
+48.05
+24.83
+24.46
+12.4
+0.502
+3.16
+5.26
+9.31
+14.28
+21.32
+64.12
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
CRC Handbook (1996)
CRC Handbook (1996)
Dean (1972)
CRC Handbook (1996)
CRC Handbook (1996)
CRC Handbook (1996)
CRC Handbook (1996)
CRC Handbook (1996)a
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Dolfing and Harrison (1992)
Dolfmg and Harrison (1992)
Dolfing and Harrison (1992)
Dolfing and Harrison (1992)
Dolfing and Harrison (1992)
Dean (1972)
Dean (1972)
B3-45
-------
Table B. 3.6
Continued.
Species
State
AG°f)298.15
(kcal/mole)
Source
Nitrogen Species
NO3"
N2
NO2
NH4+
I
g
I
aq
Sulfur S
SO42"
H2S
H2S
HS-
i
aq
g
i
-26.61
0
-7.7
-18.97
Dean (1972)
std
Dean (1972)
Dean (1972)
pecies
-177.97
-6.66
-7.9
+2.88
Dean (1972)
Dean (1972)
Dean (1972)
Dean (1972)
Iron Species
Fe2+
Fe3+
Fe2O3, hematite
FeOOH, ferric oxyhydroxide
Fe(OH)3, goethite
Fe(OH)3, goethite
FeCO3, siderite
i
i
c
c
a
c
c
-18.85
-1.1
-177.4
-117.2
-167.416
-177.148
-159.35
Dean (1972)
Dean (1972)
Dean (1972)
Naumove/a/. (1974)
Langmuir and Whittemore
(1971)
Langmuir and Whittemore
(1971)
Dean (1972)
Manganese Species
Mn2+
MnO2, pyrolusite
MnOOH, manganite
MnCO3, rhodochrosite
i
c
c
P
-54.5
-111.18
-133.29
-194
Dean (1972)
Stumm and Morgan
(1981)
Stumm and Morgan
(1981)
Dean (1972)
Chloride Species
cr
aq
-31.37
Dean (1972)
NOTES:
c = crystallized solid 1 = liquid g = gaseous aq = undissociated aqueous species
a = amorphous solid (may be partially crystallized - dependent on methods of preparation)
p = freshly precipitated solid
i = dissociated, aqueous ionic species (concentration = 1 m)
std = accepted by convention
Wherever possible multiple sources were consulted to eliminate the possibility of typographical error.
B3-46
-------
Table B.3.7 Coupled Oxidation-Reduction Reactions
Coupled Benzene Oxidation Reactions
7. 50 2 + C6H6 => 6C02,g + 3H20
Benzene oxidation /aerobic respiration
6 NO 3 + 6H+ + CeHe => 6CO2g + 6H2O + 3N2g
Benzene oxidation / denitrifwation
30H+ + 15MnQ2 + C6H6 => 6OO2g + 15Mn2+ + 18H2O
Benzene oxidation / manganese reduction
3.75 NCV + C6H6 + 7.5 H+ + 0.75 H2O =>6 CO2 + 3.75 NH4+
Benzene oxidation / nitrate reduction
60H+ + 30Fe(OH)3a + C6H6 => 6CO2 + 30Fe2+ + 78H2O
Benzene oxidation / iron reduction
7.5H+ + 3.75SO2i + C6H6 => 6CO2,S + 3.15H2S° + 3H2O
Benzene oxidation / sulfate reduction
4.5H2O + C6H6^ 225CO2,g + 3.75CH4
Benzene oxidation / methanogenesis
15 C2H2C14 + C6H6 + 12 H2O =* 6 CO2 + 15 C2H3C13 +15 H+ + 15 Cl"
Benzene oxidation/ PCA reduction
15 C2H3C13 + C6H6 + 12 H2O 6 =* CO2 + 15 C2H4C12 +15 H+ + 15 Cl"
Benzene oxidation / TCA reduction
15 C2H4C12 + C6H6 + 12 H2O =* 6 CO2 + 15 C2H5C1 +15 H+ + 15 Cl"
Benzene oxidation / DC A reduction
15C2C14 + 12H2O + C6H6 => ISCzHCh + 6CO2 + 15H+ + ISCT
Benzene oxidation/ Tetrachloroethylene reductive dehalogenation
ISCiHCh + 12H2O + C6H6 => ISCiHiCh + 6CO2 + 15H+ + 15CT
Benzene oxidation/ Trichloroethylene reductive dehalogenation
15C2H2C12 + 12H2O + C 15C2H3Cl + 6CO2 + 15H+ + ISCt
Benzene oxidation/ cis-Dichloroethylene reductive dehalogenation
15C2H3Cl + 12H2O + C6H6 => 15CiH4 + 6CO2 + 15H+ + ISCT
Benzene oxidation/ Vinyl chloride reductive dehalogenation
15C6Ck + 12H2O + C6H6 => 15C6HiCl5 + 6CO2 + 15H+ + 15CI
Benzene oxidation/ Hexachlorobenzene reductive dehalogenation
15C 15C ISCuHiCh + 6CO2 + 15lT + 15CT
Benzene oxidation/ Tetrachlorobenzene reductive dehalogenation
ISCiHsCh + 12H2O + CaHi =* 15C6H4C12 + 6CO2 + 15H+ + 15CT
Benzene oxidation/ Trichlorobenzene reductive dehalogenation
AG°r
(kcal/mole)
-765.34
-775.75
-765.45
-524.1
-560.10
-122.93
-32.40
-322.7
-372.65
-337.40
-358.55
-331.25
-297.35
-327.35
-345.68
-354.05
-324.80
-311.0
AG°r
(kJ/ mole)
-3202
-3245
-3202
-2193
-2343
-514.3
-135.6
-1349
-1558
-1410
-1499
-1385
-1243
-1368
-1445
-1480
-1358
-1300
Stoichiometric
Mass Ratio of
Electron Acceptor
or Metabolic
Byproduct to
Primary Substrate
3.07:1
4.77:1
10.56:1
2.98:1
21.5:1
4.61:1
0.77:1
32.2:1
25.6:1
19.0:1
31.8:1
25.2:1
18.6:1
12.0:1
54.7:1
48.1:1
41.5:1
34.8:1
Mass of Primary
Substrate Utilized per
Mass of Electron
Acceptor Utilized or
Metabolic Byproduct
Produced
0.326:1
0.210:1
0.095:1
0.336:1
0.047:1
0.22:1
1.30:1
0.03:1
0.04:1
0.05:1
0.03:1
0.04:1
0.05:1
0.08:1
0.02:1
0.02:1
0.02:1
0.03:1
B3-47
-------
Table B. 3.7 Continued.
Coupled Toluene Oxidation Reactions
9O2 + C6H5CH3 => 7CO2,g + 4H2O
Toluene oxidation /aerobic respiration
7.2NO3 + 7.2H* + CeHsCHs => 7CO2g + 7.6H2O + a6JV2g
Toluene oxidation / denitriftcation
36H+ + I8MnO2 + C6H5CH3 => 7CO2g + I8MR* + 22H2O
Toluene oxidation / manganese reduction
72H+ + 36FefOHj3a + C6H5CH3 => 7CO2 + 36F^ + 94H2O
Toluene oxidation / iron reduction
9H* + 4.5SO24 + C6H5CH3 => 7CO2g + 4.5H2S> + 4H2O
Toluene oxidation / sulfate reduction
5H2O + C6H5CH3 => 25CO2g + 4.5CH4
Toluene oxidation / methanogenesis
18 C2H2C14 + C6H5CH3 + 14 H2O ^7 CO2 + 18 C2H3C13 + 18H+ + 18C1"
Toluene oxidation/ PCA reduction
18 C2H3C13 + C6H5CH3 + 14 H2O ^7 CO2+ 18 C2H4C12 + 18H+ + 18C1"
Toluene oxidation / TCA reduction
18 C2H4C12 + C6H5CH3 + 14 H2O =* 7 CO2 + 18 C2H5C1 + 18 H+ +18 Cl"
Toluene oxidation / DCA reduction
18C2C14 + 14H2O + C6H5CH3 => 18C2HC13 + 7CO2 + 18H+ + 18CT
Toluene oxidation/ Tetrachloroethylene reductive dehalogenation
ISCiHCh + 14H2O + C6H5CH3 => 18C2H2Ch + 7 CO 2 + 18H+ + 18CI
Toluene oxidation/ Trichloroethylene reductive dehalogenation
18C2H2C12 + 14H2O + CtjHsCHi => ISCiHjCl + 7CO2 + 18H+ + 18CT
Toluene oxidation/ cis-Dichloroethylene reductive dehalogenation
18C2H3Cl + 14H2O + CiHsCHs => 18C2H4 + 7 CO 2 + 18H+ + 18Ct
Toluene oxidation/ Vinyl chloride reductive dehalogenation
18C6C16 + 14H2O + C6H5CH3 => ISCeHiCh + 7CO2 + 18H+ + 18CI
Toluene oxidation/ Hexachlorobenzene reductive dehalogenation
ISCtjHiCh + 14H2O + CtjHsCHj => 18C ISCaHfh + 7CO2 + 18H+ + 18CI
Toluene oxidation/ Tetrachlorobenzene reductive dehalogenation
18C6H3Ch + 14H2O + C6H5CH3 => 18C6H4Ch + 7CO2 + 18H+ + 18CI
Toluene oxidation/ Trichlorobenzene reductive dehalogenation
AG°r
(kcal/ mole)
-913.76
-926.31
-913.89
-667.21
-142.86
-34.08
-382.6
-442.5
-400.2
-425.6
-404.9
-340.1
-331.5
-410.3
-420.3
-385.2
-368.6
AG°r
(kJ/ mole)
-3823
-3875
-3824
-2792
-597.7
-142.6
-1599
-1850
-1673
-1779
-1693
-1422
-1386
-1715
-1757
-1610
-1541
Stoichiometric
Mass Ratio of
Electron Acceptor
or Metabolic
Byproduct to
Primary Substrate
3.13:1
4.85:1
10.74:1
21.86:1
4.7:1
0.78:1
32.8:1
26.1:1
19.3:1
32.4:1
25.7:1
18.9:1
12.2:1
55.6:1
48.9:1
42.2:1
35.4:1
Mass of Primary
Substrate Utilized per
Mass of Electron
Acceptor Utilized or
Metabolic Byproduct
Produced
0.32:1
0.21:1
0.09:1
0.05:1
0.21:1
1.28:1
0.03:1
0.04:1
0.05:1
0.03:1
0.04:1
0.05:1
0.08:1
0.02:1
0.02:1
0.02:1
0.03:1
B3-48
-------
Table B. 3.7 Continued.
Coupled Ethylbenzene Oxidation reactions
10.5O2 + CeHsCzHs => 8COZg + 5H2O
Ethylbenzene oxidation /aerobic respiration
8.4NO3 + 8.4H* + CeHsCsHs => 8CO2g + 9.2H2O + 4.2N2,g
Ethylbenzene oxidation / denitriflcation
46H* + 22MnO2 + C6H5C2H5 => 8CO2g + 22Mn* + 28H2O
Ethylbenzene oxidation / manganese reduction
84H* + 42Fe(OH)3a + C6H5C2H5 => 8CO2 + 42Fe* + 110H2(
Ethylbenzene oxidation / iron reduction
10.5H+ + 5.25SO24 + C6H5C2H5 => 8CO2g + 5.25H2S> + 5H2
Ethylbenzene oxidation / sulfate reduction
5.5H2O + CsH5C2H5 => 2.75CO2,g + 5.25CH4
Ethylbenzene oxidation / methanogenesis
21C2H2C14 + 16H2O + CtfiiCzHi => 21C2H3Ch + SCO 2 + 21H+ + 21 Ct
Ethylbenzene oxidation/ PCA reductive dehalogenation
21C2 H3Ch + 16H2O + C6H5C2H5 => 21C2H4Ch + SCO 2 + 21H+ + 21CI
Ethylbenzene oxidation/ TCA reductive dehalogenation
21C2H4C12 + 16H2O + C 21C2H5Cl + SCO 2 + 21H+ + 21CT
Ethylbenzene oxidation/ DCA reductive dehalogenation
21C2C14 + 16H2O + C6HSC2HS => 21C2HC13 + SCO 2 + 21H+ + 21 Ct
Ethylbenzene oxidation/ Tetrachloroethylene reductive dehalogenation
21C2HCh + 16H2O + C6H5C2H5 => 21C2H2C12 + SCO 2 + 21H+ + 21CI
Ethylbenzene oxidation/ Trichloroethylene reductive dehalogenation
21C2H2C12 + 16H2O + CtjHsCtfs => 21C2H3Cl + SCO 2 + 21H+ + 21CI
Ethylbenzene oxidation/ cis-Dichloroethylene reductive dehalogenation
21C2H3Cl + 16H2O + CiHsC^s => 21C2H4 + SCO 2 + 21H+ + 21 Cl'
Ethylbenzene oxidation/ Vinyl chloride reductive dehalogenation
21C6C16 + 16H2O + C6H5C2H5 => 21C6HiCls + 8CO2 + 21H+ + 21CI
Ethylbenzene oxidation/ Hexachlorobenzene reductive dehalogenation
21C6HiCls + 16H2O + C 21C 21C6H3C13 + SCO 2 + 21H+ + 21 Cl
Ethylbenzene oxidation/ Tetrachlorobenzene reductive dehalogenation
21C6H3Ch + 16H2O + C6H5C2H5 => 21C6H4C12 + SCO 2 + 21H+ + 21CI
Ethylbenzene oxidation/ Trichlorobenzene reductive dehalogenation
AG°r
kcal/ mole
-1066.13
-1080.76
-1066.27
-778.48
-166.75
-39.83
-446.43
-516.36
-467.01
-496.67
-484.70
-384.74
-368.79
-478.7
-490.4
-449.4
-430.1
AG°r
kJ/ mole
-4461
-4522
-4461
-3257
-697.7
-166.7
-1866
-2158
-1952
-2078
-2028
-1610
-1617
-2001
-2050
-1878
-1794
Stoichiometric
Mass Ratio of
Electron Acceptor
or Metabolic
Byproduct to
Primary Substrate
3.17:1
4.92:1
11.39:1
22.0:1
4.75:1
0.79:1
32.8:1
26.1:1
19.4:1
32.8:1
26.0:1
19.2:1
12.3:1
55.6:1
48.9:1
42.2:1
35.5:1
Mass of Primary
Substrate Utilized per
Mass of Electron
Acceptor Utilized or
Metabolic Byproduct
Produced
0.32:1
0.20:1
0.09:1
0.05:1
0.21:1
1.27:1
0.03:1
0.04:1
0.05:1
0.03:1
0.04:1
0.05:1
0.08:1
0.02:1
0.02:1
0.02:1
0.03:1
B3-49
-------
Table B. 3.7 Continued.
Coupled m-Xylene Oxidation Reactions
10.5 02+ C6H4(CH3)2 => 8C02 + 5 H2O
m-Xylene oxidation /aerobic respiration
8.4 H++ 8.4NQ-3+ C6H4(CH3)2 => 8CO2+ 4.2 N2+ 9.2 H2O
m-Xylene oxidation / denitrification
46H++ 22Mn02+ C6H4(CH3)2 =* 8CO2+ 22Mn2++ 28 H2O
m-Xylene oxidation /manganese reduction
84H++ 42Fe(OH)3-a+ C6H4(CH3)2 => 8CO2+ 42 Fe2++ 110H2O
m-Xylene oxidation /iron reduction
JO.SH++ 5.25SO42' + C6H4(CH3)2 =$ 8CO2 + 5.25 H2S°+ 5 H2O
m-Xylene oxidation / sulfate reduction
5.5H2O + C6H4(CH3)2 => 2.7 SCO 2 + 5.25CH4
m-Xylene oxidation / methanogenesis
21C2H2C14 + 16H20 + C6H4(CH3)2 => 21C2H3C13 + 8CO2 + 21H+ +
21CT
m-Xylene oxidation/ PC A reductive dehalogenation
21C2H3C13 + 16H2O + C6H4(CH3)2 => 21C2H4C12 + 8CO2 + 2JH+ +
21CT
m-Xylene oxidation/ TCA reductive dehalogenation
21C2H4C12 + 16H2O + C6H4(CH3)2 =$ 21C2H5Cl + 8CO2 + 2JH+ +
21CI
m-Xylene oxidation/ DC A reductive dehalogenation
21C2C14 + 16H20 + C6H4(CH3)2 =* 21C2HC13 + 8CO2 + 21H+ + 21CT
m-Xylene oxidation/ Tetrachloroethylene reductive dehalogenation
21C2HC13 + 16H20 + C6H4(CH3)2 => 21C2H2C12 + 8CO2 + 21H+ + 21CT
m-Xylene oxidation/ Trichloroethylene reductive dehalogenation
21C2H2C12 + 16H2O + C6H4(CH3)2 =$ 21C2H3Cl + 8CO2 + 2JH+ +
21CI
m-Xylene oxidation/ cis-Dichloroethylene reductive dehalogenation
21C2H3Cl + 16H20 + C6H4(CH3)2 => 21C2H4 + 8CO2 + 21H+ + 21CT
m-Xylene oxidation/ Vinyl chloride reductive dehalogenation
21C6C16 + 16H2O + C6H4(CH3)2 =$ 21C(jHiCh + 8CO2 + 2JH+ + 21CI
m-Xylene oxidation/ Hexachlorobenzene reductive dehalogenation
21C6H,Cls + 16H2O + C6H4(CH3)2 =$ 21C^{2C14+ 8CO2 + 2JH+ + 21CI
m-Xylene oxidation/ Pentachlorobenzene reductive dehalogenation
21C6H2C14 + 16H20 + C6H4(CH3)2 => 21C6H3C13 + 8CO2 + 21H+ + 21CI
m-Xylene oxidation/ Tetrachlorobenzene reductive dehalogenation
21C6H3C13 + 16H2O + C6H4(CH3)2 =$ 21C^{4C12 + 8CO2 + 2JH+ +
21CI-
m-Xylene oxidation/ Trichlorobenzene reductive dehalogenation
AG°r
(kcal/ mole)
-1063.25
-1077.81
-1063.39
-775.61
-163.87
-36.95
-445.70
-513.48
-464.13
-493.79
-469.59
-393.99
-383.91
-475.9
-487.5
-446.6
-426.9
AG°r
(kJ/ mole)
-4448
-4509
-4449
-3245
-685.6
-154.6
-1863
-2146
-1940
-2066
-1963
-1647
-1605
-1989
-2038
-1867
-1784
Stoichiometric Mass
Ratio of Electron
Acceptor or
Metabolic
Byproduct to
Primary Substrate
3.17:1
4.92:1
11.39:1
22:1
4.75:1
0.79:1 a/
32.7:1
26.0:1
19.3:
32.8:1
26.0:1
19.2:1
12.3:1
55.6:1
48.9:1
42.2:1
35.5:1
Mass of Primary
Substrate Utilized per
Mass of Electron
Acceptor Utilized or
Metabolic Byproduct
Produced
0.32:1
0.20:1
0.09:1
0.05:1
0.21:1
1.27:1
0.03:1
0.04:1
0.05:1
0.03:1
0.04:1
0.05:1
0.08:1
0.02:1
0.02:1
0.02:1
0.03:1
B3-50
-------
Table B. 3.7 Continued.
Coupled Naphthalene Oxidation Reactions
12O2 + C10HS^ 10CO2+ 4H2O
Naphthalene oxidation /aerobic respiration
9.6N03 + 9.6H+ + CIOH8 => 10CO2 + 8.8H2O + 4.8N2_
Naphthalene oxidation / denitrification
24MnO2 + 48H+ + C10HS =f 10CO2 + 24Mn2++ 28H2O
Naphthalene oxidation / manganese reduction
48Fe(OH)3,a + 96H+ + C10HS =t 10CO2 + 48Fe2++ 124H2O
Naphthalene oxidation / iron reduction
6S042-+ 12H+ + C10Hs => 10C02 + 6H 2S° + 4H2O
Error! Switch argument not specified. Naphthalene oxidation/
sulfate reduction
8H2O + C10HS =f 4CO2 + 6CH4
Naphthalene oxidation / methanogenesis
24C2H2C14 + 20H2O + C10HS =t 24C2H3C13 + 10CO2 + 24H+ +
24CT
Naphthalene oxidation/ PCA reductive dehalogenation
24C2H3C13 + 20H2O + C10H8 =t 24C2H4C12 + 10CO2 + 24H+ +
24CT
Naphthalene oxidation/ TCA reductive dehalogenation
24C2H4C12 + 20H2O + C10HS =3 24C2HSCI + 10CO2 + 24H+ +
24CT
Naphthalene oxidation/ DC A reductive dehalogenation
24C2C14 + 20H20 + CwHs => 24C2HC13 + 10CO2 + 24H+ + 24CT
Naphthalene oxidation/ Tetrachloroethylene reductive
dehalogenation
24C2HC13 + 20H20 + C WH 3 => 24C2H2C12 + 10CO2 + 24H+ + 24CT
Naphthalene oxidation/ Trichloroethylene reductive dehalogenation
24C2H2C12 + 20H2O + C10HS =f 24C2H3Cl + 10CO2 + 24H+ + 24CT
Naphthalene oxidation/cis-Dichloroethylene reductive
dehalogenation
24C2H3Cl + 20H20 + C10HS => 24C2H4 + 10CO2 + 24H+ + 24CT
Naphthalene oxidation/ Vinyl chloride reductive dehalogenation
24C6C16 + 20H2O + CIOHS => 24C6H,Cls + 10CO2 + 24H++24Cl
Naphthalene oxidation/ H exachlorobenzene reductive
dehalogenation
24C6H,Cli + 20H20 + C10HS ^24CSH2C14+ 10CO2 + 24H+ + 24CI
Naphthalene oxidation/Pentachlorobenzene reductive
dehalogenation
24C6H2C14 + 20H20 + CIOHS => 24C6H3C13 + 10CO2 + 24H+ +
24C1
Naphthalene oxidation/ Tetrachlorobenzene reductive
dehalogenation
24C6H3C13 + 20H2O + C10HS =t 24C6H4C12 + 10CO2 + 24H+ +
24CT
Naphthalene oxidation/ Trichlorobenzene reductive dehalogenation
AG°r
(kcal/
m ole)
-1217.40
-1234.04
-1217.57
-932.64
-196.98
-44.49
-511.68
-589.09
-532.69
-566.59
-552.91
-438.67
-441.01
-545.94
-559.33
-512.53
-490.45
AG°r
(kJ/
m ole)
-5094
-5163
-5094
-3902
-824.2
-186.1
-2139
-2462
-2227
-2371
-2313
-1835
-1843
-2282
-2338
-2142
-2050
Stoichiom etric Mass
Ratio of Electron
Acceptor or Metabolic
Byproduct to Primary
Substrate
3.00:1
4.65:1
16.31:1
40.13:1
4.50:1
1.13:1
31.1:1
24.8:1
18.4:1
31.1:1
24.6:1
18.2:1
11.6:1
52.9:1
46.5:1
40.1:1
33.8:1
Mass of Primary
Substrate Utilized per
Mass of Electron
Acceptor Utilized or
Metabolic Byproduct
Produced
0.33:1
0.22:1
0.06:1
0.02:1
0.22:1
0.88:1
0.03:1
0.04:1
0.05:1
0.03:1
0.04:1
0.05:1
0.09:1
0.02:1
0.02:1
0.02:1
0.03:1
B3-51
-------
Table B. 3.7 Continued.
Coupled 1,3,5-Trimethylbenzene (1,3,5-TMB) Oxidation Reactions
1202 + C6H3(CH3)3 =* 9C02 + 6H20
1,3,5-TMB oxidation /aerobic respiration
9.6NO3- + 9.6H+ + C6H3(CH3)3 => 9CO2 + 10.8H2O + 4.8N2_S
1,3,5-TMB oxidation / denitrification
24MnO2 + 48H+ + C6H3(CH3)3 =1 9CO2 + 30H2O + 24Mn2+
1,3,5-TMB oxidation /manganese reduction
48Fe(OH)3,a + 96H+ + C6H3(CH3)3 => 9CO2 + 48Fe2+ + 126H2O
1,3,5-TMB oxidation /iron reduction
6SO42' + 12H+ + C(jH3(CH3)3 => 9CO2 + 6H2O + 6H2S°
1,3,5-TMB oxidation / sulfate reduction
6H2O + C6H3(CH3)3 => 3CO2 + 6CH4
1,3,5-TMB oxidation / methanogenesis
24 C2H2C14 + 18H2O + C6H3(CH3)3 => 24C2H3C13 + 9CO2 + 24H+ + 24CI
1,3,5-TMB oxidation/ PCA reductive dehalogenation
24C2H3C13 + 18H2O + C6H3(CH3)3 => 24C2H4C12 + 9CO2 + 24H+ + 24CT
1,3,5-TMB oxidation/ TCA reductive dehalogenation
24C2H4C12 + 18H2O + C6H3(CH3)3 => 24C2Hfl + 9CO2 + 24H+ + 24CT
1,3,5-TMB oxidation/ DCA reductive dehalogenation
24C2C14 + 18H2O + C6H3(CH3)3 => 24C2HC13 + 9CO2 + 24H+ + 24CT
1,3,5-TMB oxidation/ Tetrachloroethene reductive dehalogenation
24C2HC13 + 18H2O + C6H3(CH3)3 => 24C2H2C12 + 9CO2 + 24H++ 24CT
1,3,5-TMB oxidation/ Trichloroethene reductive dehalogenation
24C2H2C12 + 18H2O + C6H3(CH3)3 => 24C2H3Cl + 9CO2 + 24H+ + 24CT
1,3,5-TMB oxidation/ cis-Dichloroethene reductive dehalogenation
24C2H3Cl + 18H2O + C6H3(CH3)3 => 24C2H4 + 9CO2 + 24H+ + 24CT
1,3,5-TMB oxidation/ Vinyl chloride reductive dehalogenation
24C6C16 + 18H2O + C6H3(CH3)3 => 24CffiC^ + 9CO2 + 24H+ + 24CT
1,3,5-TMB oxidation/ Hexachlorobenzene reductive dehalogenation
24C 24CaH2Cl4+ 9CO2 + 24H+ + 24CI
1,3,5-TMB oxidation/ Pentachlorobenzene reductive dehalogenation
24C6H2C14 + 18H2O + C6H3(CH3)3 => 24C6H3C13 + 9CO2 + 24H+ + 24CI
1,3,5-TMB oxidation/ Tetrachlorobenzene reductive dehalogenation
24C6H3C13 + 18H2O + C6H3(CH3)3 => 24C6H4C12 + 9CO2 + 24H+ + 24CT
1,3,5-TMB oxidation/ Trichlorobenzene reductive dehalogenation
AG°r
(kcal/ mole)
-1213.29
-1229.93
-1213.46
-928.53
-192.87
-40.39
-507.36
-584.99
-528.59
-562.48
-548.80
-434.56
-436.91
-541.84
-555.23
-508.43
-486.35
AG°r
(kJ/ mole)
-5076
-5146
-5077
-3885
-807.0
-169.0
-2121
-2445
-2210
-2353
-2296
-1818
-1826
-2265
-2321
-2125
-2033
Stoichiometric Mass
Ratio of Electron
Acceptor or Metabolic
Byproduct to Primary
Substrate
3.20:1
4.96:1
17.40:1
42.80:1
4.80:1
0.90:1
33.2:1
26.4:1
19.6:1
33.2:1
26.3:1
19.4:1
12.4:1
56.4:1
49.6:1
42.8:1
36.0:1
Mass of Primary Substrate
Utilized per Mass of
Electron Acceptor Utilized
or Metabolic Byproduct
Produced
0.31:1
0.20:1
0.06:1
0.02:1
0.21:1
1.11:1
0.03:1
0.04:1
0.05:1
0.03:1
0.04:1
0.05:1
0.08:1
0.02:1
0.02:1
0.02:1
0.03:1
B3-52
-------
Table B. 3.7 Continued.
Coupled 1,2,4-Trimethylbenzene (1,2,4-TMB) Oxidation Reactions
12O2 + C6H3(CH3)3 =$ 9CO2 + 6H2O
1,2,4-TMB oxidation /aerobic respiration
9.6NO3- + 9.6H+ + C6H3(CH3)3 =$ 9CO2 + 10.8H2O + 4.8N2_g
1,2,4-TMB oxidation / denitrification
24MnO, + 48H+ + C*H3(CH3), =$ 9COj + 30H,O + 24Mn2+
1,2,4-TMB oxidation /manganese reduction
48Fe(OH),n + 96H+ + Cf[}(CH})} =$ 9CO2 + 48Fe2+ + 126H2O
1,2,4-TMB oxidation /iron reduction
6SO42" + UH+ + C(jH3(CH3)3 =$ 9CO2 + 6H2O + 6H2S°
1,2,4-TMB oxidation / sulfate reduction
6H2O + C6H3(CH3)3 =3 3CO2 + 6CH4
1,2,4-TMB oxidation / methanogenesis
24C2H2C14 + 18H2O + C6H3(CH3)3 => 24C2H3C13 + 9CO2 + 24H+ +
24Cr
1,2,4-TMB oxidation/ PCA reductive dehalogenation
24C2H3C13 + 18H2O + C6H3(CH3)3 => 24C2H4C12 + 9CO2 + 24H+ +
24Cr
1,2,4-TMB oxidation/ TCA reductive dehalogenation
24C2H4C12 + 18H2O + C(jH3(CH3)3 =3 24C2H5Cl + 9CO2 + 24H+ + 24CT
1,2,4-TMB oxidation/ DCA reductive dehalogenation
24C2C14 + 18H2O + C6H3(CH3)3 => 24C2HC13 + 9CO2 + 24H+ + 24Ct
1,2,4-TMB oxidation/ PCE reductive dehalogenation
24C2HC13 + 18H2O + C6H3(CH3)3 => 24C2H2C12 + 9CO2 + 24H+ + 24CT
1,2,4-TMB oxidation/ ' TCE reductive dehalogenation
24C2H2C12 + 18H2O + C6H3(CH3)3 =$ 24C2H3Cl + 9CO2 + 24H+ + 24CT
1,2,4-TMB oxidation/ cis-DCE reductive dehalogenation
24C2H3Cl + 18H2O + C6H3(CH3)3 =$ 24C2H4 + 9CO2 + 24H+ + 24CT
1,2,4-TMB oxidation/ Vinyl chloride reductive dehalogenation
24C6C16 + 18H2O + C6H3(CH3)3 =$ 24C6H,C15 + 9CO2 + 24H+ + 24CT
1,2,4-TMB oxidation/ Hexachlorobenzene reductive dehalogenation
24C6H,C15 + 18H2O + CtHjfCHjJj =$ 24C6H2C14+ 9CO2 + 24H+ + 24Ct
1,2,4-TMB oxidation/ Pentachlorobenzene reductive dehalogenation
24C(jH2Cl4 + 18H2O + C6H3(CH3)3 =3 24CaH3Cl3 + 9CO2 + 24H+ + 24CT
1,2,4-TMB oxidation/ Tetrachlorobenzene reductive dehalogenation
24C(jH3Cl3 + 18H2O + C6H3(CH3)3 =$ 24CaH4Cl2 + 9CO2 + 24H+ + 24Ct
1,2,4-TMB oxidation/ Trichlorobenzene reductive dehalogenation
AG°r
(kcal/ mole)
-1212.92
-1229.56
-1213.09
-928.16
-192.50
-40.02
-507.36
-584.62
-528.22
-562.11
-548.43
-434.19
-436.54
-541.47
-554.86
-508.06
-485.98
AG°r
(kJ/ mole)
-5075
-5144
-5076
-3883
-805.4
-167.4
-2121
-2444
-2208
-2352
-2295
-1817
-1825
-2263
-2319
-2124
-2031
Stoichiometric Mass
Ratio of Electron
Acceptor or
Metabolic
Byproduct to
Primary Substrate
3.20:1
4.96:1
17.4:1
42.8:1
4.80:1
0.90:1
33.2:1
26.4:1
19.6:1
33.2:1
26.3:1
19.4:1
12.4:1
56.4:1
49.6:1
42.8:1
36.0:1
Mass of Primary Substrate
Utilized per Mass of
Electron Acceptor Utilized
or Metabolic Byproduct
Produced
0.31:1
0.20:1
0.06:1
0.02:1
0.21:1
1.11:1
0.03:1
0.04:1
0.05:1
0.03:1
0.04:1
0.05:1
0.08:1
0.02:1
0.02:1
0.02:1
0.03:1
B3-53
-------
Table B. 3.7 Continued.
Coupled Vinyl Chloride Oxidation Reactions
2. SO 2 + C2H3Cl => 2CO2 + H2O + H+ + Cl'
Vinyl Chloride oxidation /aerobic respiration
2NO3 + H+ CiHsCl => 2CO2 + 2H2O + Cl + N2,s
Vinyl Chloride oxidation / denitrification
5MnO_2 + 9H+ + CJljCl =* 2CO2 + 6H2O + 5Mn + + Cl
Vinyl Chloride oxidation / manganese reduction
10Fe(OHhn + 19H+ + C6H}(CH})} => 2CO2 + 10Fe2+ + 26H2O + Cl
Vinyl Chloride oxidation / iron reduction
1.25S042' + 1.5H+ + C2H3Cl => 2CO2 + H2O + 1.25H2Sa + Cl
Vinyl Chloride oxidation / sulfate reduction
l.5H2O + CiHiCl =* . 7 SCO 2 + 1.25CH4 + H+ + Cl'
Vinyl Chloride oxidation / methanogenesis
5C2H2Ck + 4H2O + C2H3Cl =* 5C2H3C13 + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ PCA reductive dehalogenation
5C2H3C13 + 4H2O + C2H3Cl => 5C2H4C12 + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ TCA reductive dehalogenation
5C2H4C12 + 4H2O + C2H3Cl => SCtfsCl + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ DCA reductive dehalogenation
5C2C14 + 4H2O + C2H3Cl => 5C2HC13 + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ DCE reductive dehalogenation
5C2HC13 + 4H2O + C2H3Cl => 5C2H2Ch + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ TCE reductive dehalogenation
5C2H2C12 + 4H2O + C2H3Cl =* 5C2H3Cl + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ cis-DCE reductive dehalogenation
5C6C16 + 4H2O + C2H3Cl => SCtjHiCls + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ Hexachlorobenzene reductive dehalogenation
SCtjHiCls + 4H2O + C2H3Cl => 5C 5C6H3C13 + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ Tetrachlorobenzene reductive dehalogenation
5C6H}Cl} + 4H2O + C2H3Cl => SCeH^h + 2CO2 + 6H+ + 6CI
Vinyl Chloride oxidation/ Trichlorobenzene reductive dehalogenation
2O2 + C2H2C12 =* 2CO2 + 2H+ + 2CI
DCE oxidation /aerobic respiration
AG°r
(kcal/ mole)
-288.98
-292.44
-289.01
-229.65
-76.40
-44.62
-141.90
-158.08
-146.33
-153.39
-150.54
-126.74
-144.60
-138.59
-142.13
-137.53
-256.53
AG°r
(kJ/ mole)
-1209
-1224
-1209
-960.9
-319.7
-186.7
-593.1
-661
-612
-641.8
-629.9
-530.3
-604.4
-579.3
-594.1
-574.9
-1072
Stoichiometric Mass
Ratio of Electron
Acceptor or Metabolic
Byproduct to Primary
Substrate
1.29:1
2.00:1
7.02:1
17.3:1
1.94:1
0.44:1
13.4:1
10.7:1
7.92:1
13.4:1
10.6:1
7.82:1
22.8:1
20.0:1
17.3:1
14.5:1
1.31:1
Mass of Primary Substrate
Utilized per Mass of
Electron Acceptor Utilized
or Metabolic Byproduct
Produced
0.78:1
0.50:1
0.14:1
0.06:1
0.52:1
2.27:1
0.07:1
0.09:1
0.13:1
0.07:1
0.09:1
0.13:1
0.04:1
0.05:1
0.06:1
0.07:1
0.76:1
B3-54
-------
Table B. 3.7 Continued.
Coupled Chlorobenzene Oxidation Reactions
7O2 + C6H4Cl => 6CO2 + H+ + 2H2O+ Cl
Chlorobenzene oxidation /aerobic respiration
5.6NOi + 4.6H+ + C6H4Cl => 6CO2 + 4.8H2O + 2.8N2_g + 2CT
Chlorobenzene oxidation / denitrification
14MnO, + 27H+ + CfHsCl => 6CO2 + 16H2O + 14Mn2+ + CT
Chlorobenzene oxidation / manganese reduction
28Fe(OHhn + 55H+ + C6H5Cl => 6CO2 + 72H2O + 28Fe2+ + CT
Chlorobenzene oxidation / iron reduction
3.5SO42' + 6H+ + C6H5Cl => 6CO2 + 2H2O + 3.5H2S° + CT
Chlorobenzene oxidation / sulfate reduction
5H2O + C6H5Cl => 2.5CO2 + 3.5CH4 + H+ + Cl
Chlorobenzene oxidation / methanogenesis
14C2H2C14 + 12H2O + CfHsCl => 14C2H3C13 + 6CO2 + 15H+ + 15CT
Chlorobenzene oxidation/ PCA reductive dehalogenation
14C2H3C13 + 12H2O + C 14C2HC13 + 6CO2 + 15H+ + 15CT
Chlorobenzene oxidation/ PCE reductive dehalogenation
14C2HC13 + 12H2O + C6H5Cl => 14C2H2C12 + 6CO2 + 15H+ + 15CI
Chlorobenzene oxidation/ TCE reductive dehalogenation
14C2H2C12 + 12H2O + CtHsCl => 14C2H3Cl + 6CO2 + 15H+ + 15CI
Chlorobenzene oxidation/ cis-DCE reductive dehalogenation
14C2H3Cl + 12H2O + CfHsCl => 14C2H4 + 6CO2 + 15H+ + 15CT
Chlorobenzene oxidation/ Vinyl chloride reductive dehalogenation
AG°r
(koal/
mole)
-731.62
-741.33
-731.72
-565.51
-136.38
-47.43
-320.04
-365.11
-332.21
-351.99
-344.01
-277.37
-278.73
AG°r
(kJ/ mole)
-3061
-3102
-3062
-2366
-570.6
-198.4
-1338
-1526
-1389
-1473
-1439
-1161
-1165
Stoichiometric
Mass Ratio of
Electron
Acceptor or
Metabolic
Byproduct to
Primary
Substrate
2.00:1
3.10:1
10.9:1
26.8:1
3.00:1
0.80:1
20.8:1
16.5:1
12.3:1
20.7:1
16.4:1
12.1:1
7.75:1
Mass of Primary
Substrate Utilized
per Mass of
Electron Acceptor
Utilized or
Metabolic
Byproduct
Produced
0.50:1
0.32:1
0.09:1
0.04:1
0.33:1
1.25:1
0.05:1
0.06:1
0.08:1
0.05:1
0.06:1
0.08:1
0.13:1
B3-55
-------
Table B. 3.7 Continued.
Coupled Dichlorobenzene Oxidation Reactions
6. SO 2 + C6H4C12 => 6CO2 + 2H+ + H2O+ 2Ct
Dichlorobenzene oxidation /aerobic respiration
5.2NOi + 3.2H+ + C6H4C12 => 6CO2 + 3.6H2O + 2.6N2_g + 2CT
Dichlorobenzene oxidation / denitrification
ISMnO, + 24H+ + C6H4C12 => 6CO2 + 14H2O + 13Mn2+ + 2CT
Dichlorobenzene oxidation /manganese reduction
26Fe(OHhn + 50H+ + CaH4Cl2 => 6CO2 + 66H2O + 26Fe2+ + 2CT
Dichlorobenzene oxidation / iron reduction
3. 2 5 SO 42~ + 4.5H+ + C(jH4Cl2=> 6CO2 + H2O + 3.25H2S° + 2CI
Dichlorobenzene oxidation / sulfate reduction
5.5H2O + CnH4Cl2 => 2.75CO2 + 3.25CH4 + 2H+ + 2Ct
Dichlorobenzene oxidation / methanogenesis
13C2H2C14 + 12H2O + C6H4C12=> 13C2H3C13 + 6CO2 + 15H+ + 15CT
Dichlorobenzene oxidation/ PCA reductive dehalogenation
13C2H3C13 + 12H2O + C6H4C12 => 13C2H4C12 + 6CO2 + 15H+ + 15CT
Dichlorobenzene oxidation/ TCA reductive dehalogenation
13C2H4C12 + 12H2O + C6H4C12 => 13C2HSCI + 6CO2 + 15H+ + 15Cl~
Dichlorobenzene oxidation/ DCA reductive dehalogenation
13C2C14 + 12H2O + C6H4C12 => 13C2HC13 + 6CO2 + 15H+ + 15CT
Dichlorobenzene oxidation/ PCE reductive dehalogenation
13C2HC13 + 12H2O + C6H4C12=> 13C2H2C12 + 6CO2 + 15H+ + 15CI
Dichlorobenzene oxidation/ ' TCE reductive dehalogenation
13C2H2C12 + 12H2O + C6H4C12 => 13C2H3Cl + 6CO2 + 15H+ + 15CI
Dichlorobenzene oxidation/ cis-DCE reductive dehalogenation
13C2H3Cl + 12H2O + C6H4C12 => 13C2H4 + 6CO2 + 15H+ + 15CT
Dichlorobenzene oxidation/ Vinyl chloride reductive dehalogenation
AG°r
(kcal/ mole)
-698.36
-708.76
-698.36
-521.56
-142.74
-64.22
-317.20
-358.93
-328.38
-347.10
-339.56
-277.68
-278.72
AG°r
(kJ/ mole)
-2919
-2963
-2919
-2180
-596.7
-268.4
-1326
-1500
-1373
-1450
-1419
-1161
-1165
Stoichiometric
Mass Ratio of
Electron
Acceptor or
Metabolic
Byproduct to
Primary
Substrate
1.42:1
1.64:1
7.75:1
19.05:1
2.14:1
0.33:1
14.8:1
11.8:1
8.73:1
14.6:1
11.6:1
8.55:1
5.52:1
Mass of Primary
Substrate Utilized
per Mass of
Electron Acceptor
Utilized or
Metabolic
Byproduct
Produced
0.70:1
0.61:1
0.13:1
0.05:1
0.47:1
2.99:1
0.07:1
0.09:1
0.11:1
0.07:1
0.09:1
0.12:1
0.18:1
B3-56
-------
Table B. 3.7 Continued.
Coupled Trichlorobenzene Oxidation Reactions
6O2 + CfHiCli => 6CO2 + 3H+ + 3CT
Trichlorobenzene oxidation /aerobic respiration
4.8NOf + 1.8H+ + CtHiCh => 6CO2 + 2.4H2O + 2.4N2,g + 3CT
Trichlorobenzene oxidation / denitrifwation
12MnO? + 21H+ + CfHiCh=> 6COj + 12H,O + 12Mn2+ + 3CT
Trichlorobenzene oxidation / manganese reduction
24Fe(OH)3a + 45H+ + C6H3C13=> 6CO2 + 60H2O + 24Fe2+ + 3CT
Trichlorobenzene oxidation / iron reduction
3SO42' + 3H+ + C6H3Cl3=> 6CO2 + 3H2S° + 3CI
Trichlorobenzene oxidation / sulfate reduction
6H2O + CiHiCh => 3CO2 + 3CH4 + 3H+ + 3CT
Trichlorobenzene oxidation / methanogenesis
12C2H2C14 + 12H2O + C6H3C13=> 12C2H3C13 + 6CO2 + 15H+ + 15CT
Trichlorobenzene oxidation/ PCA reductive dehalogenation
12C2H3C13 + 12H2O + C(jH3Cl3=> 12C2H4C12 + 6CO2 + 15H+ + 15CI
Trichlorobenzene oxidation/ TCA reductive dehalogenation
12C2H4C12 + 12H2O + C6H3C13 => 12C2HŁl + 6CO2 + 15H+ + 15CT
Trichlorobenzene oxidation/ DCA reductive dehalogenation
12C2C14 + 12H2O + C(jH3Cl3 => 12C2HC13 + 6CO2 + 15H+ + 15CV
Trichlorobenzene oxidation/ PCE reductive dehalogenation
12C2HC13 + 12H2O + C6H3C13 => 12C2H2C12 + 6CO2 + 15H+ + 15CT
Trichlorobenzene oxidation/ TCE reductive dehalogenation
12C2H2Cl2 + 12H2O+C(jH3Cl3 => 12C2H3CI +6CO2+ 15H++ 15CI
Trichlorobenzene oxidation/ cis-DCE reductive dehalogenation
12C2H3Cl +12H2O +C6H3C13 => 12C2H4 + 6CO2 + 15H+ + 15CT
Trichlorobenzene oxidation/ Vinyl chloride reductive dehalogenation
AG°r
(kcal/ mole)
-668.16
-677.76
-688.16
-504.96
-155.28
-82.80
-316.32
-354.82
-326.62
-343.92
-336.96
-279.58
-280.78
AG°r
(kJ/ mole)
-2793
-2833
-2793
-2111
-649.1
-346.1
-1322
-1483
-1365
-1438
-1408
-1169
-1174
Stoichiometric Mass
Ratio of Electron
Acceptor or Metabolic
Byproduct to Primary
Substrate
1.07:1
1.65:1
5.80:1
14.3:1
1.60:1
0.25:1
11.1:1
8.8:1
6.53:1
10.9:1
8.67:1
6.40:1
4.13:1
Mass of Primary
Substrate Utilized per
Mass of Electron
Acceptor Utilized or
Metabolic Byproduct
Produced
0.94:1
0.60:1
0.17:1
0.07:1
0.63:1
4.00:1
0.09:1
0.11:1
0.15:1
0.09:1
0.12:1
0.16:1
0.24:1
B3-57
-------
Table B. 3.7 Continued.
Coupled Tetrachlorobenzene Oxidation Reactions
5. SO 2 + H2O + C6H2C14 => 6CO2 + 4H++ 4CT
Tetrachlorobenzene oxidation /aerobic respiration
4.4NOf + 0.4 H+ + C6H2C14 =>6CO2 + 1.2H2O + 2.2N2_g + 4CT
Tetrachlorobenzen oxidation / denitrification
HMnOj + 18H+ + Cf,H?Cl4 => 6CO, + 10H?O + llMn* + 4CT
Tetrachlorobenzenoxidation /manganese reduction
22Fe(OH),a + 40H+ + C6H2C14 => 6CO2 + 54H2O + 22Fe2+ + 4CT
Tetrachlorobenzen oxidation / iron reduction
2. 75SO42' + 1. 75H+ + H2O + C6H2C14 => 6CO2 + 2. 75H2S° + 4CI
Tetrachlorobenzen oxidation / sulfate reduction
6.5H2O + C6H2C14 => 3.25CO2 + 2.75CH4 + 4H+ + 4CT
Tetrachlorobenzen oxidation / methanogenesis
11C2H2C14 + 12H2O + C6H2C14 => nC2H3Cl3 + 6CO2 + 15H+ + 15CT
Tetrachlorobenzen oxidation/ PCA reductive dehalogenation
UC2H3C13 + 12H2O + C(jH2Cl4 => UC2H4C12 + 6CO2 + 15H+ + 15CI
Tetrachlorobenzen oxidation/ TCA reductive dehalogenation
11C2H4C12 + 12H2O + C6H2C14 => nC2Hfl + 6CO2 + 15H+ + 15CT
Tetrachlorobenzen oxidation/ DCA reductive dehalogenation
UC2C14 + 12H2O + C(jH2Cl4 => UC2HC13 + 6CO2 + 15H+ + 15CI
Tetrachlorobenzen oxidation/ PCE reductive dehalogenation
11C2HC13 + 12H2O + C6H2C14 => 11C2H2C12 + 6CO2 + 15H+ + 15CT
Tetrachlorobenzen oxidation/ TCE reductive dehalogenation
UC2H2C12 + 12H2O + C(jH2Cl4 => UC2H3Cl + 6CO2 + 15H+ + 15CI
Tetrachlorobenzen oxidation/ cis-DCE reductive dehalogenation
11C2H3CI + 12H2O + C6H2C14 => 11C2H4 + 6CO2 + 15H+ + 15CT
Tetrachlorobenzen oxidation/ Vinyl chloride reductive dehalogenation
AG°r
(kcal/ mole)
-639.10
-647.90
-639.10
-489.50
-168.96
-102.52
-287.01
-323.64
-297.79
-313.3
-307.03
-254.67
-255.77
AG°r
(kJ/ mole)
-2671
-2708
-2671
-2046
-706.3
-428.5
-1200
-1353
-1392
-1310
-1283
-1065
-1069
Stoichiometric Mass
Ratio of Electron
Acceptor or Metabolic
Byproduct to Primary
Substrate
0.82:1
1.27:
4.47:1
11.0:1
1.23:1
0.19:1
8.53:1
6.79:1
5.04:1
8.43:1
6.68:1
4.93:1
3.19:1
Mass of Primary
Substrate Utilized per
Mass of Electron
Acceptor Utilized or
Metabolic Byproduct
Produced
1.22:1
0.78:1
0.22:1
0.09:1
0.81:1
5.19:1
0.12:1
0.15:1
0.20:1
0.12:1
0.15:1
0.20:1
0.31:1
B3-58
-------
A n th ro p o g e n ic E lectrqn
Acceptors AG,
PC E R eduction -1500
TC E R eduction -1465
c/s-1 ,2-D C E Reductioti166
For Benzene Oxidation, kJ/mole
N a tu ra I E le ctro n
Acceptors AG,'
A e ro bic R e s p ira tio n-3 2 0 2
D e n itrificatio n -3245
Manganese (IV) -3202
Reduction
Iron (III) R eduction -2343
S u Ifate Red uctio n -514
Methanogenesis -136
Figure B.3.4 Expected sequence of microbially-mediated redox reactions and Gibbs free energy of
reaction.
EQUATION
The advection-dispersion equation is obtained by adding a biodegradation term to
equation B.2.20. In one dimension, this is expressed as:
ac D a2c v ac
eq. B.3.2
, x , x H
Where:
vx = average linear ground-water velocity [L/T]
jR = coefficient of retardation
C = contaminant concentration [M/L3]
Dx = hydrodynamic dispersion [L2/T]
t = time [T]
x = distance along flow path [L]
'k = first-order biodegradation decay rate [T"1]
This equation considers advection, hydrodynamic dispersion, sorption (retardation), and biodeg-
radation. First-order rate constants are appropriate for iron (Ill)-reducing, suifate-reducing, and
methanogenic conditions. They are not appropriate under aerobic or denitrifying conditions.
B3-59
-------
SECTION B-4
ATTENUATION -
Chlorinated solvents dissolved in ground water may also be degraded by abiotic mechanisms,
although the reactions are typically not complete and often result in the formation of an intermediate
that may be at least as toxic as the original contaminant. The most common reactions affecting
chlorinated compounds are hydrolysis (a substitution reaction) and dehydrohalogenation (an elimina-
tion reaction). Other possible reactions include oxidation and reduction reactions. Butler and Barker
(1996) note that no abiotic oxidation reactions involving typical halogenated solvents have been
reported in the literature. They also note that reduction reactions (which include hydrogenolysis and
dihaloelimination) are commonly microbially mediated, although some abiotic reduction reactions
have been observed.
As Butler and Barker (1996) note, attributing changes in either the presence or absence of
halogenated solvents or the concentrations of halogenated solvents to abiotic processes is usually
difficult. For example, microbial activity is generally required to produce reducing conditions that
favor reductive dehalogenation. If such activity is taking place, chlorinated solvents may be under-
going both biotic and abiotic degradation, and discerning the relative contribution of each mecha-
nism on the field scale, if possible, would be very difficult. As another example, Butler and Barker
(1996) note that to substantiate that hydrolysis is occurring, the presence of non-halogenated break-
down products such as acids and alcohols should be established. In general, these products are more
easily biodegraded than their parent compounds and can be difficult to detect. Field evidence of this
nature has yet to be collected to demonstrate hydrolysis of halogenated solvents (Butler and Barker,
1996).
Given the difficulties of demonstrating abiotic degradation on the field scale, it may not be
practical to demonstrate that such processes are occurring and to quantitatively evaluate the contribu-
tions of those reactions (i.e., separately from biotic processes). If biodegradation is occurring at a
site, the loss of contaminant mass due to that process may dwarf the mass lost to abiotic reactions,
ruling out a cost-effective evaluation of abiotic degradation. However, while the rates of abiotic
degradation may be slow relative to biotic mechanisms, the contribution of these mechanisms may
still play a significant role in natural attenuation, depending on site conditions (e.g., a site with a
slow solute transport velocity or a long distance to the nearest receptor). Vogel (1994) describes data
patterns that may result from varying combinations of biotic and abiotic degradation of chlorinated
solvents. Moreover, because some of the by-products of these reactions are chlorinated compounds
that may be more easily or less easily degraded than the parent, the contributions of abiotic mecha-
nisms may need to be considered when evaluating analytical data from a site.
B.4.1 HYDROLYSIS AND
As discussed by Butler and Barker (1996), hydrolysis and dehydrohalogenation reactions are the
most thoroughly studied abiotic attenuation mechanisms. In general, the rates of these reactions are
often quite slow within the range of normal ground-water temperatures, with half-lives of days to
centuries (Vogel etal., 1987; Vogel, 1994). Therefore, most information about the rates of these
reactions is extrapolated from experiments run at higher temperatures so that the experiments could
be performed within a practical time frame.
B.4.1.1 Hydrolysis
Hydrolysis is a substitution reaction in which an organic molecule reacts with water or a com-
ponent ion of water, and a halogen substituent is replaced with a hydroxyl (OH") group. The hy-
droxyl substitution typically occurs at the halogenated carbon. This leads initially to the production
of alcohols. If the alcohols are halogenated, additional hydrolysis to acids or diols may occur. Also,
B4-60
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the addition of a hydroxyl group to a parent molecule may make the daughter product more suscep-
tible to biodegradation, as well as more soluble (Neely, 1985). Non-alcohol products have also been
reported by Vogel et al. (1987) and Jeffers et al. (1989), but they are apparently products of compet-
ing dehydrohalogenation reactions.
The likelihood that a halogenated solvent will undergo hydrolysis depends in part on the num-
ber of halogen substituents. More halogen substituents on a compound will decrease the chance for
hydrolysis reactions to occur (Vogel et al., 1987), and will therefore decrease the rate of the reaction.
In addition, bromine substituents are more susceptible to hydrolysis than chlorine substituents (Vogel
et al., 1987). 1,2-Dibromoethane is one compound that is subject to significant hydrolysis reactions
under natural conditions. Locations of the halogen substituent on the carbon chain may also have
some effect on the rate of reaction. The rate also may increase with increasing pH; however, a rate
dependence upon pH is typically not observed below a pH of 11 (Mabey and Mill, 1978; Vogel and
Reinhard, 1986). Rates of hydrolysis may also be increased by the presence of clays, which can act
as catalysts (Vogel et al., 1987). Hydrolysis rates can generally be described using first-order kinet-
ics, particularly in solutions in which water is the dominant nucleophile (Vogel et al., 1987). How-
ever, this oversimplifies what is typically a much more complicated relationship (Neely, 1985). As
noted in the introduction to this Appendix, reported rates of environmentally significant hydrolysis
reactions involving chlorinated solvents are typically the result of extrapolation from experiments
performed at higher temperatures (Mabey and Mill, 1978; Vogel, 1994).
Hydrolysis of chlorinated methanes and ethanes has been well-demonstrated in the literature.
Vogel (1994) reports that monohalogenated alkanes have half-lives on the order of days to months,
while poly chlorinated methanes and ethanes have half-lives that may range up to thousands of years
for carbon tetrachloride. As the number of chlorine atoms increases, dehydrohalogenation may
become more important (Jeffers el al., 1989). Butler and Barker (1996) note that chlorinated ethenes
do not undergo significant hydrolysis reactions (i.e., the rates are slow). Butler and Barker also
reported that they were unable to find any studies on hydrolysis of vinyl chloride. A listing of half-
lives for abiotic hydrolysis and dehydrohalogenation of some chlorinated solvents is presented on
Table B.4.1. Note that no distinctions are made in the table as to which mechanism is operating; this
is consistent with the references from which the table has been derived (Vogel el al., 1987; Butler
and Barker, 1996).
One common chlorinated solvent for which abiotic transformations have been well-studied is
1,1,1 -TCA. 1,1,1 -TCA may be abiotically transformed to acetic acid through a series of substitution
reactions, including hydrolysis. In addition, 1,1,1-TCA may be reductively dehalogenated to form
1,1- DCA) and then chloroethane (CA), which is then hydrolyzed to ethanol (Vogel and McCarty,
1987) or dehydrohalogenated to vinyl chloride (Jeffers et al., 1989). Rates of these reactions have
been studied by several parties, and these rates are summarized in Table B.4.1.
B.4.1.2 Dehydrohalogenation
Dehydrohalogenation is an elimination reaction involving halogenated alkanes in which a
halogen is removed from one carbon atom, followed by the subsequent removal of a hydrogen atom
from an adjacent carbon atom. In this two-step reaction, an alkene is produced. Although the
oxidation state of the compound decreases due to the removal of a halogen, the loss of a hydrogen
atom increases it. This results in no external electron transfer, and there is no net change in the
oxidation state of the reacting molecule (Vogel el al., 1987). Contrary to the patterns observed for
hydrolysis, the likelihood that dehydrohalogenation will occur increases with the number of halogen
substituents. It has been suggested that under normal environmental conditions, monohalogenated
aliphatics apparently do not undergo dehydrohalogenation, and these reactions are apparently not
likely to occur (March, 1985; Vogel et al., 1987). However, Jeffers et al. (1989) report on the
B4-61
-------
dehydrohalogenation of CA to VC. Polychlorinated alkanes have been observed to undergo
dehydrohalogenation under normal conditions and extremely basic conditions (Vogel el al., 1987).
As with hydrolysis, bromine substituents are more reactive with respect to dehydrohalogenation.
Table B.4.1 Approximate Half-Lives of Abiotic Hydrolysis andDehydrohalogenation Reactions Involving
Chlorinated Solvents
Compound
Chloromethane
Methylene Chloride
(Dichloro methane)
Trichloromethane
(Chloroform)
Carbon Tetrachloride
Chloroethane
1 , 1 -Dichloroethane
1 ,2-Dichloroethane
1,1,1 -Trichloroethane
1 , 1 ,2-Trichloroethane
1,1,1 ,2-Tetrachloroethane
1 , 1 ,2,2-Tetrachloroethane
Tetrachloroethene
Trichloroethene
1 , 1 -Dichloroethene
1 ,2-Dichloroethene
Half-Life (years)
no data
70437
3500a, 1800b
41b
0.12C
61b
72b
1.7a, l.lb
2.5d
140b, 170a
47b, 380a
0.3e
0.4b, 0.8a
0.7f*, 1.3xl06b
0.7f*, 1.3xl06b
1.2xl08b
2.1xl0lob
Products
ethanol
acetic acid
1,1 -DCE
1,1 -DCE
TCE
1,1,2-TCA
TCE
• From Mabey and Mill, 1978
b From Jeffers el al., 1989
c From Vogel el al., 1987
d From Vogel and McCarty, 1987
° From Cooper et al., 1987
f From Billing el al., 1975
* Butler and Barker (1996) indicate that these values may reflect experimental difficulties
and that the longer half-life [as calculated by Jeffers et al. (1989)] should be used.
Dehydrohalogenation rates may also be approximated using pseudo-first-order kinetics. Once
again, this is not truly a first-order reaction, but such approximations have been used in the literature
to quantify the reaction rates. The rates will not only depend upon the number and types of halogen
substituent, but also on the hydroxide ion concentration. Under normal pH conditions (i.e., near a
B4-62
-------
pH of 7), interaction with water (acting as a weak base) may become more important (Vogel et a/.,
1987). Transformation rates for dehydrohalogenation reactions is presented in Table B.4.1. 1,1,1-
TCA is also known to undergo dehydrohalogenation (Vogel and McCarty, 1987). In this case, TCA
is transformed to 1,1-DCE, which is then reductively dehalogenated to VC. The VC is then either
reductively dehalogenated to ethene or consumed as a substrate in an aerobic reaction and converted
to CO,,. In a laboratory study, Vogel and McCarty (1987) reported that the abiotic conversion of
1,1,1-TCA to 1,1-DCE has a rate constant of about 0.04 year1. It was noted that this result was
longer than indicated in previous studies, but that experimental methods differed. Jeffers el al.
(1989) reported on several other dehydrohalogenation reactions; in addition to 1,1,1-TCA and 1,1,2-
TCA both degrading to 1,1-DCE, the tetrachloroethanes and pentachloroethanes degrade to TCE and
PCE, respectively. Rates of these reactions are included in Table B.4.1. As noted previously, Jeffers
el al. (1989) also report that CA may degrade to VC, but no information on rates was encountered
during the literature search for this Appendix.
B.4.2 REDUCTION REACTIONS
Two abiotic reductive dechlorination reactions that may operate in the subsurface are
hydrogenolysis and dihaloelimination. Hydrogenolysis is the simple replacement of a chlorine (or
another halogen) by a hydrogen, while dihaloelimination is the removal of two chlorines (or other
halogens) accompanied by the formation of a double carbon-carbon bond. Butler and Barker (1996)
review work by Griddle el al. (1986), Jafvert and Wolfe (1987), Reinhard el al. (1990), and Acton
(1990) and this review suggests that while these reactions are thermodynamically possible under
reducing conditions, they often do not take place in the absence of biological activity, even if such
activity is only indirectly responsible for the reaction. While not involved in a manner similar to that
for cometabolism, microbes may produce reductants that facilitate such reactions in conjunction with
minerals in the aquifer matrix, as has been suggested by work utilizing aquifer material from the
Borden test site (Reinhard el al., 1990). Moreover, the reducing conditions necessary to produce
such reactions are most often created as a result of microbial activity. It is therefore not clear if some
of these reactions are truly abiotic, or if because of their reliance on microbial activity to produce
reducing conditions or reactants, they should be considered to be a form of cometabolism.
In some cases, truly abiotic reductive dechlorination has been observed; however, the conditions
that favor such reactions may not occur naturally. For example, Gillham and O'Hannesin (1994)
describe reductive dehalogenation of chlorinated aliphatics using zero-valent iron, in which the iron
serves as an electron donor in an electrochemical reaction. However, this is not a natural process.
Wang and Tan (1990) reported reduction of TCE to ethene and carbon tetrachloride to methane
during a platinum-catalyzed reaction between elemental magnesium and water. Given that the
metals involved in these reactions are unlikely to occur naturally in the reduced forms used in the
aforementioned work, such processes are not likely to contribute to natural attenuation of chlorinated
solvents.
B4-63
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B4-64
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APPENDIX C
and CALCULATIONS
-------
TABLE OF CONTENTS - APPENDIX C
C-l INTRODUCTION Cl-5
C-2 PREPARATION OF GEOLOGIC BORING LOGS,
HYDROGEOLOGIC SECTIONS, AND MAPS C2-6
C.2.1 PREPARATION OF LITHOLOGIC LOGS C2-6
C.2.2 PREPARATION OF HYDROGEOLOGIC SECTIONS C2-7
C.2.3 REVIEW OF TOPOGRAPHIC MAPS AND PREPARATION OF
POTENTIOMETRIC SURFACE MAPS AND FLOW NETS C2-7
C.2.3.1 Review of Topographic Maps C2-7
C.2.3.2 Preparation of Potentiometric Surface Maps C2-7
C.2.3.3 Preparation of Flow Nets C2-9
C.2.3.4 Preparation of Contaminant Isopach Maps C2-9
C.2.3.5 Preparation of Contaminant and Daughter Product Isopleth Maps C2-14
C.2.3.6 Preparation of Electron Donor, Inorganic Electron Acceptor, and
Metabolic By-product Contour (Isopleth) Maps C2-15
C-3 NATURAL ATTENUATION CALCULATIONS C3-18
C.3.1 CALCULATING HYDRAULIC PARAMETERS C3-18
C.3.1.1 Hydraulic Conductivity C3-18
C.3.1.2 Transmissivity C3-20
C.3.1.3 Hydraulic Head and Gradient C3-20
C.3.1.4 Total Porosity (n) and Effective Porosity (ne) C3-23
C.3.1.5 Linear Ground-water Flow Velocity (Seepage or Advective Velocity) C3-24
C.3.1.6 Coefficient of Retardation and Retarded Contaminant Transport Velocity C3-25
C.3.2 CONTAMINANT SOURCE TERM CALCULATIONS C3-28
C.3.2.1 Direct Measurement of Dissolved Contaminant Concentrations in
Ground Water in Contact with NAPL C3-31
C.3.2.2 Equilibrium Partitioning Calculations C3-32
C.3.2.3 Mass Flux Calculations C3-33
C.3.3 CONFIRMING AND QUANTIFYING BIODEGRADATION C3-37
C.3.3.1 Isopleth Maps C3-37
C.3.3.2 Data Set Normalization C3-37
C.3.3.3 Calculating Biodegradation Rates C3-41
C.3.4 DESIGN, IMPLEMENTATION, AND INTERPRETATION OF
MICROCOSM STUDIES C3-49
C.3.4.1 Overview C3-49
C.3.4.2 When to Use Microcosms C3-50
C.3.4.3 Application of Microcosms C3-50
C.3.4.4 Selecting Material for Study C3-50
C.3.4.5 Geochemical Characterization of the Site C3-51
C.3.4.6 Microcosm Construction C3-54
C.3.4.7 Microcosm Interpretation C3-54
C.3.4.8 The Tibbetts Road Case Study C3-55
C.3.4.9 Summary C3-58
Cl-2
-------
FIGURES
No. Title Page
C.2.1 Example hydrogeologic section C2-6
C.2.2 Example ground-water elevation map C2-8
C.2.3 Example flow net C2-9
C.2.4 Example mobile LNAPL isopach (A) and contaminant isopleth (B) maps C2-10
C.2.5 Measured (apparent) versus actual LNAPL thickness C2-11
C.2.6 Type curve for LNAPL baildown test C2-14
C.2.1 Example isopleth maps of contaminants and soluble electron acceptors C2-16
C.2.8 Example isopleth maps of contaminants and metabolic by-products C2-17
C.3.1 Range of hydraulic conductivity values C3-18
C.3.2 Hydraulic head C3-21
C.3.3 Ground-water Elevation Map C3-23
C.3.4 Location of sampling points at the St. Joseph, Michigan NPL site C3-25
C.3.5 Field rate constants for TCE as reported in literature C3-43
C.3.6 Field rate constants for PCE as reported in literature C3-43
C.3.7 Field rate constants for Vinyl Chloride as reported in literature C3-44
C.3.8 Exponential regression of TCE concentration on time of
travel along flow path C3-46
C.3.9 Regression of the TCE concentration on distance along flow path C3-48
C.3.10 Tibbetts Road study site C3-49
C3.ll TCE microcosm results C3-56
C.3.12 Benzene microcosm results C3-56
C.3.13 Toluene microcosm results C3-57
Cl-3
-------
TABLES
No. Title Page
C.2.1 Typical Values for hcaw dr C2-12
C.2.2 Surface Tensions for Various Compounds C2-13
C.2.3 Results of Example Baildown Test C2-15
C.3.1 Representative Values of Hydraulic Conductivity for Various
Sediments and Rocks C3-19
C.3.2 Representative Values of Dry Bulk Density, Total Porosity, and
Effective Porosity for Common Aquifer Matrix Materials C3-24
C.3.3 Representative Values of Total Organic Carbon
for Common Sediments C3-27
C.3.4 Example Retardation Calculations for Select Compounds C3-28
C.3.5 Attenuation of Chlorinated Ethenes and Chloride Downgradient
of the Source of TCE in the West Plume at the St. Joseph, Michigan, NPL Site C3-40
C. 3.6 Use of the Attenuation of a Tracer to Correct the Concentration of TCE
Downgradient of the Source of TCE in the West Plume at the
St. Joseph, Michigan NPL Site C3-41
C.3.7 Geochemical Parameters Important to Microcosm Studies C3-53
C.3.8 Contaminants and Daughter Products C3-53
C.3.9 Concentrations of TCE, Benzene, and Toluene in the
Tibbetts Road Microcosms C3-58
C.3.10 First-order Rate Constants for Removal of TCE, Benzene, and Toluene
in the Tibbetts Road Microcosms C3-59
C.3.11 Concentrations of Contaminants and Metabolic By-products in Monitoring
Wells along Segments in the Plume used to Estimate Field-scale Rate Constants C3-59
C.3.12 Comparison of First-order Rate for Contaminant Attenuation in Segments
of the Tibbetts Road Plume C3-60
C.3.13 Comparison of First-order Rate Constants in a Microcosm Study,
and in the Field at the Tibbetts Road NPL Site C3-60
Cl-4
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SECTION C-l
INTRODUCTION
Successful documentation of natural attenuation requires interpretation of site-specific data to
define the ground-water flow system, refine the conceptual model, quantify rates of contaminant
attenuation, and model the fate and transport of dissolved contaminants. Tasks to be completed
include preparation of lithologic logs, hydrogeologic sections, potentiometric surface maps and flow
nets, contaminant isopach and isopleth maps, electron acceptor and metabolic byproduct isopleth
maps, and calculation of hydraulic parameters, retardation coefficients, and biodegradation rate
constants. The rate and amount of partitioning of organic compounds from mobile and residual
nonaqueous-phase liquid (NAPL) into ground water should also be determined to allow estimation
of a source term. Completion of these tasks permits refinement of the conceptual model and is
necessary to successfully support remediation by natural attenuation.
This appendix consists of three sections, including this introduction. Section C-2 discusses
preparation of geologic boring logs, hydrogeologic sections, and maps. Section C-3 covers natural
attenuation calculations, including hydraulic parameter calculations, contaminant source term
calculations, confirming and quantifying biodegradation, and designing, implementing, and
interpreting microcosm studies.
Cl-5
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SECTION C-2
OF
AND
The first step after completion of site characterization field activities is to prepare geologic
boring logs, hydrogeologic sections, water table elevation (or potentiometric surface) maps, flow
nets, and maps depicting contaminant concentrations, electron acceptor and metabolic byproduct
concentrations, and mobile NAPL thickness. The construction of these items is discussed in the
following sections.
C.2.1 OF
Lithologic logs should be prepared using field data. Whenever possible, these logs should
contain descriptions of the aquifer matrix, including relative density, color, major textural constitu-
ents, minor constituents, porosity, relative moisture content, plasticity of fines, cohesiveness, grain
size, structure or stratification, relative permeability, and any significant observations such as visible
fuel or fuel odor. It is also important to correlate the results of volatile organic compound (VOC)
screening using headspace vapor analysis with depth intervals of geologic materials. The depth of
lithologic contacts and/or significant textural changes should be recorded to the nearest 0.1 foot.
This resolution is necessary because preferential flow and contaminant transport pathways may be
limited to stratigraphic units less than 6 inches thick.
C.2.2 OF
Lithologic logs should be used in conjunction with water level data to prepare a minimum of
two hydrogeologic sections for the site. One section should be oriented parallel to the direction of
ground-water flow, and one section should be oriented perpendicular to the direction of ground-water
flow. Both sections should be drawn to scale. Hydrogeologic sections are an integral part of the
conceptual model and are useful in identifying preferential contaminant migration pathways and in
modeling the site.
At a minimum, hydrogeologic sections should contain information on the relationships between
hydrostratigraphic units at the site, including the location and distribution of transmissive vs. non-
transmissive units, the location of the water table relative to these units, and the location(s) of the
contaminant source(s). Figure C.2.1 is an example of a completed hydrogeologic section.
SW A
A1 NE
Poorly to M oderately
Clayey Silt and Silty Clay
I I Sorted Sands
,, r—i Approximate Extentof
Y7\ Clayey and Sandy Silt | j R e sid u a I a n d M o b ile L N
FEET
VERTICAL EXAGGERATION = 6 7 TIME
Silty and C layey Sand
G ro u n du a te r L ev el
(August 1993)
Figure C.2.1 Example hydrogeologic section.
C2-6
-------
C.2.3 REVIEW OF OF
SURFACE FLOW NETS
Determining the direction of ground-water flow and the magnitude of hydraulic gradients is
important because these parameters influence the direction and rate of contaminant migration.
Ground-water flow directions are represented by a three-dimensional set of equipotential lines and
orthogonal flow lines. If a plan view (potentiometric surface, or water table elevation, map) or a
two-dimensional cross-section is drawn to represent a flow system, the resultant equipotential lines
and flow lines constitute a flow net. A flow net can be used to determine the distribution of hydrau-
lic head, the ground-water velocity distribution, ground-water and solute flow paths and flow rates,
and the general flow pattern in a ground-water system.
C.2.3.1 Review of Topographic Maps
Ground-water flow is strongly influenced by the locations of ground-water divides and by
recharge from and discharge to surface water bodies such as rivers, streams, lakes, and wetlands.
Topographic highs generally represent divergent flow boundaries (divergent ground-water divide),
and topographic lows such as valleys or drainage basins typically represent convergent flow bound-
aries (convergent ground-water divide). In addition, the configuration of the water table is typically a
subtle reflection of the surface topography in the area. However, topography is not always indicative
of subsurface flow patterns and should not be depended upon unless confirmed by head data. In
order to place the local hydrogeologic flow system within the context of the regional hydrogeologic
flow system, it is important to have an understanding of the local and regional topography. Included
in this must be knowledge of the locations of natural and manmade surface water bodies. This
information can generally be gained from topographic maps published by the United States Geologi-
cal Survey.
C.2.3.2 Preparation of Potentiometric Surface Maps
A potentiometric surface map is a two-dimensional graphical representation of equipotential
lines shown in plan view. Water table elevation maps are potentiometric surface maps drawn for
water table (unconfined) aquifers. Potentiometric surface maps for water table aquifers show where
planes of equal potential intersect the water table. A potentiometric surface map should be prepared
from water level measurements and surveyor's data. These maps are used to estimate the direction
of plume migration and to calculate hydraulic gradients. To document seasonal variations in ground-
water flow, separate potentiometric surface maps should be prepared using quarterly water level
measurements taken over a period of at least 1 year.
The data used to develop the potentiometric surface map should be water level elevation data
(elevation relative to mean sea level) from piezometers/wells screened in the same relative position
within the same hydrogeologic unit. For example, wells that are screened at the water table can be
used for the same potentiometric surface map. Wells screened in different hydrogeologic units or at
different relative positions within the same water table aquifer cannot be used to prepare a potentio-
metric surface map. Where possible, a potentiometric surface map should be prepared for each
hydrogeologic unit at the site. In recharge areas, wells screened at various elevations cannot all be
used to prepare the same potentiometric surface map because of strong downward vertical gradients.
Likewise, wells screened at various elevations in discharge areas such as near streams, lakes, or
springs, should not all be used because of the strong upward vertical gradients.
When preparing a potentiometric surface map, the locations of system boundaries should be
kept in mind; particularly the site features that tend to offset the shape of the contours on the map.
Such features include topographic divides, surface water bodies, and pumping wells.
In addition to, and separately from, preparation of a potentiometric surface map, water level
measurements from wells screened at different depths can be used to determine any vertical hydrau-
C2-7
-------
lie gradients. It is important to have a good understanding of vertical hydraulic gradients because
they may have a profound influence on contaminant migration.
In areas with measurable mobile LNAPL, a correction must be made for the water table deflec-
tion caused by the LNAPL. The following relationship, based on Archimedes' Principle, provides a
correction factor that allows the water table elevation to be adjusted for the effect of floating
LNAPL.
(PT)
eq. C.2.1
Where:
CDTW = corrected depth to water [L]
MDTW = measured depth to water [L]
plnapl = density of the LNAPL [M/L3]
pw = density of the water, generally 1.0 [M/L3]
PT = measured LNAPL thickness [L]
Using the corrected depth to water, the corrected ground-water elevation, CGWE, is given by:
CGWE = Datum Elevation -CDTW eq. C.2.2
Corrected ground-water elevations should be used for potentiometric surface map preparation.
Figure C.2.2 is an example of a ground-water elevation map for an unconfined aquifer. Water table
elevation data used to prepare this map were taken from wells screened across the water table.
4604,17 \ 4616.40
Colony Loop
4643.21
GROUNDWATER FLOW DIRECTION
4600.06 GROUNDWATER ELEVATION (Feet Above MSL)
4580—•*
LINE OF EQUAL GROUNDWATER
ELEVATION (Feet Above MSL)
Figure C.2.2 Example ground-wafer elevation map.
C2-8
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C.2.3.3 Preparation of Flow Nets
Where an adequate three-dimensional database is available, flow nets can be constructed to
facilitate the interpretation of the total hydraulic head distribution in the aquifer. This will help
determine potential solute migration pathways. The simplest ground-water flow system is one that is
homogeneous and isotropic. This type of hydrogeologic setting serves as a simple basis for describ-
ing the basic rules of flow net construction, despite the fact that homogeneous, isotropic media rarely
occur in nature. Regardless of the type of geologic media, the basic rules of flow net construction
must be applied, and necessary modifications must be made throughout the procedure to account for
aquifer heterogeneity or anisotropic conditions. Water level data for flow net construction should
come from multiple sets of nested wells (two or more wells at the same location) at various depths in
the aquifer. The fundamental rules of flow net construction and the important properties of flow nets
are summarized as follows:
« Flow lines and equipotential lines intersect at 90-degree angles if the permeability is
isotropic;
« The geometric figures formed by the intersection of flow lines and equipotential lines must
approximate squares or rectangles;
« Equipotential lines must meet impermeable boundaries at right angles (impermeable
boundaries are flow lines); and
« Equipotential lines must be parallel to constant-head boundaries (constant-head bound-
aries are equipotential lines).
Trial-and-error sketching is generally used to construct a flow net. Flow net sketching can be
sufficiently accurate if constructed according to the basic rules outlined above. A relatively small
number of flow lines (three to five) generally are sufficient to adequately characterize flow condi-
tions. Flow nets should be superimposed on the hydrogeologic sections. Figure C.2.3 is an example
of a completed flow net. This figure shows ground-water flow patterns in both recharge and dis-
charge areas.
C.2.3.4 Preparation of Contaminant Isopacfa Maps
If NAPL is present at the site, isopach maps showing the thickness and distribution of NAPL
should be prepared. Two maps should be prepared: one for mobile NAPL, and one for residual
NAPL. Such isopach maps allow estimation of the distribution of NAPL in the subsurface and aid in
Convergent Groundwater
Divergent Groundwater Divide
Water Table
*• Flow Line
^--^ Equipotential Line
10 Total Head (meters)
Modified from Hubbert (1 940)
Figure C.2.3 Example flow net.
C2-9
-------
fate and transport model development by identifying the boundary of the NAPL. Because of the
differences between the magnitude of capillary suction in the aquifer matrix and the different surface
tension properties of fuel and water, LNAPL thickness observations made in monitoring points are
only an estimate of the actual volume of mobile LNAPL in the aquifer. To determine the actual
NAPL thickness it is necessary to collect and visually analyze soil samples. LNAPL thickness data
also should be used to correct for water table deflections caused by the mobile LNAPL. This process
is described in Section C.2.2.3.2.
Isopach maps are prepared by first plotting the measured NAPL thickness on a base map pre-
pared using surveyor's data. Lines of equal NAPL thickness (isopachs) are then drawn and labeled.
Each data point must be honored during contouring. Figure C.2.4 is an example of a completed
isopach map. This figure also contains an example of an isopleth map.
C.2.3.4.1 Relationship Between Apparent and Actual LNAPL Thickness
It is well documented that LNAPL thickness measurements taken in ground-water monitoring
wells are not indicative of actual LNAPL thicknesses in the formation (de Pastrovich et a/., 1979;
Blake and Hall, 1984; Hall el aL, 1984; Hughes el aL, 1988; Abdul el aL, 1989; Testa and
Paczkowski, 1989; Fair elal., 1990; Kemblowski and Chiang, 1990; Lenhard and Parker, 1990;
Mercer and Cohen, 1990; Ballestero el aL, 1994; Huntley el ai, 1994a). These authors note than the
measured thickness of LNAPL in a monitoring well is greater than the true LNAPL thickness in the
aquifer and, according Mercer and Cohen (1990), measured LNAPL thickness in wells is typically 2
to 10 times greater than the actual LNAPL thickness in the formation. The difference between actual
and measured LNAPL thickness occurs because mobile LNAPL floating on the water table flows
into the well (if the top of well screen is above the base of the LNAPL) and depresses the water
table. Figure C.2.5 is a schematic that illustrates this relationship. The equation for correcting depth
A) MOBILE LNAPL ISOPACH MAP
B) CONTAMINANT ISOPLETH MAP
SOURCE AREA
0.36
1.00
IVPSUKTO WR THCJMSS (Fffl)
LMCFEOjaLLWFL
THCM\ESS(FK)
CCMCURINERvR=1FCa
1.27
-1000
TdABTKCCKMlWICNI gi)
LMCFEORCCNW*™
COCMWICNI gl) n
CCMCURINERvR=1,CCO
Figure C.2.4 Example mobile LNAPL isopach (A) and contaminant isopleth (B) maps.
C2-10
-------
to ground water caused by LNAPL in the well is given in Section C.2.3.2. Empirical relationships
relating measured LNAPL thickness to actual LNAPL thickness are presented below. Also presented
below are test methods that can be used to determine actual LNAPL thickness. There are no estab-
lished methods for determining actual DNAPL volume based on measurements taken in monitoring
wells.
,LN APLFraction at or Below
Residual Saturation
Fraction Greater Than
Residual Saturation
LEGEND Measured Water Table ^
I I Residual Hydrocarbons
^^| Free Liquid Hydrocarbons
Source: Modified from de Pastrovich and others, 1972.
Zone of LNAPL
Capillary Rise
Actual
LNAPL Thickness
Zone of Water
Capillary Rise
Figure C.2.5 Measured (apparent) versus actual LNAPL thickness.
C.2.3.4.2. Empirical Relationships
There are several empirical methods available to estimate the actual thickness of mobile
LNAPL in the subsurface based on LNAPL thicknesses measured in a ground-water monitoring
well. Such empirical relationships are, at best, approximations because many factors influence the
relationship between measured and apparent LNAPL thickness (Mercer and Cohen, 1990):
• Capillary fringe height depends on grain size and is hysteretic with fluid level fluctuations.
• LNAPL can become trapped below the water table as the water table rises and falls.
• The thickness of LNAPL is ambiguous because the interval of soil containing mobile
LNAPL is not 100-percent saturated with LNAPL.
Some empirical methods for determining actual LNAPL thickness are described below.
Method of de Pastrovich etal (1979)
Hampton and Miller (1988) conducted laboratory experiments to examine the relationship
between the actual thickness of LNAPL in a formation, h and that measured in a monitoring well,
h . Based on their research, Hampton and Miller (1988) suggest using the following relationship
(developed by de Pastrovich et al, 1979) to estimate LNAPL thickness:
C2-11
-------
Where:
hf =
h =
m
P =
jr w
J o
r Inapl
actual thickness of LNAPL in formation
measured LNAPL thickness in well
density of water (1.0 gm/cm3 for pure water)
= density of LNAPL (See Table C.3.9)
eq. C.2.3
Method of Kemblowski and Chiang (1990)
Another empirical relationship was proposed by Kemblowski and Chiang (1990) to estimate
actual LNAPL thickness based on measured LNAPL thickness. This relationship is given by:
h=H-2.2hc
dr
eq.C.2.4
Where:
h --
o
H --
o
hL
equivalent thickness of LNAPL in the formation (volume of oil per unit area of aquifer,
divided by porosity)
measured LNAPL thickness in well
dr
= capillary height of air-water interface assuming water is being displaced by oil
(typical values are given in Table C.2. 1)
This method assumes equilibrium conditions, water drainage, and oil imbibition.
Table C.2.1 Typical Values for hcaw dr (Bear, 1972)
Aquifer Matrix
Coarse Sand
Sand
Fine Sand
Silt
Clay
^*(cm)
2-5
12-35
35-70
70-150
>200-400
^L (ft)
0.066-0.16
0.39-1.15
1.14-2.30
2.30-4.92
>6.56-13.12
Method of Lenhard and Parker (1990)
Another empirical relationship was proposed by Lenhard and Parker (1990) to estimate actual
LNAPL thickness based on measured LNAPL thickness. This relationship is given by:
D=-
- Pro)
eq. C.2.5
Where:
Do= actual thickness of LNAPL in formation
H = measured LNAPL thickness in well
o
Pm= specific gravity of LNAPL (density of oil/density of water)
cr
(3ao= Air-oil scaling factor
C2-12
-------
' OW
"Oil-water scaling factor
Gaw = surface tension of uncontaminated water (72.75 dynes/cm @ 20°C)
oT = surface tension of LNAPL [25 dynes/cm @20°C for JP-4, Table C.2.2]
o = 0-0 = interfacial tension between water and LNAPL (47.75dynes/cm @ 20°C)
ow aw ao v J ^—' x
It is important to note that this method includes the capillary thickness of the hydrocarbon, and is,
therefore, likely to be an overestimate.
Table C.2.2 Surface Tensions for Various Compounds
Compound
JP-4
Gasoline
Pure Water
Surface Tension @ 20°C (dyne/cm)
25^
19-23 ^
72.75b/
a/Mar lei (1987).
b/ CRC Handbook (1956).
C.2.3.4.3. LNAPL Baildown Test
The LNAPL baildown test is applicable in areas where the hydrocarbon/water interface is below
the potentiometric surface, and the recharge rate of hydrocarbon into the well is slow (Hughes et al.,
1988).
Baildown Test Procedure (from Hughes et al., 1988):
1) Gauge the well and calculate the corrected potentiometric surface elevation using equations
C.2. land C.2.2.
2) Rapidly bail the hydrocarbon from the well.
3) Gauge the well again, and if the thickness of the hydrocarbon is acceptable (0.1 to 1 foot),
calculate the potentiometric surface elevation. The potentiometric surface elevation thus
calculated should be within 0.005 foot of the value calculated in step 1. If it is, then continue
to step 4; if it is not, repeat steps 2 and 3.
4) Record the top of the LNAPL surface in the well as it recharges until the well is fully re-
charged.
5) Plot the elevation of the top of LNAPL in the well vs. time since bailing ceased.
6) The true thickness of the mobile LNAPL layer (T ) is the distance from the inflection point to
the top of the hydrocarbon under static conditions (Figure C.2.6). Thus, Tf is picked directly
off the plot. Table C.2.3 is an example of the results of this procedure.
C2-13
-------
g
I
J>
UJ
_j
a.
<
• Inflection Point
Time
Figure C.2.6 Type curve for LNAPL baildown test.
Table C.2.3 Results of Example Baildown Test (Modified from Hughes etai, 1988)
Well
ROW- 143
ROW- 189
ROW- 129
Tw
(ft)a/
4.97
12.5
0.94
Tf
(ft)
0.61
0.29
0.0b/
Exaggeration (Tw/Tf)
8.1:1
43.0:1
N/A
a/ T - LNAPL thickness initially measured in the well, if LNAPL thickness that is actually
mobile
b/ Capillary oil only
Hughes et al, (1988) also present a recharge method that involves pumping the mobile LNAPL until
steady-state conditions are achieved, and then letting the well fully recharge.
C.2.3.5 Preparation of Contaminant and Daughter Product Isopleth Maps
Isopleth maps should be prepared for all chlorinated solvents of concern and their daughter
products and for total BTEX if present. For example, if trichloroethene and BTEX were released (as
is typical for fire training areas), then maps of dissolved trichloroethene, dichloroethene, vinyl
chloride, ethene, and total BTEX concentrations should be prepared. Isopleth maps allow interpreta-
tion of data on the distribution and the relative transport and degradation rates of contaminants in the
subsurface. In addition, contaminant isopleth maps allow contaminant concentrations to be gridded
and used for input into a solute transport model.
Isopleth maps are prepared by first plotting the concentration of the contaminant on a base map
prepared using surveyor's data. Lines of equal contaminant concentration (isopleths) are then drawn
C2-14
-------
and labeled. It is important to ensure that each data point is honored during contouring. Outliers
should be displayed and qualified, if they are not contoured. Figures C.2.4, C.2.7, and C.2.8 contain
examples of contaminant isopleth maps.
Dissolved contaminant concentrations are determined through ground-water sampling and
laboratory analysis. From these data, isopleth maps for each of the contaminant compounds and for
total dissolved contaminant should be made. Dissolved BTEX concentrations are transferred to the
fate and transport model grid cells by overlaying the isopleth map onto the model grid.
C.2.3.6 Preparation of Electron Donor, Inorganic Electron Acceptor, and Metabolic By-
product Contour (Isopleth) Maps
Isopleth maps should be prepared for any organic compound that can be used as an electron
donor. Examples of such compounds include natural organic carbon, and petroleum hydrocarbons
(and landfill leachate). These maps are used to provide visible evidence that biodegradation could
occur or is occurring. Isopleth maps also should be prepared for dissolved oxygen, nitrate,
manganese (II), iron (II), sulfate, methane, and chloride. These maps are used to provide visible
evidence that biodegradation is occurring. The electron acceptor and metabolic by-product isopleth
maps can be used to determine the relative importance of each of the terminal electron-accepting
processes (TEAPs).
Isopleth maps are prepared by first plotting the concentration of the electron donor, electron
acceptor, or metabolic by-product on a base map prepared using surveyor's data. Lines of equal
concentration (isopleths) are then drawn and labeled. It is important to ensure that each data point is
honored during contouring, unless some data are suspect.
C.2.3.6.1 Inorganic Electron Acceptor Isopleth Maps
Electron acceptor isopleth maps allow interpretation of data on the distribution of dissolved
oxygen, nitrate, and sulfate in the subsurface. Isopleth maps for these compounds provide a visual
indication of the relationship between the contaminant plume and the electron acceptors and the
Figure C.2.7 Example isopleth maps of contaminants and soluble electron acceptors,
C2-15
-------
relative importance of each TEAP. Dissolved oxygen concentrations below background levels in
areas with high organic carbon concentrations are indicative of aerobic respiration. Nitrate concen-
trations below background in areas with high organic carbon concentrations are indicative of denitri-
fication. Sulfate concentrations below background in areas with high organic carbon concentrations
are indicative of sulfate reduction.
Figure C.2.7 gives examples of completed isopleth maps for dissolved oxygen, nitrate, and
sulfate. This figure also contains isopleth maps for TCE and DCE and the total BTEX (electron
donor) isopleth map for the same period. Comparison of the total BTEX isopleth map and the
electron acceptor isopleth maps shows that there is a strong correlation between areas with elevated
organic carbon and depleted electron acceptor concentrations. The strong correlation indicates that
the electron acceptor demand exerted during the metabolism of BTEX has resulted in the depletion
of soluble inorganic electron acceptors. These relationships provide strong evidence that biodegra-
dation is occurring via the processes of aerobic respiration, denitrification, and sulfate reduction.
TOTAL BTEX (mg/L)
.IRON(II) (mg/L)
(mg/L)
(mgO.) (mg/L)
(mg/L)
\
Figure C.2.8 Example isopleth maps of contaminants and metabolic by-products.
C2-16
-------
C.2.3.6.2 Metabolic By-product Isopleth Maps
Metabolic by-product maps should be prepared for manganese (II), iron (II), methane, and
chloride. The manganese (II) map is prepared in lieu of an electron acceptor isopleth map for
manganese (IV) because the amount of bioavailable amorphous or poorly crystalline manganese (IV)
in an aquifer matrix is extremely hard to quantify. The iron (II) map is prepared in lieu of an electron
acceptor isopleth map for iron (III) because the amount of bioavailable amorphous or poorly crystal-
line iron (III) in an aquifer matrix is extremely hard to quantify. Iron (II) concentrations above
background levels in areas with BTEX contamination are indicative of anaerobic iron (III) reduction.
Methane concentrations above background levels in areas with BTEX contamination are indicative
of methanogenesis, another anaerobic process. Biodegradation of chlorinated solvents tends to
increase the chloride concentration found in ground water. Thus, chloride concentrations inside the
contaminant plume generally increase to concentrations above background. This map will allow
visual interpretation of chloride data by showing the relationship between the contaminant plume
and chloride. During anaerobic biodegradation, the oxidation-reduction potential of ground water is
lowered. Thus, the oxidation-reduction potential (or pE) inside the contaminant plume generally
decreases to levels below background.
Figure C.2.8 gives examples of completed isopleth maps for iron (II), methane, chloride, and
pE. This figure also contains the TCE, DCE and Vinyl Chloride isopleth maps, and total BTEX
(electron donor) isopleth map for the same period. Comparison of the total BTEX isopleth map and
the metabolic by-product isopleth maps and comparison of Figures C.2.7 and C.2.8 shows that there
is a strong correlation between areas with elevated organic carbon and elevated metabolic by-product
concentrations. These relationships provide strong evidence that biodegradation is occurring via the
processes of iron (III) reduction, methanogenesis, and reductive dechlorination.
C2-17
-------
C-3
NATURAL ATTENUATION CALCULATIONS
Several calculations using site-specific data must be made in order to document the occurrence
of natural attenuation and successfully implement the natural attenuation alternative. The following
sections describe these calculations.
C.3.1 CALCULATING
Hydraulic parameters necessary for adequate site characterization and model implementation
include hydraulic conductivity, transmissivity, hydraulic gradient, linear ground-water flow velocity,
hydrodynamic dispersion, and retarded solute transport velocity. Calculations for these parameters
are discussed in the following sections.
C.3.1.1 Hydraulic Conductivity
Hydraulic conductivity, K, is a measure of an aquifer's ability to transmit water and is perhaps
the most important variable governing fluid flow in the subsurface. Hydraulic conductivity has the
units of length over time [L/T]. Observed values of hydraulic conductivity range over 12 orders of
magnitude, from 3xlO12 to 3 cm/sec (3xlO9 to 3xl03 m/day) (Figure C.3.1 and Table C.3.1). In
Unconsolidated
Deposits
Gravel
coarse
flrMME
in*
Sand
rrxxSurrt
fine
Alluvial deposits
Silt
Clay
dense
w«attitrtd
Rocks
Sandstone
dans*
Limestone
dense
kareBc
Dolomite
Crystalline recks
itense
fractured
Basalt
dansa
teamed
Claystone
Volcanic tuff
Shale
dense
fiactumd
BH typical range
Modified from: Spte and Moreno, 1996.
Figure C.3.1 Range of hydraulic conductivity values.
C3-18
-------
general terms, the hydraulic conductivity for unconsolidated sediments tends to increase with in-
creasing grain size and sorting. The velocity of ground water and dissolved contaminants is directly
related to the hydraulic conductivity of the saturated zone. Subsurface variations in hydraulic con-
ductivity directly influence contaminant fate and transport by providing preferential pathways for
contaminant migration. The most common methods used to quantify hydraulic conductivity in the
subsurface are aquifer pumping tests and slug tests. The quantitative analysis of pumping and slug
test data is beyond the scope of this document. For information on the quantitative analysis of these
data, the reader is referred to the works of Kruseman and de Ridder (1991) and Dawson and Istok
(1991).
Table C.3.1 Representative Values of Hydraulic Conductivity for Various Sediments and Rocks (From
Domenico and Schwartz, 1990)
Material
UNCONSOLIDATED
SEDIMENT
Glacial till
Clay
Silt
Fine sand
Medium sand
Coarse sand
Gravel
SEDIMENTARY ROCK
Karstic limestone
Limestone and dolomite
Sandstone
Siltstone
Shale
CRYSTALLINE ROCK
Vesicular basalt
Basalt
Fractured igneous and
metamorphic
Unfractured igneous
and metamorphic
Hydraulic Conductivity
(m/day)
9xW-s-2xW~l
9xlO"7-4xlO"4
9xlCT5-2
2xlCr2-2xl01
SxlCT2 - SxlO1
SxlCT2 - SxlO2
SxlC^-SxlO3
9xlCr2-2xl03
9xlCT5 - SxlCT1
SxlCT5 - SxlCT1
9xlCT7 - IxlCT3
9xlO-9-2xlCr4
3xlCr2-2xl03
2xlO-6-3xlCr2
7xlCT4 - SxlO1
3xlO"9-2xlO"5
Hydraulic Conductivity
(cm/sec)
Ixl0-10-2xicr4
Ixicr9 - Sxicr7
Ixl0"7-2xl0"3
2xlO-5-2xlCT2
9xlCT5-6xlO-2
9xW-5-6xW~l
3xlCT2-3
lxlCT4-2
Ixl0"7-6xl0"4
3xlCT8-6xlO-4
IxlCT9 - IxlCT6
Ixl0-n-2xicr7
4xlCT5-2
2xlO-9-4xlO-5
SxlO"7 - 3xlO"2
3xlO"12-2xlO"8
C3-19
-------
C.3.1.1.1 Hydraulic Conductivity from Pumping Tests
Pumping tests generally provide the most reliable information about aquifer hydraulic conduc-
tivity. Pumping test data for geohydraulic characteristics are most commonly interpreted by graphic
techniques. The analytical method used for interpretation of the data will depend upon the physical
characteristics of the aquifer and test wells. The assumptions inherent in the analytical method used
to calculate aquifer characteristics should be evaluated to ensure acceptance of the method for the
subsurface conditions present at the site under investigation.
The interpretation of aquifer pumping test data is not unique. Similar sets of data can be ob-
tained from various combinations of geologic conditions. Field data of drawdown vs. time and/or
distance are plotted on graph paper either by hand or using programs such as AQTESOLV® or a
spreadsheet program. There are numerous methods of interpreting pumping test data. The method
to be used for each pumping test should be selected based on site-specific conditions (aquifer condi-
tions, test conditions, assumptions made, etc.). Most hydrogeology text books contain pumping test
evaluation techniques. Two publications dealing with pump test analysis are recommended
(Kruseman and de Ridder, 1991 and Dawson and Istok, 1991).
C.3.1.1.2 Hydraulic Conductivity from Slug Tests
Slug tests are a commonly used alternative to pumping tests that are relatively easy to conduct.
The biggest advantage of slug tests is that no contaminated water is produced during the test. During
pumping tests at fuel-hydrocarbon-contaminated sites, large volumes of contaminated water that
must be treated typically are produced. One commonly cited drawback to slug testing is that this
method generally gives hydraulic conductivity information only for the area immediately surround-
ing the monitoring well. If slug tests are going to be relied upon to provide information on the three-
dimensional distribution of hydraulic conductivity in an aquifer, multiple slug tests must be per-
formed, both within the same well and at several monitoring wells at the site. It is not advisable to
rely on data from one slug test in a single monitoring well. Data obtained during slug testing are
generally analyzed using the method of Hvorslev (1951) for confined aquifers or the method of
Bouwer and Rice (1976) and Bouwer (1989) for unconfined conditions.
C.3.1.2 Transmlssivlty
The transmissivity, T, of an aquifer is the product of the aquifer's hydraulic conductivity, K, and
the saturated thickness, b:
T=Kb eq. C.3.1
For a confined aquifer, b is the thickness of the aquifer between confining units. For uncon-
fined aquifers, b is the saturated thickness of the aquifer measured from the water table to the under-
lying confining layer. Transmissivity has the units of length squared over time [L2/T].
C.3.1.3 Hydraulic Head and Gradient
Determining the magnitude of hydraulic gradients is important because gradients influence the
direction and rate of contaminant migration. Hydraulic head, H, and specifically, variations in
hydraulic head within an aquifer, is the driving force behind ground-water movement and solute
migration. The total hydraulic head at one location in a system is the sum of the elevation head,
pressure head, and velocity head (Figure C.3.2):
H = hz+hp+hv eq. C.3.2
Where:
H = total hydraulic head [L]
h_ = elevation head = z = elevation relative to the reference plane [L]
h = pressure head [L]
h: = velocity head [L]
C3-20
-------
Pressure head is given by:
P Pg
Where:
p = fluid pressure
p = density
g = acceleration due to gravity
Velocity head is given by:
Where:
v = ground-water velocity
g = acceleration due to gravity
Because hv is generally assumed to be zero for most ground-water flow, the relationship for total
head is generally written:
H=z + -
Pg
eq. C.3.3
Thus, the total hydraulic head at a point measured by a piezometer is the sum of the elevation at the
base of the piezometer plus the length of the water column in the piezometer. The total hydraulic
head in a piezometer is determined by measuring the depth from a surveyed reference point (datum)
to the surface of the standing water. The elevation of the water surface is the total hydraulic head in
the piezometer. This total head is the total head at the base of the piezometer, not the water table
elevation, unless the piezometer terminates immediately below the water table or is a well screened
A B / — M easurementDatum
t
Depth to Water
P re s s u re
Hea
\
i
Elev
H e
i
dl
r
.
Total
atio n
d (z|
H ead (H
; — Open-ended Tube (Piezometer)
^ Ground Surface
1
Depth to Water
(-T -
'
Total H
i
.
ead (h
1 v Water Table
K~P re s s u re Hea dEf
)
E lev atio n
H e
' i
at (z)
Mean Sea Level (Reference Elevation
Figure C.3.2 Hydraulic head.
C3-21
-------
across the water table. Figure C.3.2 shows a pair of nested piezometers that illustrate the relation-
ships between total hydraulic head, pressure head, and elevation head. Because ground water flows
from areas with high total head (point A, Figure C.3.2) to areas with lower total head (point B), this
figure depicts a water table aquifer with a strong upward vertical gradient. This figure illustrates
how nested piezometers (or wells) are used to determine the importance of vertical gradients at a
site. This figure also illustrates the importance of using wells screened in the same portion of the
aquifer (preferably across the water table) when preparing potentiometric surface maps.
The hydraulic gradient (dH/dL) is a dimensionless number that is the change in hydraulic head
(dH) between two points divided by the length of ground-water flow between these same two points,
parallel to the direction of ground-water flow, and is given by:
dH
Hydraulic Gradient = eq C34
dL
Where:
dH = change in total hydraulic head between two points [L]
dL = distance between the two points used for head measurement [L]
In a system where flow is not occurring, the total hydraulic head, H, is the same everywhere in
the system and the hydraulic gradient is zero. To accurately determine the hydraulic gradient, it is
necessary to measure ground-water levels in all monitoring wells at the site. Because hydraulic
gradients can change over a short distance within an aquifer, it is essential to have as much site-
specific ground-water elevation information as possible so that accurate hydraulic gradient calcula-
tions can be made. In addition, seasonal variations in ground-water flow direction can have a pro-
found influence on contaminant transport. To determine the effect of seasonal variations in ground-
water flow direction on contaminant transport, quarterly ground-water level measurements should be
taken over a period of at least 1 year.
The hydraulic gradient must be determined parallel to the direction of ground-water flow.
Unless two monitoring wells screened in the same relative location within the same hydrogeologic
unit are located along a line parallel to the direction of ground-water flow, the potentiometric surface
map is generally used to determine the hydraulic gradient. To determine the hydraulic gradient, an
engineer's scale is used to draw a line perpendicular to the equal-potential lines on the potentiomet-
ric surface map (i.e., parallel to the direction of ground-water flow). Measure the distance between
the two equal-potential lines, making note of the ground-water potential at each equal-potential line.
Subtract the larger potential from the smaller potential, and divide this number by the distance
between the two equal potential lines, being sure to use consistent units. The number generated will
be a negative number because water flows from areas of higher potential to areas of lower potential.
ExampleC.3.1: Hydraulic Gradient Calculation
Given the water table elevation map shown in Figure C.3.3, calculate the hydraulic gradient
between points A and B. Assume that all wells are screened across the water table.
Solution:
The hydraulic gradient is given by dH/dL. The line connecting points A and B is parallel to the
direction of ground-water flow. The water table elevation is 4659.34 ft msl at point A and
4602.41 ft msl at point B. Therefore, because ground water flows from areas of high head to areas of
lower head:
dH= 4602.41 - 4659.34 = - 56.93 feet
The distance between the two points A and B is 936 feet. Therefore:
dL = 936 feet
C3-22
-------
\
4600.55 4^OQ.Qo
4596.69
4602.41
4617.87
'4599.66
GROUND WATER ELEVATION (MSL)
LINE OF EQUAL GROUND WATER
ELEVATION (FEET ABOVE MSL)
Figure C.3.3 Ground water elevation map.
and
dL 936ft
=-9.06 ^- = -0.06—
ft m
C.3.1.4 Total Porosity (n) Effective Porosity (ne)
Total porosity (n) is the volume of voids in a unit volume of aquifer. Specific retention is the
amount of water (volumetric) that is retained against the force of gravity after a unit volume of an
unconfmed aquifer is drained. Storativity is defined as the volume of water that a confined aquifer
takes into or releases from storage per unit surface area of the aquifer per unit change in total hydrau-
lic head. Effective porosity, ne, is the total porosity of the aquifer minus the specific retention (un-
confmed) or storativity (confined) of the aquifer:
ne=n-S eq. C.3.5
Where:
ne = effective porosity [dimensionless]
n = total porosity [dimensionless]
S = specific retention (unconfmed) or storativity (confined) [dimensionless]
Effective porosity can be estimated using the results of a tracer test. Although this is potentially the
most accurate method, time and monetary constraints can be prohibitive. For this reason, the most
common technique is to use an accepted literature value for the types of materials making up the
aquifer matrix, and then to calibrate a contaminant transport model by adjusting the value of effec-
tive porosity (in conjunction with other input parameters such as transmissivity) within the range of
C3-23
-------
accepted literature values until the modeled and observed contaminant distribution patterns match.
Because aquifer materials can have a range of effective porosity, sensitivity analyses should be
performed to determine the effect of varying the effective porosity on numerical model results.
Values of effective porosity chosen for the sensitivity analyses should vary over the accepted range
for the aquifer matrix material. Table C.3.2 presents accepted literature values for total porosity and
effective porosity.
Table C.3.2 Representative Values of Dry Bulk Density, Total Porosity, and Effective Porosity for
Common Aquifer Matrix Materials (After Walton, 1988 and Domenico and Schwartz, 1990)
Aquifer
Matrix
Clay
Peat
Glacial
Sediments
Sandy Clay
Silt
Loess
Fine Sand
Medium Sand
Coarse Sand
Gravely Sand
Fine Gravel
Medium
Gravel
Coarse Gravel
Sandstone
Siltstone
Shale
Limestone
Granite
Basalt
Volcanic Tuff
Dry Bulk
Density
(gm/cm)
1.00-2.40
—
1.15-2.10
—
—
0.75-1.60
1.37-1.81
1.37-1.81
1.37-1.81
1.37-1.81
1.36-2.19
1.36-2.19
1.36-2.19
1.60-2.68
—
1.54-3.17
1.74-2.79
2.24-2.46
2.00-2.70
—
Total
Porosity
0.34-
0.60
—
—
—
0.34-
0.61
—
0.26-
0.53
—
0.31-
0.46
—
0.25-
0.38
—
0.24-
0.36
0.05-
0.30
0.21-
0.41
0.0-0.10
0.0-50
—
0.03-
0.35
—
Effective
Porosity
0.01-0.2
0.3-0.5
0.05-0.2
0.03-0.2
0.01-0.3
0.15-0.35
0.1-0.3
0.15-0.3
0.2-0.35
0.2-0.35
0.2-0.35
0.15-0.25
0.1-0.25
0.1-0.4
0.01-0.35
—
0.01-0.24
—
—
0.02-0.35
C.3.1.5 Linear Ground-water Flow Velocity (Seepage or Advective Velocity)
The average linear ground-water flow velocity (seepage velocity) in one dimension in the
direction parallel to ground-water flow in a saturated porous medium is given by:
C3-24
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KdH
V*=^ eq-c-3
Where:
vx = average linear ground-water velocity parallel to ground-water flow direction (seepage
velocity) [L/T]
K = hydraulic conductivity [L/T]
ne = effective porosity [L3/L3]
dH
—7j- = hydraulic gradient [L/L]
The average linear ground-water flow velocity should be calculated to estimate ground-water flow
and solute transport velocity, to check the accuracy of ground-water models, and to calculate first-
order biodegradation rate constants.
Example C. 3. 2: Linear Ground-water Flow Velocity Calculation
Calculate the linear ground-water flow velocity in a medium-grained sandy aquifer. The hy-
draulic gradient as determined from the potent! ometric surface map in the previous example is -
0.06 m/m. The hydraulic conductivity is 1 .TxlO"1 m/day as determined by pumping tests.
Solution:
Because the effective porosity of this sediment is not known, it is necessary to estimate this
parameter. From Table C.3.2, the effective porosity for a medium-grained sand is approximately
23 percent.
n dL 0.23
C.3.1.6 Coefficient of Retardation Retarded Contaminant Transport Velocity
When the average linear velocity of a dissolved contaminant is less than the average linear
velocity of the ground water, the contaminant is said to be "retarded." The difference between the
velocity of the ground water and that of the contaminant is caused by sorption and is described by the
coefficient of retardation, R, which is defined as:
D Vx
R = — eq.C.3.7
r c
Where:
R = coefficient of retardation
vx = average linear ground-water velocity parallel to ground-water flow
vc = average velocity of contaminant parallel to groundwater flow
The ratio VX/YC describes the relative velocity between the ground water and the dissolved contami-
nant. When Kd = 0 (no sorption), the transport velocities of the ground water and the solute are equal
(vx = vc). If it can be assumed that sorption is adequately described by the distribution coefficient,
the coefficient of retardation for a dissolved contaminant (for saturated flow) is given by:
eq.C.3.8
n
Where:
R = coefficient of retardation
ph = bulk density (Section C. 3. 1.6.1)
Kd= distribution coefficient (Section C. 3. 1 .6.2)
n = total porosity
C3-25
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This relationship expresses the coefficient of retardation in terms of the bulk density and effective
porosity of the aquifer matrix and the distribution coefficient for the contaminant. Substitution of
this equation into equation C.3.7 gives:
vc n
Solving for the contaminant velocity, v,, gives:
v =
I | Pb^d eq. C.3.10
n
Retardation of a contaminant relative to the advective transport velocity of the ground-water flow
system has important implications for natural attenuation. If retardation is occurring, dissolved
oxygen and other electron acceptors traveling at the advective transport velocity of the ground water
sweep over the contaminant plume from the upgradient margin. This results in greater availability of
electron acceptors within the plume for biodegradation of fuel hydrocarbons. In addition, adsorption
of a contaminant to the aquifer matrix results in dilution of the dissolved contaminant plume.
C.3.1.6.1 Bulk Density
The bulk density of a soil, pb, as used in most ground-water models, expresses the ratio of the
mass of dried soil to its total volume (solids and pores together).
_ Ms _ Ms
Where:
pb = bulk density
M= mass of solid in the system
VT = total volume in the system
F = volume of solid in the system
= volume of air (or gas) in the system
= volume of water (or liquid) in the system
Bulk density is related to particle density by:
pb=(\-n)ps eq.C.3.12
Where:
pb = bulk density
n = total porosity
ps = density of grains comprising the aquifer
The bulk density is always less than the particle density, ps; for example, if pores constitute half
the volume, then pb is half of ps. The bulk density of a soil is affected by the structure of the soil
(looseness and degree of compaction), as well as by its swelling and shrinking characteristics, both
of which depend on clay content and soil moisture. Even in extremely compacted soil, the bulk
density remains appreciably lower than the particle density. This is because the particles can never
interlock perfectly, and the soil remains a porous body, never completely impervious. In sandy soils,
pb can be as high as 1.81 gm/cm3. In aggregated loams and clayey soils, pb can be as low as
l.lgm/cm3. Table C.3.2 contains representative values of dry bulk density for common sediments
and rocks.
C.3.1.6.2 Distribution Coefficient and Total Organic Carbon Content
The distribution coefficient is described in Section B.4.3. Recall equation B.4.10, which gives
the relationship between / and K :
r J oc oc
C3-26
-------
Kd=Kocfoc eq.C.3.13
Where:
Kd= distribution coefficient [L3/M]
Ko= soil adsorption coefficient for soil organic carbon content [L3/M]
foc = fraction soil organic carbon (mg organic carbon/mg soil) [M/M]
Representative Koc values are given in Table B.4.1. The fraction of soil organic carbon must be
determined from site-specific data. Representative values of total organic carbon (TOC) in common
sediments are given in Table C.3.3. Because most solute transport occurs in the most transmissive
aquifer zones, it is imperative that soil samples collected for total organic carbon analyses come from
these zones in background areas. To be conservative, the average of all total organic carbon concen-
trations from sediments in the most transmissive aquifer zone should be used for retardation calcula-
tions.
Table C.3.3 Representative Values of Total Organic Carbon for Common Sediments
Texture
medium sand
Fine sand
Fine to coarse sand
organic silt and peat
silty sand
silt with sand, gravel and
clay (glacial till)
medium sand to gravel
oess (silt)
fine - medium sand
fine to medium sand
fine to coarse sand
sand
coarse silt
medium silt
fine silt
silt
fine sand
medium sand to gravel
Depositional Environment
fluvial-deltaic
jack-barrier (marine)
glacial (lacustrine)
glaciofluvial
glacial moraine
glaciofluvial
eolian
glaciofluvial or
glaciolacustrine
glaciofluvial
glaciofluvial
fluvial
Fluvial
fluvial
fluvial
acustrine
glaciofluvial
glaciofluvial
Fraction Organic
Carbon
0.00053 -0.0012
0.0006 - 0.0015
0.00026 - 0.007
0.10 -0.25
0.0007 - 0.008
0.0017 - 0.0019
0.00125
0.00058 - 0.0016
< 0.0006 -0.0061
0.00021 -0.019
0.00029 -0.073
0.0057
0.029
0.020
0.0226
0.0011
0.00023 -0.0012
0.00017 -0.00065
Site Name
Hill AFB, Utah
Boiling AFB, D.C.
Patrick AFB, Florida
Elmendorf AFB, Alaska
Elmendorf AFB, Alaska
Elmendorf AFB, Alaska
Elmendorf AFB, Alaska
Offutt AFB, Nebraska
Iruax Field, Madison
Wisconsin
King Salmon AFB, Fire
Training Area, Alaska
Dover AFB, Delaware
Battle Creek ANGB, Michigan
Oconee River, Georgia3
Oconee River, Georgia3
Oconee River, Georgia3
Oconee River, Georgia3
Wildwood, Ontario
Various sites in Ontario
Various sites in Ontario
a/Karickhoff, 1981
b/ Domenico and Schwartz (1990)
Example C.3.3: Retarded Solute Transport Velocity Calculation
For ground-water flow and solute transport occurring in a shallow, saturated, well-sorted, fine-
grained, sandy aquifer, with a total organic carbon content of 0.7 percent, a hydraulic gradient of -
0.015 m/m, and an hydraulic conductivity of 25 m/day, calculate the retarded contaminant velocity
for trichloroethene.
Solution:
Because the total porosity, effective porosity, and the bulk density are not given, values of
these parameters are obtained from Table C.3.2. The median values for total porosity, effective
C3-27
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porosity, and bulk density are approximately 0.4, 0.2, and 1.6 kg/L, respectively.
The first step is to calculate the average linear ground-water velocity, v .
v =-•
0.2
= 1.9™,
day
The next step is to determine the distribution coefficient, K,. Values of K for chlorinated
r ' d oc
solvents and BTEX are obtained from Tables B.2.1 and B.2.2, respectively, and are listed in
TableC.3.4.
For trichloroethene KOC = 87 L/kg, and (using equation C.3.13):
. .
I kg} kg
The retarded contaminant velocity is given by (equation C.3. 10):
1.
v =-
day
1 +
- = 0.55™,
day
0.4
Table C.3.4 presents the estimated coefficient of retardation contaminant velocity for a number of
contaminants under the conditions of Example C.3.3. This example illustrates that contaminant
sorption to total organic carbon can have a profound influence on contaminant transport by signifi-
cantly slowing the rate of dissolved contaminant migration.
Table C.3.4
Example Retardation Calculations for Select Compounds
Confound
Benzene
Toluene
EthylbenzEne
m-xylene
Tetrachloroethene
Triehloroethene
cis- 1 ,2-Diehloroethene
Vmyl Chloride
1 ,3,5-trirrEthylbenzEne
„
79
190
468
405
209
87
49
2.5
676
Fraction
Organic
Carbon
0.007
0.007
0.007
0.007
0.007
0.007
0.007
0.007
0.007
Distribution
Coefficient
0.553
1.33
3.276
2.835
1.463
0.609
0.343
0.0175
4.732
Bulk
Density
1.60
1.60
1.60
1.60
1.60
1.60
1.60
1.60
1.60
Total
Porosity
0.40
0.40
0.40
0.40
0.40
0.40
0.40
0.40
0.40
Coefficient of
Retardation
3.21
6.32
14.10
12.34
6.85
3.44
2.37
1.07
19.93
Advective
Ground-water
Velocity (rrfday)
1.90
1.90
1.90
1.90
1.90
1.90
1.90
1.90
1.90
Contaninant
Velocity
(irfday)
0.59
0.30
0.13
0.15
0.28
0.55
0.80
1.78
0.10
C.3.2 CONTAMINANT CALCULATIONS
NAPLs present in the subsurface represent a continuing source of ground-water contamination.
N APLs may be made up of one compound, or more likely, a mixture of compounds. Concentrations
of dissolved contaminants and the lifetime of NAPL source areas and associated ground-water
plumes are ultimately determined by the rate at which contaminants dissolve from the NAPL. When
sufficient quantities of NAPL are present, the unsaturated zone may initially be saturated with
C3-28
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NAPL, and the NAPL may migrate under the influence of gravity. After a period of time the NAPL
may drain from the pores under the influence of gravity, leaving a thin coating of NAPL. Depending
on the surface area of the subsurface materials, the surface tension of the NAPL, and the porosity and
permeability of the subsurface materials, some NAPL also may be held between the grains by capil-
larity. NAPL adhering to the grains of the aquifer matrix or retained by capillarity is herein referred
to as residual NAPL. In residual zones, NAPL will be present in immobile blobs or ganglia that may
occupy 10 percent or less of the pore space (Feenstra and Guiguer, 1996). If the NAPL is at satura-
tion and is mobile within and among the pores of the aquifer matrix, the NAPL is referred to as
mobile NAPL. Mobile NAPL may occupy as much as 50 to 70 percent of the pore space and can
reduce flow of water through these zones.
In the unsaturated zone, dissolution from residual or mobile NAPL into downward-migrating
precipitation (recharge) will occur, as well as migration and dissolution of vapors. In the saturated
zone, dissolution of contaminants from residual NAPL occurs as ground-water flows through the
residual zone. Dissolution from mobile NAPL mostly takes place along the tops, bottoms, or lateral
margins of the NAPL bodies, because ground-water (or recharge) flow through the NAPL is re-
stricted. Because the distribution of residual NAPL results in a greater surface area of product in
contact with ground water and does not restrict ground-water velocities, concentrations of contami-
nants entering ground water will typically be closer to the compounds' equilibrium solubilities than
in the case of mobile NAPL bodies. The equilibrium solubility of the compound(s) of interest will
depend on the composition of the NAPL (i.e., the molar fraction of the NAPL represented by the
compound).
In general, residual and mobile NAPL may be present above or below the water table, but direct
dissolution into ground water will only occur when NAPL is at or below the capillary fringe. In
either case, quantifying the flux of contamination entering ground water from above or below the
water table is a difficult proposition. The processes governing dissolution from NAPLs are complex
and depend upon many variables (Feenstra and Guiguer, 1996). Among these variables (in the
saturated zone) are the shape of a mobile NAPL body, the contact area between the NAPL and the
ground water, the velocity of the ground water moving through or past the NAPL, the effect of
residual NAPL on the effective porosity of the contact zone, the solubility of the compounds of
interest, the relative fractions of the compounds in the NAPL, the diffusion coefficients of the com-
pounds, and the effects of other compounds present in the NAPL. This will be further complicated
by any processes in the vadose zone (e.g., volatilization, dissolution from residual NAPL into re-
charge, or dissolution of vapors into recharge) that also will add contaminant mass to ground water.
Further, as the mass of the NAPL body changes over time, the rate of dissolution will also change.
Clearly, given the number of variables that affect the transfer of contaminant mass to ground water, it
is difficult to accurately estimate the flux of contaminants into ground water. Depending on the
intended use of the flux estimate, different approaches can be used.
If one desires to estimate a source term for a contaminant fate and transport model, one can
attempt to estimate the mass loading rate and use that estimate as an input parameter. However, this
often does not yield model concentrations (dissolved) that are similar to observed concentrations. As
a result, the source in the model often becomes a calibration parameter (Mercer and Cohen, 1990;
Spitz and Moreno, 1996). This is because the effects of the source (i.e., the dissolved contaminant
plume) are easier to quantify than the actual flux from the source. The frequent need for such a
"black box" source term has been borne out during modeling associated with evaluations of natural
attenuation of fuel hydrocarbons [following the AFCEE technical protocol (Wiedemeier et
a/.,1995d)] at over 30 U.S. Air Force sites. Use of other methods to calculate source loading for
those models often produced model concentrations that differed from observed concentrations by as
C3-29
-------
much as an order of magnitude. From the model, the flux estimate then can be used for estimating
source lifetimes or other such calculations.
For other purposes, one can estimate flux using several methods, as summarized by Feenstra
and Guiguer (1996). For bodies of mobile LNAPL, this is more practical, because the area of NAPL
in contact with ground water can be estimated from plume/pool dimensions. Where most NAPL is
residual, the surface area can be highly variable, and cannot be measured in the field. Laboratory
studies to understand and quantify mass transfer from residual NAPL in porous media are in the
early stages, and when such mass transfer is modeled, surface area is a calibration parameter with
great uncertainty (Abriola, 1996). Most methods of estimating NAPL dissolution rates require an
estimate of the contact area and, therefore, will contain a great deal of uncertainty. This is one of the
main reasons why, for purposes of modeling, the "black box" source term is more commonly used.
One reason practitioners want to estimate mass transfer rates is to provide a basis for estimating
contaminant source lifetimes, which can affect regulatory decisions and remedial designs. To deter-
mine how long it will take for a dissolved contaminant plume to fully attenuate, it is necessary to
estimate how fast the contaminants are being removed from the NAPL. In general, it is difficult to
estimate cleanup times, so conservative estimates should be made based on NAPL dissolution rates.
Predicting the cleanup time for sites with mobile NAPL is especially difficult because residual
NAPL will remain after the recoverable mobile NAPL has been removed. Of course, this is all
complicated by the many factors that affect dissolution rates as discussed above. Moreover, most
methods do not account for changing dissolution rates as a result of NAPL volume loss (and subse-
quent surface area decrease), preferential partitioning from mixed NAPLs, and the change in porosity
(and, therefore, ground-water velocity) resulting from NAPL dissolution. Finally, the mass of the
NAPL present in the subsurface must also be estimated, lending further uncertainty to any calcula-
tion of source lifetime.
There are several ways to quantify the mass loading rate from a body of mobile or residual
NAPL. Feenstra and Guiguer (1996) present a good summary of some common methods. As noted
above, transfer rates calculated from these methods are all dependent upon several parameters, many
of which cannot be measured or derived from the literature. This is especially true for residual
NAPL. Johnson and Pankow (1992) present a method for estimating dissolution rates from pools of
NAPL which contact ground water over an area that is essentially two-dimensional. Many other
dissolution models may be available; however, as noted before, the experimental evidence to support
dissolution models is really just starting to be collected. Despite these limitations, some of these
models can prove useful, and a selected few are presented (in limited detail) in the following subsec-
tions.
If estimating mass flux rates is less important, one can use direct measurement or equilibrium
concentration calculations to estimate contaminant source area concentrations. The first method
involves directly measuring the concentration of dissolved contaminants in ground water near the
NAPL plume. The second method involves the use of partitioning calculations. These approaches
are described in the following sections. This type of data can be useful if it can be demonstrated that
the source is not capable of introducing concentrations of compounds of concern that exceed regula-
tory limits, or that with slight weathering the same results can be expected. Source area concentra-
tions, whether measured or calculated, also may be used to provide calibration targets for transport
models in which a "black box" source term is used.
If contaminant concentrations in the residual and mobile NAPL are not decreasing over time, or
if they are decreasing very slowly, extremely long times will be required for natural attenuation of the
dissolved contaminant plume. This will likely make natural attenuation less feasible and will reduce
the chance of implementation. In order for natural attenuation to be a viable remedial option, the
C3-30
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source of continuing ground-water contamination must be decreasing over time (decaying), either by
natural weathering processes or via engineered remedial solutions such as mobile NAPL recovery,
soil vapor extraction, bioventing, or bioslurping. Because natural weathering processes can be fairly
slow, especially in systems where the NAPL dissolves slowly or is inhibited from volatilizing or
biodegrading, it will generally be necessary to implement engineered remedial solutions to remove
the NAPL or reduce the total mass of residual and dissolved NAPL.
A discussion of estimating source terms for sites contaminated solely with fuel hydrocarbons is
presented by Wiedemeier et al. (1995a). In general, estimating dissolution rates of individual com-
pounds from fuels is simpler than estimating rates of dissolution from other NAPL mixtures because
there is a great deal of experimental evidence regarding partitioning and equilibrium solubilities of
individual compounds from common fuel mixtures. Methods presented in the following subsections
can use such data to reduce some of the uncertainty in source term calculations.
Typical uses of chlorinated solvents (e.g.., degreasing or parts cleaning) and past disposal
practices that generally mixed different waste solvents or placed many types of waste solvents in
close proximity have resulted in complex and greatly varying NAPL mixtures being released at sites.
For mixtures containing other compounds (e.g., either DNAPLs containing multiple chlorinated
compounds, or fuel LNAPLs containing commingled chlorinated compounds), the equilibrium
solubility of the individual compounds of interest must first be calculated, then that information can
be used in the common mass transfer rate calculations. Except in the case of pure solvent spills,
therefore, the estimation of dissolution rates is then further complicated by this need to estimate
equilibrium solubilities from the mixture.
Because this work focuses largely on saturated-zone processes, vadose zone dissolution pro-
cesses will not be discussed in any detail. However, this discussion will provide a starting point for
estimating source terms for ground-water contaminant fate and transport modeling, as well as for
estimating source and plume lifetimes. As a starting point, two basic methods of estimating or
measuring equilibrium dissolved contaminant concentrations in the vicinity of NAPL bodies are
presented. In addition, methods for estimating fluxes summarized by Feenstra and Guiguer (1996)
and presented by Johnson and Pankow (1992) will be briefly summarized.
C.3.2.1 Direct Measurement of Dissolved Contaminant Concentrations In Ground Water in
Contact with NAPL
Two methods can be used to determine the dissolved concentration of contaminants in ground
water near a NAPL plume. The first method involves collecting ground-water samples from near a
NAPL lens in monitoring wells. The second method involves collecting samples of mixed NAPL
and water from monitoring wells.
C.3.2.1.1 Collecting Ground-water Samples from Near the NAPL
This method involves carefully sampling ground water beneath a floating LNAPL lens or near a
DNAPL lens. One way of collecting a ground-water sample from beneath a lens of floating LNAPL
or above/adjacent to a DNAPL body involves using a peristaltic pump. For LNAPL, the depth to the
base of the mobile LNAPL is measured, a length of high-density polyethylene (HDPE) tubing that
will reach 1 to 2 feet beneath the LNAPL is lowered into the well, and the sample is collected. For
DNAPL, the tube would be cut to reach 1 to 2 feet above the NAPL. Another useful technique for
obtaining such samples where the depth to ground water is too deep to allow use of a peristaltic
pump is to use a Grundfos® pump. If a Grundfos® pump is used to collect a water sample from
beneath LNAPL, it is imperative that the pump be thoroughly cleaned after each use, and that good
sampling logic be used (e.g., sample less contaminated wells first). Also, dedicated bladder pumps
that are being used for long-term monitoring (LTM) in wells with NAPL can be used to collect water
samples from beneath or above the NAPL.
C3-31
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C.3.2.1.2 Collecting Mixed Ground-water/NAPL Samples
This method involves collecting a sample of ground water and NAPL from a monitoring well,
placing the sample in a sealed container used for volatile organics analysis being careful to ensure
there is no headspace, allowing the sample to reach equilibrium, and submitting the water above or
below the floating NAPL to a qualified laboratory for analysis. A disposable bailer generally works
best for collection of this type of sample. Smith et al. (1981) has information on how to conduct
such a test for LNAPL. Two or three samples should be collected from different monitoring wells
containing NAPL at the site. This test should only be done when it is not possible to collect a dis-
crete sample from above or below the NAPL.
C.3.2.2 Equilibrium Partitioning Calculations
The NAPL present at a site represents a continuing source of contamination because chlorinated
solvents, BTEX, and other compounds will partition from the NAPL into the ground water. In such
cases, it is generally necessary to estimate the dissolved concentration of contaminants expected in
ground water near the LNAPL. Partitioning calculations can be performed for sites with NAPL to
quantify contaminant loading from the NAPL into the ground water at the time the ground water or
NAPL samples are collected. Such calculations allow a crude estimation of the impact of continuing
sources of contamination on dissolved contaminant concentrations. The results of partitioning
calculations may show that even if the NAPL is allowed to remain in the ground, dissolved contami-
nant concentrations will remain below regulatory guidelines. This is especially true when weathered
NAPLs with initially low contaminant concentrations are present. Partitioning calculations made by
Wiedemeier el al. (1993) showed that NAPL present in the subsurface at a fueling facility near
Denver, Colorado, was incapable of producing dissolved contaminant concentrations in ground water
above regulatory standards. Such partitioning calculations should be confirmed with an LTM pro-
gram.
On the other hand, if partitioning calculations indicate that continued dissolution will produce
contaminant concentrations exceeding regulatory guidelines, further work will be needed. The
contaminant concentrations calculated by equilibrium methods will clearly not provide mass flux
estimates that can be used in modeling; again, the "black box" methods will be more useful. More-
over, there is no estimation of the actual mass flux across the entire body of NAPL and, therefore,
source lifetimes and weathering rates cannot be estimated directly from partitioning data. More
advanced calculations, such as those that will be discussed in later sections, are then required, keep-
ing in mind that greater uncertainties will be introduced.
When found in the saturated zone, residual NAPL is extremely difficult to remove. Maximum
contaminant concentrations resulting from such partitioning will occur when the ground water and
NAPL reach equilibrium. Assuming that equilibrium is reached gives the most conservative model-
ing results.
C.3.2.2.1 Equilibrium Partitioning of Contaminants from Mobile NAPL into Ground Water
Because most NAPLs will be a mixture of compounds, the solubilities of those compounds will
be lower than the solubility of the individual compound (which is what is most commonly found in
the literature). For an organic NAPL mixture, the dissolved concentration of each compound (in
equilibrium with the mixture) can be approximated by:
Csat,m=XmCsatip eq. C.3.14
Where:
Csat m = solubility of compound from mixture
Xm = mole fraction of compound in the mixture
CSat,P = solubility of pure compound
C3-32
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This equilibrium concentration may also be referred to as the effective solubility of the compound
from the mixture. Experimental evidence (Banerjee, 1984; Broholm and Feenstra, 1995) have
suggested that eq. C.3.14 produces reasonable approximations of effective solubilities for mixtures
of structurally similar compounds, and that the relationship works best for binary mixtures of similar
compounds. For other mixtures, the error is greater due to the complex solubility relationships
created; however, the method is appropriate for many environmental studies for which there are
many other uncertainties (Feenstra and Guiguer, 1996).
For complex mixtures (e.g., multiple identified and unidentified solvents, or mixed fuels and
solvents), it will be necessary to estimate the weight percent and an average molecular weight of the
unidentified fraction of the NAPL before the calculation can be completed. In doing so, it should be
remembered that increasing the average molecular weight for the unidentified fraction will produce
greater estimated effective solubilities for the identified contaminants. A higher molecular weight
for the unidentified fraction will result in a lower mole fraction for that fraction and, therefore,
higher mole fractions (and solubilities) for the known compounds. Feenstra and Guiguer (1996)
provide an example of these calculations for a mixture of chlorinated and nonchlorinated com-
pounds.
In the case of fuel hydrocarbon mixtures, experimental partitioning data has been collected and
used to develop individual-compound solubility calculations, largely because fuel mixtures are
somewhat consistent in their makeup. The fuel-water partitioning coefficient, K is defined as the
ratio of the concentration of a compound in the fuel to the compound's equilibrium concentration in
water in contact with the fuel:
_CL
Kfw -— eq. C.3.15
w
Where:
K = fuel-water partitioning coefficient [dimensionless]
C = concentration of the compound in the fuel [M/L3]
Cw = concentration of the compound dissolved in ground water [M/L3]
A summary of values of K for BTEX and trimethylbenzenes (TMB) in jet fuel and gasoline are
presented by Wiedemeier et al. (1995d), along with the relationships relating Kfw to the aqueous
solubility of a pure compound in pure water, S, which can be used to estimate K for compounds for
which there is no experimental data.
Using the definition of.K presented above, the maximum (equilibrium) total dissolved BTEX
concentration resulting from the partitioning of BTEX from NAPL into ground water is given by:
---
uw r eq. C.3.16
Kfa>
This relationship predicts the concentration of dissolved BTEX in the ground water if the LNAPL is
allowed to remain in contact with the ground water long enough so that equilibrium between the two
phases is reached. Further discussion and example calculations for this method are presented by
Wiedemeier et al. (1995d).
C.3.2.3 Mass Flux Calculations
In general, the rate of mass transfer from a NAPL can be given as the product of a mass transfer
coefficient, a concentration difference, and a contact area. As Feenstra and Guiguer (1996) note, the
driving force for mass transfer is the concentration difference across a boundary layer between the
NAPL and the ground water. The concentration difference can be approximated using the effective
solubility of a compound (eq. C.3.14) and either the measured concentration of the compound in
ground water adjacent to the NAPL, or a calculated (theoretical) ground-water concentration. How-
C3-33
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ever, the contact area and the mass transfer coefficient incorporate a great deal of uncertainty and are
typically calibration parameters for modeling dissolution, as discussed previously.
Once these parameters have been estimated, one can use them in a variety of models. In gen-
eral, models for dissolution of NAPL in porous media either assume local equilibrium between
phases, or assume that dissolution is a first-order process governed by the variables discussed above
(Feenstra and Guiguer, 1996). Abriola and Finder (1985a), Baehr and Corapcioglu (1987), and
Kaluarachchi and Parker (1990) developed two-dimensional NAPL migration models that account
for dissolution using the local equilibrium assumption (LEA). As noted by Abriola (1996), these
studies generally were computer modeling studies for which follow-up laboratory work is ongoing
and uncovering additional factors to consider. For single-component NAPLs, models utilizing a
first-order reaction have been developed by Miller et al. (1990), Powers et al. (1992), Brusseau
(1992), Guiguer (1993), and Guiguer and Frind (1994). For multi-component NAPLs, a model
developed by Shiuetal. (1988) andMackay et al. (1991) may be of use.
Due to approximate nature of flux calculations and the inherent uncertainty in those calcula-
tions, we have chosen to omit a detailed discussion of such efforts. The numerical modeling using
LEA methods is beyond the scope of this work, and may not be practical for use at most sites. In-
stead, we will present a brief review of ideas presented by Feenstra and Guiguer (1996) and Johnson
and Pankow (1992) in order to illustrate some of the concepts involved in estimating flux terms.
Should further detail or other methods be desired, both of those works provide excellent background
and references to start with, including many of the works referenced in this discussion of source term
calculations.
C.3.2.3.1 General Mass Transfer Models
Using concepts from the field of chemical engineering, Feenstra and Guiguer (1996) note that
for a single-component NAPL, simple dissolution of the compound may be described by:
N=Kc(Cw-Csat) eq.C.3.17
Where:
N = flux of the species of interest (M/L2T)
Kc = mass transfer coefficient (L/T)
Cw = concentration of compound in bulk aqueous phase (M/L3)
Cmt = concentration of compound at NAPL-water interface (taken as the solubility of the
compound) (M/L3)
The mass transfer coefficient may be calculated various ways, but in all cases, the diffusivity of the
species of interest is a factor. Feenstra and Guiguer (1996) present three methods for determining a
mass transfer coefficient.
In a porous media, the mass transfer rate per volume of porous medium can be defined by
multiplying the mass flux by the ratio of NAPL surface contact area to the unit volume of porous
medium, yielding:
JV=A(CW-CJ eq.C.3.18
Where:
7\r* = flux of the species of interest per unit volume of porous medium (M/L2T)
k = lumped mass transfer coefficient (L/T)
= concentration of compound in bulk aqueous phase (M/L3)
t = concentration of compound at NAPL-water interface (taken as the solubility of the
compound) (M/L3)
The lumped mass transfer coefficient is the product of Kc and the ratio of the NAPL surface contact
C3-34
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area and the unit volume of the porous media. This can further be extended for multicomponent
NAPLs:
eq.C.3.19
Where:
N*m = flux of component m per unit volume of porous medium (M/L2T)
Am = lumped mass transfer coefficient for component m (L/T)
Cw m = concentration of component m in bulk aqueous phase (M/L3)
Csat m = concentration of component m at N APL-water interface (calculated using eq. C.3 . 1 4)
(M/L3)
Further complicating all of these relationships is the fact that as dissolution continues, Am will vary
over time as the amount of NAPL changes. This can be accounted by using the following first-order
relation:
eq.C.3.20
Where:
Nm = flux of component m per unit volume of porous medium (M/L2T)
Sw = average fraction of pore volume occupied by water
Am = lumped mass transfer coefficient for component m (L/T)
Cwm = concentration of component m in bulk aqueous phase (M/L3)
Csat m = concentration of component tn at NAPL-water interface (calculated using eq . C . 3 . 1 4))
(M/L3)
Again, it bears repeating that on the field scale, measurement of many of the parameters used
for these calculations is not possible, and, therefore, great uncertainty is introduced. Source terms
calculated using these or any other methods should be presented in that light, and if used for solute
transport modeling, should be accompanied with a sensitivity analysis.
C.3. 2.3. 2 Nonequilibrium Partitioning Model of Johnson and Pankow (1992)
The steady-state, two-dimensional dissolution of contaminants from a pool of NAPL floating on
the water table into ground water (assumed to be a semi-infinite medium) can be described by the
steady-state, two-dimensional, advection-dispersion equation (Hunt etal, 1988):
~\j~i -\2/^t
vx— = Dz—T x,z>0 eq. C.3.21
dx dz
Where:
C = contaminant concentration dissolved in water
vx = average linear ground-water velocity
Dz= vertical dispersion coefficient
If it is assumed that:
• The time required for total NAPL dissolution is exceedingly long in comparison to the
contact time between the NAPL pool and the flowing ground water
• The NAPL pool is wide compared to the horizontal transverse mixing process
• The NAPL pool can be approximated as a rectangle
• The NAPL lens width does not affect the dissolution rate
• The elevation of the NAPL lens is taken as z=0, with z measured positively upward
• The boundary conditions are:
C3-35
-------
\_/^X.,j Z ) U
C(x, z = 0) = Ce 0 < x > L
Where:
C = contaminant concentration dissolved in water
Ce = effective water solubility
L = horizontal length of NAPL pool
then the rate of dissolution of constituents from an LNAPL lens into ground water flowing beneath
the lens can be calculated as two-dimensional, steady-state dissolution, and the surface area averaged
mass transfer rate, Mq, is calculated as (Johnson and Pankow, 1992; Hunt el a/., 1988):
[4^T
C.3.22
\ n L
Where:
ne = effective porosity
L = length of NAPL lens parallel to ground-water flow direction
vx = average linear ground-water flow velocity
Ce = effective water solubility (proportional to a compound's pure phase solubility and mole
fraction in the NAPL)
D^= vertical dispersion coefficient
The vertical dispersion coefficient, Dz, results from a combination of molecular diffusion and me-
chanical dispersion and is defined as (Johnson and Pankow, 1992):
Dz=De+vxaz eq. C.3.23
Where:
De = effective molecular diffusivity (corrected for porosity and tortuosity)
ar = vertical dispersivity (typically 0.01 of longitudinal dispersivity)
vx = average linear ground-water flow velocity
A typical value of De for a nonpolar organic compound is 1 x 10~5 cm2/sec (Sellers and Schreiber,
1992).
"At very low flow velocities where molecular diffusion dominates, the average concentration
decreases with increasing flow velocity because of decreasing contact time. At higher groundwater
flow velocities where dispersion dominates over diffusion, average percent solubility becomes
independent of velocity. This is because the transverse dispersion coefficient is proportional to flow
velocity, and Dz/v is constant. At typical groundwater flow velocities, an effluent concentration far
less than the solubility limit is expected. For example, for a flow velocity of 1 m/day and a=W~4 m,
less than 1 percent of solubility is predicted, and considerable pumping would be required to remove
the contaminant. The analysis predicts a constant contaminant concentration dissolved in the ex-
tracted water as long as the separate phase covers the boundary" (Hunt elal, 1988, pp. 1253 and
1254).
C3-36
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C.3.3 CONFIRMING AND QUANTIFYING BIODEGRADATION
Chemical evidence of two types can be used to document the occurrence of biodegradation.
The first type of evidence is graphical and is provided by the electron acceptor and metabolic
byproduct maps discussed in Section C.2. The second line of evidence involves using a conservative
tracer.
C.3.3.1 Isopleth Maps
The extent and distribution of contamination relative to electron acceptors and metabolic
byproducts can be used to qualitatively document the occurrence of biodegradation. Depleted
dissolved oxygen concentrations in areas with fuel hydrocarbon contamination indicates that an
active zone of aerobic hydrocarbon biodegradation is present. Depleted nitrate and sulfate concen-
trations in areas with fuel hydrocarbon contamination indicate that an active zone of anaerobic
hydrocarbon biodegradation is present and that denitrification and sulfate reduction are occurring.
Elevated iron (II) and methane concentrations in areas with fuel hydrocarbon contamination indicate
that an active zone of anaerobic hydrocarbon biodegradation is present and that iron reduction and
methanogenesis are occurring. Isopleth maps of contaminants, electron acceptors, and metabolic
byproducts can be used as evidence that biodegradation of fuel hydrocarbons is occurring.
Figures C.2.7 and C.2.8 show how these maps can be used to support the occurrence of biodegrada-
tion. Figure C.2.7 shows that areas with depleted dissolved oxygen, nitrate, and sulfate correspond
with areas having elevated BTEX concentrations. Figure C.2.8 shows that areas with elevated
iron (II) and elevated methane concentrations also coincide with areas having elevated BTEX con-
centrations. These figures suggest that aerobic respiration, denitrification, iron reduction, sulfate
reduction, and methanogenesis are all occurring at the example site.
C.3.3.2 Data Set Normalization
In order to calculate biodegradation rates accurately, measured contaminant concentrations must
be normalized for the effects of dispersion, dilution, and sorption. A convenient way to do this is to
use compounds or elements associated with the contaminant plume that are relatively unaffected or
predictably affected by biologic processes occurring within the aquifer. At sites where commingled
fuel hydrocarbon and chlorinated solvent plumes are present, the trimethylbenzene isomers (TMB),
which can be biologically recalcitrant under some geochemical conditions have proven useful when
estimating biodegradation rates for BTEX and chlorinated solvents. At sites where TMB data are
not available, the chloride produced as a result of biodegradation or the carbon nucleus of the chlori-
nated compound can be used as a tracer.
Measured concentrations of tracer and contaminant from a minimum of two points along a flow
path can be used to estimate the amount of contaminant that would be expected to remain at each
point if biodegradation were the only attenuation process operating to reduce contaminant concentra-
tions. The fraction of contaminant remaining as a result of all attenuation processes can be com-
puted from the measured contaminant concentrations at two adjacent points. The fraction of con-
taminant that would be expected to remain if dilution and dispersion were the only mechanisms for
attenuation can be estimated from the tracer concentrations at the same two points. The tracer is
affected by dilution and dispersion to the same degree as the contaminant of interest and is not
affected by biologic processes. The following equation uses these assumptions to solve for the
expected downgradient contaminant concentration if biodegradation had been the only attenuation
process operating between two points along the flow path:
CBiCorr=CB\ — \ eq. C.3.24
C3-37
-------
Where:
CB = corrected contaminant concentration at a point B downgradient
fi.corr r °
CB = measured contaminant concentration at point B
TA = tracer concentration at a point A upgradient
TB = tracer concentration at point B downgradient
This equation can be used to estimate the theoretical contaminant concentration that would result
from biodegradation alone for every point along a flow path on the basis of the measured contami-
nant concentration at the origin and the dilution of the tracer along the flow path. This series of
normalized concentrations can then be used to estimate a first-order rate of biodegradation as de-
scribed in Section C.3.3.3.
C.3.3.2.1 Normalization Using Organic Compounds as Tracers
A convenient way of estimating biodegradation rate constants is to use compounds present in
the dissolved contaminant plume that that are biologically recalcitrant. One such compound that is
useful in some, but not all, ground-water environments is Trimethylbenzene (TMB). The three
isomers of this compound (/,2,3-TMB, 1,2,4-TMB, and /,3,5-TMB) are generally present in suffi-
cient quantities in fuel mixtures to be readily detectable when dissolved in ground water. When
chlorinated solvents enter the subsurface as a mixture with petroleum hydrocarbons, the TMB
compounds can be useful tracers. The TMB isomers are fairly recalcitrant to biodegradation under
anaerobic conditions; however, the TMB isomers do not make good tracers under aerobic conditions
(because they are readily biodegraded in aerobic environments). The degree of recalcitrance of TMB
is site-specific, and the use of this compound as a tracer must be evaluated on a case-by-case basis.
Nevertheless, if any TMB mass is lost to biodegradation, equation C.3.24 will be conservative
because the calculated mass losses and the attenuation rate constants calculated on the basis of those
losses will be lower than the actual losses and attenuation rates. Another compound of potential use
as a conservative tracer is tetramethylbenzene; however, detectable dissolved tetramethylbenzene
concentrations are generally less common than detectable dissolved TMB concentrations.
An ideal tracer would have Henry's Law and soil sorption coefficients identical to the contami-
nant of interest; however, TMB is more hydrophobic than BTEX, chlorinated ethenes, and chlori-
nated ethanes, resulting in a higher soil sorption coefficient than the compound of interest. As a
result, use of TMB as a tracer is often conservative, and the biodegradation rates can be underesti-
mated. It is best, whenever possible, to compare several tracers to determine whether they are
internally consistent.
C.3.3.2.2 Normalization Using Inorganics as Tracers
Inorganic compounds also can serve as tracers for the contaminant of interest as long as their
presence is in some way associated (either directly or indirectly) with the dissolved contaminant
plume. For many chlorinated solvent plumes, the sum of ionic chloride and organic chloride associ-
ated with the solvents can be considered a conservative tracer. Note that the following discussion
assumes that the background chloride concentration is negligible in comparison to the source area
concentration of total chloride plus chlorine. If background chloride is more than approximately 10
percent of the total source area chloride plus chlorine concentration, then background concentrations
will need to be accounted for prior to performing the tracer normalization.
Total chlorine can easily be calculated by multiplying the measured concentration of a chlori-
nated organic compound by the mass fraction of chlorine in the molecule, then summing that quan-
tity for all the chlorinated organic compounds represented in the plume. The stoichiometry for
chlorinated ethenes is presented in the following paragraphs.
C3-38
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As PCE is reduced to ethene, 4 moles of chloride are produced:
C,C1 ,-»C,H, + 4C1-
2 4 L 4
On a mass basis, the ratio of chloride produced to PCE degraded is given by:
Molecular weights: PCE 2(12.011)+ 4(35.453)= 165.83 gm
Chloride 4(35.453)= 141.81 gm
M ass Rati oof Chloride to PC E= 141.81:165.83 =0.86:1
Similarly, as TCE is reduced to ethene, 3 moles of chloride are produced:
C2C13H-»C,H4 + 3C1-
On a mass basis, the ratio of chloride produced to TCE degraded is given by:
Molecular weights: TCE 2(1.2.011) + 3(35.453) + 1(1.01)= 131.39 gm
Chloride 3(35.453) = 106.36 gm
Mass Ratio of Chloride to TCE =106.36:131.39 = 0.81:1
Likewise, as DCE is reduced to ethene, 2 moles of chloride are produced:
C2C12H2^C2H4 + 2C1-
On a mass basis, the ratio of chloride produced to DCE degraded is given by:
Molecular weights: DCE 2(12.011) +2(35.453)+ 2(1.01)= 96.95 gm
Chloride 2(35.453) = 70.9 gm
Mass Ratio of Chloride to DCE = 70.9:96.95 = 0.73:1
As VC is reduced to ethene, 1 mole of chloride is produced:
Cl-
On a mass basis, the ratio of chloride produced to VC degraded is given by:
Molecular weights: VC 2(12.011)+ 1(35.453) + 3(1.01)= 62.51 gm
Chloride 1(35.453) = 35.453 gm
Mass Ratio of Chloride to VC = 35.453:62.51 = 0.57:1
Therefore, the amount of total chloride plus chlorine for a spill undergoing reductive dechlorination
would be estimated as:
[ClTota]] = 0.86[PCE] + 0.81 [TCE] + 0.73[DCE]) + 0.57[VC]) eq. C.3.25
Example C.3.4: Calculating Total Concentration of Chloride and Organic Chlorine
The approach is illustrated in the following data set from the West TCE Plume at the St. Joseph,
Michigan NPL site.
A series of discrete vertical water samples were taken in transects that extended across the
plume at locations downgradient of the source of TCE. The locations of the samples are depicted in
Figure C.3.5 as circles. At each sampling location, water samples were acquired using a hollow-
stem auger. The leading auger was slotted over a five-foot interval. After a sample was collected,
the auger was driven five feet further into the aquifer and the next sample was collected. At any one
location, the water samples were collected in a sequential and contiunuous series that extended from
the water table to a clay layer at the bottom of the aquifer. The concentrations of contaminants at
each location were averaged in water samples that extend across the entire vertical extent of the
plume. The location with the highest average concentration of chlorinated ethenes in a particular
transect was selected to represent the centerline of the plume. The locations of the sample locations
in the centerline of the plume are depicted in Figure C.3.5 as open circles. Each centerline location
is labelled in an oval.
C3-39
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Lake Michigan
N
598
Source
Figure C.3.4 Location of sampling points at the St. Joseph, Michigan, NPL site.
The concentrations of chlorinated ethenes and chloride in the centerline of the TCE plume at St.
Joseph, Michigan, are presented in Table C.3.5.
Table C.3.5 Attenuation of Chlorinated Elhenes and Chloride Downgradienl of the Source of TCE in the
West Plume at the St. Joseph, Michigan, NPL Site.
Compound
PCE
TCE
Total DCE
Vinyl
Chloride
Total
Organic
Chloride
Chloride
Tracer (Total
Chloride plus
Chlorine)
Sampling Locations
T-2-5 | T-l-4
T-4-2
T-5-3
55AE
Distance Downaradient ffeef)
0
200 1 1.000
1.500
0.0
12.1
37.6
2.3
38.5
89.7
128.2
0.0
3.4
11.7
3.7
13.4
78.6
92.0
0.0
1.3
2.4
0.51
3.2
98.9
102.1
0.0
0.035
0.23
0.063
0.2
63.6
63.8
2.000
0.0
0.022
0.45
0.070
0.4
54.7
55.1
C3-40
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At the monitoring point closest to the source of the plume (see location T-2-5 in Table C.3.5 and
Figure C.3.4) the concentrations of TCE, total DCE, vinyl chloride and chloride were 12.1, 37.6, 2.3
and 89.7 mg/L, respectively. This results in an upgradient tracer concentration of
TCE chlorine +
DCE chlorine +
Vinyl chloride chlorine +
Chloride +
Total chloride plus chlorine =
At the downgradient location 55AE, which is 2,000 feet from the source, the concentrations of
TCE, total DCE, vinyl chloride, and chloride were 0.022, 0.45, 0.070, and 54.70 mg/L, respectively.
This results in a downgradient concentration of
TCE chlorine +
DCE chlorine +
Vinyl chloride chlorine +
Chloride +
Total chloride plus chlorine =
The computed series of total chloride plus chlorine concentrations can be used with equation
C.3.24 to estimate a normalized data set for contaminant concentrations.
(0.809)02.1 mg/L)
(0.731)(37.6 mg/L)
(0.567)(2.3 mg/L)
89.7 mg/L
128.2 mg/L
(0.809)(0.022mg/L)
(0.731)(0.45 mg/L)
(0.567)(0.070mg/L)
54.7 mg/L
55.1 mg/L
Example C.3.5: Normalizing Contaminant Concentrations Along a Flow Path
Equation 3.24 will be used to calculate a normalized concentration for TCE at the locations
depicted in Figure C.3.4 and Table C.3.5. Given are the observed concentrations of TCE and tracer
(Table C.3.5) for five points that form a line parallel to the direction of ground-water flow (Figure
C.3.4) To calculate normalized concentrations of TCE using the attenuation of the tracer, the dilu-
tion of the tracer is caculated at each location by dividing the concentration of tracer at the source (or
most contaminated location) by the concentration of tracer at each downgradient location. Then the
measured concentration of TCE downgradient is multiplied by the dilution of the tracer. The cor-
rected concentrations of TCE are presented in Table C.3.6. This information will be used in sections
C.3.3.3 to calculate the rate of natural biodegradation of TCE.
Table C.3.6. Use of the Attenuation of a Tracer to Correct the Concentration of TCE Downgradient of the
Source of TCE in the West Plume at the St. Joseph, Michigan, NPL Site
Compound
TCE
Tracer
Dilution of Tracer
Corrected TCE
Sampling Locations
T-2-5
T-l-4
T-4-2
Distance Down Gradient
0
200
1.000
12.1
128.2
128.2/28.2
12.1
3.4
92.0
128.2/92.0
4.7
T-5-3
(feet)
1.500
1.3
102.1
128.27 102
1
1.6
55AE
2.000
0.035
63.8
128.2/63.8
0.070
0.022
55.1
128.2/55.1
0.051
C.3.3.3 Calculating Biodegradation Rates
Several methods, including first- and second-order approximations, may be used to estimate the
rate of biodegradation of chlorinated compounds when they are being used to oxidize other organic
compounds. Use of the first-order approximation can be appropriate to estimate biodegradation rates
C3-41
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for chlorinated compounds when the rate of biodegradation is controlled solely by the concentration
of the contaminant. However, the use of a first-order approximation may not be appropriate when
more than one substrate is limiting microbial degradation rates or when microbial mass is increasing
or decreasing. In such cases, a second- or higher-order approximation may provide a better estimate
of biodegradation rates.
C.3.3.3.1 First-Order Decay
As with a large number of processes, the change in a solute's concentration in ground water
over time often can be described using a first-order rate constant. A first-order approximation, if
appropriate, has the advantage of being easy to calculate and simplifies fate and transport modeling
of complex phenomenon. In one dimension, first-order decay is described by the following ordinary
differential equation:
dC ,
—kt eq. C.3.26
Where:
C = concentration at time t [M/L3]
k = overall attenuation rate (first-order rate constant) [1/T]
Solving this differential equation yields:
C = C0e~h eq. C.3.27
The overall attenuation rate groups all processes acting to reduce contaminant concentrations and
includes advection, dispersion, dilution from recharge, sorption, and biodegradation. To determine
the portion of the overall attenuation that can be attributed to biodegradation, these effects must be
accounted for, and subtracted from the total attenuation rate.
Aronson and Howard (1997) have compiled a large number of attenuation rate constants for
biodegradation of organic compounds in aquifers. This information is supplied to provide a basis for
comparison of rate constants determined for at a particular site to the general experience with natural
attenuation as documented in the literature. It is not intended to provide rate constants for a site in a
risk assessment or exposure assessment. The rate constants used to describe behavior of a particular
site must be extracted from site characterization information particular to that site.
The distribution of the rate constants reported for TCE is presented in Figure C.3.5. Notice that
the average rate is near 1.0 per year, and that most of the rates cluster in a relatively narrow range
between 3.0 per year and 0.3 per year. Some of the published rates are very low, less than 0.1 per
year. The report compiles data from sites where rates are published. The general bias against pub-
lishing negative data suggests that there are many plumes where TCE attenuation was not detectable
(Type 3 behavior), and that data on these plumes is not found in the literature. The data from
Aronson and Howard (1997) reflect the behavior of plumes where reductive dechlorination is an
important mechanism (Type 1 and Type 2 sites). Rate constants for PCE and Vinyl Chloride are
presented in Figures C.3.6 and C.3.7. The average rate for dechlorination of PCE is somewhat faster
than for TCE, near 4.0 per year, and the rate for Vinyl Chloride is slower, near 0.6 per year.
Two methods for determining first-order biodegradation rates at the field scale are presented.
The first method involves the use of a normalized data set to compute a decay rate. The second
method was derived by Buscheck and Alcantar (1995) and is valid for steady-state plumes.
Wiedemeier et al. (1996b) compare the use of these two methods with respect to BTEX biodegrada-
tion.
C3-42
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100
10
g
o
U
CD
0.01
Figure C.3.5. Field rate constants for TCE as reported in literature.
Field Rate Constants for PCE as Reported in
Literature
1 1
1000
100
0.'
0.01
(0
o
o
ra
Figure C.3.6 Field rate constants for PCE as reported in literature.
C3-43
-------
100
10
§
^—>
CO
a
o
U
0)
0.1
Figure C.3.7 Field rate constants for vinyl chloride as reported in literature.
C.3.3.3.2 Use of a Normalized Data Set
In order to ensure that observed decreases in contaminant concentrations can be attributed to
biodegradation, measured contaminant concentrations must be corrected for the effects of advection,
dispersion, dilution from recharge, and sorption, as described in Section C.3.3.2 using equation
C.3.24. The corrected concentration of a compound is the concentration that would be expected at
one point (B) located downgradient from another point (A) if the processes of dispersion and dilution
had not been occurring between points A and B.
The biodegradation rate can be estimated between any two points (A and B) of a normalized
data set (where point A is upgradient of point B) by substituting the concentration at point A for C0,
and the normalized concentration at point B, CB coir, for C in equation C.3.27. The resulting relation-
ship is expressed as:
CB.eorr = CAe-*' eq.C.3.28
Where:
C
I
t
normalized contaminant concentration at downgradient point B (from eq. C.3.25)
contaminant concentration at upgradient point A that if point A is the first point in
the normalized data set, then C = C,
^ A A,corr
first-order biological decay rate (first-order rate constant) [1/T]
time of contaminant travel between points A and B
The rate constant in this equation is no longer the total attenuation rate, k, but is the biological
decay rate, A, because the effects of advection, dispersion, dilution from recharge, and sorption have
been removed (Section C.3.3.2). This relationship can be used to calculate the first-order biological
decay rate constant between two points by solving equation C.3.28 for A:
C3-44
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eq. C.3.29
The travel time, t, between two points is given by:
_ x
t = ~ eq. C.3.30
Where:
x = distance between two points [L]
vc = retarded solute velocity [L/T]
Example C.3.6: First-Order Decay Rate Constant Calculation Using Normalized Data Set
Equation C.3.30 and C.3.29 can be used to calculate rate constants between any two points
along a flow line. For travel from locations T-2-5 and and 55AE in Figure C.3.4 and Table C.3.6,
the upgradient concentration of TCE is 12.1 mg/1, the corrected downgradient concentration is 0.051
tng/1, and the distance between the locations is 2,000 feet.
From Figure C.3.4, the water table drops 10 feet as the plume moves 1,300 feet from transect 1
to transect 5. The site has a hydraulic gradient of 0.008 feet per foot. Aquifer testing at the site
predicts an average hydraulic conductivity of 50 feet per day. If the effecive porosity of the sandy
aquifer is assumed to be 0.3, the seepage velocity (V ) would be (Equation C.3.6):
0.4/^x0 =
0.3
The average organic matter content of the aquifer matrix material is less than the detection limit
of 0.001 g/g. We will assume the organic matter content is equal to the detection limit. If the Koc of
TCE is 120 ml/g, the porosity is 0.3, and the bulk density is 1.7 gm/cm3, the distribution of TCE
between ground water and aquifer solids is the product of the Koc, the fraction organic carbon, the
bulk density, divided by the porosity, or 0.3. The retarded velocity of TCE compared to water (R)
would be (Equation C.3.8 and Equation C.3.13):
R = 1 + 120 (ml/g) * 0.001(g/g) * 1.7 (g/cm3)/ 0.3 (ml/ ml) = 1.7
The velocity of TCE in the aquifer would be equal to the velocity of water in the plume divided by
the retardation of TCE. The TCE velocity (vr) would be:
vc = 1.3 feet per day/1.7 = 0.8 feet per day
If the distance between the wells is 2,000 feet, and the retarded velocity of TCE is 0.8 feet per
day, by equation C.3.30 the travel time is:
t = 2,000 feet/ 0.8 feet per day = 2,500 days = 6.8 years
From equation C.3.29, the rate of biotransformation between locations T-2-5 and 55AE is:
A, = In (0.055/12.1)1 6.8 per year = 0.79 per year
If a number of sampling locations are available along a flow path, all the locations should be
included in the calculation of the biotransformation rate. The simplest way to determine the first-
order rate constant from an entire set of normalized data is to make a log-linear plot of normalized
contaminant concentrations versus travel time. If the data plot along a straight line, the relationship
is first-order and an exponential regression analysis can be performed.
The exponential regression analysis gives the equation of the line of best fit for the data being
regressed from a log-linear plot and has the general form:
C3-45
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= ben
eq. C.3.31
Where:
y =
b =
m =
x =
y axis value
y intercept
slope of regression line
x-axis value
When using normalized data, x is the contaminant travel time to the downgradient locations and m is
the first-order rate of change equal to the negative. The correlation coefficient, R2, is a measure of
how well the regression relationship approximates the data. Values of R2 can range from 0 to 1; the
closer R2 is tol, the more accurate the equation describing the trend in the data. Values of R2 greater
than 0.80 are generally considered useful; R2 values greater than 0.90 are considered excellent.
Several commonly available spreadsheets can be used to facilitate the exponential regression analy-
sis. The following example illustrates the use of this technique.
Figure C.3.8 depicts a regression of normalized TCE concentration against travel time
downgradient. The slope of the exponential regression is —0.824jc where x is travel time in years,
corresponding to a first-order rate of change of-0.824 per year and a first-order rate of biodegrada-
tion of 0.824 per year. In Figure C.3.8, an exponential regression was performed on the normalized
concentrations of TCE against time of travel along the flow path. An alternative approach would be
to perform a linear regression of the natural logarithm of the normalized concentration of TCE
against travel time along the flow path.
100
Travel Distance (feet)
500 1000 1500
2000
-0.824x
10 -'
bfl
u
H
0.1
0.01
0
y=11.332e
Rz = 0.9332
4 6
Travel Time (years)
8
Figure C.3.8 Exponential regression of TCE concentration on time of travel along flow path.
C3-46
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C.3.3.3.3. Method of Buscheck and Aicantar (1995)
Buscheck and Aicantar (1995) derive a relationship that allows calculation of first-order decay
rate constants for steady-state plumes. This method involves coupling the regression of contaminant
concentration (plotted on a logarithmic scale) versus distance downgradient (plotted on a linear
scale) to an analytical solution for one-dimensional, steady-state, contaminant transport that includes
advection, dispersion, sorption, and biodegradation. For a steady-state plume, the first-order decay
rate is given by (Buscheck and Aicantar, 1995):
eq. C.3.32
— * yj_ v-xyj y
Where:
A = first-order biological rate constant
vc = retarded contaminant velocity in the x-direction
ax = dispersivity
k/vx = slope of line formed by making a In-linear plot of contaminant concentration versus
distance downgradient along flow path
Example C.3.7: First-Order Rate Constant Calculation Using Method of Buscheck and Aicantar
(1995)
The first step is to confirm that the contaminant plume has reached a steady-state configuration.
This is done by analyzing historical data to make sure that the plume is no longer migrating
downgradient and that contaminant concentrations are not changing significantly through time. This
is generally the case for older spills where the source has not been removed. The next step is to
make a plot of the natural logarithm of contaminant concentration versus distance downgradient (see
Figure C.3.9). Using linear regression, y in the regression analysis is the contaminant concentration,
x is the distance downgradient from the source, and the slope of the In-linear regression is the ratio k/
VK that is entered into equation C.3.32.
The slope is -0.0028 feet. As calculated above, the retarded TCE velocity in the plume vc is 0.8
feet per day. If a. = 5% of the plume length, then ax = 100 feet. Inserting these values for a., k/vx,
and v, into equation C.3.32, the estimated value of A = -0.0016 per day or -0.59 per year.
C. 3.3.2.2.3 Comparison of First-Order Methods
If the data are available, concentrations of tracers should be used to normalize concentrations of
contaminants prior to calculation of rate constants. If tracer data is not available, the method of
Buscheck and Aicantar (1995) can be used if a value for longitudinal dispersion is available, or if
one is willing to assume a value for longitudinal dispersion. Whenever possible, more than one
tracer should be used to normalize the concentrations of contaminants. If the normalized concentra-
tions agree using several different tracers, the approach can be accepted with confidence. In addition
to chloride and trimethylbenzene, methane, and total organic carbon dissolved in ground water are
often useful tracers in plumes of chlorinated solvents undergoing natural attenuation.
C3-47
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100
10 J
D)
E,
LJJ
o
0.1 -
0.01
0 500 1000 1500 2000 2500
Distance Downgradient (feet)
Figure C.3.9 Regression of (he TCE concentration on distance along flow path.
C3-48
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C.3.4 DESIGN, IMPLEMENTATION, AND INTERPRETATION OF
STUDIES
C.3.4.1 Overview
If properly designed, implemented, and interpreted, microcosm studies can provide very con-
vincing documentation of the occurrence of intrinsic bioremediation. They are the only "line of
evidence" that allows an unequivocal mass balance on the biodegradation of environmental contami-
nants. If the microcosm study is properly designed, it will be easy for decision makers with non-
technical backgrounds to interpret. The results of a microcosm study are strongly influenced by the
nature of the geological material submitted to study, by the physical properties of the microcosm, by
the sampling strategy, and the duration of the study. In addition, microcosm studies are time con-
suming and expensive. A microcosm study should only be undertaken at sites where there is consid-
erable uncertainty concerning the biodegradation of contaminants based on soil and ground-water
samples alone.
Material for a microcosm study should not be selected until the geochemical behavior of the site
is well understood. Contaminant plumes may consume oxygen, nitrate, or sulfate, and produce iron
(II), manganese (II), or methane. These processes usually operate concurrently in different parts of
the plume. Regions where each process prevails may be separated in directions parallel to ground-
water flow by hundreds of meters, in directions perpendicular to ground-water flow by tens of
meters, and vertically by only a few meters. Rate constants and constraints for petroleum hydrocar-
bon biodegradation will be influenced by the prevailing geochemistry. Material from microcosms
must be acquired for depth intervals and locations that have been predetermined to be representative
of the prevailing geochemical milieu in the plume.
Contaminant biodegradation supported by oxygen and nitrate cannot be adequately represented
in microcosm. In the field, organisms that use oxygen or nitrate proliferate until they become limited
by the supply of electron acceptor. After that time, the rate of hydrocarbon degradation is controlled
by the supply of electron acceptor through diffusion or hydrodynamic dispersion. Microcosms have
been used successfully to simulate sulfate-reducing, iron-reducing, and methanogenic regions of
plumes. Oxygen is toxic to sul fate-reducing and methanogenic microorganisms. Material should be
collected and secured in a manner that precludes oxygenation of the sample.
Batch microcosms that are sacrificed for each analysis usually give more interpretable results
than column microcosms or batch microcosms that are sampled repetitively. For statistical reasons,
at least three microcosms should be sampled at each time interval. If one assumes a first-order rate
law, and no lag, a geometrical time interval for sampling should be the most efficient. An example
would be sampling after 0 weeks, 2 weeks, 1 month, 2 months, 4 months, and 8 months. As a
practical matter, long lags frequently occur, and the rate of bioremediation after the lag is rapid. A
simple linear time scale is most likely to give interpretable results.
The batch microcosms should have approximately the same ratio of solids to water as the
original material. Most of the microbes are attached to solids. If a microcosm has an excess of
water, and the contaminant is mostly in the aqueous phase, the microbes must process a great deal
more contaminant to produce the same relative change in the contaminant concentration as would be
obtained at field scale. The kinetics at field scale would be underestimated.
Microcosms are inherently time consuming. At field scale, the residence time of a plume may
be several years to decades. Slow rates of transformation may have a considerable environmental
significance. A microcosm study that lasts only a few weeks or months may not have the resolution
to detect slow changes that are still of environmental significance. Further, microcosms often show a
pattern of sequential utilization, with toluene and the xylenes degrading first, and benzene and
ethylbenzene degrading at a later time. Degradation of benzene or ethylbenzene may be delayed by
as much as a year.
C3-49
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As a practical matter, batch microcosms with an optimal solids-to-water ratio, sampled every 2
months in triplicate for up to 18 months, can resolve biodegradation from abiotic losses with a rate
detection limit of 0.001 to 0.0005 per day. Many plumes show significant attenuation of contamina-
tion at field-calibrated rates that are slower than the detection limit of today's microcosm technology.
The most appropriate use of microcosms is to document that contaminant attenuation is largely a
biological process. Rate constants for modeling purposes are more appropriately acquired from
field-scale studies.
Microcosm studies are often used to provide a third line of evidence. The potential for biodeg-
radation of the contaminants of interest can be confirmed by the use of microcosms, through com-
parison of removals in the living treatments with removals in the controls. Microcosm studies also
permit an absolute mass balance determination based on biodegradation of the contaminants of
interest. Further, the appearance of daughter products in the microcosms can be used to confirm
biodegradation of the parent compound.
C.3.4.2 When to Use Microcosms
There are two fundamentally different applications of microcosms. They are frequently used in
a qualitative way to illustrate the important processes that control the fate of organic contaminants.
They are also used to estimate rate constants for biotransformation of contaminants that can be used
in a site-specific transport and fate model of a plume of contaminated groundwater. This paper only
discusses microcosms for the second application.
Microcosms should be used when there is no other way to obtain a rate constant for attenuation
of contaminants, in particular, when it is impossible to estimate the rate of attenuation from data
from monitoring wells in the plume of concern. There are situations where it is impossible to com-
pare concentrations in monitoring wells along a flow path due to legal or physical impediments. In
many landscapes, the direction of ground-water flow (and water table elevations in monitoring wells)
can vary over short periods of time due to tidal influences or changes in barometric pressure. The
direction of ground-water flow may also be affected by changes in the stage of a nearby river or
pumping wells in the vicinity. These changes in ground-water flow direction do not allow simple
snap-shot comparisons of concentrations in monitoring wells because of uncertainties in identifying
the flow path. Rate constants from microcosms can be used with average flow conditions to estimate
attenuation at some point of discharge or point of compliance.
C.3.4.3 Application of Microcosms
The primary objective of microcosm studies is to obtain rate constants applicable to average
flow conditions. These average conditions can be determined by continuous monitoring of water
table elevations in the aquifer being evaluated. The product of the microcosm study and the continu-
ous monitoring of water table elevations will be a yearly or seasonal estimate of the extent of attenu-
ation along average flow paths. Removals seen at field scale can be attributed to biological activity.
If removals in the microcosms duplicate removal at field scale, the rate constant can be used for risk
assessment purposes (B.H. Wilson et a/., 1996; Bradley, el a/., 1998).
C.3.4.4 Selecting Material for Study
Prior to choosing material for microcosm studies, the location of major conduits of ground-
water flow should be identified and the geochemical regions along the flow path should be deter-
mined. The important geochemical regions for natural attenuation of chlorinated aliphatic hydrocar-
bons are regions that are actively methanogenic; regions that exhibit sulfate reduction and iron
reduction concomitantly; and regions that exhibit iron reduction alone. The pattern of biodegrada-
tion of chlorinated solvents varies in different regions. Vinyl chloride tends to accumulate during
reductive dechlorination of TCE or PCE in methanogenic regions (Weaver etal., 1995; J.T. Wilson
el a/., 1995); it does not accumulate to the same extent in regions exhibiting iron reduction and
C3-50
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sulfate reduction (Chapelle, 1996). In regions showing iron reduction alone, vinyl chloride is con-
sumed but dechlorination of PCE, TCE, or DCE may not occur (Bradley and Chapelle, 1996;1997).
Core material from each geochemical region in major flow paths represented by the plume must be
acquired, and the hydraulic conductivity of each depth at which core material is acquired must be
measured. If possible, the microcosms should be constructed with the most transmissive material in
the flow path.
Several characteristics of ground water from the same interval used to collect the core material
should be determined. These characteristics include temperature, redox potential, pH, and concen-
trations of oxygen, sulfate, sulfide, nitrate, iron II, chloride, methane, ethane, ethene, total organic
carbon, and alkalinity. The concentrations of compounds of regulatory concern and any breakdown
products for each site must be determined. The ground water should be analyzed for methane to
determine if methanogenic conditions exist and for ethane and ethene as daughter products from
reductive dechlorination of PCE and TCE. A comparison of the ground-water chemistry from the
interval where the cores were acquired to that in neighboring monitoring wells will demonstrate if
the collected cores are representative of that section of the contaminant plume.
Reductive dechlorination of chlorinated solvents requires an electron donor to allow the process
to proceed. The electron donor could be soil organic matter, low molecular weight organic com-
pounds (lactate, acetate, methanol, glucose, etc.), H,, or a co-contaminant such as landfill leachate or
petroleum compounds (Bouwer, 1994; Sewell and Gibson, 1991; Klecka etal., 1996). In many
instances, the actual electron donor(s) may not be identified.
Several characteristics of the core material should also be evaluated. The initial concentration
of the contaminated material (on a mass per mass basis) should be identified prior to construction of
the microcosms. Also, it is necessary to know if the contamination is present as a nonaqueous phase
liquid (NAPL) or in solution. A total petroleum hydrocarbon (TPH) analysis will determine if any
hydrocarbon-based oily materials are present. The water-filled porosity is a parameter generally used
to extrapolate rates to the field. It can be calculated by comparing wet and dry weights of the aquifer
material.
To insure sample integrity and stability during acquisition, it is important to quickly transfer the
aquifer material into ajar, exclude air by adding ground water, and seal the jar without headspace.
The material should be cooled during transportation to the laboratory. Incubate the core material at
the ambient ground-water temperature in the dark before the construction of microcosms.
At least one microcosm study per geochemical region should be completed. If the plume is
over one kilometer in length, several microcosm studies per geochemical region may need to be
constructed.
C.3.4.5 Geochemical Characterization of the Site
The geochemistry of the subsurface affects behavior of organic and inorganic contaminants,
inorganic minerals, and microbial populations. Major geochemical parameters that characterize the
subsurface encompasses (1) pH; (2) ORP; (3) alkalinity; (4) physical and chemical characterization
of the solids; (5) temperature; (6) dissolved constituents, including electron acceptors; and (7) micro-
bial processes. The most important of these in relation to biological processes are redox potential,
alkalinity, concentration of electron acceptor, and chemical nature of the solids.
Alkalinity: Field indications of biologically active portions of a plume may be identified by
increased alkalinity, compared to background wells, from carbon dioxide due to biodegradation of
the pollutants. Increases in both alkalinity and decrease in pH have been measured in portions of an
aquifer contaminated by gasoline undergoing active utilization of the gasoline components
(Cozzarelli et a/,., 1995). Alkalinity can be one of the parameters used when identifying where to
collect biologically active core material.
C3-51
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pH: Bacteria generally prefer a neutral or slightly alkaline pH level with an optimum pH range
for most microorganisms between 6.0 and 8.0; however, many microorganisms can tolerate a pH
range of 5.0 to 9.0. Most ground waters in uncontaminated aquifers are within these ranges. Natural
pH values may be as low as 4.0 or 5.0 in aquifers with active oxidation of sulfides, and pH values as
high as 9.0 may be found in carbonate-buffered systems (Chapelle, 1993). However, pH values as
low as 3.0 have been measured for ground waters contaminated with municipal waste leachates
which often contain elevated concentrations of organic acids (Baedecker and Back, 1979). In ground
waters contaminated with sludges from cement manufacturing, pH values as high as 11.0 have been
measured (Chapelle, 1993).
ORP: The ORP of ground water is a measure of electron activity that indicates the relative
ability of a solution to accept or transfer electrons. Most redox reactions in the subsurface are
microbially catalyzed during metabolism of native organic matter or contaminants. The only ele-
ments that are predominant participants in aquatic redox processes are carbon, nitrogen, oxygen,
sulfur, iron, and manganese (Stumm and Morgan, 1981). The principal oxidizing agents in ground
water are oxygen, nitrate, sulfate, manganese (IV), and iron (III). Biological reactions in the subsur-
face both influence and are affected by the redox potential and the available electron acceptors. The
redox potential changes with the predominant electron acceptor, with reducing conditions increasing
through the sequence oxygen, nitrate, iron, sulfate, and carbonate. The redox potential decreases in
each sequence, with methanogenic (carbonate as the electron acceptor) conditions being most reduc-
ing. The interpretation of redox potentials in ground waters is difficult (Snoeyink and Jenkins,
1980). The potential obtained in ground waters is a mixed potential that reflects the potential of
many reactions and cannot be used for quantitative interpretation (Stumm and Morgan, 1981). The
approximate location of the contaminant plume can be identified in the field by measurement of the
redox potential of the ground water.
To overcome the limitations imposed by traditional redox measurements, recent work has
focused on the measurement of molecular hydrogen to accurately describe the predominant in situ
redox reactions (Chapelle etal., 1995; Lovley etal., 1994; Lovley and Goodwin, 1988). The evi-
dence suggests that concentrations of Ft, in ground water can be correlated with specific microbial
processes, and these concentrations can be used to identify zones of methanogenesis, sulfate reduc-
tion, and iron reduction in the subsurface (Chapelle, 1996).
Electron Acceptors: Measurement of the available electron acceptors is critical in identifying
the predominant microbial and geochemical processes occurring in situ at the time of sample collec-
tion. Nitrate and sulfate are found naturally in most ground waters and will subsequently be used as
electron acceptors once oxygen is consumed. Oxidized forms of iron and manganese can be used as
electron acceptors before sulfate reduction commences. Iron and manganese minerals solubilize
coincidently with sulfate reduction, and their reduced forms scavenge oxygen to the extent that strict
anaerobes (some sulfate reducers and all methanogens) can develop. Sulfate is found in many
depositional environments, and sulfate reduction may be very common in many contaminated
ground waters. In environments where sulfate is depleted, carbonate becomes the electron acceptor
with methane gas produced as an end product.
Temperature: The temperature at all monitoring wells should be measured to determine when
the pumped water has stabilized and is ready for collection. Below approximately 30 feet, the
temperature in the subsurface is fairly consistent on an annual basis. Microcosms should be stored at
the average in situ temperature. Biological growth can occur over a wide range of temperatures,
although most microorganisms are active primarily between 10°C and 35°C (50°F to 95°F).
Chloride: Reductive dechlorination results in the accumulation of inorganic chloride. In
aquifers with a low background of inorganic chloride, the concentration of inorganic chloride should
C3-52
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increase as the chlorinated solvents are degraded. The sum of the inorganic chloride plus the chlo-
ride in the contaminant being degraded should remain relatively consistent along the ground water
flow path.
Tables C.3.7 and C.3.8 list the geochemical parameters, contaminants, and daughter products
that should be measured during site characterization for natural attenuation. The tables include the
analyses that should be performed, the optimum range for natural attenuation of chlorinated solvents,
and the interpretation of the value in relation to biological processes.
Table C.3.7 Geochemical Parameters Important to Microcosm Studies
Analysis
Redox Potential
Sulfate
Nitrate
Oxygen
Oxygen
Iron (II)
Sulfide
Hydrogen
Hydrogen
pH
Range
<50 millivolt against
Ag/AgCl
<20 mg/L
<1 mg/L
<0.5 mg/L
>1 mg/L
>1 mg/L
>1 mg/L
>1 nM
<1 nM
5 < pH < 9
Interpretation
Reductive pathway possible
Competes at higher concentrations with
Competes at higher concentrations with
Tolerated, toxic to reductive pathway at
reductive pathway
reductive pathway
higher concentrations
Vinyl chloride oxidized
Reductive pathway possible
Reductive pathway possible
Reductive pathway possible, vinyl chloride may accumulate
Vinyl chloride oxidized
Tolerated range
Table C.3.8
Contaminants and Daughter Products Important to Microcosm Studies
Analysis
PCE
TCE
1,1,1-TCA
c/s-l,2-DCE
fra«5-l,2-DCE
Vinyl Chloride
Ethene
Ethane
Methane
Chloride
Carbon Dioxide
Alkalinity
Interpretation
Material spilled
Material spilled or daughter product of PCE
Material spilled
Daughter product of TCE
Daughter product of TCE
Daughter product of dichloroethylenes
Daughter product of vinyl chloride
Daughter product of ethene
Ultimate reductive daughter product
Daughter product of organic chlorine
Ultimate oxidative daughter product
Results from interaction of carbon dioxide with aquifer minerals
C3-53
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C.3.4.6 Microcosm Construction
During construction of the microcosms, it is best if all manipulations take place in an anaerobic
glovebox. These gloveboxes exclude oxygen and provide an environment where the integrity of the
core material may be maintained, since many strict anaerobic bacteria are sensitive to oxygen. Strin-
gent aseptic precautions not necessary for microcosm construction. It is more important to maintain
anaerobic conditions of the aquifer material and solutions added to the microcosm bottles.
The microcosms should have approximately the same ratio of solids to water as the in situ
aquifer material, with a minimum or negligible headspace. Most bacteria in the subsurface are
attached to the aquifer solids. If a microcosm has an excess of water, and the contaminant is prima-
rily in the dissolved phase, the bacteria must consume or transform a great deal more contaminant to
produce the same relative change in the contaminant concentration. As a result, the kinetics of
removal at field scale will be underestimated in the microcosms.
A minimum of three replicate microcosms for both living and control treatments should be
constructed for each sampling event. Microcosms sacrificed at each sampling interval are preferable
to microcosms that are repetitively sampled. The compounds of regulatory interest should be added
at concentrations representative of the higher concentrations found in the geochemical region of the
plume being evaluated. The compounds should be added as a concentrated aqueous solution. If an
aqueous solution is not feasible, dioxane or acetonitrile may be used as solvents. Avoid carriers that
can be metabolized anaerobically, particularly alcohols. If possible, use ground water from the site
to prepare dosing solutions and to restore water lost from the core barrel during sample collection.
For long-term microcosm studies, autoclaving is the preferred method for sterilization. Nothing
available to sterilize core samples works perfectly. Mercuric chloride is excellent for short-term
studies (weeks or months). However, mercuric chloride complexes to clays, and control may be lost
as it is sorbed over time. Sodium azide is effective in repressing metabolism of bacteria that have
cytochromes, but is not effective on strict anaerobes.
The microcosms should be incubated in the dark at the ambient temperature of the aquifer. It is
preferable that the microcosms be incubated inverted in an anaerobic glovebox. Anaerobic jars are
also available that maintain an oxygen-free environment for the microcosms. Dry redox indicator
strips can be placed in the jars to assure that anoxic conditions are maintained. If no anaerobic
storage is available, the inverted microcosms can be immersed in approximately two inches of water
during incubation. Teflon®-lined butyl rubber septa are excellent for excluding oxygen and should
be used if the microcosms must be stored outside an anaerobic environment.
The studies should last from one year to eighteen months. The residence time of a plume may
be several years to tens of years at field scale. Rates of transformation that are slow in terms of
laboratory experimentation may have a considerable environmental significance. A microcosm
study lasting only a few weeks to months may not have the resolution to detect slow changes that are
of environmental significance. Additionally, microcosm studies often distinguish a pattern of se-
quential biodegradation of the contaminants of interest and their daughter products.
C.3.4.7 Microcosm Interpretation
As a practical matter, batch microcosms with an optimal solids/water ratio, that are sampled
every two months in triplicate, for up to eighteen months, can resolve biodegradation from abiotic
losses with a detection limit of 0.001 to 0.0005 per day. Rates determined from replicated batch
microcosms are found to more accurately duplicate field rates of natural attenuation than column
studies. Many plumes show significant attenuation of contamination at field calibrated rates that are
slower than the detection limit of microcosms. Although rate constants for modeling purposes are
more appropriately acquired from field-scale studies, it is reassuring when the rates in the field and
the rates in the laboratory agree.
C3-54
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The rates measured in the microcosm study may be faster than the estimated field rate. This
may not be due to an error in the laboratory study, particularly if estimation of the field-scale rate of
attenuation did not account for regions of preferential flow in the aquifer. The regions of preferential
flow may be determined by use of a downhole flow meter or by other methods for determining
hydraulic conductivity in one- to two-foot sections of the aquifer.
Statistical comparisons can determine if removals of contaminants of concern in the living
treatments are significantly different from zero or significantly different from any sorption that is
occurring. Comparisons are made on the first-order rate of removal, that is, the slope of a linear
regression of the natural logarithm of the concentration remaining against time of incubation for both
the living and control microcosm. These slopes (removal rates) are compared to determine if they are
different, and if so, extent of difference that can be detected at a given level of confidence.
C.3.4.8 The Tlbbetts Case Study
The Tibbetts Road Superfund Site in Barrington, New Hampshire, a former private home, was
used to store drums of various chemicals from 1944 to 1984. The primary ground-water contami-
nants in the overburden and bedrock aquifers were benzene and TCE, with respective concentrations
of 7,800 |ig/L and 1,100 |ig/L. High concentrations of arsenic, chromium, nickel, and lead were also
found.
Material collected at the site was used to construct a microcosm study evaluating the removal of
benzene, toluene, and TCE. This material was acquired from the most contaminated source at the
site, the waste pile near the origin of Segment A (Figure C.3.10). Microcosms were incubated for
nine months. The aquifer material was added to 20-mL headspace vials, dosed with 1 mL of spiking
solution, capped with a Teflon®-lined, gray butyl rubber septa, and sealed with an aluminum crimp
cap. Controls were prepared by autoclaving the material used to construct the microcosms over-
night. Initial concentrations for benzene, toluene, and TCE were, respectively, 380 |ig/L, 450 (ig/L,
and 330 |ig/L. The microcosms were thoroughly mixed by vortexing, then stored inverted in the
dark at the ambient temperature of 10°C.
The results (Figures C.3.11, C.3.12, and C.3.13; Table C.3.9) show that significant biodegrada-
tion of both petroleum aromatic hydrocarbons and the chlorinated solvent had occurred. Significant
removal in the control microcosms also occurred for all compounds. The data exhibited more
variability in the living microcosms than in the control treatment, which is a pattern that has been
observed in other microcosm studies. The removals observed in the controls are probably due to
sorption; however, this study exhibited more sorption than typically seen.
The rate constants determined from the microcosm study for the three compounds are shown in
Table C.3.10. The appropriate rate constant to be used in a model or a risk assessment would be the
first-order removal in the living treatment minus the first-order removal in the control, in other words
the removal that is in excess of the removal in the controls.
The first-order removal in the living and control microcosms was estimated as the linear regres-
sion of the natural logarithm of concentration remaining in each microcosm in each treatment against
time of incubation. Student's t-distribution with n-2 degrees of freedom was used to estimate the
95% confidence interval. The standard error of the difference of the rates of removal in living and
control microcosms was estimated as the square root of the sum of the squares of the standard errors
of the living and control microcosms, with n-4 degrees of freedom (Glantz, 1992).
Table C.3.11 presents the concentrations of organic compounds and their metabolic products in
monitoring wells used to define line segments in the aquifer for estimation of field-scale rate con-
stants. Wells in this aquifer showed little accumulation of/ra«s-l,2-DCE; 1,1-DCE; vinyl chloride;
or ethene, although removals of TCE and c/s-l,2-DCE were extensive. This can be explained by the
observation (Bradley and Chapelle, 1996) that iron-reducing bacteria can rapidly oxidize vinyl
chloride to carbon dioxide. Filterable iron accumulated in ground water in this aquifer.
C3-55
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Waste Pile
Ground Water
Flow Segment
Figure C.3.10 Tihbetts Road study site.
1000
100 -
10 -
10
15 20 25 30
Time (Weeks)
35 40 45
Figure C.3.11 TCE microcosm results.
C3-56
-------
1000
100 -
CO
zt
aBenzene Microcosm
mBenzene Control
10
15 20 25 30
Time (Weeks)
35
40
45
Figure C.3.12 Benzene microcosm results.
1000
CO
n Toluene Microcosm
• Toluene Control
10 15 20 25 30 35
40
45
Time (Weeks)
Figure C.3.13 Toluene microcosm results.
C3-57
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Table C.3.9
Concentrations (ng/L) ofTCE, Benzene, and Toluene in the Tibhetts Road Microcosms
Compound
ICE
Mean±
Standard
Deviation
Benzene
Mean±
Standard
Deviation
Toluene
Mean±
Standard
Deviation
Time Zero
Microcosm
328
261
309
299 ± 34.5
366
280
340
329 ± 44. 1
443
342
411
399 ±51.6
Time Zero
Control
337
394
367
366 ± 28.5
396
462
433
430 ± 33. 1
460
557
502
506 ± 48.6
Week 23
Microcosm
1
12.5
8.46
7.32 ± 5.83
201
276
22.8
167 ± 130
228
304
19.9
184 ± 147
Week 23
Control
180
116
99.9
132 ± 42.4
236
180
152
189 ± 42.8
254
185
157
199 ± 49.9
Week 42
Microcosm
2
2
2
2.0 ± 0.0
11.1
20.5
11.6
14.4 ± 5.29
2
2.5
16.6
7.03 ± 8.29
Week 42
Control
36.3
54.5
42.3
44.4 ± 9.27
146
105
139
130 ±21.9
136
92
115
114 ±22.0
The extent of attenuation from well to well listed in Table C.3.11, and the travel time between
wells in a segment (Figure C.3.4) were used to calculate first-order rate constants for each segment
(Table C.3.12). Travel time between monitoring wells was calculated from site-specific estimates of
hydraulic conductivity and from the hydraulic gradient. In the area sampled for the microcosm study,
the estimated Darcy flow was 2.0 feet per year. With an estimated porosity in this particular glacial
till of 0.1, this corresponds to a plume velocity of 20 feet per year.
C.3.4.9 Summary
Table C.3.13 compares the first-order rate constants estimated from the microcosm studies to
the rate constants estimated at field scale. The agreement between the independent estimates of rate
is good; indicating that the rates can appropriately be used in a risk assessment. The rates of biodeg-
radation documented in the microcosm study could easily account for the disappearance of trichloro-
ethylene, benzene, and toluene observed at field scale. The rates estimated from the microcosm
study are several-fold higher than the rates estimated at field scale. This may reflect an underestima-
tion of the true rate in the field. The estimates of plume velocity assumed that the aquifer was
homogeneous. No attempt was made in this study to correct the estimate of plume velocity for the
influence of preferential flow paths. Preferential flow paths with a higher hydraulic conductivity
than average would result in a faster velocity of the plume, thus a lower residence time and faster
rate of removal at field scale.
C3-58
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Table C.3.10 First-order Rate Constants for Removal ofTCE, Benzene, and Toluene in the Tibbetts Road
Microcosms
Parameter
Living Microcosms
Autoclaved
Controls
Removal Above
Controls
First-order Rate of Removal (per year)
ICE
95% Confidence Interval
Vlinimum Rate Significant at 95%
Confidence
Benzene
95% Confidence Interval
Vlinimum Rate Significant at 95%
Confidence
Toluene
95% Confidence Interval
Vlinimum Rate Significant at 95%
Confidence
6.31
±2.50
3.87
± 1.96
5.49
±2.87
2.62
±0.50
1.51
±0.44
1.86
±0.45
3.69
±2.31
1.38
2.36
± 1.83
0.53
3.63
±2.64
0.99
Table C3.ll Concentrations of Contaminants and Metabolic By-products in Monitoring Wells along
Segments in the Plume used to Estimate Field-scale Rate Constants
Parameter
Monitoring
Well
TCE
cis-l,2-DCE
trans- 1,2-DCE
1,1-DCE
Vinyl Chloride
Ethene
Benzene
Toluene
o-Xylene
m-Xylene
p-Xylene
Ethylbenzene
Methane
Iron
Segment A
806
79S
Upgradient Downgradient
Segment B
70S
Up
gradient
52S
Down
gradient
Segment C
70S
Upgradient
(lid A\far\
IMs/11^1-)
200
740
0.41
0.99
<1
<4
510
10000
1400
2500
1400
1300
353
13.7
10.9
<1
<1
<1
<4
2.5
<1
8.4
<1
22
0.7
77
710
220
0.8
<1
<1
7
493
3850
240
360
1100
760
8
67
270
0.3
1.6
<1
<4
420
900
71
59
320
310
3
710
220
0.8
<1
<1
7
493
3850
240
360
1100
760
8
53S
Down
gradient
3.1
2.9
<1
<1
<1
<4
<1
<1
<1
<1
<1
<1
<2
27000
C3-59
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Table C.3.12 First-order Rate Constants for Contaminant Attenuation in Segments of the Tibbetts Road
Plume
Flow Path Segments in Length and Time of Ground-water Travel
Compound
TCE
cis-l,2-DCE
Benzene
Toluene
o-Xylene
m-Xylene
p-Xylene
Ethylbenzene
Segment A
130 feet =6.5 years
Segment B
80 feet =4.0 years
Segment C
200feet= 10 years
First-order Rate Constants in Segments ( per year)
0.41
0.65
0.82
>1.42
0.79
>1.20
0.64
1.16
0.59
produced
0.04
0.36
0.30
0.45
0.31
0.22
0.54
0.43
>0.62
>0.83
>0.55
>0.59
>0.70
>0.66
Table C.3.13 Comparison of First-order Rate Constants in a Microcosm Study, and in the Field, at the
Tibbetts Road NPL Site
Parameter
rrichloroethylene
Benzene
Toluene
Microcosms Corrected for
Controls
Average
Rate
3.69
2.36
3.63
Minimum Rate
Significant at 95%
Confidence
TTirct n
1.38
0.53
0.99
Field Scale
Segment A
nder Rate (per
0.41
0.82
>1.42
Segment B
Segment C
ycdij
0.59
0.04
0.36
0.54
>0.62
>0.83
C3-60
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