SEPA
EPA/600/R-20/346 | July 2020 | www.epa.gov/research
United States
Environmental Protection
Agency
Technical Considerations for
Evaluating the Environmental
Emissions from RCRA Subtitle
D Landfills Beyond the 30-Year
Post-Closure Care Period
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Center for Environmental Solutions and Emergency Response
Homeland Security and Materials Management Division
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EPA/600/R-20/346
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Technical Considerations for Evaluating
the Environmental Emissions from
RCRA Subtitle D Landfills Beyond the
30-Year Post-Closure Care Period
by
Thabet Tolaymat, PhD, PE
US EPA/Center for Environmental Solutions and Emergency
Response, Cincinnati, OH 45268
David Carson
US EPA/ Center for Environmental Solutions and Emergency
Response, Cincinnati, OH 45268
Contract Number
EP-C-15-012
Project Officer
Thabet Tolaymat, PhD, PE
Center for Environmental Solutions and Emergency Response
Cincinnati, Ohio, 45268
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Notice/Disclaimer
The research described in this report has been funded wholly or in part by the US Environmental
Protection Agency's contract number EP-C-15-012 to CSRA LLC.. The report has been subjected to the
Agency's peer and administrative review and has been approved for publication as an EPA document.
Mention of trade names or commercial products does not constitute endorsement or recommendation for
use.
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EPA/600/R-20/346
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Foreword
The US Environmental Protection Agency (EPA) 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 this
mandate, EPA's research program is providing data and technical support for solving
environmental problems today and building a science knowledge base necessary to manage our
ecological resources wisely, understand how pollutants affect our health, and prevent or reduce
environmental risks in the future.
The Center for Environmental Solutions and Emergency Response (CESER) within the Office of
Research and Development (ORD) conducts applied, stakeholder-driven research and provides
responsive technical support to help solve the Nation's environmental challenges. The Center's
research focuses on innovative approaches to address environmental challenges associated with
the built environment. We develop technologies and decision-support tools to help safeguard
public water systems and groundwater, guide sustainable materials management, remediate sites
from traditional contamination sources and emerging environmental stressors, and address
potential threats from terrorism and natural disasters. CESER collaborates with both public and
private sector partners to foster technologies that improve the effectiveness and reduce the cost of
compliance, while anticipating emerging problems. We provide technical support to EPA regions
and programs, states, tribal nations, and federal partners, and serve as the interagency liaison for
EPA in homeland security research and technology. The Center is a leader in providing scientific
solutions to protect human health and the environment.
Gregory Sayles, Director
Center for Environmental Solutions and Emergency Response
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Acknowledgements
«mJr
This research was funded by the Center for Environmental Solutions and Emergency Response (CESER)
of the US Environmental Protection Agency (EPA), Office of Research and Development (ORD) under
the Sustainable and Healthy Communities Research program. This report was prepared by Innovative
Waste Consulting Services, LLC, under subcontract to CSRA LLC, a General Dynamics Information
Technology company. Work was performed in accordance with a Work Assignment issued by EPA under
Contract EP-C-15-012.
The project team would like to acknowledge the following state regulators, companies, and individuals
for the information and time they contributed to this report or for providing the data and information used
for this report. Please note that for confidentiality, the site owners/operators and government agencies
responsible for regulatory oversight of the nine study sites are not presented here. However, the
environmental monitoring data provided by these individuals were critical in the development of this
report.
Pradeep Jain, PhD, PE, Innovative Waste Consulting Services, LLC
Timothy G. Townsend, PhD, PE, Innovative Waste Consulting Services, LLC
Shrawan Singh, PhD, PE, Innovative Waste Consulting Services, LLC
Justin L. Smith, PE, Innovative Waste Consulting Services, LLC
James Wally, PE, Innovative Waste Consulting Services, LLC
Justin Roessler, PhD, Innovative Waste Consulting Services, LLC
William Balcke, CSRA, LLC
Jaime M. Colby, NH Department of Environmental Services
Michael R. Gerchman, Section Chief, NJ Department of Environmental Protection
Clement DeLattre, PA Department of Environmental Protection
Scott McWilliams, AR Department of Environmental Quality
Scott Walker, CalRecycle
Solid Waste Association of North America
Waste Management, Inc.
Morton Barlaz, PhD, North Carolina State University
David Daniel, PhD, University of Texas
Edward Barth, PhD, Environmental Protection Agency
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Table of Contents
Notice/Disclaimer ii
Foreword iii
Acknowledgements iv
Table of Contents v
List of Figures x
List of Tables xiii
Acronyms and Abbreviations xv
Executive Summary i
1. Introduction 1
1.1. Background 1
1.2. Objectives and Scope 3
1.3. Report Organization 4
2. Background 5
2.1. Landfill Operation and Design 5
2.2. Monitoring 6
2.3. Closure and Post-Closure Care 6
2.4. Financial Assurance 7
2.5. NSPS and NESHAP Regulations 8
2.6. Existing Guidance on PCC Period Evaluation 9
2.6.1. Overview 9
2.6.2. Leachate Management 10
2.6.3. Landfill Gas Management 11
2.6.4. Groundwater Monitoring 11
2.6.5. Final Cover Management and Maintenance 12
3. Closed Landfill Case Studies 13
3.1. Study Sites Selection Process 13
3.2. Data Collected 14
3.4. Descriptions of Selected Sites 16
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3.5. Data Presentation 19
4. Post-Closure Care Cost 20
4.1. Overview 20
4.2. Leachate Management Cost 24
4.3. GCCS Management Cost 25
4.4. Monitoring Cost 26
4.5. Final cover Maintenance Cost 27
4.6. Summary 29
5. Landfill Gas 31
5.1. Overview 31
5.2. Methodol ogy 32
5.2.1. Data Sources 32
5.2.2. Methane Collection Rate for the Study Sites 34
5.2.3. Methane and NMOCs Collection Rate Estimation 35
5.2.4. Site-Specific Decay Rate and Methane Collection Potential Estimation .... 37
5.2.5. Potential of Elevated NMOCs Generation Rate after GCCS Termination.. 38
5.2.6. Remaining Methane and NMOCs Generation Potential 39
5.3. Results and Discussion 40
5.3.1. Subsurface Methane Monitoring Data 40
5.3.2. Methane Collection Rate Trends 42
5.3.3. Estimated Site-Specific First-Order Decay Rates and Methane Collection
Potentials 43
5.3.4. Timeframes for Achieving Annual NMOCs Collection Rate below 50/34 Mg
and LFG Flow Rate Below 5% and 10% of the Peak Rates 45
5.3.5. Impact of In-Place Waste Amount and Decay Rate on Timeframes to Achieve
Annual NMOCs Collection Rate Reduction below the NSPS Thresholds.. 47
5.3.6. Potential of Elevated NMOCs Generation after GCCS Termination 50
5.3.7. Estimated Remaining Methane and NMOCs Generation Potential 53
5.4. Summary 55
5.5. Limitations 58
6. Landfill Leachate 60
6.1. Overview 60
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6.2. Data Sources 62
6.3. Methodologies for Prediction of Long-Term Leachate Collection after
Closure 64
6.3.1. Overview 64
6.3.2. First-Order Decay Modeling 64
6.3.3. Unsaturated Flow Modeling 65
6.3.4. HELP Modeling 66
6.4. Results and Discussion-Leachate Collection Rate 68
6.4.1. Measured Leachate Collection after Closure 68
6.4.2. Modeled Leachate Collection Rates 72
6.4.3. An Evaluation of the Performance of the Primary Liner System 81
6.5. Leachate Quality 84
6.5.1. Leachate Quality Data Available for the Study Sites 84
6.5.2. Impact of Leachate Quality on Groundwater 87
6.5.3. Temporal Analysis of Leachate Quality 95
6.6. Impacts of Leachate Collection System Failure and Subsequent Leachate
Leakage 110
6.7. Considerations for Assessing and Mitigating Leachate Impacts on HHE. Ill
6.8. Summary 112
6.8.1. Data Availability 112
6.8.2. Measured Leachate Collection Rate 112
6.8.3. Leachate Collection Rate Modeling Approach Evaluation 113
6.8.4. Hydraulic Efficiency of Primary Liner 113
6.8.5. Available Leachate Quality Data 113
6.8.6. Contaminants of Potential Concern 114
6.8.7. Temporal Trends of the Contaminants of Potential Concern 114
6.9. Limitations 115
7. Groundwater Monitoring Data 116
7.1. Overview 116
7.2. Data Sources 117
7.3. Challenges to Isolating and Understanding Study Cell(s) Groundwater
Impacts 119
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7.4. Groundwater Impacts at the Studied Sites 119
7.5. Impact of Biogeochemical Changes on Groundwater Quality 128
7.6. Variation in Groundwater Flow 132
7.7. Integrity and Performance of Monitoring Wells 132
7.8. Summary 132
7.9. Limitations 133
8. Final Cover Performance 135
8.1. Surface Emissions Monitoring 135
8.2. Settlement 136
9. Summary and Considerations 139
9.1. Summary of Findings 139
9.1.1. Overview 139
9.1.2. Post-Closure Care Cost 139
9.1.3. Landfill Gas 140
9.1.4. Landfill Leachate 142
9.1.5. Groundwater Monitoring Data 144
9.1.6. Final Cover Performance 146
9.2. Considerations for Assessing and Mitigating Long-term Impacts of MSWLFs
146
9.2.1. Operating Considerations 146
9.2.2. Monitoring Considerations 147
10. References 150
Attachments
Attachment A -Site A Description and Data Evaluation
Attachment B -Site B Description and Data Evaluation
Attachment C -Site C Description and Data Evaluation
Attachment D -Site D Description and Data Evaluation
Attachment E -Site E Description and Data Evaluation
Attachment F -Site F Description and Data Evaluation
Attachment G -Site G Description and Data Evaluation
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Attachment H -Site H Description and Data Evaluation
Attachment I -Site I Description and Data Evaluation
Attachment J - Leachate Quality Parameters Evaluated
Attachment K - Maximum Contaminant Levels for Subtitle D Landfills
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Figure 1-1. Number of Closed MSWLFs by Reported Closure Year and the Projected 30-Year
PCC Completion Year based on the EPA Greenhouse Gas Reporting Program Database (EPA
2019b) 1
Figure 1-2. Locations of All Closed MSWLFs Included in the EPA Greenhouse Gas Reporting
Program Database (EPA 2019b) 2
Figure 3-1. Box-and-Whisker Plot Definition Sketch 19
Figure 4-1. Distribution of Total Annual PCC Cost at Each Study Site 22
Figure 4-2. Average Annual Cost Distribution at Each Study Site/Cell After Closure 23
Figure 4-3. Distribution of Leachate Management Cost at Six Study Sites after Closure (Site G
Has Two Study Cells) 25
Figure 4-4. Distribution of Annual GCCS Management Cost at Seven Study Sites After Closure
26
Figure 4-5. Distribution of Annual Groundwater and Subsurface Gas Monitoring Cost After
Closure at Each Study Sites 27
Figure 4-6. Distribution of Annual Final Cover Maintenance Cost During PCC at the Eight Study
Sites (Site G has two study cells) 28
Figure 4-7. Annual Final Cover Maintenance Cost and Annual Volume Loss at Site 1 29
Figure 5-1. Temporal Variation of Methane Flow Data for Study Sites 43
Figure 5-2. Distribution of Capacities of Active and Closed MSWLFs in the US 48
Figure 5-3. Estimated Annual NMOCs Generation Rate for Typical Small MSWLF in a Moderate
Precipitation Zone with 100% Increase in Waste Decay Rate 10 Years after GCCS
Termination 51
Figure 5-4. Estimated Annual NMOCs Generation Rate for Typical Large MSWLF in an Arid
Area with 100% Increase in Waste Decay Rate 10 Years after GCCS Termination 52
Figure 5-5. Annual NMOCs Generation Rate as a Function of the Percent Remaining Potential for
Small MSWLF in Moderate Precipitation Zone and Large MSWLF in Arid Area 54
Figure 6-1. Temporal Variation of Annual Leachate Collection Rates at Study Cell(s) After
Closure 69
Figure 6-2. Temporal Variation of Monthly Leachate Collection Rate and Precipitation for Site F
70
Figure 6-3. Annual Leachate Collection Rates as a Percentage of Average Annual Precipitation of
Study Cells after Closure 72
Figure 6-4. Measured and Modeled Leachate Collection Rates from Site B 74
Figure 6-5. Measured and Modeled Annual Leachate Collection Rates from Site CI 74
Figure 6-6. Measured and Modeled Annual Leachate Collection Rates from Site C2 75
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Figure 6-7. Measured and Modeled Annual Leachate Collection Rates from Site D 75
Figure 6-8. Measured and Modeled Annual Leachate Collection Rates from Site E 76
Figure 6-9. Measured and Modeled Annual Leachate Collection Rates from Site F 76
Figure 6-10. Measured and Modeled Annual Leachate Collection Rates from Study Cell G3.... 77
Figure 6-11. Measured and Modeled Annual Leachate Collection Rates from Study Cell G4.... 77
Figure 6-12. Measured and Modeled Annual Leachate Collection Rates from Site H 78
Figure 6-13. Measured and Modeled Annual Leachate Collection Rates from Site 1 78
Figure 6-14. Temporal Variation of LDS Collection Rates at Study Cell(s) of Each Study Site
After Closure 81
Figure 6-15. Chloride Concentration Distribution of Primary (or Composite) Leachate, Secondary
Leachate, and Background Groundwater Quality for at Site I and Site H 83
Figure 6-16. Organic Compound Concentrations in Primary (or Composite) and Secondary
Leachate at Site H and 1 84
Figure 6-17. Number of §258 App I and App II Parameters Analyzed for Leachate at Eight of the
Study Sites 86
Figure 6-18. Distribution of MCL-normalized Concentrations of Frequently Detected Parameters
94
Figure 6-19. Distribution of SMCL-normalized Concentrations of Frequently Detected Parameters
94
Figure 6-20. Distribution of Leachate Arsenic Concentration at Eight Study Sites Since Closure
96
Figure 6-21. Distribution of Arsenic Concentration in Leachate from Two Sumps of the Study Cell
C2 96
Figure 6-22. Temporal Variability in Leachate Arsenic Concentration of Sites B, D, and H Since
Closure 97
Figure 6-23. Temporal Variability in Arsenic Release Rate of Site H Since Closure 98
Figure 6-24. Distribution of Leachate Iron Concentration at Seven Study Sites Since Closure.. 99
Figure 6-25. Temporal Variability in Leachate Iron Concentration of Sites H and I Since Closure
99
Figure 6-26. Distribution of Leachate Manganese Concentration at Seven Study Sites Since
Closure 100
Figure 6-27. Temporal Variability in Leachate Manganese Concentration of Sites B and H Since
Closure 100
Figure 6-28. Distribution of Leachate TDS Concentration at Six Study Sites Since Closure.... 101
Figure 6-29. Temporal Variability in Leachate TDS Concentration of Site C Since Closure.... 102
Figure 6-30. Distribution of Leachate Chloride Concentration at Seven Study Sites Since Closure
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103
Figure 6-31. Temporal Variability in Leachate Chloride Concentration of Sites C and I Since
Closure 103
Figure 6-32. Distribution of Leachate pH at Eight Study Sites Since Closure 104
Figure 6-33. Temporal Variability in Leachate pH of Sites A and F Since Closure 105
Figure 6-34. Distribution of Leachate BOD at Seven Study Sites Since Closure 106
Figure 6-35. Temporal Variability in Leachate BOD of Sites B and C Since Closure 107
Figure 6-36. Distribution of Leachate COD at Seven Study Sites Since Closure 107
Figure 6-37. Temporal Variability in Leachate COD of Sites B and C Since Closure 108
Figure 6-38. Distribution of Leachate Ammonia-N Concentration at Five Study Sites Since
Closure 109
Figure 6-39. Temporal Variability of Leachate Ammonia-N Concentration of Sites C and H Since
Closure 110
Figure 7-1. Temporal Variation of Arsenic Concentrations Observed Above the MCL at Four
Downgradient Wells at Site I After Closure 123
Figure 7-2. Distribution of (a) Chloride, and (b) Ammonia for Leachate and Four Downgradient
Wells and an Upgradient Well at Site 1 125
Figure 7-3. Temporal Variation of Detected 1,4-Dioxane Concentrations Observed at three
Downgradient Wells of Site I After Closure 126
Figure 7-4. Distribution of Vanadium Concentration at Upgradient (U) and Downgradient (D)
Groundwater Monitoring Wells of Site A After Closure 127
Figure 7-5. Distribution of Iron Measured at Various Groundwater Monitoring Wells
(identification labels with MW and CW prefixes are groundwater monitoring wells), and
Leachate at an Active Lined MSWLF in Florida (IWCS 2010) 129
Figure 7-6. Distribution of Iron Concentrations Measured in Leachate and in Downgradient
Groundwater Monitoring Wells with Elevated Arsenic Concentrations at Site I after Closure
130
Figure 7-7. Distribution of (a) Arsenic-to-Chloride Ratio, and (b) Iron-to-Chloride Ratio for
Leachate and Four Downgradient Wells at Site I after Closure 131
Figure 8-1. Distribution of Yearly Point-to-Point Settlement Rate at Site 1 137
Figure 8-2. Site I Estimated Annual Volume Loss Based on Topographical Data and Modeled
Volume Loss 138
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Table 3-1. Environmental Monitoring and Operational Data Collected for Each Study Site 15
Table 3-2. Environmental Monitoring and Operational Data Analyzed for Each Study Site 15
Table 3-3. Design Summary for the Studied Cell(s) at Each Site 17
Table 3-4. Operational Summary for the Studied Cell(s) at Each Site 18
Table 4-1. Summary of Availability of PCC Cost Data from the Nine Study Sites 21
Table 5-1. Key Attributes Relevant to the GCCS of Each Study Site 34
Table 5-2. Modeling Parameters for the Estimation of the Future Methane and NMOCs Collection
Rate for the Study Sites 36
Table 5-3. Initial Values and Constraints for Parameters used for Modeling 38
Table 5-4. Number of Annual Methane Monitoring Probe Methane Exceedances at the Site as a
Function of Time Since Closure 42
Table 5-5. Estimated First-Order Decay Rate and Methane Collection Potential, and Average
Annual Precipitation for the Study Sites 45
Table 5-6. Estimated Timeframes for NMOCs Collection Rate and LFGFlow Rate Reduction 46
Table 5-7. Capacity, Lifespan, and Annual Disposal Amounts of the Example MSWLFs 49
Table 5-8. Impact of In-Place Waste Amount and Decay Rate on the Estimated Timeframes for
Achieving Annual NMOCs Collection Rate Below 50 Mg per Year 50
Table 5-9. Estimated Percent Remaining Methane Generation Potential of the Study Sites at
Closure and 30 Years after Closure 53
Table 5-10. Estimated Remaining NMOCs Generation Potential of the Study Sites at Closure and
30 Years after Closure 55
Table 6-1. Summary of Availability of Leachate Quantity and Quality Data at the Nine Study Sites
Since Closure 63
Table 6-2. Initial Conditions and Constraints for Calculating Kui 66
Table 6-3. Leachate Collection Rate Summary at the Nine Study Sites Since Closure 71
Table 6-4. Values Used for Different Modeling Approaches 80
Table 6-5. Estimated Leachate Collection Rate at the End of 30 Years after Closure Based on
Modeling Approaches Used 80
Table 6-6. Summary of Leachate Parameters that Exceeded the MCL or SMCL at Least Once at
the Study Sites 90
Table 6-7. Summary of Leachate Quality Parameters that were Detected in more than 50% of
Samples and Exceeded the Respective Drinking Water Standard at least Once at the Study
Sites 92
Table 7-1. Summary of Groundwater-Related Features for Study Cells of Each Studied Site.. 118
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Table 7-2. Site Summary of App I and II Parameters with Exceedances Above the Respective
Maximum Contaminant Level After Closure 121
Table 8-1. Number of Annual Surface Emissions Monitoring Exceedances Over Study Cells as a
Function of Time after Closure 135
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Acronyms and Abbreviations
AP-42 Compilation of Air Pollutant Emission Factors
App I 40 CFR 258 Appendix I
App II 40 CFR 258 Appendix II
ASTSWMO Association of State and Territorial Solid Waste Management Officials
BLS US Bureau of Labor Statistics
BOD Biochemical Oxygen Demand
CDD Construction and Demolition Debris
CFR Code of Federal Regulations
CH4 Methane
cm Centimeter
COD Chemical Oxygen Demand
COPC Contaminant of Potential Concern
C.U. Color Units
EPA United States Environmental Protection Agency
EPACMTP EPA's Composite Model for Leachate Migration with Transformation
Products
EREF Environmental Research and Education Foundation
ERG Eastern Research Group, Inc.
FAC Florida Administrative Code
FDEP Florida Department of Environmental Protection
ft Foot
ft3 Cubic Feet
GPAD Gallons Per Acre Per Day
GCCS Gas Collection and Control System
GCL Geosynthetic Clay Liner
GHG Greenhouse Gas
GM Geomembrane
HAP Hazardous Air Pollutant
HDPE High-Density Polyethylene
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HELP
HHE
IDNR
ITRC
IWCS
IWEM
KDHE
k
ki
kg
L
lb
LCRS
LCS
LDS
LFG
LFGTE
LLDPE
m2
m3
Hg
MCL
MDEP
MDL
mg
Mg
mm
MSW
MSWLF
NESHAP
NCSU
Hydrologic Evaluation of Landfill Performance
Human Health and the Environment
Iowa Department of Natural Resources
Interstate Technology and Regulatory Council
Innovative Waste Consulting Services, LLC
Industrial Waste Management Evaluation Model
Kansas Department of Health and Environment
Landfill Gas Decay Rate (year"')
Leachate Collection Decay Rate (year)
Kilogram
Liter
Pound
Leachate Collection and Removal System
Leachate Collection System
Leachate Detection System
Landfill Gas
Landfill Gas-to-Electricity or Energy
Linear Low-Density Polyethylene
Square Meter
Cubic Meter
Microgram
Maximum Contaminant Level
Massachusetts Department of Environmental Protection
Method Detection Limit
Milligram
Megagram
Millimeter
Municipal Solid Waste
Municipal Solid Waste Landfill
National Emission Standards for Hazardous Air Pollutants
North Carolina State University
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NMOC
Nonmethane Organic Compound
NO A A
National Oceanic and Atmospheric Administration
NRMRL
National Risk Management Research Laboratory
NSPS
New Source Performance Standards
NTU
Nephelometric Turbidity Unit
OH EPA
Ohio Environmental Protection Agency
ORD
Office of Research and Development
PCB
Polychlorinated Biphenyls
PCC
Post-Closure Care
PE
Polyethylene
ppmv
Parts Per Million by Volume
PVC
Polyvinyl Chloride
RCRA
Resource Conservation and Recovery Act
SCFM
Standard Cubic Feet per Minute
SEAs
State Environmental Agencies
SEM
Surface Emissions Monitoring
SMCL
Secondary Maximum Contaminant Level
SSE
Sum of Squared Errors
SST
Total Sum of Squares
STP
Standard Temperature and Pressure
S.U.
Standard Units
TDS
Total Dissolved Solids
TOC
Total Organic Carbon
UCL
Upper Confidence Limit
UDSHW
Utah Division of Solid and Hazardous Waste
US
United States
USGS
United States Geological Survey
VDEQ
Virginia Department of Environmental Quality
VLDPE
Very Low-Density Polyethylene
VTANR
Vermont Agency of Natural Resources
WADOC
Washington Department of Ecology
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WDEQ Wyoming Department of Environmental Quality
WWTP Wastewater Treatment Plant
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Executive Summary
Title 40 of the Code of Federal Regulations, 40 CFR Part 258 (herein referred to as §258) includes
standards for the design, operation, closure, post-closure care (PCC), monitoring and other
requirements for municipal solid waste landfills (MSWLFs or Subtitle D landfills) under the
authority of Subtitle D of the Resource Conservation and Recovery Act (RCRA). These
regulations require the owner or operator of each MSWLF unit to conduct PCC for 30 years unless
an extended or reduced period is demonstrated to be necessary or sufficient, respectively, to protect
human health and the environment (HHE). In addition to §258, the owners/operators of MSWLF
units may be required to meet additional regulations during the PCC period pertaining to the
management of landfill gas (LFG). These additional regulations include the New Source
Performance Standards (NSPS) for MSWLFs and the National Emission Standards for Hazardous
Air Pollutants (NESHAP). In general, MSWLFs required to meet these regulations must construct,
operate, and monitor an active LFG collection and control system (GCCS). The GCCS system
must be operated for at least 15 years (or sooner due to lack of adequate LFG flow per Subpart
XXX) and until the annual nonmethane organic compounds (NMOCs) emissions are less than 50
Mg/year (or 34 Mg/year if required to meet the requirements of 40 CFR 60 Subpart XXX).
RCRA Subtitle D PCC requirements (including the 30-year period) were promulgated in 1991 and
1992, and it is expected that several MSWLFs will reach or surpass a PCC period of 30 years in
the next 5-10 years. Unlike NSPS thresholds for terminating GCCS, RCRA Subtitle D regulations
do not specify numeric criteria for determining the PCC extensions or termination. This
determination is dependent on the current and potential HHE impacts of MSWLFs. The
availability of various site-specific data (e.g., in-place waste characteristics, LFG and leachate
collection rate and quality, groundwater quality, surface and subsurface gas emissions, closure cap
performance, landfill settlement) is vital for assessing HHE impacts of MSWLFs. This report
presents an assessment of the nature and prevalence of available data that can be used for
evaluating the HHE impacts of closed MSWLFs based on a review of data from nine closed
MSWLFs (or MSWLF with closed cells) located in different regions of the United States. It also
presents data gaps, approaches to identify the contaminants of potential concern and technical
approaches to estimate emission rates of these contaminants, and operating and monitoring
considerations for robust evaluation and mitigation of long-term HHE impacts of MSWLFs.
The analyses provided in this report are intended to be relevant to state environmental agencies,
MSWLF owners/operators, community decision-makers, and other stakeholders interested in
understanding approaches for assessing the HHE impacts of modification or termination of
MSWLF PCC. Nine sites that contained at least one MSW cell that has been closed for at least
five years, and has environmental monitoring records (e.g., groundwater monitoring, LFG
collection rate and quality, leachate collection rate and quality) and located in different climate
zones of the US were selected for detailed evaluation. The study cell(s) footprint ranged from 6 to
69 acres and contained 0.63 to 4.40 million metric tons of MSW, respectively.
The available actual and estimated cost data of different PCC activities were analyzed to develop
an understanding of the financial impacts associated with PCC. Study-cell specific PCC cost data
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were available only for three sites; the average annual PCC cost for the cells at these sites was
$5,300 per acre (in 2017 dollars). The average annual PCC cost for the remaining sites, for which
only site-wide cost data were available, was estimated to be $6,450 per acre (in 2017 dollars).
Landfill gas collection rate and methane content data were available for review from eight of the
nine sites. As expected, the methane collection rates from the closed sites exhibited a declining
trend, which is potentially attributed to the first-order decay kinetics of the anaerobic waste
decomposition process. Site-specific decay rates were estimated by best-fitting the first-order
waste decomposition rate equation, which is typically used for estimating the methane and
NMOCs generation rate from landfills to the measured methane collection rates. The annual
NMOCs collection rates - calculated based on the site-specific decay rate and methane collection
potential estimates and waste placement data - are estimated to decline below 34 Mg/year for
seven of the study sites within 30 years of closure. These findings may not be applicable to other
MSWLFs in the US as the study sites are smaller than approximately 75% of MSWLFs in the
United States.
In order to estimate GCCS operation timeframes for typical size MSWLFs, NMOCs and methane
collection rates were modeled for MSWLFs containing approximately 3.35 (small MSWLF), 7.85
(medium MSWLF), and 19.1 (large MSWLF) million metric tons of MSW, which correspond to
the 25th, 50th, and 75th percentile capacity of MSWLFs in the United States, respectively. Decay
rates ranging from 0.02 to 0.22 year were used to represent conditions from arid weather (slow
waste decomposition) conditions to bioreactor operation (fast waste decomposition). A methane
generation potential of 100 m3 per Mg waste, NMOCs content of 4,000 ppmv (as hexane), and a
gas collection efficiency of 100% was assumed for this analysis. The results suggest that annual
NMOCs collection rates for small MSWLFs located in arid to moderate precipitation areas are not
likely to decline below 50 Mg per year within 30 years after closure. NMOCs collection rates for
medium and large MSWLFs are not expected to decrease below 50 Mg per year within 30 years
after closure irrespective of location. The operation of MSWLFs to promote more rapid waste
stabilization (e.g., bioreactor landfill operation) has the potential to significantly reduce the
timeframes needed for annual NMOCs collection rates to decline below the NSPS threshold of 50
Mg/year and required GCCS operating timeframe.
The analysis also suggests that the mass-based remaining methane (Mg methane) and NMOCs
generation potential (Mg NMOCs) are more appropriate indicators of the HHE impacts than the
percent remaining methane generation potential, which are currently used by some state as a PCC
period evaluation criterion. The results also suggest that a potential increase in the decay rate can
increase the annual NMOCs collection rates above the NSPS threshold of 50 Mg per year after
GCCS operation termination if the in-place waste has considerable remaining NMOCs generation
potential; the decay rate was assumed to increase by 100% ten years after GCCS termination for
this analysis. The final cover, therefore, should be rigorously maintained even after GCCS
termination until the NMOCs generation potential and the leaching potential of the in-place waste
has declined to levels that are unlikely to pose a risk to HHE. In addition, landfill owners and
regulators should also continue surface and subsurface emissions and odor monitoring to
proactively identify signs of an increase in LFG generation rate after GCCS operation termination
and have provisions in place to resume GCCS operation, if needed, to control these issues.
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Leachate collection records were reviewed and analyzed to identify approaches for estimating
post-closure leachate collection rates at MSWLFs. In general, except for those landfill cells that
recirculated leachate after closure, the most recently-measured leachate collection rate from all of
the study cells (as of the time of this study) was less than 100 gallons per acre per day (GPAD).
None of the landfill cells examined exhibited trends indicative of having reached a steady-state
leachate collection/generation rate. Several sites exhibited a general declining trend in leachate
collection rates after closure. The hydraulic efficiency of the primary liner was estimated for four
of the study cells equipped with a double bottom liner system based on the available leachate
collection system (LCS) and leak detection system (LDS) flow rates. The sum of leachate flow
rates from the LCS and LDS was assumed to represent the leachate generation rate for this analysis.
The primary liner efficiency was estimated to range from 96.8% to 99.6%. Substantial differences
in several leachate indicator parameters (chloride, trace organics, total organic carbon) among
LCS, LDS, and groundwater suggest groundwater intrusion might be a significant source of liquids
collected from the LDS.
Three modeling methodologies were evaluated to estimate PCC leachate collection rates from each
of the cells of the study sites: first-order decay modeling, unsaturated flow modeling, and the
Hydrologic Evaluation of Landfill Performance (HELP) model. All three approaches provided
relatively similar approximations to measured leachate collection rates for a majority of the study
cells.
The comprehensiveness of available leachate quality data at the study sites was analyzed for the
parameters specified in the federal regulations for groundwater monitoring at Subtitle D landfills
(Appendix I and Appendix II of §258). In the current study, data for a total of 272 constituents
were evaluated to assess the leachate quality at eight of the study sites. The selected constituents
either have a primary or secondary maximum contaminant level (MCL or SMCL), or are listed in
40 CFR 258 Appendix I or II, or were used by EPA (2017a) for leachate quality evaluation of
Subtitle C landfills. Only three sites reported leachate constituent data for every Appendix I
parameter, and only one of the study sites reported at least one measurement for all but three of
the Appendix II parameters. The leachate characteristic data for less than half of the Appendix I
parameters were available for two study sites. More than half of the study sites reported leachate
constituent concentration data for ten or fewer Appendix II parameters (excluding Appendix I
parameters). The available data suggest that apart from the lack of the data for a large number of
groundwater monitoring parameters, the small number of measurements available at the study sites
may limit a reliable HHE impact assessment for several constituents.
A screening analysis was conducted to identify the contaminants frequently measured in leachate
above respective risk-based protection standards after closure to identify the contaminants of
potential concern. Parameters never measured above their risk-based standards are not expected to
present a risk to HHE in the future. The federal primary and secondary drinking water standards
were used as the thresholds for this evaluation. Fifteen out of 68 primary/secondary drinking water
parameters monitored at least once were measured at concentrations above the respective method
detection limit (MDL) in more than 50% of the samples. Six of these parameters (i.e., arsenic, total
dissolved solids (TDS), iron, manganese, chloride, and color) were measured above their
MCL/SMCL in more than 94% of the samples. Among all of the constituents with MCL, arsenic
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and turbidity were the only primary MCL parameters that consistently exceeded the MCL. Arsenic
exhibited a declining and an increasing trend at four and three study cells, respectively. The most
recent arsenic measurements at three of the sites, which exhibited a declining trend, were above
the MCL. A majority of organic compounds measured at the sites were undetected in most of the
samples. The analysis presented in the report is, by no means, represents a comprehensive
evaluation as it was limited to the constituents that were measured at the study sites and excluded
various emerging contaminants (e.g., PFAS) that may have HHE impacts.
It should be noted that the leachate quality is typically reflective of the decomposition status of the
bottom-most waste layer and does not necessarily represent the degree of stabilization of the entire
landfill. A well-decomposed waste layer above the LCS may attenuate the concentration of
parameters such as biological oxygen demand (BOD) and chemical oxygen demand (COD) that
are commonly used to assess leachate and waste stability. Moreover, a relatively lower
concentration of a large number of contaminants than the respective MCL/SMCL may also be due
to partial stabilization/mineralization of landfilled waste because of the lack of exposure to
adequate moisture. After the LCS operation termination, leachate might accumulate within the
landfill and eventually discharge into the environment as leakage through liner defects and/or side
slope seeps resulting from the leachate build-up. The stabilization/mineralization of the unstable
waste constituents and ensuing elevated contaminants concentrations in leachate after LCS
operation termination due to potential moisture intrusion into the landfill may pose a risk to HHE.
The impacts of any future moisture intrusion on potential leachate emissions can be evaluated and
mitigated by actively recirculating leachate or adding other liquids sources (e.g., stormwater) while
the site is actively monitored and maintained by owners/operators and regulators. The liquids
addition, however, has design, operating, and monitoring challenges such as unavailability of
moisture source especially in arid areas, the complexity of adding liquids to achieve uniform
moisture distribution in the landfill, flooding of gas collection devices, and a need to collect and
manage excess leachate at the end of bioreactor operation.
Groundwater quality data were reviewed and analyzed to identify impacts to groundwater for three
sets of parameters: those with MCL, those with SMCL, and for some parameters without MCL or
SMCL. Occasional MCL exceedances were observed following cell closure. Three sites exhibited
recurring groundwater exceedances. Due to observed exceedances above the respective MCL at
several upgradient wells, the recurring exceedances at two of these sites could not be conclusively
attributed to the lined cells. Monitoring data from the third site indicates that the elevated levels of
arsenic observed at downgradient wells may be a result of subsurface geochemical changes below
the liner system. The laboratory detection limit was greater than the respective MCLs for several
measurements (e.g., several organics at one site, arsenic and thallium measurements at another
site). Only one site with silver and zinc (only Appendix II parameters with SMCL) measurement
data had a single silver exceedance. As of the time of this evaluation, all three of the sites under
assessment monitoring and/or corrective action had instances where parameters without an MCL
(vanadium, 1,1-dichloroethane, and 1,4-dioxane) contributed to groundwater impacts; these
parameters had a state-specified risk-based standard.
A comprehensive review of the data collected from the study sites suggests several data gaps or
data quality issues that could hinder a robust and quantitative HHE impact assessment of PCC
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reduction or termination at the study sites. The ability for the owner and/or operator of an MSWLF
unit to demonstrate the protection of HHE following completion of PCC is dependent on having
quality environmental monitoring data specific to the MSWLF unit of interest. MSWLFs
owners/operators should consider collecting the following data that are important for assessing
HHE impacts:
1. Waste Tonnage and Composition - Waste-specific (e.g., MSW, industrial waste,
construction and demolition debris) disposal tonnages were available for several study
sites. Detailed waste characterization data (e.g., plastics, paper, food waste, household
hazardous waste etc.) estimated based on composition studies were available only for one
site. Both composition and tonnage information, especially of industrial waste, are valuable
for HHE impact assessment.
2. Leachate Quality Data - Although RCRA Subtitle D regulations do not require routine
leachate chemical characterization, leachate quality data are essential for the HHE impact
evaluation. MSWLF owners/operators should consider harmonizing the list of monitoring
parameters for groundwater and leachate for a comprehensive HHE impact assessment.
Leachate quality should be monitored for the required groundwater monitoring parameters,
constituents that occur at elevated levels in leachate (e.g., chloride, ammonia, total
dissolved solids), and those that are commonly used for assessing the waste stability (e.g.,
biochemical oxygen demand and chemical oxygen demand). Furthermore, monitoring of
emerging contaminants such as per- and polyfluoroalkyl substances and pharmaceuticals,
which have been reported to be present in leachate, should be considered for
comprehensive HHE impact assessment. The laboratory reporting limits of the monitored
parameters should be equal to or lower than the respective groundwater protection
standard.
3. Groundwater Monitoring System - A periodic review of changes such as surrounding
land use/zoning changes that can impact the groundwater flux and flow direction should
be considered while assessing the long-term impacts of modifying or terminating PCC.
Monitoring of groundwater quality with respect to leachate indicator parameters such as
chloride, ammonia, BOD, and COD should be considered.
4. Settlement - Differential settlement of the landfill surface represents one of the more
probable risks to the integrity of the final cover. The compromises in the final cover system
might result in moisture intrusion, which could subsequently increase gas and leachate
generation rates and the fugitive gas emissions. Routine settlement monitoring data can be
used to estimate the future settlement rate. In addition, settlement data, when used in
conjunction with a temporal analysis of LFG collection and leachate quality, can provide
an indication of waste stabilization status. Settlement data were not available for several
study sites. It should be noted that landfill owners routinely conduct topographic surveys
during landfill operating life to assess airspace usage and availability. Continuation of these
surveys after closure would provide valuable data for evaluating the waste stabilization rate
and the magnitude of total and differential settlement.
5. Monitoring Records - Some monitoring data (e.g., perimeter probes monitoring for
tracking subsurface gas migration and surface emissions monitoring for identifying
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fugitive LFG emissions), which are typically required to be routinely monitored for
MSWLFs, were not available for review/analysis for several study sites. MSWLFs
owners/operators and regulators should consider implementing documentation systems for
cataloging monitoring data for prompt retrieval and analysis.
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1. Introduction
1.1. Background
The disposal in landfills has been the predominant method of managing municipal solid waste
(MSW) generated in the United States (US), and approximately 139 million tons, or 52.1% of all
MSW, were landfilled in 2017 (EPA 2019a). Landfills receiving MSW are required to comply
with federal, state, and local, if applicable, regulations. The requirements found under Title 40 of
the Code of Federal Regulations (CFR), Part 258 (herein referred to as §258): Criteria for
Municipal Solid Waste Landfills (referred herein to as MSWLFs), specify the performance-based
design, operating, monitoring, closure and post-closure care (PCC) criteria for these landfills. The
PCC criteria require the operation, maintenance, and monitoring of MSWLFs for 30 years after
closure or as necessary to ensure that the MSWLFs do not pose a threat to human health and the
environment (HHE). The federal PCC criteria allow the reduction (or extension) of the 30-year
PCC period if the modified period is sufficient (or necessary) for protecting HHE. However,
currently, there is no federal guidance or specific direction on approaches that an MSWLF
owner/operator can use for making a demonstration supporting the reduction or termination of
PCC activities.
Many MSWLFs have closed since the promulgation of §258 in the early 1990s and are quickly
approaching the end of the 30-year PCC period. Figure 1-1 shows the number of MSWLFs closed
and the corresponding 30-year PCC timeline based on the closure date reported to the US
Environmental Protection Agency (EPA) (EPA 2019b) under the federal mandatory greenhouse
gas (GHG) reporting program, as described in Title 40 of the CFR Part 98 (40 CFR 98).
Closure Year
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Figure 1-2. Locations of All Closed MSWLFs Included in the EPA Greenhouse Gas Reporting
Program Database (EPA 2019b)
There are 330 MSWLFs that are listed as closed in the EPA Greenhouse Gas Reporting Program
database (EPA, 2019b) as of 2018 - the locations of these closed MSWLFs are presented in Figure
1-2. As a point of comparison, there are approximately 980 operating sites included in the EPA
greenhouse gas database. It should be noted that landfills presented in Figures 1-1 and 1-2 are only
a subset of the total number of closed MSWLFs in the US as only MSWLFs that emit GHG above
a regulations-specified amount are required to report data to the GHG reporting program. The
number of operating and closed landfills in the US is greater than the numbers presented above.
The Environmental Research and Education Foundation and the Interstate Technology and
Regulatory Council have published a performance-based methodology to assess ITHE impacts of
closed MSWLFs (EREF 2006, lTRC 2006). In addition, a few state environmental agencies
(SEAs) have developed criteria for evaluating the PCC period of closed MSWLFs. The
Association of State and Territorial Solid Waste Management Officials (ASTSWMO) conducted
a survey in 2011 to gather information on states' policies and/or regulations with respect to post-
closure requirements beyond 30 years for Subtitle C and D landfills (ASTSWMO 2013). The
survey showed considerable variation among the SEAs regarding the approaches, processes, and
procedures that could be used to adjust the PCC period. ASTSWMO recommended that EPA
develop guidance on the review and consideration for adjusting the PCC period for MSWLFs.
EPA (2017a) recently published a guidance document to evaluate the performance of Subtitle C
landfills under PCC. This report presents the application of various approaches and criteria for
evaluating long-term environmental emissions potential of closed Subtitle D landfills.
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. Objectives and Scope
The primary objectives of this study were to (1) assess the nature and prevalence of the design,
operation, and monitoring data available for MSWLFs that can be used for HHE impacts
evaluation, (2) present application of approaches that site owners/operators and engineers can use
to evaluate monitoring data to identify contaminants of potential concern (COPCs) and develop
potential emission rates of these contaminants to the atmosphere and groundwater/surface water,
(3) identify data gaps, and (4) present operating and monitoring considerations for MSWLFs
owners to evaluate and mitigate HHE impacts of MSWLFs. To provide real-world context, data
from nine MSWLF sites scattered across the US were compiled and evaluated to assess data
availability, demonstrate the use of the analysis approaches, and identify typical data gaps.
Specifically, this report discusses an evaluation of environmental monitoring data related to
leachate quantity and quality, landfill gas (LFG) subsurface migration, LFG surface emissions,
LFG quantity and quality, groundwater monitoring data, and the final cover system settlement and
maintenance that can be used to assess the HHE impacts of MSWLFs. The PCC cost data for the
study sites were also reviewed to understand the cost associated with PCC activities and PCC
extension beyond the 30-year period. Several of the approaches proposed by EREF (2006), ITRC
(2006), and EPA (2017a) were used for analyzing data from these study sites.
The EPA ORD collected, reviewed, and analyzed environmental monitoring data from nine
MSWLF sites with closed cells. Sites that were selected had at least one closed cell with a
containment system (liner and cap) that included a geomembrane, had/have an active gas
collection system, and had monitoring data/records readily available. Initially, EPA attempted to
identify sites with cells that had been in PCC for a minimum of 10 years. However, due to the
scarcity of closed MSWLFs with all the monitoring data listed above, sites closed before 2010
(with at least five years of PCC data) were also considered for the study.
It should be noted that PCC, as discussed with respect to cells located at the nine MSWLF sites
selected for this study, refers to the period after the cell had been capped (with a geomembrane)
and no longer received waste. The final cover of the study cells may not necessarily meet the
regulatory definition of "closed" (e.g., one site has an exposed geomembrane cap, which only
meets the state's definition of an intermediate closure system); the cover systems evaluated in this
study may be permitted as a final or intermediate cover system.
This study primarily focuses on the requirements of 40 CFR Part 258: Criteria for MSWLFs.
Additional federal regulations that MSWLFs may be required to comply with are the Standards of
Performance for MSWLFs (i.e., 40 CFR Part 60 Subpart WWW and XXX) and the National
Emission Standards for Hazardous Air Pollutants (NESHAP) (i.e., 40 CFR Part 63, Subpart
AAAA). These requirements are discussed in more detail in the next chapter of this report. It
should be noted that MSWLFs operation and monitoring is regulated by the states (with EPA-
approved regulatory programs) requirements, which may be more stringent than the federal
regulations.
The data evaluation and analysis approaches presented in this report are expected to be useful to
all stakeholders, including site owners, operators, regulators, and engineers associated with
MSWLFs design, permitting, operation, monitoring, closure, and PCC activities. The evaluations
presented in the report have limitations and do not represent a comprehensive HHE risk
assessment. The stakeholders should critically evaluate the appropriateness of the assumptions and
limitations of the analysis presented in the report before electing to use these for their sites.
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Furthermore, the application of approaches for evaluating the study sites is imperfect due to the
gaps in LFG, leachate, and groundwater data available for the study sites. This report is not meant
to serve as a standalone manual for conducting a comprehensive assessment of HHE risks
associated with closed RCRA Subtitle D landfills for evaluating appropriate PCC duration.
1 ort Organization
This report is organized into ten chapters. Chapter 1 introduces the objectives and scope of the
study and describes the organization of this report. Chapter 2 describes the federal regulations
pertaining to PCC of MSWLFs and summarizes several sets of state and industry guidance for
evaluating and modifying PCC at MSWLFs. The study sites' selection criteria, the types of data
collected, and a summary of the key design and operating features of each of the cell(s) at the nine
MSWLFs chosen for this study are presented in Chapter 3. Chapter 4 summarize the cost of
different PCC activities at the study sites. Chapter 5 presents a detailed review of site LFG data
and the methane monitoring probes surrounding the sites, and example approaches to estimate
long-term gas emissions potential of a site at any given point of time. Chapter 6 summarizes
temporal trends in leachate quantity and quality from the sites and describes approaches to estimate
long-term leachate generation rate and to identify the COPCs with respect to HHE. Groundwater
monitoring data from the sites are reviewed in Chapter 7. Chapter 7 also presents approaches to
assess sources of impacts to groundwater quality around the MSWLF unit. Chapter 8 presents an
evaluation of settlement and surface emissions monitoring (SEM) data. A summary of the key
findings of this study and operating and monitoring considerations for MSWLF's stakeholders are
provided in Chapter 9, along with recommendations for PCC data collection. A list of the data and
information sources used in the development of this report is included in Chapter 10.
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2. Background
The EPA promulgated minimum national criteria under Subtitle D of the Resource Conservation
and Recovery Act (RCRA) for MSWLFs codified under §258 (referred herein to as §258 or
Subtitle D or MSWLF regulations). MSWLF units that ceased receiving waste by October 9, 1991,
were exempted from these criteria. The MSWLFs that received flood-related waste or received
less than 100 tons of waste per day were also exempted from all these criteria except the final
cover requirements if they stopped accepting waste on or before April 9, 1994 (or October 9, 1994,
under certain conditions). The regulations also conditionally exempt another subset of MSWLFs
from some of the requirements. For example, MSWLFs accepting 20 tons of waste per day or less
(based on annual average) are conditionally exempt from the design and groundwater monitoring
criteria. Because these MSWLFs are exempted from requirements such as liner design,
groundwater monitoring, and PCC, these MSWLFs were not considered for review in this study.
This chapter summarizes the federal requirements promulgated in §258 for the design, operation,
monitoring, closure, and PCC of MSWLFs. For a detailed description of Subtitle D regulations,
readers should refer to the CFR, which is electronically available at https://www.ecfr.gov. The
requirement not specific to PCC are also presented as many of these must be met throughout the
PCC period, and a number of states require compliance with these requirements as a starting point
to evaluate an adjustment to the frequency of PCC activities.
lion and Design
Subpart C of MSWLF regulations (§258.20-§258.29) lists the operational criteria of an MSWLF
unit. The following is a summary of the operational criteria:
1. The concentration of methane gas generated at the facility shall not exceed 25% of the
lower explosive limit for methane in the facility structures (excluding gas collection and
control system (GCCS) components) and shall not exceed the lower explosive limit at the
facility boundary. The criteria require routine monitoring to ensure compliance.
2. Landfill owners or operators shall meet the applicable standards developed under a State
Implementation Plan pursuant to section 110 of the Clean Air Act. The open burning of
waste (with limited exceptions) is prohibited.
3. Landfill owners or operators shall have a run-on and run-off control system to manage
stormwater run-off resulting from a 24-hour 25-year storm event and shall prevent the
discharge of pollutants into waters of the US.
4. Bulk or noncontainerized liquid waste may not be placed in the MSWLF unless it is
household waste (other than septic waste), or it is leachate or gas condensate from the
MSWLF and the MSWLF has a composite liner and leachate collection system.
5. All the records pertaining to design, operation, inspection, training, notification, PCC plan,
financial assurance, etc. must be retained near the facility or as approved by the state
authority.
MSWLF design criteria are listed in Subpart D of §258. A new MSWLF or lateral expansion of
an MSWLF shall have a composite liner with a leachate collection system that can maintain less
than a 30-cm depth of leachate over the liner. The composite liner shall contain a minimum 30-
mil thick flexible membrane liner (60-mil thick if high-density polyethylene) overlain by a two-
foot thick layer of compacted soil with hydraulic conductivity no more than lxlO"7 cm/sec. The
federal regulations do not specifically require a leak detection system for MSWLFs. Alternative
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liner designs, which ensure that the concentration of specific parameters would not exceed their
maximum contaminant level (MCL) in the uppermost aquifer at the point of compliance, are
allowed. The parameters and their MCL are presented in Table K-l of Attachment K.
2.2. Monitoring
Subpart E of §258 provides requirements for groundwater quality monitoring at MSWLFs. A
groundwater monitoring system shall be installed with a sufficient number of wells at appropriate
locations and depths to evaluate the impact of the MSWLF unit on the uppermost aquifer.
Groundwater monitoring is performed under a detection monitoring (§258.54) or assessment
monitoring (§258.55) program. The detection monitoring shall be performed at all the groundwater
monitoring wells for the parameters listed in §258 Appendix I (hereafter referred to as 'App I').
App I has a total of 62 parameters consisting of 15 inorganics and 47 organic parameters. The
regulation allows the states to modify this list of parameters. Unless a demonstration is made
showing the need for an alternative monitoring frequency, detection monitoring is required at least
semi-annually during the MSWLF's active life and the PCC period.
Assessment monitoring is required when a statistically significant increase over background
concentrations has been detected for any parameter listed in App I or a state-approved alternative
list. The groundwater samples must be analyzed for all the constituents listed in Appendix II of
§258 (hereafter referred to as 'App II') within 90-days of starting an assessment monitoring
program and annually thereafter. The regulations allow the states to modify the list of assessment
monitoring parameters. App II has a total of 215 parameters; all App I parameters are also listed
in App II. App I and II parameters are listed in Table J-2 of Attachment J.
Within 90 days of finding that a constituent listed in App II is exceeding the groundwater
protection standard as defined under §258.55 (h) or (i), an assessment of corrective measures and
remedial action is required. If all App II parameters are measured at or below background values
for two consecutive sampling events, the owner or operator may return to detection monitoring.
Additionally, §258.23 requires the quarterly monitoring of methane concentrations in site
structures and at the property boundary. The methane gas shall not accumulate at levels equal to
or more than 25% of the lower explosive limit in site structures and shall not exceed the lower
explosive limit at the property boundary. Subsurface methane monitoring locations must be
selected around the periphery of the site at appropriately selected depths based on the site
hydrogeologic and hydraulic conditions. The landfill owners shall implement a remediation plan
for the methane gas releases if methane is detected above the levels mentioned above.
2.3. CI cm mi ' nil'" II 1 i II 1;m ' ' -ire
Closure and PCC criteria and requirements are listed in Subpart F of §258. Based on the closure
criteria as listed in §258.60, MSWLFs are required to be capped with a final cover system designed
to minimize infiltration and erosion. The final cover shall have permeability less than or equal to
the permeability of any bottom liner system or natural subsoil present, or a permeability no more
than lxlO"5 cm/sec, whichever is less. The final cover design should include an earthen material
infiltration layer with a minimum 18-inch thickness, and a 6-inch earthen material layer capable
of sustaining native vegetative growth. The Director of an approved State may approve an alternate
final cover design that achieves an equivalent reduction in infiltration and equivalent protection
from wind and water erosion. The regulations require the preparation of a closure plan that lists
the steps necessary to close all MSWLF units at any point during their active life.
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Unless the Director of an approved State accepts an alternative PCC period (based on a
demonstration by the owner or operator of an MSWLF unit that the alternative PCC period is
sufficiently protective of HHE), §258.61 requires the MSWLFs owner or operator to perform PCC
for 30 years following the closure. Specifically, the PCC of each MSWLF unit must be performed
to:
1. Maintain the integrity and effectiveness of the final cover. This includes repairing the cover
as needed to address the effects of settlement, subsidence, erosion, or other events and
prevent the final cover system erosion and damage from stormwater run-on and run-off;
2. Maintain and operate the leachate collection system in accordance with §258.40. Leachate
management can be discontinued if it is demonstrated that the leachate no longer poses a
threat to HHE;
3. Monitor groundwater in accordance with §258 Subpart E and maintain the groundwater
monitoring system;
4. Maintain and operate the explosive gas monitoring system in accordance with §258.23.
In addition to these operation and monitoring requirements, the landfill owners may be required to
implement an active gas collection and control system per the New Source Performance Standards
(NSPS). The NSPS operations and monitoring requirements are presented in Section 2.5. The
owner or operator must prepare a PCC plan and place it in the operating record no later than the
initial receipt of waste (or October 9, 1993 - whichever was later). At a minimum, the PCC plan
must include:
1. A description of monitoring and maintenance activities and the associated frequency for
each MSWLF unit;
2. Contact information of the person or office to contact during PCC period;
3. A description of the planned use of the property during PCC.
Following the completion of the PCC period, the State Director shall be notified by the owner or
operator that a certification (signed by an independent professional engineer or approved by the
director of an approved state) verifying that the PCC has been completed has been placed in the
operating record.
incial Assurance
Subpart G of §258 provides requirements for the financial assurance for closure, post-closure care,
and corrective action, if applicable. The owner or operator is required to have a detailed written
estimate (in current dollars) of the cost of hiring a third party to close the largest area of all MSWLF
units ever requiring a final cover (as required under §258.60) at any time during the active life in
accordance with the closure plan (§258.61).
The owner or operator is also required to have a detailed written estimate, in current dollars, of the
cost of hiring a third party to conduct PCC for the MSWLF unit in compliance with the post-
closure plan developed under §258.61 of this part. The PCC estimate used to demonstrate financial
assurance must account for the total costs of conducting PCC, including annual and periodic cost,
as described in the post-closure plan over the entire PCC period.
The cost estimates must be based on the most expensive closure and PCC costs and must be
annually adjusted for inflation. The estimated costs must be increased if the changes in closure or
post-closure care plan increase the maximum costs. The owner or operator of each MSWLF unit
must establish financial assurance for closure and PCC and must provide continuous coverage
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until released from financial assurance requirements. The regulation also specifies the allowable
mechanisms such as trust fund, surety bond, insurance for demonstrating financial assurance.
lions
NSPS for MSWLFs are applicable to all MSWLFs that commenced construction, reconstruction,
or modification on or after May 30, 1991. The NSPS are promulgated under 40 CFR §60, Subpart
WWW. MSWLFs that commenced construction, reconstruction, or modification after July 17,
2014 must also meet the additional standards promulgated under 40 CFR §60, Subpart XXX. In
general, the NSPS describe the following:
• Whether an MSWLF is required to install a GCCS and conduct SEM
• Design of the GCCS and the parts of the MSWLF that the GCCS needs to collect gas
from
• Standards for operating and monitoring the GCCS
• Standards for conducting SEM
• Timeline of GCCS operation termination
• Approved test methods for meeting numerical standards
• Provisions and deadlines to remain in compliance if standards are not met
• Reporting and recordkeeping requirements
Under 40 CFR §60.752 (b)(2)(v), the NSPS includes specific provisions for when the GCCS can
be capped, removed, or decommissioned. For this to occur, all the following conditions must be
met:
• The MSWLF must be closed.
• The GCCS must have been in operation for at least 15 years (or shown that it would be
unable to operate in 15 years due to declining gas flow if meeting the requirements of
Subpart XXX).
• The non-methane organic compound (NMOC) emission rate must be less than 50
megagrams per year (or 34 megagrams per year if meeting the requirements of Subpart
XXX) on three successive test dates. Tests entail collecting and analyzing samples of
LFGto determine the concentration ofNMOCs.
NESHAP for MSWLFs are promulgated under 40 CFR §63, Subpart AAAA, and are applicable
to MSWLFs that accepted waste after November 8,1987, or that have additional capacity to accept
waste and meet any one of the following:
• Are considered a major source or are collocated with a major source of hazardous air
pollutants (HAPs)
• Are required to follow NSPS
• Are operated as or including a bioreactor that has a design capacity greater or equal to
2.5 million megagrams and 2.5 million cubic meters of waste, and that was not
permanently closed as of January 16, 2003
As generally described in §63, a major source is a stationary source that annually emits ten or more
tons of any individual HAP or 25 or more tons of any combination of HAPs. A full list of the 187
8
-------
HAPs is included in Section 112(b) of the Clean Air Act. A bioreactor landfill is defined
(§63.1990) as "a MSWLF where any liquid other than leachate (leachate includes landfill gas
condensate) is added in a controlled fashion into the waste mass (often in combination with
recirculating leachate) to reach a minimum average moisture content of at least 40 percent by
weight to accelerate or enhance the anaerobic (without oxygen) biodegradation of the waste."
The MSWLFs that must meet the requirements of NSPS would typically also need to meet the
requirements of Subpart AAAA. However, Subpart AAAA also requires that MSWLFs operated
as or including a bioreactor with a design capacity greater or equal to 2.5 million megagrams and
2.5 million cubic meters of waste (and which were not permanently closed as of January 16, 2003)
to also comply with NSPS, irrespective of NMOCs emissions.
Besides requiring MSWLFs operated as or including a bioreactor to comply with NSPS, Subpart
AAAA also requires the development of a Startup, Shutdown, and Malfunction Plan for the gas
control system for all MSWLFs that must meet the requirements of NSPS.
As described in §63.1992(b), MSW bioreactor landfills do not need to comply with Subpart AAAA
if they are a closed landfill, have permanently ceased adding liquids to the bioreactor and have not
added liquids to the bioreactor for at least one year. The owners or operators of MSWLFs are not
required to comply with the requirements of Subpart AAAA once they are no longer required to
apply controls as specified in 40 CFR §60.752(b)(2)(v) of Subpart WWW (§63.1950).
<' II rM ii i -,, mlidc ii 1 ii II II 'i i-v.! II i> 'Illation
2.6.1. Overview
PCC at MSWLF sites is conducted according to the general conditions specified in §258.61, which
include operating, maintenance and monitoring requirements for the final cover, leachate
collection, groundwater monitoring, and gas monitoring systems. The regulatory default PCC
period is 30 years. However, once the MSWLF owner or operator demonstrates that the site no
longer poses a significant threat to HHE, the owner or operator can present a request for a reduction
in the frequency of PCC monitoring and maintenance activities, or for early PCC termination, to
an approved State Director. The 30-year period included in §258 was adopted from the Subtitle C
PCC requirement that was based on an EPA estimate that".. .it might take as long as 30 years for
material leaching from hazardous wastes to migrate to groundwater..." (ITRC 2006). EPA (2017a)
recently published a guidance document to evaluate the performance of Subtitle C landfills under
PCC.
While the federal regulations allow the reduction (or extension) of the PCC period if the owner or
operator can demonstrate the protection of HHE, detailed processes and specific criteria/provisions
that can be used to make such demonstrations are not provided in the federal regulations. The
Association of State and Territorial Solid Waste Management Officials (ASTSWMO) conducted
a survey in 2011 to gather information on states' policies and/or regulations with respect to post-
closure requirements beyond 30 years for Subtitle C and D landfills (ASTSWMO 2013). The
survey showed considerable variation among the SEAs regarding the approaches, processes, and
procedures that could be used to adjust the PCC period. ASTSWMO recommended that EPA
develop guidance on the review and consideration for adjusting the PCC period for MSWLFs. The
ASTSWMO survey data suggested a lack of experience pertaining to adjusting or concluding the
PCC period; only three out of 26 SEAs that responded to the survey had an MSWLF that had been
in PCC for at least 30 years. Only fiveSEAs reported having established specific and two SEAS
9
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reported having mandatory criteria in the state regulations that may be used for extending and
reducing the PCC period, respectively. Nine additional states reported publishing specific PCC
period evaluation criteria as guidance, policy, or other types of documents.
The Environmental Research and Education Foundation (EREF 2006) and the Interstate
Technology and Regulatory Council (ITRC 2006) have proposed an iterative and modular
performance-based methodology (also referred to as functional stability approach) to evaluate
closed MSWLFs. The ITRC (2006) approach appears to be similar to that of EREF (2006). The
approach entails a modular assessment of the leachate collection and control system, the GCCS,
the groundwater monitoring system, and the cap system; and the associated pathways that these
systems can impact HHE. The impacts to HHE associated with a less-frequent maintenance and
monitoring program (or even termination) of each of these four systems are sequentially evaluated
to assess if the proposed modifications provide sufficient protection to HHE. Following the
implementation of the change(s), the approach calls for confirmatory and surveillance monitoring
to verify that such modifications do not result in an inadvertent risk to HHE. This approach of
incremental reduction in PCC monitoring program does not rely on a complete stabilization of
waste and assume that some de minimis level of control such as maintenance of the final cover
would continue beyond the PCC to manage HHE impacts at the point of exposure.
The complete stabilization approach (also referred to as organic stability approach) of long-term
landfill management, on the other hand, necessitates monitoring of the landfill until it is completely
stable with respect to chemical, biological, and physical characteristics containment system
(Morris and Barlaz 2011). Once a landfill is completely stable, a failure of the containment system
would not result in HHE impacts. Based on an analysis of the monitoring data from a closed
MSWLF in NY, O'Donnell et al. (2018) reported that the organic stability approach would require
longer LFG and leachate management period than the functional stability approach. Morris and
Barlaz (2011) indicated that this approach might not be practical for managing closed landfills due
to factors such as the lack of a mechanism for ammonia transformation in the anaerobic
environment of the landfill and the potential presence of trace organic compounds.
Published guidelines for the termination of PCC at MSWLFs for eight states were publicly
available as of 2018. These state guidance documents are summarized in the following sections,
arranged by the environmental containment systems most commonly evaluated during PCC, to
provide a sense of the type and specificity of the available guidelines. In general, these guidelines
provide criteria pertaining to monitoring data, duration and frequency of data collection, and trends
needed for PCC reduction or termination application.
2.6.2 "hate Management
Five of the eight state guidance documents identified have criteria for evaluating leachate quantity
(FDEP 2016; VDEQ 2007; UDSHW 2012; WADOC 2011; WDEQ 2000). Three of these specify
no leachate generation or leachate generation at a historically low rate as a criterion for PCC
termination (FDEP 2016; UDSHW 2012; WDEQ 2000). One state's guidance advises that leachate
should not be produced for five years prior to petitioning to end PCC (WDEQ 2000). Most
guidance documents suggest the review of historical leachate collection records for proposing the
termination of PCC. Three state guidance documents suggest the review of biological and chemical
oxygen demand in the leachate (FDEP 2016; IDNR 2016; KDHE 2014). Two of these state
documents also recommend the review of ammonia and total suspended solids leachate
concentrations (IDNR 2016; KDHE 2014).
10
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The duration of leachate quality records recommended for inclusion in the application to terminate
PCC varies among the guidance documents. Three of the eight states do not specify a leachate
monitoring duration in their PCC termination guidance (VDEQ 2007; WADOC 2011; WDEQ
2000). The state with the most extensive leachate quality records review recommends the review
of three consecutive five-year demonstration periods that show key parameters of leachate quality
to be in equilibrium or decreasing (KDHE 2014). One state suggests ten years (three years of data
before closure and seven years of data after closure) of annual sampling records before PCC
termination (FDEP 2016). Two state guidance documents recommend presenting at least five years
of monitoring data (IDNR 2016; UDSHW 2012), and one of these documents also recommends
an additional five years of monitoring records if the leachate collection system is decommissioned
before terminating PCC (IDNR 2016). Two state guidance documents do not provide a specific
leachate quality record duration (VDEQ 2007; WDEQ 2000).
2.61 ttfill Gas Management
Similar to the guidance on leachate monitoring, the duration and frequency of landfill gas
measurement suggested for the PCC termination petition varies among the state guidance
documents. Three of the eight guidance documents reviewed list the gas monitoring timelines the
same as for leachate characterization (FDEP 2016; IDNR 2016; KDHE 2014). One state guidance
document recommends the review of five years of gas monitoring data (VTANR 2013), while
another suggests eight consecutive sampling events (WADOC 2011). One state's guidance
provides the option of including either twelve consecutive months or three consecutive years of
quarterly methane monitoring data in the petition for PCC termination (WDEQ 2000). Two states
do not specify gas monitoring data periods in their PCC termination guidance (VDEQ 2007;
UDSHW 2012).
Seven of the eight state guidance documents reviewed suggest that the MSWLF operator provides
evidence that the landfill gas generation rate has either stabilized or shown a declining trend in
methane production (FDEP 2016; IDNR 2016; KDHE 2014; UDSHW 2012; VTANR 2013;
WADOC 2011; WDEQ 2000). One of these state guidance documents specifically suggests
achieving methane production that is below 10% of the peak rate flow rate or achieving remaining
methane generation potential of less than 10% of the total generation potential before considering
changes to the active LFG system (FDEP 2016). Five documents also mention that methane
concentrations in structures built on the landfill site be below 25% of the lower explosive limit of
methane to avoid explosion and toxicity risks (VDEQ 2007; UDSHW 2012; VTANR 2013;
WADOC 2011; WDEQ 2000).
2.6.4. Groundwater Monitoring
MSWLF operators attempting to reduce the frequency of or termination of PCC should show a
history of compliance with groundwater quality standards when petitioning to the approved State
Director. In the collected guidance documents for eight states, three states recommend that the
operator provides five years of groundwater monitoring records after closure (FDEP 2016; IDNR
2016; VTANR 2013), and one additional state recommends a three-year monitoring period (VDEQ
2007). The remaining four states do not specify a timeline, though one state's guidance suggests
the MSWLF operator can terminate groundwater monitoring once records of environmental
monitoring and control systems have demonstrated the facility closure is protective of HHE
(WDEQ 2000). All state guidance reviewed suggests no significant risk to HHE should be present
at the site, and parameters of concern must be below existing standards. One state guidance
11
-------
document specifies that concentrations should be less than one half the state's groundwater
protection standards (FDEP 2016). Four other state guidance documents recommend that a
decreasing trend in parameter concentrations should be observed (IDNR 2016; VDEQ 2007;
VTANR 2013; WADOC 2011). Based on a national survey, ASTSWMO (2013) reported that 25
states (out of 27 states responded to survey) do not allow termination of PCC if corrective action
is still necessary at the site.
2.6.5. Final Cover Management and Maintenance
The final cover integrity and performance would be dependent on the magnitude of its settlement
and differential settlement. Excessive differential settlement may result in final cover irregularities
that can enhance its erosion and/or cause stormwater ponding, which would increase the potential
for moisture percolation into the landfill. Of the eight state guidance documents describing the
content of applications for reducing or terminating PCC, one state recommends including ten years
of settlement monitoring data (FDEP 2016). Three other states suggest collecting and analyzing
settlement data for at least five years (IDNR 2016; UDSHW 2012; WDEQ 2000). Four state
guidance documents specify a qualitative criterion of very low or negligible final cover settlement
rate for PCC termination ( IDNR 2016; UDSHW 2012; VTANR 2013; WADOC 2011). FDEP
(2016) recommends achieving an annual settlement rate of less than 5% of the total post-closure
settlement before reducing or eliminating the final cover maintenance. Additional final cover
considerations include qualitative assessments of vegetation and cover membrane integrity (FDEP
2016; WDEQ 2000).
12
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3. " II " II vl i" 'll IIIII 1 v! < ' ' '! I ' I« '
i Hi":11, H ilectk-h Ti
The following data sources were used to identify closed MSWLF:
1. Greenhouse gas reporting database. The website https://www.epa.gov/ehgreporting lists
all the MSWLF sites in the US that reported GHG estimates to EPA. Based on the
closure date listed in this database, approximately 300 MSWLFs were closed from 1993
through 2013.
2. State environmental agencies (SEAs) websites. Several SEAs maintain a list of closed
MSWLF sites. For example, the Florida Department of Environmental Protection (FDEP
2018) and the Massachusetts Department of Environmental Protection (MDEP 2018)
provide a list of all the active and inactive MSWLF sites along with details such as
contact information, amount of waste deposited, and closure year.
3. EPA regional offices, facility engineer contacts, and institutional knowledge of project
team members.
The scope of the study was limited to the evaluation of the data from only nine closed sites due to
time constraints. The following criteria were used for selecting nine sites (referred herein to as the
study sites) with closed MSWLF cells from over 300 closed MSWLFs identified from the sources
listed above. The site search was ended once nine sites meeting the following criteria were
identified:
1. Active GCCS - Only sites with active GCCS were selected for the study.
2. Leachate collection and quality data - Sites with routinely-tracked leachate collection
rates and chemical characterization data were selected.
3. Geographic location - Due to variation in the average annual precipitation, temperature,
and evapotranspiration rates, the geographic location of a landfill is expected to have a
significant impact on leachate and gas generation rates. A 2017 EPA study of Subtitle C
landfills assessed sites located in four geographical regions of the US: Northeast,
Northwest, Southeast, and Southwest (EPA 2017a). An attempt was made to select
MSWLF sites from each of these four regions in the same proportion as the regional
population to achieve an approximate regional representation.
4. Liner Type -An evaluation of the long-term leachate generation rate was one of the
primary objectives of this study. The leachate collection rate from the sites with an
impervious liner component are expected to be more representative of the actual leachate
generation rate than the sites without a geomembrane liner. Therefore, only sites/cells
with a bottom liner configuration that included a geomembrane were selected.
5. PCC Data Duration - Initially, sites closed before 2005 were targeted for selecting sites
with at least ten years of PCC data as long-term data are vital for meaningful analysis of
trends over time. However, due to scarcity of closed MSWLFs meeting all of the above
criteria, sites closed before 2010 (with at least five years of PCC data) were also
considered for the study.
13
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3.2. Data led
For each of the nine study sites selected for detailed evaluation, information was gathered on the
design of the bottom liner and final (or intermediate) cover system of the cell(s) of interest. The
data pertaining to the hydrogeography and topography of the site were also collected. These data
were typically available through site construction permit applications, site environmental impact
assessments, and construction quality assurance documentation.
Table 3-1 presents a summary of the environmental monitoring and operational data that were
collected and reviewed for each site. An "X" indicates that a substantially complete dataset was
obtained, a "P" indicates that the dataset was extremely limited or missing key pieces of
information, and an "N/A" indicates that the dataset is not applicable to the site. It should be noted
that MSWLFs commonly accept non-MSW materials (beyond commercial or residential MSW)
that may pose a different level of HHE risk than MSW (e.g., yard waste, some industrial wastes,
land clearing debris, waste-to-energy or coal ash). An understanding of the amounts and nature of
waste materials such as industrial waste and contaminated soils deposited in the landfill is valuable
for HHE impact assessment. Waste-specific (e.g., MSW, industrial waste, construction and
demolition debris) disposal tonnages were available for several study sites. However, detailed
waste characterization data (e.g., plastics, paper, food waste, household hazardous waste etc.)
estimated based on composition studies were available only for one site (Table 3-1).
The cell-specific data were analyzed where available. The data from all the cells were used for
the analysis presented in the report for Site H as all the cells at the site met the liner criteria listed
in Section 3.1. For Site C, the leachate collection rate and quality for Cells CI and C2 were
individually available for analysis, whereas only cumulative LFG flow rate and composition data
were available from these cells. The leachate flow rate and composition were for individual cells
CI and C3 were analyzed. Only the collective LFG data for these were analyzed. Similarly,
leachate and LFG data from the entire site, including the cells that did not meet the liner criterion,
were analyzed for Site E as the study-cell-specific data were not available. For Site G, leachate
collection rate and post-closure cost data were available and analyzed for individual study cells
(G3 and G3). However, only cumulative LFG flow rate data were available from several cells
(including G3 and G4) and used for the analysis for Site G. Table 3-2 summarizes the data
available/analyzed for each site. More details about the available data and corresponding cells are
presented in the report chapters and Attachments A-I.
14
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Table 3-1. Environmental Monitoring and Operational Data Collected for Each Study Site
Data Type
Site
Category
Subcategory
A
B
C
D
E
F
G
H
I
Landfill Gas
Landfill Gas Flow
X
X
X
X
X
X
X
X
No
Landfill Gas Composition
X
X
X
X
X
X
X
P
No
Surface Emissions
Monitoring
No
No
p
X
X
X
No
X
X
Methane Monitoring Probes
X
X
No
X
X
X
No
X
X
Leachate
Collection Rate
X
X
X
X
X
X
X
X
X
Quality
X
X
X
X
X
X
P
X
X
Recirculation Quantity
X
N/A
X
N/A
X
N/A
N/A
N/A
No
Groundwater
Monitoring Wells Locations
and Groundwater Quality
Data
X
X
X
X
X
X
X
X
X
Final Cover
Topographic Survey or
Settlement Data
No
P
X
No
p
p
No
No
X
Maintenance Issues
P
P
p
X
X
X
No
No
X
Operations
Disposal Quantity
X
P
X
X
X
X
P
X
X
Waste Composition
No
No
No
X
No
No
No
No
No
Post-Closure Care
Actual Costs
No
X
No
X
X
No
No
No
No
Notes:
X- a substantially complete dataset was obtained;
P- dataset is extremely limited or missing key pieces of information;
N/A- dataset is not applicable to the site.
No - dataset not available
Table 3-2. Environmental Monitoring and Operational Data Analyzed for Each Study Site
Data Type
Site
Category
Subcategory
A
B
c
D
E
F
G
H
I
Landfill Gas
Landfill Gas Flow
S
S
c
S
s
s
S
C
s
Landfill Gas Composition
S
S
c
s
s
s
s
C
s
Surface Emissions
Monitoring
No
No
c
s
s
c
No
c
c
Methane Monitoring Probes
S
S
No
s
s
s
No
c
c
Leachate
Collection Rate
c
S
c
c
s
c
C
c
c
Quality
c
s
c
c
s
c
s
c
c
Recirculation Quantity
c
N/A
c
N/A
s
N/A
N/A
N/A
No
Groundwater
Monitoring Wells Locations
and Groundwater Quality
Data
s
S
c
c
s
c
C
c
c
Final Cover
Topographic Survey or
Settlement Data
No
P
c
No
p
p
No
No
c
Maintenance Issues
P
P
c
C
p
c
No
No
c
Operations
Disposal Quantity
c
c
c
c
c
c
C
C
c
Waste Composition
No
No
No
c
c
No
No
No
No
Post-Closure Care
Costs
S
S
s
s
s
s
C
C
c
Notes:
S- sitewide/multiple cell data were analyzed as the study cell-specific data were not available;
C- study cell-specific data were available and analyzed;
P- dataset is extremely limited or missing key pieces of information;
N/A- dataset is not applicable to the site.
No - dataset not available
15
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3.4. Des ii ii | i iions of Self ¦ i ¦'¦! ¦ 11 ¦'
The area of the cell(s) selected from each site ranged from 6-69 acres, and five of the sites had
standalone lined cells (i.e., the lined cell(s) was not piggybacked over an unlined cell(s)).
Additional information on the capacity, types of waste accepted (i.e., in addition to MSW), design
of the final cover and bottom liner system, annual precipitation, and annual rainfall for each of the
cell(s) is provided in Table 3-3. Five of the study sites are publicly owned and operated, and the
rest four are privately owned and operated.
Material-specific disposal amounts were available for several sites (e.g., Site G, E, and I). The
materials listed in the Table 3-3 are major waste categories that were accepted at each of the study
sites as explicitly described in the site documents reviewed - they are not intended to be an
exhaustive list of the materials received at each site. The "composite liner" description, unless
otherwise noted, refers to a 60-mil high-density polyethylene (HDPE) geomembrane (GM) in
direct contact with an underlying natural or geosynthetic clay layer.
Annual precipitation data was derived from the closest weather station to each site - the values
represent an average of annual station precipitation data collected from 1981 to 2010 (NOAA 2017
http://bit.ly/lLFSEqM). On average, each site was approximately 5 miles from the closest weather
station; Site A was the furthest at approximately 12 miles from the nearest station.
Table 3-4 presents an operational summary for the studied cell(s) at each of the sites. At the time
of the development of this report, two sites (i.e., Site F and Site I) have been in PCC for 20 years.
All of the sites have or have had a GCCS and, except for two of the sites (i.e., Site A and Site G),
GCCS was operating at the time of this study. It should be noted that the actual start date for gas
collection at Site G is unknown - the start year of 2008 listed in Table 3-1 represents the
commencement of the landfill gas-to-electricity (LFGTE) system at this site. Four of the sites have
had LFGTE projects, and these projects are still ongoing at two of these sites. A fifth site (i.e., Site
D) is beneficially using landfill gas in a direct thermal application as a process fuel at a nearby
chemical and pharmaceutical facility.
Three of the sites reported recirculating leachate with the annual average rate ranging from 6 to
1,009 GPAD (annual leachate recirculation volumes were divided by 365 and the corresponding
cell area). Two of the three sites conducted leachate recirculation during the PCC period. For sites
with two studied cells that recorded cell-specific leachate data (i.e., Site C and Site G), the range
of values presented in the table represents the minimum and maximum values for both cells.
Additional details on leachate collection rate trends are discussed in Section 6.
Each study site's construction details including timeframe, liner construction, site's hydrogeology,
the final cover system details, waste placement and composition, and monitoring details including,
groundwater monitoring, leachate collection rate and quality monitoring, landfill gas monitoring,
the final cover maintenance, landfill settlement, and PCC cost were summarized. Individual site-
specific summaries of all the studied cells at each site are presented in Attachments A to I.
16
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Table 3-3. Design Summary for the Studied Cell(s) at Each Site
Site
A
B
C
D
E
F
G
H
I
US Region
Southeast
Northwest
Northeast
Southeast
Northeast
Southwest
Northeast
Northeast
Northeast
Ownership
Public
Private
Public
Public
Public
Private
Public
Private
Private
Area of Study Cell(s)
(acres)
28
20
40.8
69
6
38
15
60
51
Standalone Lined Cell(s)
No
No
Yes
Yes
No
Yes
No
Yes
Yes
Waste Quantity in Study
Cell(s) (million MT)
1.20
0.88
1.40
4.40
1.38
0.79
0.63
3.46
3.48
Site Capacity (million
MT)
2.05
3.27
Not
available
5.56
3.97
1.85
1.64
3.46
Not available
Additional Wastes
Accepted
Additional
waste info not
available
Industrial,
CDD, coal ash,
sewage sludge,
asbestos
Additional
waste info
not
available
Industrial,
CDD, land
clearing debris
Pulp and paper
mill waste,
industrial,
sludges, utility
ash, foundry
waste, treated
contaminated
soil
Additional waste
info not available
Petroleum-contaminated
soil, wastewater treatment
sludge, whey, ash
Sludge, CDD, non-
hazardous industrial
waste, ash, asbestos
Alternative daily cover,
special wastes,
contaminated soil, CDD,
sludge, asbestos, ash
Study Cell(s) Final
Cover (layers are listed
from the bottom to the
top)
40-mil LLDPE
geomembrane
(GM) with 18-
inch granular
drainage and
6-inch topsoil
layers
60-mil HDPE
GM with
geocomposite
drainage, 18-
inch soil, and
6-inch topsoil
layers
Exposed
3 5-mil
scrim-
reinforced
polypropyle
neGM
40-mil LLDPE
GM with
geocomposite
drainage, 18-
inch protective
soil, and 6-inch
vegetative soil
layers
40-mil VLDPE
GM with 12-
inch sand
drainage, 18-
inch soil, and 6-
inch topsoil
layers
40-mil HDPE GM
with geonet drainage
(side slopes). 40-mil
HDPE GM with
GCL (top deck); GM
covered with 12-inch
vegetative soil layers
Cell G3: 40-mil VLDPE
GM overlain by geogrid,
24-inch protective soil, and
6-inch topsoil layers
Cell G4: GCL, 40-mil
VLDPE, geotextile, 24-
inch protective soil, and 6-
inch topsoil layers
30-mil PVC/40-mil
HDPE GM with
geonet
drainage/geocompos
ite, geotextile, and
24-inch vegetative
layers
40-mil PE GM with
geonet drainage,
geotextile, 18-inch
granular cover, and 4-inch
vegetative layers
Study Cell(s) Bottom
Liner
Composite
with 1 ft of 10"
7 cm/ s clay
Composite
with 2 ft of 10"
7 cm/ s clay
Double 30-
mil PVC
(primary
and
secondary)
with 12-18-
inch
secondary
drainage
layer
Composite with
2 ft of 10"7 cm/s
clay
Composite with
4 ft of 10"7 cm/s
clay
Phase I: 60-mil
HDPE GM
Phase II: Composite
with 1 ft of 10"6
cm/ s clay
Phase III:
Composite with 2 ft
of 10"7 cm/s clay
Cell G3 (part): 36-mil
hypalon GM with
secondary collection layer
over 2 ft of 10"7 cm/s clay
of secondary liner
Cell G4: Double composite
with primary GM underlain
by an intermediate barrier
layer of 6 inches of 10"7
cm/s clay overlaying 1 ft of
10"5 cm/s clay over
geotextile over geonet over
secondary geomembrane
over 2 ft of 10"7 cm/s clay
Double liner system
with geonet between
the primary and
secondary GMs.
Cells 1-7 secondary
GM underlain with
6 inches of 10"5 cm/s
(max) soil overlying
5-ft compacted base.
Cell 8 secondary
GM underlain by
GCL.
Double composite liner
with primary 60-mil
HDPE underlain by GCL
and secondary GM
underlain by clay subbase.
Primary and secondary
drainage layer consists of
a sand drainage layer,
geotextile, and a geonet
from top to bottom.
Average Annual
Precipitation (inches)
51
56
47
46
30
22
46
47
51
Approximate
Groundwater Depth (ft)
25-75
140-260
1-8
7-40
5-40
12-26
12-65
2-13
0-55
17
-------
Table 3-4. Operational Summary for the Studied Cell(s) at Each Site
Site
A
B
C
D
E
F
G
H
I
Study Cell(s) Start-
Closure Year
1988-1998
1996-2004
1984-1998
1996-2009
1997-2004
1989-1997
1987-2000
1991-2008
1990-1997
Years since Closure (as
of 2016)
18
12
18
7
12
19
16-23
8
19
Gas Collection Start-
Stop Year
1999-2011
1991-N/A
1994-N/A
1999-N/A
1991-N/A
1999-N/A
2008a-
2012
1996-N/A
1994-N/A
Landfill Gas (or
Methane) Collection
Rate and Composition
Data Available
1999-2011
2007-2016
1995-2015
2009-2016
1997-2016
1999-2015
2008-2012
2005-2017
Not
Available
Observed Methane
Collection Rate
(standard cubic feet per
minute)
12-924
40-302
268-798
515-1333
63-924
36-367
0-104
328-1218
Not
Available
Landfill Gas-to-
Electricity Project
Start-Stop Year
2002-2005
N/A
2007-N/A
N/A
2009-N/A
N/A
2008-
Unknown
N/A
N/A
Leachate Recirculation
Start-Stop Year
2003-N/A
N/A
1986-1994
N/A
2011-N/A
N/A
N/A
N/A
N/A
Post-Closure Leachate
Collection Rate Range
(gallons per acre per
day)
320-1009
49-180
25-183
3-21
118-427
14-64
38-2,070
78-302
56-122
Leachate Recirculation
Rate Range (gallons per
acre per day)
89-1009
N/A
6-245
N/A
0-183
N/A
N/A
N/A
N/A
Notes:
N/A- dataset is not applicable to the site.
a Start date of GCCS not known. LFGTE (Landfill Gas-to-Electricity) project commenced in 2008.
18
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3.5. Data Presentation
Throughout this report, the pertinent data are summarized using box-and-whisker plots. This type
of plot provides a visual portrayal of the statistical distribution of the data. Figure 3-1 presents a
definition sketch of the box-and-whisker plot. The top, middle, and bottom of the box represent
the 75th, 50th (i.e., the median), and 25th percentiles, respectively. The lines that extend upward and
downward (whiskers) from the box represent the 90th and 10th percentiles, respectively. The values
less than the 10th percentile or more than the 90th percentile are presented individually outside the
whiskers. These plots were prepared using Sigmaplot 11 (Systat Software, Inc) or Excel (Microsoft
Corporation) software.
> 90th Percentile ZZt§
90th Percentile
75th Percentile
50th Percentile
25th Percentile
< 10 Percentile ^
Figure 3-1. Box-and-Whisker Plot Definition Sketch
10th Percentile
19
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4. Post-Closure Care Cost
rview
The termination or extension of the PCC period may have considerable financial implications for
the site owners. An understanding of the cost for various operations, maintenance, and monitoring
activities helps in the identification of the major cost centers. The site owners may consider
prioritizing the assessment of HHE impacts associated with the termination of cost-intensive
activities. As a means of assessing the financial impact of MSWLF maintenance and monitoring
beyond the 30-year PCC period, available PCC cost data for all the nine study sites were evaluated.
PCC costs were organized into the following six major cost categories:
(1) leachate management including monitoring and treatment,
(2) GCCS management,
(3) final cover maintenance including revegetation, mowing, and cover soil and runoff
control device maintenance,
(4) groundwater and subsurface gas monitoring and maintenance,
(5) administrative expenditures (e.g., permitting), and
(6) other expenditures (e.g., utilities, surface water monitoring and maintenance, site
security).
Available PCC cost data had limitations such as unavailability of actual PCC cost data for each
year since the closure, and lack of cost data exclusive to the studied cell(s). Furthermore, as
discussed later in this section, the available data suggest that the cost varied over a wide range due
to several factors, including inconsistencies in cost categorization, mixed availability of actual cost
and cost estimates, and the necessity for occasional capital-intensive system upgrades. Because of
these limitations, landfill owners and engineers should consider tracking and using the site-specific
cost data for evaluating the financial impacts of PCC activities instead of using the data presented
in the report as proxies.
A summary of the availability of PCC cost data for the study sites is presented in Table 4-1. Among
the nine study sites, the actual costs associated with PCC activities were available only for three
sites (Sites B, D, and E) for a limited number of years. For the other sites, the available PCC cost
data were estimated values. For Sites H and I, the available PCC cost data were specific to the
study cell(s). Site G provided estimated PCC costs for the entire site, and the estimated cost for
individual cells (G3 and G4) calculated based on the design capacity of each cell and total sitewide
PCC cost. Site A estimated PCC cost data for was available for years 2, 6, and 13 after closure;
however, each of these year's cost represented different waste footprint area of the site (28, 69,
and 84 acres, respectively), including the study cell(s). For other sites, the available PCC cost data
was associated with the maintenance of other site features (e.g., groundwater monitoring,
subsurface gas monitoring, cover maintenance) that may not necessarily be exclusive to the study
cell(s).
20
-------
Table 4-1. Summary of Availability of PCC Cost Data from the Nine Study Sites
Site
A
B
C
D
E
F
G
H
J(2)
Cell
G3
Cell
G4
Closure Year
1998
2004
1998
2009
2004
1997
1993
2000
2008
1997
Data Available
for Years after
Closure
2, 6,
15
8-11
14
3-7
0-11
2, 4, 6,
15, 18
8, 12,
15-17
1,5,8-
10
0-
30(1)
0-1,3,
5-12,
14, 16
Actual Post-
Closure Care
Cost
V
V
V
Exclusive to
Study Cell(s)
V
V
V
V
Representing
Area (Acre)
28, 69,
84(3)
65
127
110.5
61
105
7.1
7.9
60
51
Notes:
(1) Site H data included a cost estimate of 31 years (closure year and 30 years after closure).
(2) For Site I, the cost data were included only for the years in which new estimates were available.
(3) For Site A, PCC cost of year 2, 6, and 15 after closure represented the waste footprint of 28, 69, and 84
acres, respectively.
Figure 4-1 shows the distribution of available annual PCC costs per unit acre waste footprint for
each site. For consistency, the available PCC cost at each site was adjusted to 2017 dollars based
on consumer price index values (BLS 2018). Figure 4-1 shows only one data point for Site C and
Site H. As shown in Table 4-1, Site C has only one year of data available. Site H provided a
summary of total projected expenditures for 31 years of PCC (including the closing year and 30
PCC years); the total PCC cost was divided by 31 to obtain the average annual PCC cost.
Annual PCC costs for the cells studied at Sites G, H, and I varied over a wide range from
approximately $1,200 to $11,000 per acre, with an average of the cost of approximately $5,300
per acre. For the other sites that included PCC cost for the entire site, including the study cell(s),
the annual PCC costs ranged between approximately $1,550 to $37,000 per acre, with an average
cost of approximately $6,450 per acre and a median cost of $3,900 per acre. This wide range is
due to many factors, including inconsistencies in cost categorization, mixed availability of actual
cost and cost estimates, and the necessity for occasional capital-intensive system upgrades. For
example, this range includes one year of high cost for Site E for constructing a sewer connection
to pump leachate directly to the local wastewater treatment plant (WWTP) (i.e., Year 2 of PCC).
As a point of comparison, Morris and Barlaz (2011) estimated annual PCC cost for a hypothetical
MSWLF in PCC years 10 to 30 to be approximately $2,000 per acre based on 2009 dollars, which
corresponds to approximately $2,300 per acre in 2017 dollars.
21
-------
100
B
C
D
G3
G4
H
Site
Figure 4-1. Distribution of Total Annual PCC Cost at Each Study Site
Figure 4-2 presents the average percent contribution of different cost categories for the cell(s) at
each of the study sites based on the available data. The wide variation in the cost and relative
fraction of different activities among the sites is due to inconsistent cost accounting practices used
for tracking the cost. For example, a couple of sites (e.g., Sites D and E) tracked leachate
monitoring and management cost separately, whereas others likely included it in the total leachate
management cost. GCCS management costs were not available for Sites C and G. Also, Site G
documents presenting PCC cost data did not have an 'other' cost category. Cover maintenance
cost was not separately available for Site H, and administrative support costs were not separately
available for Site E and are not included in Figure 4-2.
22
-------
AB CDEFG3G4H I
Site
Leachate Management
' ' GCCS Management
K8888SI Cover Maintenance
Groundwater and Subsurface Monitoring
V////A Administration
I I Other
Figure 4-2. Average Annual Cost Distribution at Each Study Site/Cell After Closure
Overall, leachate management costs during PCC ranged between approximately 3% and 68% of
the total average annual PCC cost among the studied sites. Relative to other PCC activities,
leachate management appeared to be one of the most cost-intensive activities for all of the study
sites except for Sites D, E, and F. Although Sites C and G showed leachate management costs in
the range of 56% to 67% of total annual average PCC cost, the absence of GCCS management
cost data for these sites potentially inflated the relative fraction of other cost categories including
leachate management. In addition, the leachate management cost for both Sites C and G is
overestimated as these include the cost of management of leachate from the entire site and not
from the study cells.
The leachate management cost at Sites A, H, and I, which treated/pre-treated leachate onsite,
constituted approximately 23%, 35%, and 68% of the overall cost, respectively. Site I had an on-
site leachate pretreatment plant, and the pretreated leachate is discharged to a local sewer system
connected to a nearby WWTP. Site H has an on-site leachate treatment plant that discharges its
effluent to a nearby surface water body. Site A treated a fraction of the leachate and recirculated
the rest into the landfill. Treated leachate was spray irrigated over an area of the top deck of the
Site A study cell.
23
-------
Site D appears to have a substantially lower average annual PCC costs for leachate management,
possibly due to relatively lower leachate collection rates at Site D compared to the other sites, as
discussed in Chapter 6. Site D started direct discharge to a WWTP starting a year before closure,
the PCC leachate management cost included in this analysis does not include the construction cost
of connection to the sewer system, which was completed before closure. Site E constructed a
connection to the local sewer system to allow direct discharge of leachate to a WWTP in PCC year
2. The construction cost is included in the leachate management cost, which is a reason for the
high leachate management cost at Site E as compared to other sites.
Based on the available data of seven sites, the GCCS management cost ranged from approximately
11% to 44% of the total cost of PCC among the sites. GCCS management represented the highest
PCC cost at two sites and the second-highest cost at three sites. As discussed before, GCCS
management cost data for Sites C and G were not available. The GCCS management expenditures
of Sites B, D, E, and F included the cost associated with GCCS management for the entire site.
GCCS management cost data for Site A included the cost associated with the study cell and an
unlined cell and that for Site I was exclusive to the closed studied cell at the site.
The reported groundwater and subsurface gas monitoring cost ranged between approximately 5%
to 26% of the total average annual PCC cost for the sites where the PCC cost was reported for the
entire site. For Sites G, H, and I (for which cell-specific PCC cost data were available), the
groundwater and subsurface gas monitoring cost ranged from approximately 6% to 16% of the
total average annual PCC cost. The average annual administrative costs and other expenditures
among the sites for which the PCC costs were reported for the entire site ranged between
approximately 5% to 19% and 2% to 47%, respectively. Annual administrative costs for Site E
were not available. For Sites G, H, and I for which cell-specific PCC cost data were available, the
average annual administrative and other cost varied in the range of approximately 3% to 22% and
4% to 24%), respectively. PCC costs associated with the 'Other' category for Site G were not
available.
Annual final cover maintenance costs ranged between approximately 0.2% to 14% of the total
average annual PCC cost for sites where the PCC costs were reported for the entire site. For Sites
G and I (for which cell-specific cap maintenance PCC cost data were available), the annual closure
cap maintenance expenditures ranged between approximately 2% to 9%. As discussed above, Site
H cover maintenance costs were not available.
1 . II chate Managt in in 1 • I
Leachate management cost includes the cost associated with leachate collection system operation
and maintenance such as cleaning (when exclusively available), leachate hauling for off-site
treatment (if applicable), treatment (on-site and/or off-site), and leachate sampling and analysis
(when exclusively available). As discussed earlier, the categories used for leachate management
costs were not consistent among the sites and were not necessarily consistent for a site among the
years of available cost data.
Post-closure annual leachate management cost was evaluated for six study sites (excluding Site A
and C) based on the available data. All the collected leachate was recirculated for a majority of
duration after closure at Site A, and the annual sitewide leachate collection rates were not available
for Site C for the years leachate management cost was available. As leachate management cost is
expected to depend on the leachate collection rate, the annual leachate management costs were
24
-------
normalized with the annual leachate collection rate for each site for a consistent comparison among
the sites. Figure 4-3 shows the distribution of the annual leachate management cost per gallon of
leachate collected at each site since closure. The leachate management cost ranged between
approximately $5.5 to $219 per 1,000 gallons of leachate collected. On an area scale, annual
leachate management cost among the sites (except Sites A and C) ranged from $23 to $25,162 per
acre per year, with a median of approximately $2,300 per acre per year. The median annual
leachate management cost varied over an order of magnitude among the sites. The annual cost
varied over an order of magnitude for Sites D and E. The data suggest that a significant variation
in the leachate management cost can occur over time due to various factors including
implementation of a capital-intensive project(s), e.g., construction of a sewer connection for
pumping leachate to the local WWTP at Site E. In addition, the cost is expected to decline over
time as the leachate collection rate decreases.
£
u -2
-------
Site E had a considerably higher GCCS management cost for PCC years 2 to 5 (i.e., with an
average of approximately $449,000 per year) as compared to other PCC years (0 to 1 and 6 to 11)
(i.e., with an average of approximately $155,000 per year). The higher cost appears to be related
to the expansion of the site's GCCS for LFGTE system installation. Site B GCCS management
cost includes the cost of a contractor to operate and maintain the GCCS, which probably is the
reason for the elevated cost for this site. The annual GCCS management cost ranged from
approximately $2,700 to $593,000 among the study sites (excluding Sites C, G, and I), which is
equivalent to $96 to $9,731 per acre per year among the study sites (median of approximately
$1,770 per acre per year).
T3
CD
+->
O
-------
Groundwater and subsurface gas monitoring costs among the sites are expected to vary depending
on the number of monitoring locations, frequency of monitoring, and the number of groundwater
quality parameters being monitored. Figure 4-5 presents a distribution of annual groundwater and
subsurface gas monitoring costs during the PCC period for the study sites. The annual groundwater
and subsurface gas monitoring cost among the sites ranged from approximately $2,400 to
$169,600, which is equivalent to $40 to $2,600 per acre per year among the study sites (median of
approximately $490 per acre per year). It should be noted that the cost presented in Figure 4-5 for
the study cells G3 and G4 are specific to individual cells and are associated with the entire site for
the other sites. This probably is one of the reasons for the lower cost for G3 and G4 compare to
those for the other sites.
Site B had substantially higher monitoring costs as compared to other sites. The average annual
monitoring cost at Site B was approximately $144,000, whereas the average of the remaining sites
was approximately $27,000. The elevated monitoring costs at Site B might be associated with
groundwater impacts observed at the site with respect to the state groundwater quality standards.
The Subtitle D regulations (§258.55) and equivalent state regulations require enhanced
groundwater monitoring (more parameters and more frequent analysis) in the event of observed
groundwater impacts. The site may need to assess and implement corrective measures depending
on the nature of impacts. The implementation of enhanced monitoring and corrective measures at
Site B likely resulted in increased monitoring costs.
180
~0
c
!-h
(L)
C/3
o
u ^
oo £
S J3
o o
a °
£ o f-
c o
8 « ^
So g
i—i CD c/5
ctf O 3
3 ,ca o
a ¦ -
<
3 H
GO
B
C
D
G3 G4
H
Site
Figure 4-5. Distribution of Annual Groundwater and Subsurface Gas Monitoring Cost After
Closure at Each Study Sites
4.5. Final cover Maintenance Cost
The final cover maintenance costs are generally associated with revegetation, mowing, grading to
accommodate differential settlement, geomembrane repair, and/or stormwater control device
maintenance. Figure 4-6 shows the distribution of annual the final cover maintenance costs per
area of waste footprint during PCC at eight of the study sites (Site G had two cells, and data for
Site H were not available). Data for the total surface area of the final cover were not available;
27
-------
therefore, the cover maintenance cost was normalized based on the waste footprint area. Sites G
and I reported study cell-specific annual final cover maintenance costs, which ranged from
approximately $56 to $595 per acre. For the other sites that had sitewide cover maintenance cost
data available, the cover maintenance cost varied in the range of approximately $24 to $1,450 per
acre, excluding one year (year 1 after closure) of cover maintenance cost of approximately $3,400
per acre at Site E.
Site E presumably reported cap maintenance cost in a category listed as land surface care cost.
The annual land surface care cost for Site E during year 1 after closure ($206,000) was
substantially higher than the other years (varied between $2,400 to $83,300). The reasons for high
land surface care cost for year 1 after closure appears to be related to regrading (clay spreading).
Site C, which had only one year (i.e., Year 14) of PCC costs available, reported an annual site-
wide final cover maintenance cost of approximately $92 per acre.
o
U
'3 Ctf
S £
Oh
o
-------
As discussed above, Site C has an exposed geomembrane cap, although the cap had been repaired
after closure, the details of these repairs were not available.
The final cover maintenance cost is expected to be dependent on the magnitude of differential
settlement, which is expected to decline over time. Figure 4-7 shows a distribution of the final
cover maintenance costs and the annual volume loss observed at Site I during PCC. The annual
volume loss represents the yearly change in in-place waste volumes, which were using the annual
topographic data and approximate landfill bottom. During the initial closure years (Years 1 to 5),
greater settlement and corresponding higher cap maintenance costs were observed for Site I. The
annual settlement volume and the annual final cover maintenance costs decreased substantially
following Year 5 of PCC. A decrease in differential settlement potentially reduces the need for
regrading and revegetation, thereby reducing the cover maintenance cost. As presented in Table
4-1, for Site I, annual PCC cost estimates were available from a specific PCC year up to 30 years
into PCC and in the current analysis, PCC cost data were included only for the years in which new
estimates were available. New estimates of cost for PCC years 2, 4, 13, and 15 were not available.
pj CO
5 o
3 u
4 -
2 -
0
Site I
Annual Volume Loss
H Closure Cap Maintenance Cost
t—i——i—"r
1 2 3 4 5
100
- 80
- 60
- 40
- 20
I I I I 1 1 "T
9 10 11 12 13 14 15 16
O0
VI
o
J
c3
-------
For the other sites, the available cost data included the cost of maintaining other site cells as well.
All the available PCC cost data were adjusted to 2017 dollars based on consumer price index
values. For the sites with available study cell-specific data, the annual PCC cost varied from
approximately $1,200 to $11,000 per acre of waste footprint, with an average of $5,300 per acre.
For the sites where PCC cost data represented the entire site, annual PCC cost ranged between
$1,550 to $37,000 per acre with an average of $6,450 per acre. Inconsistencies in cost categories
used for tracking PCC cost, and the necessity for occasional system upgrades (e.g., GCCS
expansion at Site E, and construction of a sewer connection for pumping leachate to the local
WWTP) appears to be one of the primary reasons for such a wide variation in annual PCC cost at
the study sites.
Leachate management costs include those expenses associated with leachate collection, hauling
(as applicable), treatment (on-site and/or off-site), sampling and analysis (when exclusively
available), and leachate collection system maintenance (when exclusively available). The average
annual leachate management cost after closure ranged from approximately 3% to 68% among all
of the sites, and it represented the greatest cost at six of the nine sites. The leachate management
cost among sites with available cost and leachate collection rate data ranged from $5.5 to $219 per
1,000 gallons of leachate collected with an average of $79 per 1,000 gallons of leachate collected.
This is equivalent to $23 to $25,162 per acre per year, with a median of approximately $2,300 per
acre per year.
Among the seven sites for which GCCS management cost data were available, GCCS management
cost ranged from approximately 11% to 44% of the annual average PCC cost. The annual GCCS
management cost at these seven sites ranged from approximately $2,700 to $593,000 per year.
LFGTE systems were implemented at two of the study sites (Sites E and I). Installation of an
LFGTE project substantially reduced the cost associated with power purchase at Site E from an
annual average of $24,000 before to $1,800 after the commencement of the LFGTE project.
The average annual groundwater and subsurface gas monitoring cost varied between
approximately 5% to 26% of the total average annual PCC cost. The annual groundwater and
subsurface gas monitoring cost among the sites ranged from approximately $2,400 to $169,600,
which is equivalent to $40 to $2,600 per acre per year among the study sites (median of
approximately $490 per acre per year). The specific PCC cost distribution for each type of
monitoring activity was not available.
Annual final cover maintenance cost among the sites (except Site H) ranged between
approximately 0.2% to 14% of the total average annual cost. Two sites (Sites G and I) for which
study cell-specific cover maintenance cost data were available resulted in annual cover
maintenance costs ranged from approximately $56 to $595 per acre of the waste footprint. For the
remaining sites, the annual cover maintenance cost ranged from approximately $24 per acre to
$3,400 per acre. The final cover maintenance cost at Site I was observed to decline with the amount
of differential settlement, which was observed to reduce at Site I over time.
30
-------
5. Landfill Gas
rview
As discussed previously, LFG, if not controlled, is one of the primary sources of potential impact
to HHE. RCRA Subtitle D regulations require the operation and maintenance of the gas monitoring
system to ensure that the concentration of the methane generated by the facility does not exceed
25 percent of the lower explosive limit for methane in facility structures and the lower explosivity
limit for methane at the facility property boundary (§258.23). The landfill owner/operator is
required to implement a remediation plan for controlling the methane gas releases if methane gas
levels exceeding these limits are detected in the facility structures or at the property boundary. The
first objective of this chapter is to assess the frequency of methane detection above these thresholds
in the facility structure and property boundary at the study sites.
Objective 1. Assess the frequency of subsurface methane migration to the facility structures and
property boundary of the study sites.
Although the Subtitle D regulations for MSWLFs do not specifically require installation and
operation of a GCCS, NSPS require construction and operation of a GCCS at MSWLFs that
generate more than 50 Mg (or 34 Mg if regulated under Subpart XXX) of NMOCs annually. NSPS
regulations (Subparts WWW and XXX) specify numerical thresholds for GCCS implementation
and termination timeline at MSWLFs. NSPS (§60.752 (b)(2)(v)) allows for the removal or
decommissioning of a GCCS at MSWLFs after closure as long as (a) the GCCS has been in
operation for at least 15 years (or sooner due to lack of adequate LFG flow per Subpart XXX), and
(b) the annual NMOCs generation rate is less than 50 Mg per year (or 34 Mg per year if meeting
the requirements of Subpart XXX). NSPS specifies using the measured LFG collection rate and
NMOCs concentration to estimate the annual NMOCs rate for comparison to these thresholds for
sites with active GCCS. The NMOCs rates calculated using the measured LFG collection rates are,
therefore, referred herein as the NMOC collection rate.
The blower/flare system operating constraints are additional considerations that should be taken
into account for GCCS termination as NSPS Subpart XXX allows GCCS termination sooner than
15 years due to a lack of adequate flow rate. The blower/flare system is typically designed to
handle the estimated peak LFG collection rate. The lower end of the LFG flow rate for the safe
operation of flare ranges from 5 to 10% of the design (or peak) flow rate. The blower/flare system
may need to be replaced or retrofitted to combust LFG below these flow rates if the NMOCs
collection rate corresponding to these LFG flow rates is greater than 50 or 34 Mg/year. An
understanding of these durations would allow assessment of whether the blower/flare system
would need to be retrofitted or replaced before NSPS Subpart WWW or XXX allows its
termination or decommissioning. The purpose of the second objective of this chapter, presented
below, is to evaluate (a) whether the annual NMOCs collection rate from closed MSWLFs would
be less than 50 (or 34) Mg within the minimum required GCCS operating period of 15 years or
within 30 years after closure, and (b) whether the LFG flow rate would decline below 5% or 10%
of the peak LFG flow rates before achieving annual NMOCs collection rates of 50 (or 34) Mg.
The study sites sizes, however, are not representative of the size of approximately 75% of
MSWLFs in the US as six of the study sites contain less than 4 million metric tons of waste. This
analysis was performed for the study sites as well as hypothetical landfills that are representative
of typical landfills in the US.
31
-------
Objective 2. Assess the timeframe needed for the annual NMOCs collection rate from study sites
to drop below the NSPS threshold of 50 (or 34) Mg per year and the timeframe for
the LFG flow rate to decline below 5% and 10% of the peak flow rates.
Although NSPS allows termination of GCCS operation upon demonstration that annual NMOCs
collection rate is below 50 (or 34) Mg, there is a possibility of a spike in the annual NMOCs
generation rate and the associated potential collection rate in the future above the NSPS thresholds.
The NMOCs generation rate is dictated by the LFG generation rate, which is expected to
substantially increase with moisture intrusion if the in-place waste has substantial remaining LFG
generation potential at the time of GCCS operation termination. The purpose of the third objective,
presented as follows, is to address the question of whether the closed MSWLFs can generate
NMOCs above the NSPS threshold of 50 and 34 Mg/year after termination of GCCS operation.
Objective 3. Assess whether MSWLFs would have sufficient NMOCs generation potential to
generate 50 (or 34) Mg of NMOCs annually after the termination of GCCS operation.
In addition to the NSPS, there may be state-specific required/suggested thresholds that may impact
GCCS termination. For example, the FDEP (FDEP 2016) suggests achieving a methane production
rate that is below 10% of the peak rate or achieving remaining methane generation potential of less
than 10%) of the total generation potential before considering changes to the active LFG system.
The goal of the fourth objective is to estimate the remaining methane generation potential of the
in-place waste and the associated timeframes to achieve the targeted reduction of methane
generation potential.
Objective 4. Assess the remaining methane generation potential at 30 years after closure and
assess the timeframes needed for the remaining methane potential to drop below 25%
and 10%) of the total generation potential.
Robust analysis methods and approaches are critical for reliably estimating the remaining methane
and NMOCs generation/collection potential. An additional objective of this chapter is to present
and discuss the approaches that can be used for estimating the remaining methane and NMOCs
generation potential and timeframes to achieve annual NMOCs collection rate of 50 (or 34) Mg
for a closed MSWLF. The analysis presented in the chapter should not be considered as a
comprehensive evaluation of HHE impacts with respect to LFG due to various assumptions and
limitations. These assumptions and limitations are presented along with the analysis. These
limitations are also summarized in the last section of this chapter.
. Methodology
5.2.1. Data Sources
The analysis presented in this Chapter is based on the following two data sources: the first is the
nine case-study sites, and the second is the Greenhouse Gas Reporting Database
(https ://www3. epa. gov/enviro/).
1) Case-Study sites
Perimeter monitoring probe data for methane were available for review from seven of the nine
study sites. Some carbon dioxide and oxygen data were also available for all sites except Site H.
A summary of the number of probes surrounding the study cells, as well as the total number of
data points analyzed from these probes for each of the sites are presented in Table 5-1. Facility
32
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structures methane concentration data were only available from three sites (Sites B, F, and H). The
data from all the probes at the study sites were analyzed for all the sites, except for Site G and I.
Probes data were not available for Site G and study cell-specific probe data were available for Site
I. A total of 7,598 methane monitoring probe readings were compiled from the sites with available
data; approximately 56% of these readings are for Sites E and I.
The presence of an active GCCS and the availability of LFG flow rate and composition data were
part of the site selection criteria. Therefore, the LFG flow rate and methane content data were
available for all the study sites/cells. Table 5-1 summarizes the key attributes relevant to each site's
GCCS. LFG flow rate and methane content data were available for all of the sites except Site I;
although the sitewide LFG flow rate data were available for Site I, the study cell-specific LFG
flow rate and composition data were not available in data sources reviewed for the closed study
cell. More details about GCCS and historical LFG/methane collection rate trends are presented in
the individual site descriptions included in Attachments A-I.
2) Greenhouse Gas Reporting Database
As shown in Table 5-1, at least three of the study sites are relatively small (i.e., contain less than
2.5 million MT of waste) and were not regulated under the NSPS rules. Therefore, the results of
some of the analysis may not be representative of typical MSWLFs in the US. Data from the GHG
database were primarily used to identify the representative size of MSWLFs in the US.
The capacity of the MSWLFs, reported as part of the annual GHG reports, were downloaded from
the Envirofacts database (EPA 2019b). This database contains data from all the MSWLFs that are
required to report annual methane generation and emissions amounts per federal regulations (40
CFR 98 Subpart HH) (referred hereinto to as GHG reporting regulations); the GHG reporting
regulations require the MSWLFs owners to report annual methane generation and emissions that
accepted waste on or after January 1, 1980, and are estimated to generate more than 25,000 metric
tons CO2 equivalent GHGs. Data obtained from the database were used to estimate:
1) Disposal capacity of MSWLFs that are representative of the 25th, 50th, and 75th percentile
of the capacity of MSWLFs in the US; the MSWLF containing in-place waste amounts
corresponding to 25th, 50th, and 75th percentile MSWLF capacity in the US is referred
herein to as small, medium, and large MSWLF, respectively.
2) Operating lifespan for each MSWLF included in the GHG reporting regulations database
was calculated based on the first year of waste acceptance and the actual or estimated
closure year included in the GHG database. The median lifespan of MSWLFs with
capacities ranging from 15th to 35th percentile capacity was used as the representative
lifespan of the small MSWLF. Similarly, the median lifespan of MSWLFs with capacity
in the 40th-60th and 65th-85th percentile capacity ranges were used as the representative
lifespans of medium and large MSWLFs, respectively.
33
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Table 5-1. Key Attributes Relevant to the GCCS of Each Study Site
Data
Type
Site
A
B
C
D
E
F
G
H
I
Timeline and In-place
Waste Amount
Study Cell(s)
Start-Closure
Year
1988-
1998
1996-
2004
1984-
1998
1996-
2009
1997-
2004
1989-
1997
1987-
2000
1991-
2008
1990-
1997
Years after
Closure (as of
2016)
18
12
18
7
12
19
16-23
8
19
Total Sitewide
In-Place Waste
Amount
(Million
Metric Tons)
2.05
3.27
X
5.56
3.97
1.85
1.64
3.46
X
Subsurface Monitoring
Probes Data
Years (after
Closure) of
Available
Probe Data
2007,
2009,
2011,
2015
2005-
2015
X
2010-
2015
2004-
2016
1997,
2000,
2002-
2011
X
2008-
2016
1997-
2009,
2011,
2013-
2015
Number of
Probes
54
23
X
26
19
13
X
10
11
Total Number
of Data Points
Available
194
689
X
678
2,206
926
X
870
2,035
Landfill Gas Collection and Control System
Operating
Period3
1999-
2011
1991-
N/A
1994-
N/A
1999-
N/A
1991-
N/A
1999-
N/A
2008b-
2012
1996-
N/A
1994-
N/A
Methane
Collection
Rate Years
Modeled for
Site-Specific
Decay Rate
Estimation
1999-
2011
2010-
2016
1998-
2002
2009-
2016
2005-
2016
2000-
2009
2008-
2012
2007-
2017
X
Methane
Collection
Rate (standard
cubic feet per
minute)
12-
924
40-
302
268-
798
515-
1333
63-
924
36-
367
0-104
328-
1218
X
Landfill Gas-
to-Electricity
Project Start-
Stop Year
2002-
2005
N/A
2007-
N/A
N/A
2009-
N/A
N/A
2008-
N/A
N/A
N/A
N/A- Not applicable to the site.
X - data not available
a N/A- GCCS was operating at the time of this study
b Start date of GCCS not known. LFGTE (landfill gas-to-electricity) project commenced in 2008.
5.2.2. Methane Collection Rate for the Study Sites
The blower/flare systems at MSWLFs are typically equipped with a flow meter and a datalogger
to record the LFG flow rate several times per hour (e.g., once every 15 mins). LFG composition is
typically analyzed using a handheld monitor. The frequency of the available LFG flow and
composition readings varied from once per year to multiple times per day. For Site G, the methane
flow rates were available. For four of the sites (Sites A, B, C, and F), an associated LFG
34
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composition concentration was available for each LFG flow measurement; the methane flow rate
was calculated by multiplying the LFG flow rate with the methane content for these sites. The
methane flow rate was calculated by using Eqn. 5-1.
Methane flow (scfm) = Total LFG flow (scfm) x Methane Fraction (Eqn. 5-1)
For Site E, the reported methane mass (in kilograms) was converted to the volumetric methane
flow rate at standard temperature and pressure using Eqn. 5-2. NSPS (§60.2) specifies a
temperature of 68 °F and 101.3 kilopascals as standard conditions. The methane density at standard
conditions was calculated using the ideal gas law and GHG reporting regulations (§98.343)
specified methane density of 0.0423 lb methane per cubic feet at 60 °F and 1 atm (101.3
kilopascals).
Methane flow (scfm) = Methane mass (——) x 2.2 — x 1 cubw ^eet
min kg 0.0417 lb methane @ 68 °F and 1 atm
(Eqn. 5-2)
For Site H, LFG flow rate data were available on an approximately monthly basis, but only annual
LFG composition data were available. The monthly LFG flow rates were multiplied by the
associated annual methane composition value to estimate the monthly methane flow (Eqn. 5-1).
For Site D, the average of the methane content data available was used for the analysis presented
in this Chapter; weekly to monthly gas composition data were available only for a limited period
(April 2015 through December 2016) for this site. The average of the available methane
concentration measurements was multiplied by all the LFG flow values to estimate methane flow
rates for Site D.
5.2.3. Methane and NMOCs Collection Rate Estimation
The estimation of LFG, methane, and NMOCs generation rate using the EPA's LandGEM model
is standard practice for estimating the LFG rates for GCCS design and the annual NMOCs
generation rate for compliance purposes (EPA, 2005). The LandGEM model uses the first-order
decay model, as presented in Eqn. 5-3.
Qcha(0 = YS=\MtL0k e tt) (Eqn 5.3)
Where,
Qcm(t) = modeled methane generation rate in m3 CH4 yearat time t.
Mi = waste placed in year i in metric tons,
L0 = the waste methane generation potential in m3 CH4 per metric ton of waste placed in the
landfill,
k = decay rate on a per-year basis,
n= lifespan of landfill
ti = placement time, in years, of Mi.
t= time, in years
LandGEM estimates the LFG generation rate based on the specified methane content of LFG. It
further estimates the NMOCs generation rate using the LFG generation rate and NMOCs content
of LFG. LandGEM allows the selection of one of several default values or user-specified values
for L0, k, and NMOCs content for modeling LFG and NMOCs generation rate.
35
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NSPS allows the use of the measured LFG collection rate to estimate the annual NMOCs
generation/emission rate for the sites with active GCCS for the purpose of determining when the
system can be removed. As all of the sites have or had a GCCS, the measured collection rates were
used to estimate the site-specific decay rate and the methane potential using inverse first-order
decay modeling. The site-specific methane collection potential and decay rates were subsequently
used in LandGEM model to estimate the future LFG and NMOCs collection rates. A description
of the methodology to estimate the methane collection potential and decay rate for the study sites
is presented in Section 5.2.4. As the methane potential estimated using this approach is based on
the methane collection rate and not the generation rate, it is referred herein to as the methane
collection potential (Lc).
Table 5-2 list the values of modeling parameters used for the estimate of the future methane and
NMOCs collection rates. The annual disposal amounts were available for all of the sites and were
used for LandGEM modeling. The site-specific estimates of the methane collection potential and
decay rates were used for the modeling.
NSPS allows the use of a site-specific measurement, if available, for estimating the annual NMOCs
emission rate and comparison to the NSPS threshold (50 (or 34) Mg per year). Due to a lack of the
site-specific NMOCs measurements, the NSPS-default of 4,000 parts per million volume as
hexane was used. The results of LandGEM modeling were used to estimate the post-closure period
for the annual NMOCs collection rate to decline below the NSPS thresholds of 50 (or 34) Mg per
year for the study sites.
Table 5-2. Modeling Parameters for the Estimation of the Future Methane and NMOCs Collection
Rate for the Study Sites
Parameter
Value Used
Waste Mass*
Annual site-specific disposal mass (presented in Attachments
A through I)
Decay Rate, k (year_1)
Site-specific estimate for the study sites
Methane Collection Potential, Lc (m3 methane
per MT waste)
Site-specific estimate for the study sites
Nonmethane Organic Compounds (NMOCs)
(ppmv)
4000
* The waste placement data only for the study cells CI and C2 were used as these data were not
available for the other cells at Site C. Sites G and C were actively accepting waste at the time of this
study. Site G was assumed to close in 2016 for the analysis.
As discussed later, the study sites sizes are not representative of the size of approximately 75% of
MSWLFs in the US as six of the study sites contain less than 4 million metric tons of waste. In
addition to the study sites, the methane and NMOCs collection rates were also estimated for the
hypothetical small, medium, and large MSWLF representative of the 25th, 50th, and 75th percentile
capacity of the landfills in the GHG reporting database, respectively, as described in Section 5.2.1.
In order to estimate the impact of decay rate on the methane and NMOCs collection rate trend for
these landfills, five scenarios each with a different k value were simulated using the first-order
decay model for each of the three examples MSWLFs. The decay rates used were 0.02, 0.038,
0.057, 0.17, and 0.22 year _1. The first three decay rate values selected for this analysis are the
GHG Reporting Regulations-specified decay rate for the precipitation zones with an annual rainfall
of less than 20 inches (arid area), 20-40 inches (moderate precipitation zone), and more than 40
inches (wet zone). The decay rates of 0.17 and 0.22 year _1 are the median and average,
36
-------
respectively, of the decay rates reported for bioreactor landfills (EPA 2006, Yazdani et al. 2006,
Kim and Townsend 2012, Wang et al. 2013).
As mentioned earlier, NSPS allows the use of the measured LFG collection rate for the sites with
active GCCS to estimate the annual NMOCs generation/emission rate for the purpose of
determining when the system can be removed. NSPS also explicitly allows the use of AP-42-
suggested L0 for estimating the LFG collection rate for GCCS design. The AP-42 suggested value
of 100 m3 CH4 per Mg was used for the first-order decay modeling to estimate the future annual
methane and NMOCs generation rates for these hypothetical MSWLFs analyses. It should be noted
that LandGEM model (Eqn 5-3) calculates the LFG generation rate and not the collection rate. An
LFG collection efficiency of 100% was assumed for these example sites to calculate the LFG
collection rate from the estimated generation rate for a conservative analysis. In other words, the
collection rates were assumed to be equal to the generation rate.
A spreadsheet-based first-order decay model was developed and used for estimating the methane
collection rate of these hypothetical cases. LandGEM was not used for these cases as the estimated
annual NMOCs collection rate for some of the cases did not decline below 50 Mg/year within the
150-year duration (since the landfill start) used for LandGEM model. The annual NMOCs
collection rate was calculated from the estimated methane collection rate using the following
equation (adapted from the equation used by NSPS Subparts WWW and XXX); methane was
assumed to constitute 50% (by volume) of LFG:
Mnmoc = 2 X Qch4 x Cnmoc x 3-6 X 10 9 (Eqn. 5-4)
where:
Mnmoc = Annual NMOCs collection rate (Mg year_1)
Qch4 = CH4 collection rate (m3 year_1); and
Cnmoc = Concentration of NMOCs in LFG (ppmv) (4,000 ppmv as hexane used)
3.6xl0 9 = Conversion factor (m3 hexane to Mg hexane)
2 = factor for calculating LFG flow rate using methane flow rate (volume LFG/volume CH4)
5.2.4. Site-Specific Decay Rate and Methane Collecti tential Estimation
A site-specific k and methane potential values can be estimated by conducting an inverse first-
order decay modeling (also referred herein as regression analysis) to best-fit the rate modeled using
the first-order decay model to the available methane collection rate. Given the availability of long-
term historical methane collection data from eight of the study sites, a regression analysis was
conducted to estimate the site-specific methane potential and decay rate for each site. As the
methane potential estimated using this approach is based on the methane collection rate and not
the generation rate, it is referred herein to as the methane collection potential (Lc). Because
LandGEM does not include a regression analysis feature, a first-order decay model (Eqn. 5-3) was
developed for each site to estimate the monthly methane collection rate for the period for which
the measured methane flow data were available. Monthly disposal amounts were estimated by
dividing the annual disposal amounts by 12 for the sites for which monthly disposal data were not
available.
The sum of squared errors (SSE) presented in Eqn. 5-5 is a measure of the relative goodness of fit.
The Microsoft® Excel function "Solver" was used to minimize the SSE by changing k and Lc
values. The initial values to initiate the Solver function and constraints used for the regression
analysis are presented in Table 5-3. For most sites, all three model simulations (corresponding to
three different initial values for Lc) provided the same best-fit Lc and k. The best-fit Lc and k that
37
-------
resulted in the lowest SSE were selected from the regression analyses corresponding to the three
initial values of Lc (20, 100 and 230 m3/Mg) for each site if the three model runs provided varying
Lc and k. Site-specific Lc and k estimates are presented in Section 5.3.3.
SSE — X j (Q mode led, j ~ Q measured,]) (Eqn. 5-5)
Where,
Qmodeiedj = the modeled monthly methane flow rate for month j,
Qmeasuredj = the monthly methane flow rate calculated using the reported flow rate and methane
content for month j, and
n = number of months that monthly recovered methane quantity data are reported for the site.
Table 5-3. Initial Values and Constraints for Parameters used for Modeling
Parameter
Initial Value
Lower-Upper constraints
Decay Rate, k (year_1)
0.05
0.001-2.2 (Faouretal. 2007)
Methane Collection Potential, Lc (m3 methane
per MT waste)
20, 100, 230
20-230 (Krause et al. 2016)
Furthermore, the coefficient of correlation (r2 value) (Eqn. 5-6) was calculated using the SSE, and
the total sum of squares (SST) was calculated as presented in equation (Eqn. 5-7). The SST is a
measure of the total variation of the site methane flow rates with respect to the average of the
monthly/annual recovered methane flow rates (i.e., Qmeasured,avg) available for the site.
r2 = 1 — — (Eqn. 5-6)
sst v n '
Where,
SST — X j (Qmeasured,avg ~ Qmeasuredj) (Eqn. 5-7)
All available methane collection data were included in this analysis for Sites A and D. As described
in Attachments B, C, E, F, G, and H, LFG data for only selected periods for these sites were used
for regression analysis. The methane collection data only for the period where the data appeared
to follow a first-order decay declining trend, were used for regression analysis. Several factors
such as the expansion of GCCS to new areas (e.g., Sites C and F), the progressive decommissioning
of wells (e.g., Site G), or aggressive wellfield adjustment during the GCCS start-up phase (e.g.,
Site B) result in deviation of the measured collection rate from the first-order decay trend. The
data, which were indicative of major changes to the GCCS, were excluded from the regression
analysis. It should be noted the GCCS changes only impacts the LFG collection rate and not the
LFG generation rate. For the years for which the measured/calculated methane collection rates
were not available, the modeled data points were not included in the SSE calculation.
5.2.t 1 mtial of Elevated NMOCs Generation Rate aftt mination
The potential for a spike in the annual NMOCs generation rate after the termination of GCCS
operation was evaluated for a small MSWLF in the moderate precipitation zone (20-40 inches
annual precipitation) and a large MSWLF in the arid area (less than 20 inches annual precipitation).
The objective was to assess whether a closed MSWLFs can generate NMOCs above the NSPS
threshold of 50 and 34 Mg per year after termination of GCCS operation. A decay rate of 0.02 and
0.038 year _1 was used for arid and moderate precipitation zones, respectively, as discussed in
Section 5.2.3. It was assumed that the decay rate would increase by 100% to 0.04 and 0.076 year
for arid and moderate precipitation zones, respectively, ten years after GCCS decommissioning
38
-------
(i.e., ten years after the annual NMOCs collection rate decline below 50 Mg and the resultant
GCCS operation termination). The decay rate can potentially increase due to moisture intrusion
into the landfill through a compromised final cover (assuming PCC is terminated and the final
cover is not monitored and maintained at this point). Because LandGEM does not allow the
specification of a time-varying decay rate, a spreadsheet-based first-order decay model was
developed and used for this analysis.
5.2.6. Remaining Methane and NMOCs Generation Potential
%>#
The annual methane generation rate from LandGEM or the developed first-order decay model
spreadsheet were aggregated to calculate the cumulative amount of methane generated over time.
The remaining methane generation potential was calculated by deducting the cumulative methane
generated from the total methane generation potential as depicted by the following equation (Eqn.
5-8):
VcH4 remaining,) ~ 2j£=i
MiL0 - QcH4,t (Eqn. 5-8)
Where,
VCH4 remaining,j Total remaining methane generation potential (m3 CH4) at the end of year tj
QcH4,t= modeled annual methane generation rate in m3 CH4 yearduring year t.
Mi = waste placed in year i in metric tons,
L0 = the waste methane generation potential in m3 CH4 per metric ton of waste placed in the
landfill,
n= lifespan of the landfill
t, tj= time, in years
The remaining methane generation potential per unit in-place waste was calculated by dividing the
total remaining methane generation potential by the in-place waste mass as depicted by the
following equation (Eqn. 5-9):
I — VcH4 remaining,j (Vnn S
'-'CHA remaining,]' ~ yn M. * nqn. J-y)
2-ii=il 'i
Where,
LcH4 remaining,j Remaining methane generation potential per unit in-place waste (m3 CH4 per MT
waste) at the end of year tj
The percent remaining methane generation was calculated by dividing the remaining methane
generation potential with the total methane generation potential as depicted by the following
equation (Eqn. 5-10):
Pern remain,n„,i = * 100 (Eq" 5_10>
Where,
P CH4 remaining,j Percent remaining methane generation potential (%) at the end of year tj
The annual NMOCs generation rate from LandGEM model or the developed first-order decay
model spreadsheet were aggregated to calculate the cumulative amount of NMOCs generated over
time. The remaining NMOCs generation potential was calculated by deducting the cumulative
NMOCs generated from the total NMOCs generation potential as depicted by the following
equation (Eqn. 5-11):
39
-------
MnmoCs remaining, j — 2 X Cjvmoc X 3.6 X 10 X 2j£=i ^i^o "Yi^Q^NMOCs.t (Eqn. 5-11)
Where,
M^NMOCs remaining,j Total remaining NMOCs generation potential (Mg as hexane) at the end of year
tj
Mnmocs,t = modeled annual NMOCs generation rate in Mg yearduring year t, and
other terms as defined above.
The site-specific Lc estimates were used as a proxy for L0 for estimating the percent and mass-
based remaining methane and NMOCs generation potential. It should be noted that the percent
remaining methane and NMOCs generation potential would be the same irrespective of whether
L0 or Lc value used for the estimation. The use of Lc instead of L0 would, however, result in an
underestimation of the mass-based remaining generation potential.
and Discussion
surface Methane Monitorinq Data
%>#
The regulations in §258 require the installation of subsurface methane monitoring probes around
the periphery of the site and the quarterly monitoring of these probes to identify the subsurface
migration of LFG. Site operators are also required to monitor the methane concentration in on-site
structures to verify that LFG is not migrating to the facility property boundary above the lower
explosivity limit and accumulating in the structures beyond 25% of the lower explosive limit (i.e.,
1.25% by volume in air). All the available perimeter methane monitoring probe data were
analyzed.
A summary of the number of probes surrounding the study cells, as well as the total number of
data points analyzed from these probes for each of the sites is presented in Table 5-4. Methane was
detected in 542 instances (i.e., approximately 7.1%) out of these 7,598 readings and exceeded the
lower explosive limit of methane (i.e., 5% methane by volume in air) in 103 instances (i.e.,
approximately 1.4% of all readings included in the evaluation). More than 55% of the detected
methane measurements occurred at Site E.
Table 5-4 summarizes the number of exceedances where the methane concentration was measured
above the lower explosive limit of 5% as a function of years since closure. The available data
suggest that the subsurface methane exceedances occurred relatively infrequently over the periods
with data available for review. As presented in Table 5-4, only 103 methane exceedances were
observed at five of the seven sites. Almost 80% of these methane concentrations above the lower
explosivity limit of methane occurred at Site E, and 10% of exceedances occurred at Site B. Both
these sites contained either unlined cells or cells lined with a compacted clay liner. Ten probes at
Site E exhibited methane measurements greater than its LEL at least once; the latest exceedance
was observed in 2011. Methane concentration was above its LEL for 80 measurements at ten
probes. Sixty-six of these 80 methane measurements greater than the LEL were observed at three
probes located near the cells lined with a compacted clay liner. The cell lined with geomembrane
is not likely to cause the elevated methane concentrations observed at these probes. At Site B, ten
of 12 measurements above the LEL of methane occurred at a single probe located near the area
where the lined cell(s) adjoin the unlined cell, and the exceedances are likely caused by the unlined
cell.
All of the observed exceedances were recorded within the first five years after closure for all the
sites except for Site E. Relatively infrequent methane detection/exceedances at probes in the
40
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vicinity of the study cells are like due to active gas collection and a geomembrane bottom liner
system. The GCCS at Site A was terminated during the 13th year since closure. The available
perimeter probe monitoring data do not show any methane exceedance since the termination of
GCCS for Site A.
The structure methane concentration data (2,105 measurements) were also available for Sites B,
F, and H. The methane monitoring data (70 measurements) were available for 15 structures at Site
B for year 1 and 2 since the closure. Methane was detected in two measurements at Site B. All
methane measurements were lower than 25% of its LEL. The maximum detected methane
concentration was 0.1% (volume basis).
Methane monitoring data were available for 41 facility structures at Site F for years 3 and 5 through
14 since the closure. A total of 1,155 measurements were available. Five of the measurements
exhibited (at four structures) methane concentration greater than 25% of its LEL. The maximum
methane concentration at these locations was 8.5%. These structures are located near the unlined
cells of the Site F; therefore, the study cell was not likely the cause of these exceedances.
Based on 880 measurements at ten facility structures at Site H, methane was detected at a
pumphouse on eight sampling events during years 1 and 2 since closure. None of the methane
measurements in structures at Site H were above 25% of its LEL- the maximum methane
concentration observed was 1% (volume basis).
41
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Table 5-4. Number of Annual Methane Monitoring Probe Methane Exceedances at the Site as a Function of
Time Since Closure
Site
A
B
D
E
F
H
I
Years of
Available
Probe Data
2007, 2009,
2005-
2010-
2004-
1997, 2000,
2008-
1997-2009,
2011,
2013-2015
2011,2015
2015
2015
2016
2002-2011
2016
Years
Since
Number of
Probes
54
23
26
19
13
10
11
Closure
Total Number
of Data Points
194
689
678
2,206
926
870
2,035
Available
Closure Year
1998
2004
2009
2004
1997
2008
1997
-
-
-
6
3
0
4
0
-
2
0
16
-
0
0
1
-
4
1
23
-
0
0
2
-
4
1
17
0
0
0
3
-
2
0
5
-
0
0
4
-
0
2
5
0
0
0
5
-
0
0
7
0
0
0
6
-
0
1
0
0
0
7
Exceedances
per Year
-
0
0
0
0
0
8
0
0
0
0
0
9
-
0
0
0
0
10
0
0
0
0
0
11
-
0
0
0
12
0
0
-
13
-
0
0
14
-
-
-
15
-
-
0
16
0
-
0
17
-
0
18
indicates that probe or facility structure methane data was not available for the site for that year.
Shaded cells indicate future years.
5.3.2. Methane Collection Rate Trends
Figure 5-1 presents the temporal variation of the methane flow rate for the study sites after closure.
The methane flow rate for all of the sites shows a declining trend after closure. An increase in the
LFG flow rate at Sites C and F is associated with the expansion of the GCCS to collect LFG from
new active cells (at Site C) and old cells (at Site F). The GCCS expansion results in an increase in
the LFG collection rate but does not impact the LFG generation rate. GCCS operation at Site A
was terminated 12 years after closure due to an inadequate LFG flow rate. It should be noted that
due to a lack of study-cell specific LFG data, data from several or all the closed cells at the site
were used for the analysis presented in the report.
42
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1400
1200
1000
800
600
400
200
0
i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i i
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18
Years after Closure
Figure 5-1. Temporal Variation of Methane Flow Data for Study Sites
5.3.3. Estimated Site-Specific First-Order Decay Rates and Methane Collection
Potentials
The waste decay rate has a significant impact on future LFG/methane/NMOCs generation
potential. Table 5-5 presents the average annual precipitation, the site-specific decay rate, and the
methane collection potential estimates based on the regression analysis. The estimates based on
the first-order decay model fit the measured data with an r2 > 0.7 for six of the eight study sites
modeled. An r2 of 0.7 means that the modeling approach captures or explains 70% of the variations
in the methane collection rate over time at a given site, and the rest 30% is due to factors (e.g.,
changes in GCCS operations) that are not accounted for by the model.
Based on the approach described in Section 5.2.4, the estimated k for the study Sites A-H ranged
from 0.06 to 0.24 year _1 with an average of 0.16 year _1; LFG data for Site I were not available.
The estimated decay rates for four out of eight sites were within a range of 0.18-0.21 year _1. As
expected, k estimated for Site A is highest among all of the sites, potentially due to the extensive
leachate recirculation conducted at the site. Leachate was also recirculated at study cells C and E.
However, the fraction of the collected leachate that was recirculated and the duration of
recirculation for these cells was substantially lower than that of study cell A. Nonetheless, the
estimated k for the site was approximately 20% lower than that recommended by AP-42 of 0.3
year_1 for wet MSWLFs (US EPA 2008). The estimated k was lowest for Site F, which is located
in the driest area among all the study sites. It should be noted that active LFG collection from the
study cell at Site F did not start until 2010 (13th year after closure). The LFG collection rate from
another cell (lined with a clay liner) was used for the regression modeling for this site. The
43
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estimated k values are within the range of those reported for MSW (Pelt 1993, Faour et al. 2007,
Tolaymat et al. 2010, Barlaz et al. 2010, Amini et al. 2012, Amini et al. 2013, Wang et al. 2013,
Zhao et al. 2013, Vu et al. 2017), but substantially greater than the NSPS default of 0.05 year _1.
The estimated k values are also greater than the ones recommended by AP-42 (0.02 and 0.04 year
"' for regions with less than 25 inches and more than 25 inches of precipitation, respectively) as
well as values specified by GHG reporting regulations (0.02, 0.038, and 0.057 year_1 for less than
20, 20-40, and more than 40 inches precipitation zones, respectively). The estimated k values for
these sites were approximately three to four times the k specified by the GHG reporting regulations
and three to five times the AP-42 recommended k (except for Site A). These high values suggest
that the MSW at the study sites is decomposing at a substantially faster rate than the AP-42
recommended rate as well as the rate specified by the GHG reporting regulations.
The Lc estimates ranged from 36-152 m3 CH4 Mg"1 with an average of approximately 94 m3 CH4
Mg"1. Krause et al. (2016) presented a comprehensive compilation and a critical review of MSW
L0 and reported values to range from 20-223 m3 CH4 Mg"1 based on modeled values and those
calculated based on waste composition and biodegradability of individual waste components. A
range of 35-167 m3 CH4 Mg"1 was reported based on experimentally measured values of mixed
MSW (Krause et al. 2016). NSPS prescribes a L0 of 170 m3 CH4 Mg"1 for determining whether or
not a landfill will be subjected to its requirements. This value has been reported to be conservative
and not reflective of the actual generation (Krause et al. 2016). Krause et al. (2016) reported an
average L0 of 94 m3 CH4 per Mg for North America. The EPA Compilation of Air Pollutant
Emissions Factors (AP-42) suggests L0 of 100 m3 CH4 per Mg for MSWLFs, which is very close
to the average L0 reported by Krause et al. (2016) for North America.
The site-specific estimates for all the sites were in the range of experimentally measured of L0
values of mixed MSW compiled by Krause et al. (2016). The relatively low Lc values for the Sites
E and G are potentially due to the disposal of considerable amounts of inerts in the study cells at
these sites. As discussed in Attachments E and G, non-MSW materials (primarily inerts including
contaminated soil, sludges, ashes, industrial waste) constituted over 30% (by total weight) of the
in-place waste at these sites. In addition, limited gas collection infrastructure in one of the cells at
Site G results in poor gas collection efficiency and lower Lc estimate; LFG was only collected
from the leachate collection pipes cleanouts at one of the cells at Site G.
The use of site-specific Lc for estimating the future annual NMOCs collection rate and assessing
the timeframe for the NMOCs rate to decline below the NSPS threshold is appropriate as the NSPS
regulations allow using the measured LFG collection rate for estimating the NMOCs
generation/emission rate for landfills with a GCCS. However, it should be noted that not all the
LFG generated is collected even at sites with GCCS. The LFG and NMOCs generation rates are,
therefore, expected to be greater than those estimated using Lc. The magnitude of the difference
between the collection and generation rates would depend on the GCCS collection efficiency,
which is typically not measured. It should be further noted, the percent remaining methane and
NMOCs generation potentials and timeframes to achieve the LFG flow rate below a given percent
of the peak flow rate estimates would not depend on the L0 (or Lc) value used. However, the use
of Lc would result in an underestimation of the mass-based remaining methane and NMOCs
generation potential.
44
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Table 5-5. Estimated First-Order Decay Rate and Methane Collection Potential, and Average
Annual Precipitation for the Study Sites
Site-Specific Methane
Coefficient of
Annual
Site
Site-Specific Decay
Rate, k (year _1)
Collection Potential, Lc
(m3 CH-t per Mg Waste)
Correlation (r2)
Precipitation
(inches/year)
A
0.24
152
0.93
51
B
0.21
87
0.89
56
C
0.20
127
0.86
47
D
0.09
69
0.86
46
E
0.18
58
0.86
30
F
0.06
132
0.21
22
G
0.21
36
0.71
46
H
0.11
94
0.55
47
5.3.4. Timeframes for Achieving Annual NMOCs Collection Rate below 50/34 Mg
and LFG Flow Rate Below 5% and 10% of the Peak Rates
This section addresses the questions of whether the annual NMOCs collection rate from closed
MSWLFs would be less than 50 (or 34) Mg within the minimum NSPS-required GCCS operating
period of 15 years or within 30 years after closure. The analysis also addresses whether the LFG
collection rate would decline below 5% or 10% of the peak LFG flow rate before achieving annual
NMOCs collection rates of 50 or 34 Mg. As mentioned earlier, the NSPS-default NMOCs
concentration in combination with the site-specific k and Lc estimates (Table 5-5) were used for
the first-order decay modeling to estimate the future annual methane and NMOCs collection rate.
The use of site-specific Lc estimates, which are representative of the LFG collection rate, is
appropriate for this analysis since NSPS allows the use of the LFG collection rate for estimating
the annual NMOCs generation rates for determining whether or not GCCS operation can be
terminated.
Table 5-6 presents the estimated period after closure for an annual NMOCs collection rate to
decline below the NSPS thresholds of 50 (or 34) Mg per year for the study sites. It also presents
the waste amounts used for LandGEM modeling, the estimated years after closure for LFG flow
rates to decline to 5% and 10% of the peak flow rates, as well as the estimated GCCS operating
duration before the NMOCs collection rate decline below 50 Mg per year. The GCCS operating
duration was calculated based on the actual GCCS start-up year and the year in which the NMOCs
collection rate is expected to decline below 50 Mg per year; it does not represent the actual total
GCCS operating duration. As expected, the duration for the annual NMOCs collection rate to
decline below 50 Mg is lower than those associated with 34 Mg per year.
The GCCS at all of the sites except for the study Sites A, C, and G would have operated for more
than 15 years before the annual NMOCs collection rate declines below 50 Mg year. The analysis
suggests that the annual NMOCs collection rate threshold is the limiting NSPS constraint for the
termination of GCCS operations. The expeditious decline of annual NMOCs collection rate below
50 Mg per year at Sites A, C, and G is potentially due to the relatively smaller amount of waste
placed at these sites and extensive bioreactor operation of Site A and C. As mentioned earlier,
entire waste mass deposited at the study site was used for the analysis presented in this section;
only waste data from the study cells CI and C2 were used for the analysis due to lack of waste
disposal data for the other cells at Site C. The timeframes to achieve annual NMOCs collection
rate reduction below 50 and 34 Mg per year are expected to be greater than those presented in
45
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Table 5-6 for Site C if waste placement amounts for all the cells at this site were included. Site G
was assumed to close in 2016.
Table 5-6. Estimated Timeframes for NMOCs Collection Rate and LFG Flow Rate Reduction
Landfill Gas
Duration after Closure (Years)
Collection and
Annual Nonmethane Organic
Waste
Control System
Operating
Compounds (NMOCs)
Collection Rate
Landfill Gas Flow Rate
Site
Amount
Duration Until
(MT)
Annual NMOCs
Collection Rate <
< 50 Mg per
Year
< 34 Mg per
Year
<10% of
Peak LFG
<5% of Peak
LFG Flow
50 Mg per Year
Flow Rate
Rate
(Years)
A
2.05
10
11
13
11
13
B
3.27
21
8
10
10
13
CI
1.40
13
10
12
11
14
and
C2
D
5.56
36
26
30
24
31
E
3.97
23
10
13
13
16
F
1.85
25
27
34
42
55
G
1.64
<0
<0
<0
7
10
H
3.46
32
21
24
20
26
The Sites A, F, and G contain less than 2.5 million metric tons of waste; MSWLFs containing less
than 2.5 million metric tons of waste are exempt from the LFG control requirements of NSPS. The
implementation of GCCS at Sites A, F and G were not dictated by NSPS. The termination of GCCS
operation at these sites would, therefore, not be controlled by NSPS provisions. The GCCS at Site
F was implemented to control subsurface migration of LFG, and energy generation from LFG
appears to be the motive behind implementing a GCCS at Site G.
The GCCS operation at Site A was terminated after 13 years of operation due to flare operation
issues once the LFG flow rate declined to 5% of the peak flow rate observed at the site. This
suggests that GCCS operation may need to be terminated before the NSPS (Subpart WWW) 15-
year operating duration requirement due to lack of adequate methane flow rate needed to sustain
flare operation, especially for small MSWLFs such as Site A. At this point, the blower/flare system
may need to be replaced with either a smaller capacity system or with alternative options such as
biofilters to control any potential odor issues and/or to oxidize methane and NMOCs.
The analysis suggests that the annual NMOCs collection rate would be less than 34 Mg per year
by the end of 30 years after closure for all of the study sites except Site F. The annual NMOCs
collection rate for Site F is estimated to decline below 34 Mg per year 34 years after closure.
However, the provisions of Subpart XXX are only applicable to the MSWLFs that commenced
construction, reconstruction, or modifications after July 17, 2014. As mentioned earlier, Site F is
not subjected to the requirements of NSPS (Subpart WWW or XXX). Given that all the study sites,
except for Sites C and G, were closed before 2014, Subpart XXX provisions are not applicable to
these study sites.
The timeframe to achieve LFG flow rates that are 5-10% of the peak flow rates are greater than
those needed to achieve an annual NMOCs collection rate of 34-50 Mg per year for all the study
sites. The analysis suggests that no major modification of the blower/flare system capacity would
46
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be needed at any of the study sites to continue operating the GCCS until the annual NMOCs
collection rate decline below the NSPS thresholds. The blower/flare systems at Sites D and H may
also need to be replaced as the total estimated GCCS operating duration at these sites is over typical
service life (15-25 years) of the blower/flare system.
5.3.5. Impact of In-Place Wat ' fount and Decay Rate on Timeframes hieve
Annual NMOCs Collection Rate Reduction below the NSPS Thresholds
The annual NMOCs collection rate is dependent on the LFG flow rate, which in turn is dependent
on the in-place waste amounts, waste placement timing, and decay rate. The estimated timeframe
for annual NMOCs collection rate reduction below 50 Mg per year for Site D is substantially
greater than the other sites, which is potentially due to considerably higher in-place waste amounts
compared to the other study sites (Table 5-6). The impact of the decay rate (k) on the duration for
NMOCs collection rate reduction below the NSPS thresholds can be assessed by comparing these
timeframes within study sites of similar size. For example, although Sites A and F are similar in
size, the timeframes for annual NMOCs collection rate to decline below 50 and 34 Mg for Site A
are lower than those for Site F. This distinction is potentially due to a considerably higher waste
decay rate (k) at Site A than that of Site F. Similarly, the timeframe for the annual NMOCs
collection rate to decline below the NSPS thresholds for Site B is substantially lower than that of
the Site H, which is similar in size to Site B, potentially due to a substantially higher waste decay
rate at Site B than that of Site H.
Figure 5-2 presents the distribution of the capacities of approximately 1,300 active and closed
MSWLFs that reported annual GHG emissions to the EPA at least once during the 2010-2016
timeframe. Over 75% and 60% of the MSWLFs have a capacity of more than 4 and 5 million
metric tons, respectively. The study sites sizes, therefore, are not representative of the size of
approximately 75% of MSWLFs in the US, given that six of the study sites contain less than 4
million metric tons of waste. The total capacity of Site C, which was actively receiving waste at
the time of this study, is unknown.
47
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10000
C/3
c
o
H
o
c
o
o
S3
S3
U
Ph
h-I
GO
1000 -
100 -
10 -
1 -
0.1 -
T
0.01
Figure 5-2. Distribution of Capacities of Active and Closed MSWLFs in the US
In order to assess the post-closure timeframes needed to achieve NMOCs reduction below the
NSPS threshold for MSWLFs of varying sizes, the methane and NMOCs collection rates were
estimated for three hypothetical MSWLFs that are representative of the 25th, 50th, and 75th
percentile of the capacity of MSWLFs in the US. The 25th, 50th, and 75th percentile of the capacity
of MSWLFs in the GHG database is approximately 3.35, 7.85, and 19.1 million metric tons,
respectively. As mentioned earlier, the MSWLF containing in-place waste amounts corresponding
to 25th, 50th, and 75th percentile MSWLF capacity in the US, is referred herein to as small, medium,
and large MSWLF, respectively.
The annual disposal amounts for each of the three examples MSWLFs were calculated by dividing
the total capacity by the respective median lifespan, then rounded to the nearest thousand. For the
analysis presented in this section, the waste placement rate is assumed to be consistent over the
life of the MSWLF. Table 5-7 summarizes the capacity, lifespan, and annual waste placement rate
of the three example MSWLFs used for assessing the impact of the in-place waste volume and
decay rate on the post-closure timeframe for achieving annual NMOCs reduction below 50 Mg
per year. First-order decay modeling was conducted to estimate the annual methane and NMOCs
collection rates for small, medium, and large MSWLFs using the lifespans and annual placement
rate values for these MSWLFs as presented in Table 5-7.
48
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Table 5-7. Capacity, Lifespan, and Annual Disposal Amounts of the Example MSWLFs
MSWLF Size
Category
MSWLF Capacity
(Million Metric
Tons)
Capacity percentile
Range for Lifespan
Estimation
Median Lifespan
(years)
Estimated Annual
Placement Rate
(Mg/year)
Small
3.35
15%—35%
43
78,000
Medium
7.85
40%-60%
58
135,000
Large
19.1
65%-85%
71
269,000
Table 5-8 presents the post-closure period to achieve annual NMOCs collection rate reduction
below 50 Mg per year. The estimates suggest that the annual NMOCs collection rate of only small
MSWLFs in the wet zone would decline below 50 Mg per year within 30 years after closure. The
annual NMOC collection rates from the medium MSWLFs in moderate precipitation and arid
zones are expected to decline below 50 Mg per year in 50 and 80 years after closure, respectively.
As expected, the post-closure duration for annual NMOCs collection rate to decrease below 50 Mg
per year is the longest for MSWLFs in the arid zone. It should be further noted site-specific
conditions such as subsurface methane migration and odor issues may necessitate GCCS operation
beyond the NSPS-required timeframe to ensure compliance with pertinent local, state, and federal
regulations. The impact of the local and state regulations on GCCS operating duration was not
assessed in this study.
The annual NMOC collection rates for all three MSWLFs sizes is expected to decline below 50
Mg per year within 20 years if these landfills are operated as a bioreactor. The bioreactor operation
substantially reduces the post-closure duration for the annual NMOCs collection rate to decline
below 50 Mg from 124 years to 13-17 years for large size MSWLFs in arid areas. As expected,
the analysis suggests that the bioreactor operation can substantially reduce the GCCS operating
duration for MSWLFs of all sizes, especially those located in arid and moderate climatic zones.
Based on a review of data from ten Wisconsin MSWLFs that recirculated leachate and added
external liquids waste, Bareither et a. (2018) reported that LFG generation rate is expected to
decline below 5% of the peak rate and more than 75% of the LFG generation potential would be
realized within 40 years of PCC at these sites.
The LFG flow rate declined to less than 10% of the peak LFG flow rate after the annual NMOCs
collection rate reduced below 50 Mg per year for all simulation scenarios for small and medium
MSWLFs. For large MSWLFs, the NMOCs collection rate declined to less than 50 Mg per year
after the LFG flow rate declined to 10% of the peak LFG flow rate but before the LFG flow rate
declined to 5% of the peak LFG flow rate. The analysis suggests that a blower/flare system that is
designed based on the peak LFG flow rate can be used to collect and combust LFG until the annual
NMOCs collection rate declines below 50 Mg. However, the service life (typically 15-25 years)
of a blower/flare system is expected to be much shorter than the duration the GCCS would need
to operate for NSPS compliance for the MSWLFs that are not operated as bioreactors.
The NSPS-default NMOCs concentration of 4,000 parts per million by volume as hexane was used
for the analysis. NMOCs content of LFG has been reported to vary over a wide range. As a point
of comparison, AP-42 suggests using NMOCs concentration of 2,420 parts per million volume as
hexane for landfills known to have the co-disposal of MSW with non-residential waste and
recommends 595 parts per million volume as hexane for landfill containing only MSW (EPA
1998). EPA (2007) reported NMOCs to range from 233 to 5,870 parts per million by volume as
hexane based on measurements from five MSWLFs in the US. The annual NMOCs collection rate
would decline below 50 Mg per year over a shorter duration than the estimates presented in the
49
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chapter if the actual NMOCs concentration is lower than 4,000 part per million by volume as
hexane. In addition, a constant NMOCs concentration was used for the entire duration as the data
pertaining to the variability of NMOCs concentration in LFG over time are lacking.
Table 5-8. Impact of In-Place Waste Amount and Decay Rate on the Estimated Timeframes for
Achieving Annual NMOCs Collection Rate Below 50 Mg per Year
Precipitation Zone/Operating
Condition
Decay
Rate, k
(Year_1)
Timeframe after Closure for Annual Nonmethane
Organic Compounds (NMOCs) Collection Rate <50
Mg per Year (Years)
Small MSWLF
Medium MSWLF
Large MSWLF
Arid (< 20 Inches Annual Rainfall)
0.02
49
85
124
Moderate (20-40 Inches Annual
Rainfall)
0.038
35
52
71
Wet (>40 Inches Annual Rainfall)
0.057
26
36
48
Bioreactor-Median of Reported k
0.17
9
13
17
Bioreactor-Average of Reported k
0.22
7
10
13
5.3.6. intial of Elevated NMOCs Generation aft ^ rmi nation
As described earlier, NSPS allows termination of GCCS operation once the annual NMOCs
generation rate has declined below 50 Mg (or 34 Mg) per year. The gas collection wells are,
typically, retrofitted to passively vent the generated gas into the atmosphere once the GCCS
operation is terminated. The annual NMOCs generation rate is expected to continue declining after
GCCS termination until the methane generation potential of the waste is exhausted. Precipitation
intrusion into the landfill due to a compromised final cover would increase the in-situ waste
moisture content, and enhance the waste decomposition process and resultant methane and
NMOCs generation/emission rates if the MSWLF at this point still contains undecomposed
biodegradable organics with significant methane and NMOCs generation potential.
Figure 5-3 presents the estimated annual NMOCs generation rate for a small MSWLF in a
moderate precipitation zone with no change and a 100% increase in decay rate ten years after the
annual NMOCs generation rate declines below 50 Mg NMOCs per year. With an assumed 100%
increase in decay rate to 0.076 year _1, NMOCs generation rate increases to 70 Mg NMOCs per
year and stays above the NSPS-threshold of 50 Mg NMOCs per year for four years. The assumed
increased decay rate of 0.076 year"1 might not be unreasonable given that all the study sites except
Site F, which is in an arid region, are estimated to have a decay rate of more than 0.076 year"1.
50
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200
180
160
140
120
100
80
60
40
20
0
k = 0.038 year
1
k = 0.038 and 0.076 year
20 40
Years After Closure
60
Figure 5-3. Estimated Annual NMOCs Generation Rate for Typical Small MSWLF in a Moderate
Precipitation Zone with 100% Increase in Waste Decay Rate 10 Years after GCCS Termination
Figure 5-4 shows the estimated annual NMOCs generation rate after closure for a typical large
MSWLF in an arid area. The annual NMOCs generation rate is estimated to decline below 50 Mg
per around 124 years after closure. Per NSPS, GCCS operation can be terminated at this point.
Assuming that PCC is also terminated and the final cover is no longer monitored and maintained
at this point, any deterioration of the final cover system would result in an increase of moisture
intrusion into the landfill. The NMOCs generation rate is estimated to increase to 81 Mg NMOCs
per year and remain above 50 Mg NMOCs per year for 13 subsequent years if the waste decay rate
doubles at the 134th year after closure due to moisture intrusion into the MSWLF. The oxidation
of NMOCs and methane through the final cover soil is expected to be insignificant as the
geomembrane underlying the final cover soil layer at most of the study cells would minimize gas
migration through the final cover soil.
The decay rate was assumed to increase by 100% ten years after GCCS termination for this
analysis. The likelihood of this increase within ten years of GCCS termination may be questionable
due to relatively low permeability of MSW, and the presence of the final cover, even if
compromised due to potential lack of maintenance, is expected to restrict the infiltration of
moisture into the landfill. The intent of this analysis was to illustrate a scenario where the annual
NMOCs emission rate can exceed the NSPS thresholds after GCCS termination with an increase
in waste decay rate if the in-place waste has considerable remaining NMOCs and methane
generation potential.
51
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Years After Closure
Figure 5-4. Estimated Annual NMOCs Generation Rate for Typical Large MSWLF in an Arid
Area with 100% Increase in Waste Decay Rate 10 Years after GCCS Termination
The analysis presented above suggests that the annual NMOCs generation rate can spike back
above the NSPS in years after GCCS operation is terminated. The final cover, therefore, should be
rigorously maintained even after GCCS termination until the NMOCs generation potential and the
leaching potential of the in-place waste has declined to levels that are unlikely to pose a risk to
HHE. In addition, landfill owners and regulators should also continue surface and subsurface
emissions and odor monitoring to proactively identify signs of an increase in LFG generation rate
and have provisions in place to resume GCCS operation, if needed, to control these issues.
The reliance on the importance of the final cover to alleviate the long-term HHE impacts associated
with GCCS operation termination can be mitigated by implementing operating strategies such as
bioreactor that can stabilize waste while the final cover is actively monitored and maintained
(Morris and Barlaz 2011, Bareither et al. 2018).
In addition to the NSPS criteria, the landfill owners and regulators should also consider evaluating
the remaining methane and NMOCs generation potential of the in-place waste and assessing the
likelihood of future emissions based on site-specific factors including but not limited to provisions
for the final cover monitoring and maintenance, sub-surface and surface methane emission
monitoring, and plan to resume GCCS operation any future odor/emission issues for decisions
pertaining to GCCS termination.
52
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1 . mated Remaining Methane and M1 > neratii tential
The magnitude of the hypothetical spike and duration of elevated annual NMOCs generation rate
would depend on the remaining methane and NMOCs generation potential of the in-place waste.
An assessment of the remaining methane generation potential for a landfill at the end of 30 years
after closure, therefore, is important. Table 5-9 presents the estimated percent of total methane
generation potential remaining at closure and at the end of 30 years after closure. It also presents
the methane generation potential remaining at closure and at the end of 30 years after closure. The
remaining methane generation potential at closure ranged from 2.2 m3 CH4 per Mg of waste (for
Site G) to 73.6 m3 CH4 per Mg of waste (56% of the total of 132 m3 CH4 per Mg of waste for Site
F) for the study sites; Site G is an active site and was assumed to close in 2016 for the purpose of
the analysis. The remaining methane potential at Site G declined below 10% before 2016 (assumed
closure year) due to a significant decline in the MSW disposal amounts at the site since 2007.
The remaining methane generation potential was estimated to range from less than 0.1 m3 CH4 per
Mg of waste to approximately 13.9 m3 CH4 per Mg of waste for the study sites 30 years after the
closure. The remaining methane potential is estimated to be less than 2 m3 CH4 per Mg of waste
for all of the sites except for Site F at the end of 30 years after closure. Site F, which is located in
the arid area, is estimated to have the greatest remaining methane generation potential at closure
and 30 years after closure due to the lowest estimated decay rate among the study sites.
Table 5-9. Estimated Percent Remaining Methane Generation Potential of the Study Sites at
Closure and 30 Years after Closure
Remaining Methane Generation Potential at
Remaining Methane Generation
Closure
Potential 30 Years after Closure
Site
%
m3 CH4 per Mg
%
m3 CH4 per Mg
A
26%
39.2
0.02%
<0.1
B
13%
11.3
0.02%
<0.1
C
29%
36.8
0.07%
<0.1
D
48%
32.9
2.9%
2.0
E
23%
13.2
0.11%
<0.1
F
56%
73.6
11%
13.9
G
6%
2.2
0.01%
<0.1
H
43%
40.5
1.4%
1.3
The remaining methane generation potential of 25% and 10% are used by some states as a
benchmark for GCCS operation modifications. All the study sites, except Site F, are expected to
have less than 10% remaining methane generation potential within the timeframes estimated for
NMOCs collection rates to decline below 50 Mg per year threshold. Site F is expected to achieve
a remaining methane potential of 25% and 10% in 15 and 31 years after closure, respectively.
The percent remaining methane or NMOCs generation potential is not a reliable measure for
assessing potential HHE impacts. Figure 5-5 presents the annual NMOCs generation rate as a
function of the percent remaining methane generation potentials for a small MSWLF in a moderate
precipitation zone and a large MSWLF in an arid zone. For the same percent remaining methane
generation potential, the annual NMOCs generation/emission rate is greater for the large MSWLF
compared to the small MSWLF. The percent remaining methane potential corresponding to the
annual NMOCs generation rate of 50 Mg per year for the small MSWLF and the large MSWLF is
approximately 13% and 4.6%, respectively.
53
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Remaining Methane Potential (%)
Figure 5-5. Annual NMOCs Generation Rate as a Function of the Percent Remaining Potential for
Small MSWLF in Moderate Precipitation Zone and Large MSWLF in Arid Area
The mass-based remaining methane (Mg methane) and NMOCs generation potential (Mg
NMOCs) are more appropriate indicators of the HHE impacts than the percent remaining methane
and/or NMOCs generation potential. Table 5-10 presents the mass-based methane and NMOCs
generation potential remaining at closure and 30 years after closure, and assuming an NMOCs
concentration of 4,000 parts per million by volume as hexane. All of the sites except Sites D, F,
and H are estimated to have less than 10 Mg of NMOCs generation potential at the end of the 30-
year post-closure period. The analysis suggests that Sites A, B, C, E, and G would not have the
potential to generate NMOCs above 50 Mg per year 30 years after closure. Sites D, F, and H are
estimated to have approximately 321, 7380, and 134 Mg, respectively, of the total remaining
NMOCs generation potential at the end of 30 years after closure.
The remaining NMOCs generation potential is dependent on the in-place waste amounts and the
decay rate. Although the in-place waste volume at Site F (located in an arid area) is smaller than
that of Site D, the remaining NMOCs generation potential for Site F is more than that of Site D.
Based on the timeframes presented in Table 5-6, the GCCS at Site F can be terminated around the
27th year after closure as annual NMOCs generation rate declines below 50 Mg per year at this
point. However, without active LFG control, Site F has a potential for HHE impacts from methane
NMOCs release if all the remaining NMOCs are released within a short period after 30 years of
the post-closure period.
54
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Table 5-10. Estimated Remaining NMOCs Generation Potential of the Study Sites at Closure and
30 Years after Closure
Remaining Methane Generation Potential (MT
Remaining NMOCs Generation
methane)
Potential (MT NMOCs)
30 Years after Closure
30 Years after
Site
at Closure
at Closure
Closure
A
54,641
38
2,329
2
B
25,119
40
1,069
2
C
34,975
80
1,492
3
D
124,265
7,582
5,265
321
E
35,629
168
1,510
7
F
92,727
17,482
3,913
738
G
2,220
4
105
<1
H
94,998
3,158
4,034
134
It should be noted that the mass-based remaining methane and NMOCs generation potential are
estimated based on site-specific Lc estimates, which corresponds to the LFG fraction that is
captured by the GCCS. The use of Lc may result in an underestimation of the actual methane
generation rate for the cases where part of the waste may not be decomposing due to conditions
such as inadequate moisture content and for sites with poor gas collection efficiency. The LFG
collection efficiency is not expected to be the case with NSPS compliant systems. For the cases
with Lc estimate that is substantially lower than AP-42 recommended value of 100 m3 CH4 Mg"1
waste, the landfill owners and regulators should consider a comprehensive evaluation of site-
specific data such as amounts of inerts in the landfill, level and frequency of surface and sub-
surface methane emissions, landfill surface settlement rate and the volume loss rate to assess the
appropriateness of Lc for estimating the remaining methane generation potential. For the sites that
are evaluated to contain waste that is not decomposing and/or have substantial fugitive emissions,
the use of AP-42 default L0 value (in instead of site-specific Lc estimate) should be considered for
estimating the remaining methane and NMOCs generation potentials.
The remaining methane and NMOCs generation potential can also be estimated by collecting
samples of deposited waste and measuring the biochemical methane and NMOCs generation
potential in a laboratory. As the remaining methane generation potential depends on waste age and
composition, samples across the landfill area and depth should be collected to account for the
variability in the composition and the age of the deposited waste. Several studies (e.g., Townsend
et al. 1996, Mehta et al. 2002, Tolaymat et al. 2010, Kim and Townsend 2012) have reported
methane generation potential based on this approach. Potential damage to the final cover
geomembrane and landfill infrastructure (e.g., horizontal LFG collector, buried LFG header and
lateral), extensive sample collection and analysis, and cost are the major disadvantages of this
approach.
5.4. Summary
LFG emissions are one of the primary pathways for the HHE impacts of MSWLFs. The RCRA
Subtitle D regulations require the installation of subsurface monitoring probes around the
periphery of the site and the quarterly monitoring of these probes and structures at the facility for
methane. Subsurface methane monitoring probe data from seven study sites suggest relatively few
exceedances detected in the subsurface and structural methane monitoring. Methane exceeded the
55
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lower explosive limit of methane (i.e., 5% methane by volume in air) in 103 instances (i.e.,
approximately 1.4% of all 7,598 measurements included in the evaluation). Approximately 80%
of these exceedances were observed at Site E and 10% of exceedances occurred at Site B. Both
these sites contained either unlined cells or cells lined with a compacted clay liner, which are likely
the cause of the observed methane exceedances at these sites. All the exceedances correspond to
three sites during the first five years after closure for all the sites except for Site E. The GCCS at
Site A was terminated in the 13th year of closure. The available perimeter probe monitoring data
do not show any methane exceedance since the termination of GCCS at Site A. The structure
methane concentration data (2,105 measurements) were also available for Sites B, F, and H. None
of the methane measurements in structures at Sites B and H were above 25% of its LEL. Only 5
out to the total 1,155 methane measurements in Site F structures were above 25% of its LEL.
The federal NSPS requires active collection and control of LFG from MSWLFs that produce more
than 50 Mg of NMOCs (as hexane) annually. The GCCS at the NSPS-regulated MSWLFs should
be operated for at least 15 years, and until the annual NMOCs generation rate is more than 50 Mg
per year. The future NMOCs generation rate can be estimated using the LandGEM model based
on waste decay rate, in-place waste amount and placement timing, methane generation potential
of waste, and NMOCs content of LFG. NSPS allows using the measured LFG collection rate and
the site-specific NMOCs concentration, if available, for NMOCs generation rate estimation.
All of the study sites have/had a GCCS. The GCCS monitoring data were available for review only
for eight of the case-study sites. A regression analysis was conducted to estimate a site-specific
methane collection potential and decay rate that provided the best fit to the available monthly
methane flow rate data when used with the first-order decay model. The methane flow rates were
calculated using the LFG flow rate and methane content unless the methane flow rates were
directly available. The estimated site-specific decay rates suggest that waste decomposition at all
of the eight study sites (with GCCS data) is occurring more rapidly than the decay rates
specified/recommended by NSPS, AP-42, and GHG Reporting regulations. The estimated decay
rate for four out of eight sites was within a range of 0.18-0.21. As expected, k estimated for Site
A is the highest among all of the sites, potentially due to the extensive leachate recirculation
conducted at the site. The estimated k was lowest for Site F, which is located in the driest area
among all the study sites. The estimated methane collection potential ranged from 36-152 m3 CH4
per Mg waste. Sites E and G had a relatively low methane collection potential due to the disposal
of considerable inert materials at the study cells at these sites. Limited LFG collection
infrastructure in one of the study cells also contributed to low methane collection potential at Site
G.
The future annual methane and NMOCs collection rates were estimated for the study sites using
the first-order decay model based on site-specific decay rate and methane collection potential,
study site disposal amounts, and an NSPS-default NMOCs concentration of 4,000 parts per million
by volume as hexane. The estimated NMOCs and methane collection rate data were then used to
assess the post-closure period needed to achieve (1) the annual NMOCs collection rate below 50
(and 34) Mg per year, and to (2) LFG flow rates that are 5% and 10% of the peak LFG flow rate.
The annual NMOCs collection rates from all of the study sites are expected to decline below 50
Mg per year within 30 years after closure. The annual NMOCs collection rate estimated to decline
below 34 Mg per year within 30 years after closure for all the study sites except Site F. The
remaining methane generation potential at all of the study sites, except Site F, is expected to be
56
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less than 10% of the total generation potential before the NMOCs collection rates decline below
50 Mg per year.
The study site sizes, however, are not representative of the size of approximately 75% of MSWLFs
in the US as six of the study sites contain less than 4 million metric tons of waste. The NMOC
generation/emission rate is dependent on the landfill size and, therefore, the required GCCS
operating duration estimated for the study site may underestimate the duration for typical
MSWLFs in the US. Over 75% of the MSWLFs included in the EPA GHG database have a
capacity of more than 4 million metric tons. In order to assess the post-closure timeframes to
achieve NMOCs reduction below the NSPS threshold for typical size MSWLFs in the US, NMOCs
and methane flow rates were estimated for MSWLFs containing approximately 3.35 (small
MSWLF), 7.85 (medium MSWLF), and 19.1 (large MSWLF) million metric tons MSW. These
sizes correspond to the 25th, 50th, and 75th percentile of the capacity of MSWLFs in the US,
respectively.
In order to assess the impact of decay rate on the annual NMOCs collection rate modeling was
conducted for each of the three example MSWLFs for the following decay rates: 0.02
(representative of arid areas with less than 20 inches annual precipitation), 0.038 (representative
of moderate precipitation regions with 20-40 inches annual precipitation), 0.057 (representative of
wet regions with more than 40 inches annual precipitation), 0.17 (reported median for bioreactor
landfills), and 0.22 year(reported average for bioreactor landfills).
The analysis suggests that the annual NMOCs collection rate for MSWLFs containing more than
3.35 million MT waste and located in arid and moderate precipitation areas are not expected to
decline below 50 Mg per year within 30 years after closure. The NMOCs collection rate for
MSWLFs containing more than 7.85 million MT waste is not expected to decline below 50 Mg
per year within 30 years after closure irrespective of its location. The annual NMOC collection
rate of MSWLFs that are operated as a bioreactor (with a decay rate of 0.17 yearor 0.22 year)
is expected to decline below 50 Mg per year within 20 years after closure. The bioreactor operation
can substantially shorten the GCCS operating timeframe, especially for medium and large
MSWLFs. The LFG flow rate is expected to decline below 5% of the peak flow rate after the
annual NMOCs collection rate reduces below 50 Mg per year for all of the scenarios modeled.
The analysis suggests that the annual NMOCs generation rate can spike above the NSPS threshold
of 50 Mg per year after GCCS operation termination with an adequate increase in the decay rate;
the decay rate was assumed to increase by 100% for the analysis presented in this report. The
deterioration of the final cover and ensuing moisture intrusion may result in an enhanced waste
decay rate. The stakeholders should consider monitoring and maintaining the final cover even after
GCCS termination until the remaining generation potential and leaching potential of the in-place
have declined to a level that is not likely to pose a risk to HHE in the event of the final cover
failure. In addition, landfill owners and regulators should also continue surface and subsurface
emissions and odor monitoring after GCCS termination to proactively identify signs of an increase
in LFG generation rate, surface and subsurface emissions, and odor issues, and have provisions in
place to resume GCCS operation, if needed.
Several states use percent remaining methane generation potential as a criterion for assessing the
impact of modifications to GCCS operation. The analysis suggests that the percent remaining
methane potential, however, is not an appropriate metric to assess the HHE impacts. A smaller
percent remaining methane generation potential at a large MSWLF may pose a greater risk than a
57
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small MSWLF with relatively higher percent methane generation potential. A mass-based
threshold (e.g, NMOCs threshold used by NSPS) would be a more appropriate metric than a
percent-based criterion for GCCS modifications/termination.
In addition to the evaluation of the study sites, this chapter also presents approaches that the landfill
owners/engineers can use to estimate the site-specific decay rate and remaining methane and
NMOCs generation potential using the LFG flow rate and composition data that are typically
measured at MSWLFs in the US.
'• '• II linn .lioni
The analysis presented in this chapter has the following limitations:
1. The study sites/cells size (with respect to in-place waste amounts) are smaller compare to
the size of currently operating MSWLFs in the US. Three of the sites contain less than 2.5
million metric tons of waste; MSWLFs containing less than 2.5 million metric tons of waste
are exempt from the requirements of NSPS. The landfill gas assessments pertaining to these
landfills may not be applicable to a majority of the MSWLFs operating in the US.
2. The available subsurface probes and facility structures data suggest infrequent detection
and exceedances at subsurface probes and in facility structures. However, monitoring data
were not available for all of the monitoring events conducted after closure. The facility
structure monitoring data were available for only three sites.
3. Limited LFG methane content data were available for a few of the study sites. For example,
methane content data were available for a limited duration for Site D. The methane content
of the LFG can vary considerably (typically varies between 40% to 55% by volume) over
time depending on modifications to the operating conditions of GCCS.
4. The site-specific estimates of k are based on the assumption that the LFG collection
efficiency is constant over the timeframe that the LFG data are used for regression analysis.
The actual measurement of LFG collection efficiency is typically not conducted for
MSWLFs. Any variation in the collection efficiency over the duration of the data used for
regression analysis would impact the accuracy of the decay rate estimates.
5. Site-specific methane collection potential (Lc) estimate was used for estimating the
remaining methane and NMOCs potential for the study sites. It should be noted that Lc
values are representative of the LFG collection rate and not the generation rate. The actual
remaining methane and NMOCs generation potential are expected to be greater than the
one estimated based on Lc values. AP-42 default L0 value of 100 m3 per Mg waste was
used to model LFG generation and corresponding NMOC generation at examples sites.
Variations in the methane generation potential would lead to a concomitant change in total
NMOC generation rate, which is modeled as a fraction of total LFG in this work. In
addition, an LFG collection efficiency of 100% was assumed. A lower collection efficiency
would result in proportionally smaller timeframes to achieve the NSPS thresholds for
GCCS termination than estimated in this study.
6. The NSPS-default NMOCs concentration of 4,000 parts per million by volume as hexane
was used for the analysis. NMOCs content of LFG has been reported to vary over a wide
range. As a point of comparison, AP-42 suggests using NMOCs concentration of 2,420
parts per million volume as hexane for landfills known to have the co-disposal of MSW
with non-residential waste and recommends 595 parts per million volume as hexane for
landfill containing only MSW (EPA 1998). EPA (2007) reported NMOCs to range from
233 to 5,870 parts per million by volume as hexane based on measurements from five
58
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MSWLFs in the US. The annual NMOCs collection rate would decline below 50 Mg per
year over a shorter duration than the estimates presented in the chapter if the actual NMOCs
concentration is lower than 4,000 part per million by volume as hexane. In addition, a
constant NMOCs concentration was used for the entire duration as the data pertaining to
the variability of NMOCs concentration in LFG over time are lacking. A change in
NMOCs concentration over time would concomittantly change the NMOCs emission rate
and GCCS operating duration.
7. The analysis only estimated the methane and NMOCs collection rates and the remaining
generation potential for methane and NMOCs. These estimates can be used for an
assessment of the HHE impacts. However, an HHE impact assessment was not conducted
in this study.
8. NSPS annual NMOCs generation rate thresholds were used for the analysis. However,
NMOCs emissions below the regulatory threshold may not necessarily suggest the absence
of HHE impacts from the long-term emission of NMOCs. Modeling approaches such as
life cycle assessment and contaminant fate and transport modeling coupled with risk
assessment can be used to estimate the HHE impacts of methane and NMOC emissions
associated with the termination of GCCS operation.
9. Only methane and NMOCs generation rates were modeled in this study. The LFG has been
reported to contain other trace contaminants (e.g., mercury) that may have significant HHE
impacts.
10. It should be noted that the literature-reported LFG data and/or values observed at other
sites cannot be used to reliably assess site-specific impacts of terminating GCCS operation
on HHE due to the large magnitude of variation reported in the literature. Therefore, the
data from the study sites should not be used as a proxy for conducting a reliable site-specific
post-closure period assessment. The approaches presented in this chapter can be used to
estimate the site-specific remaining methane and NMOCs generation rate potential and
rates that can be used as inputs for a more reliable assessment of the HHE impacts.
11. The site-specific estimates suggest that the decay rates recommended/specified by AP-42,
and NSPS and GHG Reporting regulations underestimate the waste decomposition rates
even at the study sites that were not operated as a bioreactor. Due to a small number of
sites analyzed in this study, a statistical evaluation could not be performed to estimate the
representative decay rate for different precipitation zones of the US. A majority of the
MSWLFs are required to report the collected methane amounts along with other data such
as annual disposal amounts to the EPA. Future research efforts should consider using the
reported data from these MSWLFs to estimate a site-specific decay rate of a large number
of MSWLFs in the US.
12. Site-specific conditions such as subsurface methane migration and odor issues may
necessitate GCCS operation beyond the NSPS-required timeframe to ensure compliance
with pertinent local, state, and federal regulations. The impact of the local and state
regulations on GCCS operation duration was not assessed in this chapter. Furthermore,
GCCS may be operated beyond the timeframe to comply with NSPS to support the ongoing
LFG beneficial use project, if applicable. The GCCS operation in these cases would,
probably, be not regulated by the NSPS and NESHAP regulations.
59
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6. Landfill Leachate
rview
Based on the PCC cost information gathered for the sites reviewed in this study, leachate
management cost constituted a significant fraction of the total PCC cost for several of the study
sites. The termination of leachate collection and management appears to represent a major cost
saving for the closed sites. Subtitle D regulations require maintaining and operating the leachate
collection system (referred herein to as LCS) in accordance with §258.40 until the "owner or
operator demonstrates that leachate no longer poses a threat to human health and the
environmentBoth the leachate generation rate and chemical characteristics dictate the potential
loading rate of contaminants to groundwater or surface water systems associated with leakage
through the bottom liner or seepage through landfill side slopes or toe after the LCS operation is
terminated. This evaluation, thus, assessed the study sites with respect to both leachate generation
rate (e.g, gallons per acre-day) and leachate chemical characteristics.
The federal Subtitle D landfill regulations do not specify numeric criteria for when a site's leachate
generation rate is sufficiently low enough to terminate PCC. However, the range of state-specified
criterion for LCS system operation termination/modification varies from no leachate generation to
historically low or steady leachate generation rate (FDEP 2016; VDEQ 2007; UDSHW 2012;
WADOC 2011; WDEQ 2000). The first objective of this chapter is to analyze the leachate
collection rate data from the study sites to assess whether any of the study sites have achieved no
or historically low and steady leachate generation rate after closure. As only collection rates and
not generation rates are monitored at MSWLFs, The reported collection rates were used as a proxy
for the leachate generation rate throughout the analyses presented in this chapter.
Objective 1. Analyze leachate collection rate data to determine whether any of the study sites have
achieved a zero or historically low and stable leachate collection rate after closure.
One of the desired outcomes from landfill closure is a long-term reduction in leachate production
from a site. As part of the process of developing PCC cost estimates, the long-term leachate
generation rate for a site must be estimated. Estimates of future leachate generation rates can also
be used to evaluate HHE impacts and necessary timeframes to achieve no or historically low
leachate generation rate thresholds used by some states for LCS operation termination. The second
objective of this chapter is to evaluate different approaches that can be used to estimate the long-
term leachate generation rate from a closed MSWLF.
Objective 2. Assess approaches that can be used to estimate a site's long-term leachate generation
rate. These approaches can also be used for estimating the timeframes for leachate
generation to decline below a state-specific threshold.
As described above, HHE impacts associated with leachate emissions after LCS termination
depend in part on the chemical characteristics of a landfill's leachate. Both the types of chemicals
occurring in the leachate and their concentrations are important for HHE impact assessment. The
Subtitle D regulations do not require routine characterization of leachate chemical composition,
and only some state regulatory programs require the collection and reporting of such data. As
groundwater contamination is the primary pathway for landfill leachate to impact HHE, some
states recommend that MSWLF owners conduct a comprehensive characterization of leachate
composition with respect to groundwater quality monitoring parameters (e.g., FDEP 2016,
60
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UDSHW 2012). The third objective of this chapter is to evaluate the comprehensiveness of the
available leachate quality data with respect to the parameters specified in the federal regulations
for groundwater monitoring.
Objective 3. Evaluate the comprehensiveness of the available leachate quality data with respect
to the parameters specified by the federal regulations for groundwater monitoring.
Leachate generated by a Subtitle D landfill, even at the end of the PCC period, is not expected to
be of sufficient quality for human consumption. The typical practice for assessing HHE impacts
from a closed landfill is to evaluate the quality of the water expected at the point of compliance
(for example, a groundwater monitoring well at the perimeter of the landfill site). This process
involves using the leachate generation rate and chemical characteristics as inputs into a chemical
fate and transport model to estimate concentrations of chemicals of concern at the compliance
point(s), and comparing these estimated concentrations to risk-based protection standards. Fate
and transport models require much more site information than leachate generation rate and quality.
A common screening step is to compare leachate concentrations directly with risk-based protection
standards to assess the degree of dilution and attenuation necessary to meet HHE objectives at the
point of compliance. This screening process also allows an assessment of which leachate
constituents (referred herein to as the contaminants of potential concern [CPOCc]) are likely to be
most limiting with respect to terminating LCS operation. The fourth objective of this chapter is to
conduct a screening analysis to identify leachate contaminants at the study sites with
concentrations greater than the respective risk-based threshold.
Objective 4. Conduct a screening analysis to identify the contaminants that have been frequently
measured in leachate above the respective human health risk-based protection
standards at the study sites after closure to identify the COPCs.
The concentrations of contaminants and the associated HHE impacts are expected to vary over
time. The screening analysis described above for identifying the COPCs does not account for
temporal variation of concentrations. The contaminants that were initially measured above the
respective protection standard but have declined below the respective risk-based protection
standards over time are not expected to be an HHE concern. Some state guidance documents
recommend demonstration of declining or steady concentration for the COPCs as a criterion for
evaluating termination or scaling back of LCS operation. The fifth objective of this chapter is to
evaluate temporal trends of the COPCs concentrations to assess whether the concentration of these
contaminants has declined below the respective risk-based threshold over time after closure.
Objective 5. Assess whether the concentration of the COPCs identified based on the screening
analysis has declined below the respective protection standard over time after
closure.
Robust analysis methods and approaches are critical for a reliable estimation of emissions. An
additional objective of this chapter is to present and discuss the approaches that can be used for
estimating the long-term leachate generation rate after closure. The analysis presented in the
chapter should not be considered as a comprehensive evaluation due to various assumptions and
limitations. These assumptions and limitations are presented along with the analysis. These
limitations are also summarized in the last section of this chapter.
61
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. Data Sources
A review of peer-reviewed literature and government publications (Moody and Townsend 2017;
Masoner et al. 2016; Townsend et al. 2015b; Masoner et al. 2014; Andrews et al. 2012; NCSU and
ERG 2011; Barlaz et al. 2010; Sizirici et al. 2010; US EPA 2006; Benson et al. 2005; Statom et al.
2004; Ward et al. 2002; Ecobalance 1999) was performed to understand the extent of existing data
on leachate quality from MSWLFs in the US. In summary, the majority of the sources identified
did not present data from closed landfills. Moreover, the data for each study were limited to a small
geographic region of the US or a relatively small timeframe.
An understanding of long-term leachate generation rate and quality are critical for assessing site-
specific HHE impacts of terminating or scaling-back of LCS operation. The availability of leachate
quantity and quality data was one of the criteria for selecting sites for this study. The study sites
are located in different regions of the US, and each has leachate quantity and quality data available
for several years since closure. While leachate quality data were available from all nine sites
reviewed for this study, Site G did not separately track the quality of leachate from its closed cells.
As mentioned earlier, federal regulations do not require routine leachate generation rate and quality
monitoring for MSWLFs. Some SEAs require routine leachate quantity and quality data tracking.
In addition, leachate collection volumes and quality from MSWLFs are commonly monitored
despite the absence of requirements to meet the contractual requirements of WWTPs in instances
where leachate is managed through disposal at a WWTP. Although national-scale data are lacking,
a handful of state-specific studies suggest that WWTP treatment of MSWLF leachate is the most
prevalent management option for leachate management (Reinhart 2017; Townsend et al. 2015b;
Wilson et al. 2011).
Table 6-1 shows a summary of the availability of leachate quality and quantity data from the nine
study sites starting from the closure year (i.e., Year 0) to the latest available data (i.e., 2016) at the
time of data collection for this study. Sites C and Site G had leachate data available from two
separate cells. Of the nine sites, four are constructed with a double liner with a secondary leachate
collection and removal system (herein referred to as leak detection system or LDS), and the rest
were constructed with a single/composite liner system. Six of the sites had leachate quantity data
available for more than ten years since closure. Two sites documented leachate recirculation after
closure and the amount of leachate recirculated at these sites since the closure was available.
Leachate quality data were available for eight of the sites, with most of the sites having at least ten
years of post-closure leachate quality data. The leachate quality data for these eight sites except
Site I were available for a majority of duration after closure. LDS leachate quality data were
available for two sites.
62
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Table 6-1. Summary of Availability of Leachate Quantity and Quality Data at the Nine Study Sites Since
Closure
Site
A
B
C
D
E
F
G
H
I
CI
C2
G3
G4
Geographical
Regions of the US (1)
SE
NW
NE
NE
SE
NE
SW
NE
NE
NE
NE
Closure Year
1998
2004
1998
1998
2009
2004
1997
1993
2000
2008
1997
Years since Closure
(as of 2016)
18
12
18
18
7
12
19
23
16
8
19
Has LDS <2>
No
No
Yes
Yes
No
No
No
Yes
Yes
Yes
Yes
Area associated with
Leachate Quantity
and Quality Data
(acres)
28
45
17.5
23.3
69
61
38
7.1
7.9
60
51
Years Data
Available for
LCS <3-4>
0-16
4-12
0-18
0-18
0-7
0-12
1-6,
8-19
7-20,
22-23
0-13,
15-16
0-8
0-12
Leachate Quantity
Years Data
Available for
LDS (2-4)
N/A
N/A
No
No
N/A
N/A
N/A
3-23
0-16
0-8
0-12
Was Leachate
Recirculated
After Closure?
Yes
No
No
No
No
Yes
No
No
No
No
No
Years Leachate
Recirculated
after Closure*4'
5-16
N/A
N/A
N/A
N/A
7, 9-
11
N/A
N/A
N/A
N/A
N/A
Amount
Leachate
Recirculated
74%
N/A
N/A
N/A
N/A
8%
N/A
N/A
N/A
N/A
N/A
(%)
Are Data
Available for
LCS/composite
leachate for
Yes
Yes
Yes
Yes
Yes
Yes
Yes
No
No
Yes
Yes
13
Closed Cell(s)?
(3,6)
s
©
0>
"ea
JS
u
a
0>
-J
Years Data
Available for
LCS/composite
leachate <3'4'6)
0-16
0-11
0-18
0-18
0-7
0-12
0-19
N/A
N/A
0-8
0-12
Data Available
for LDS(2)
N/A
N/A
No
No
N/A
N/A
N/A
No
No
Yes
Yes
Years Data
Available for
LDS (2-4)
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
3-8
0-12
Notes:
(1) SE: Southeast; NW: Northwest; NE: Northeast; SW: Southwest
(2) LDS: Leak detection system or secondary leachate collection and removal system
(3) LCS: Primary leachate collection and removal system
(4) Considering closure year as year zero (0)
(5) Percentage of leachate recirculated of the leachate collected since the closure
(6) Composite leachate represented mixed leachate collected from LCS and LDS
N/A Not applicable
63
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<' I lethodol -,,ir i i II i ¦ "i :tic 11 >i II >ii\, II "'in ill ' :hate Collection
'I
BW
A reliable estimate of the long-term leachate generation rate is needed to assess the impacts on
HHE from termination of LCS operation; leachate would not be collected and treated after
terminating the LCS operation. This section presents a description of the following three
approaches that were used to estimate the future leachate generation rate from the study sites: first-
order decay modeling, unsaturated flow modeling, and Hydrologic Evaluation of Landfill
Performance (HELP) modeling.
6.3.2. First-Order Decay Modeling
The leachate generation rate is dependent on the moisture-holding capacity of the in-place MSW
and in-place moisture content, which in turn depends on various factors, including the nature of
the waste and the annual infiltration during operation. A bulk of the moisture-holding capacity of
MSW is due to the presence of paper, cardboard, and other paper products. The moisture-holding
capacity of MSW is expected to decline over time as these materials biodegrade over time at a rate
that typically follows a first-order decomposition rate equation. The leachate generation rate is
expected to mimic the moisture-holding capacity loss rate. Following the first-order decay model,
the mass-loss rate and the associated rate of the moisture-holding capacity loss are expected to
decline, which in turn would result in a concomitant decrease in the leachate generation rate over
time. As presented later, leachate collection rate data for six of the study sites visually appears to
follow a first-order decay pattern. The following first-order decay equation was used to model
leachate generation for all the study sites except Site A, which recirculated 100% of the leachate
for a majority of the period after closure. EPA (2017) used a similar exponential-decay modeling
approach for estimating the future leachate collection rates based on the historical leachate
collection rates for Subtitle C landfills.
Qt = kt V0M0e~klt (Eqn. 6-1)
Where,
Qt = leachate flow rate at a given time (m3/year)
V0 = total leachate generation potential of MSW after closure (m3/Mg). The total leachate
generation potential represents the total leachate volume that would be generated over an
infinite time horizon per unit waste mass. It is not expected to be significantly impacted by
weather conditions such as rainfall and evapotranspiration after installation of the final
cover as it is designed to minimize infiltration into the landfill.
Mo = total in-place mass of waste in place (Mg, wet weight basis)
ki = first-order leachate collection decay rate (year"')
t = time since closure (years)
A regression analysis was conducted using the Excel Solver add-in to minimize the sum of the
squared difference between the modeled and measured data to find V0 and k corresponding to the
first-order decay curve that fit the reported leachate collection rate the best. Monthly leachate
64
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collection rate data, if available, were used for the analysis. A V0 of 0.5 m3/Mg and a ki of 0.05
yearwere used as the initial conditions for Solver to initiate the calculations.
6.3.3. Unsaturated Flow Modeling
%>#
Once the landfill is closed, only the stored moisture and hydraulic properties of MSW are expected
to impact the leachate percolation rate within the landfill. The weather conditions (specifically
rainfall and evapotranspiration) are not expected to have a significant impact as the final cover
minimizes precipitation infiltration and evapotranspiration out of the landfill. Assuming a unit
gradient (i.e., flow is only driven by gravity and not leachate pressure in the landfill), the following
equations can be used to model the annual/monthly leachate vertical percolation rate (leachate
generation rate) through MSW. These equations are the basis of the approach used by the HELP
model to simulate vertical percolation of moisture in the landfill (Schroeder et al., 1994). A similar
approach for simulating vertical percolation in landfills has been used by other models (e.g., US
EPA 1987).
Qi = Kui X A (Eqn. 6-2)
Where,
Qi = leachate percolation rate (leachate generation rate) (m3/s)
Kui = unsaturated hydraulic conductivity (m/s) at moisture content ft, calculated using
Eqn. 6-3 below)
A = landfill footprint (m2)
Km = Ks [^]3+I (Eqn. 6-3)
Bid = 6i ~ dr= drainage moisture content (v/v) (Eqn. 6-4)
(pd = (p — 6r = drainable porosity (v/v) (Eqn. 6-5)
Where, Ks = saturated hydraulic conductivity (m/s)
ft = total moisture content (v/v)
(p = total MSW porosity (v/v),
X = MSW pore-size distribution index (-)
ft = MSW residual volumetric water content of the landfill (v/v)
A spreadsheet model was developed to conduct an iterative regression analysis to estimate the
initial drainable moisture content (did), pore-size distribution index (1), drainable porosity (
-------
0l+i = 0l-|x9Oo£| (Eqn.6-6)
Where,
6i+i = total moisture content at the start of time step i+1 (v/v)
M= Mass of waste in the study cell (in kg)
Qi = leachate generation rate for time step i (m3/time)
The leachate generation rates were calculated for all timesteps (month or year) for which the actual
measurements from the site were available. The square of differences between each point of
modeled flow and actual site flow data are summed to calculate the SSE. The Excel function Solver
was used to minimize the SSE by changing initial drainable moisture content, pore-size
distribution index, drainable porosity, and saturated hydraulic conductivity values. The initial
drainable moisture content, drainable porosity, pore-size distribution, and saturated hydraulic
conductivity values that resulted in the lowest SSE provide the leachate generation rates that best-
fit the measured data for the site. This analysis was conducted for all sites except Site A, which
recirculated all of the collected leachate volume for a majority of the period after closure. Monthly
leachate collection rates, if available, were used for calculating the SSE for each iteration.
Table 6-2. Initial Conditions and Constraints for Calculating Km
Parameter
HELP Defaults
Initial, Min and Max
Values used for Modeling
Range Reference
MSW
MSW With Channeling
Initial
Min
Max
Drainable
Porosity (cp- 0r)
0.671
0.168
0.50
0.45
0.62
Townsend et al.
(2015a)
I
0.47
0.55
0.50
0.26
0.70
Calculated from default
HELP 0wp and 0fc
Drainable
Moisture
Content (0i=o - Or)
Based on Climate
0.13
0.04
0.62
Townsend et al.
(2015a)
Saturated
Hydraulic
Conductivity, Ks
(cm/s)
0.001
10"5
3xl0"7
0.25
Townsend et al.
(2015a)
6,3,4, HELP Modeling
Neither the first-order decay nor the unsaturated flow modeling approach accounts for the impacts
of moisture influx through the final cover defects or leachate recirculation and leakage through the
liner on the leachate collection rate. The HELP model offers capabilities to assess the impacts of
moisture fluxes associated with the final cover and bottom liner defects/damages, and leachate
recirculation on leachate generation/collection rates (US EPA 1997). The HELP model is a
commonly used model by landfill engineers and regulators to estimate the long-term leachate
generation rate and head on the liner. Due to the prevalence of its use by landfill designers and its
additional capabilities, the use of HELP for estimating the future leachate collection rate from
closed landfills was evaluated.
For all the study sites, with the exception of Site A, simulations were performed with the HELP
model to estimate leachate collection rates after closure. At Site A, leachate was extensively
recirculated. Although HELP allows leachate recirculation modeling, it is difficult to specify time-
varying leachate recirculation rates to simulate the site-specific operating conditions.
66
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Perfect geomembrane-clay liner contact without any pinhole/defect was assumed for the final
cover and bottom liners. The final cover and the bottom liner configuration for the study sites were
used for HELP modeling. The actual study cell areas were used for the simulations. The waste
height was estimated by dividing the in-place mass by an assumed density of 900 kg/m3 (Townsend
et al. 2015a) and the landfill area; the estimated waste height was used for the simulations.
Precipitation and temperature data were obtained from the National Oceanic and Atmospheric
Administration (NOAA) weather station located closest to the site. The default values of other
climate data (humidity, wind speed, and growing season duration) were used from the nearest
HELP model default city to the site, and the site latitude was provided to estimate solar radiation.
A fair grass condition (or bare grass condition for an exposed geomembrane cap), a site-specific
evaporative zone depth, and maximum leaf area index was also used as a weather data input based
on the HELP default values for geographic locations corresponding to the study sites. These inputs
were used to simulate rainfall, temperature, and solar radiation for a 50-year period after closure
using the HELP model. These weather and vegetative layers (overlying the final cover
geomembrane) data are not expected to impact the modeled leachate generation rate as the
geomembrane was assumed to be free of pinholes and defects. The impact of these inputs and
compromises in the final cover can be modeled by specifying geomembrane defects in the HELP
model.
HELP does not have a field to enter the pore size distribution index but rather calculates it based
on the entered wilting point and field capacity of the waste (Eqn. 6-7). A HELP default field
capacity for MSW with the channeling of 0.073 (v/v) was used to estimate the wilting point of
MSW using Eqns. 6-7 and 6-8 below using the pore-size distribution index estimated based on the
unsaturated flow model for each site.
The drainable porosity and initial drainable moisture content, estimated based on the unsaturated
modeling, were used to estimate the total porosity and the initial moisture content using Eqn. 6-5
and 6-4, respectively, for each site; residual moisture content calculated using Eqns. 6-7 and 6-8,
and the field capacity value described above were used for these estimations. Similarly, the best-
fit saturated hydraulic conductivity, porosity, initial moisture content, and field capacity were
estimated using the unsaturated flow model described above were used for the HELP model run
for each site. The initial moisture content was varied, if needed, until the HELP model output most
closely resembled the measured data. The initial moisture content of all the other layers apart from
waste was specified to be the lesser of five times the field capacity or half of the porosity (presented
in Table 6-4).
Where,
6fc = field capacity of MSW (v/v)
6wp = wilting point of MSW (v/v)
0r = MSW residual volumetric water content of the landfill (v/v)
X = MSW pore-size distribution index (dimensionless)
(Eqn. 6-7)
er = [
0.014 + 0.25 6WP for 6WP > 0.04
0.6 6WP for 6WP < 0.04
(Eqn. 6-8)
67
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<' ! |v i h and Discussicii I h r III 1 ii;iil te
Measured Leachate Collection after Closure
Annual leachate collection volumes were available for all of the sites; however, these data values
were not necessarily available every year after closure. Also, the leachate collection rates for some
of the sites were not available for the entire year for some years. The available data were added
for each year for each site and divided by the number of days for which data were available to
calculate an annual average leachate collection rate (gallons per day). Although the total collection
volume from an MSWLF is expected to be dependent on the in-place volume, the leachate
collection rate is driven by its footprint and the vertical hydraulic conductivity of the in-place
waste. The annual average leachate collection rates were, therefore, normalized by the cell area
to calculate the annual average leachate collection rate per unit area of the landfill (gallons per acre
per day (GPAD)) for an equitable comparison of leachate collection rates among the study sites
with varying footprints. Figure 6-1 and Table 6-3 below presents the annual leachate collection
rate per unit area for the study cells (i.e., 11 total cells) from all nine sites. It should be noted that
the leachate collection rates for two study cells are presented for Site C and Site G. Since the
closure, the annual leachate collection rate across all 11 cells has varied from 3-2070 GPAD, with
a median of 92 GPAD and an average of 190 GPAD; 90% of the annual collection rate estimates
are less than 500 GPAD. Bonaparte et al. (2002) reported an average post-closure LCS leachate
collection rate from 11 closed cells at 3 MSWLFs (located in the northeast region of the US) based
on data collected over periods ranging from 8 to 64 months within ten years after closure. The
average leachate collection rate from these cells was reported to range from 439 to 7,480 GPAD.
With the exception of Sites A and E, which recirculated leachate after closure, the most recent
annual leachate collection rates for all the site cells were below 100 GPAD. The median annual
leachate collection rate among the non-leachate recirculating sites varied between approximately
8 and 268 GPAD. The wide variation of the leachate collection rates among the sites is potentially
due to differences in site-specific precipitations and operating conditions (e.g., leachate
recirculation). Site A has consistently managed leachate through recirculation over a majority of
the period since closure. As shown in Table 6-1, Site E recirculated leachate a substantially lower
fraction of leachate and for a shorter period than Site A. Only 8% of the leachate collected since
the closure has been recirculated at Site E.
The reasons for a spike in leachate collection rate from the study cell G3 in the 12th and 13th year
after closure and from G4 in the 10th and 11th year after closure is unknown potentially due to
complexity of monitoring the conditions that can spike leachate collection rates. These conditions
include the final cover geomembrane damages/defects, and storm water run-off channeling into the
leachate collection system. The site records documented multiple instances of the flow meter
drifting out of calibration at Site D. The leachate collection rates reported for Site D might,
therefore, not be representative of the actual flow rates.
68
-------
Site A
Site B
Site CI
Site C2
SiteD
SiteE
Site F
Site G3
Site G4
SiteH
Site I
0 2 4 6 8 10 12 14 16 18 20 22 24
Year After Closure
Figure 6-1. Temporal Variation of Annual Leachate Collection Rates at Study Cell(s) After Closure
Table 6-3 presents the average annual leachate collection rates for the study sites after closure. As
expected, the leachate collection rate from Site A was greatest among all the study sites, potentially
due to extensive leachate recirculation at the site. The leachate collection rates for each site except
for Sites A, CI, and F show a generally declining trend. The leachate collection rate at Site A
showed a declining trend for five years after closure and showed an increasing trend during the 68
years after closure following the start of leachate recirculation during the 6th year after closure.
The leachate collection rate at Site A shows a general declining trend since the 8th year after
closure; approximately 74% of the leachate collected since closure has been recirculated between
year 616 since closure. The leachate collection rate from the study cell CI declined from 183 to
82 GPAD within a year after closure and has fluctuated between 48 and 109 GPAD since the year
after closure. None of the landfill cells examined exhibited trends indicative of reaching a steady
state leachate collection/generation rate.
The leachate collection rate for Site F, in general, has declined during the first nine years after
closure but has shown a gradually increasing trend during the subsequent 9-year period. The most
recent annual leachate collection rate for Site F represents the maximum rate since closure. The
monthly leachate collection rates were analyzed with respect to the monthly rainfall data to assess
the impact of rainfall on the leachate collection rate at Site F. Figure 6-2 presents the monthly
leachate collection rates and the monthly precipitation for the most recent two to three years for
Site F. The monthly leachate collection rate during rainy seasons over this period was much greater
(as high as -120 GPAD) than those during dry periods (~ 20-30 GPAD). A spike in leachate
collection rate with no or minimal time lag with rainy season is a strong indicator of stormwater
channeling into the leachate collection infrastructure at the site. The leachate quality data collected
at appropriate points of time (during the dry and rainy season), specifically the concentration of
P
<
%
O
0)
03
P4
e
o
o
O
o3
o
03
0)
H-l
1000 -
mm ~ T - t
o
~
A
¦
~
~
-O-
~
V
69
-------
leachate indicator parameters (e.g., chloride), can also be used to assess stormwater channeling
into leachate collection infrastructure. The chloride concentration of leachate from one of the
leachate collection sumps/wet wells in December 2014 (rainy season) was less than 100 mg/L
compared to 1,000 mg/L during June 2015 (dry season). This suggests that leachate constituted
less than 10% of the liquid extracted from this sump in December 2014, and stormwater constituted
the rest; leachate quality at the site was characterized twice per year. The location or the cause of
stormwater infiltration into the LCS could not be assessed based on the data available for the site
for this study.
<10"7 cm/s (1.34 inches/year). As a point of comparison, the
leachate leakage rate through a compacted clay liner (without a geomembrane) with a saturated
hydraulic conductivity of 10"7 cm/s is estimated (using Darcy's equation) to be 92.4 GPAD
assuming a unit gradient and saturated flow conditions. The compacted clay liner underlying a
geomembrane in a composite liner system is typically required to have a maximum hydraulic
conductivity of 10"7 cm/s. All the leachate would leak through a compacted clay liner with a
hydraulic conductivity of 10"7 cm/s and would not result in leachate accumulation over the liner
once the leachate generation rate declines below 92.4 GPAD if the site is just lined with a
compacted clay liner. The overlying geomembrane, due to its significantly lower equivalent
hydraulic conductivity of ~10"13 cm/s (for HDPE liner) (equivalent to 9.2x 10"5 GPAD), is expected
to be the limiting factor for the leakage rate through a composite liner. Based on the equations used
by the HELP model, leakage rate from a 1-cm2 defect in geomembrane underlain by 10"7 cm/s
hydraulic conductivity clay liner for good and poor liner contact case is approximately 0.1314 and
0.7234 gallons per day, respectively. The leachate is expected to accumulate above the liner and
may cause elevated pore pressure if LCS operation is terminated before achieving a leachate
collection rate of less than equal to the leakage rate through the primary liner. The leakage rate is
70
-------
expected to increase over time with an increase in the head on the liner after LCS operation
termination.
Excluding Site D data and the collection rate for Site G4 for Year 0 (2,070 GPAD), the annual
leachate collection rate (GPAD) varies over almost two orders of magnitude, as shown in Figure
6-1 and Table 6-3. The leachate generation rate would be primarily dependent on precipitation
intrusion into the landfill during its operation. The daily leachate collection rate per unit area
(GPAD) was converted to an equivalent rate in inches per year (which is the unit used for annual
precipitation) and divided by annual precipitation to assess the impact of precipitation on the
variability of leachate collection rates among the study sites.
Table 6-3. Leachate Collection Rate Summary at the Nine Study Sites Since Closure
Site
A
B
CI
C2
D
E
F
G3
G4
H
I
Precipitation (in/year)
51
56
47
47
46
30
22
46
46
47
51
Years Since Closure as of 2016
(Considering closure year as
year zero)
18
12
18
18
7
12
19
23
16
8
19
Years Since Closure
Annual Average Leachate Collection Rate (gallons per acre per day)
0
715
-
183
137
7
427
-
-
2070
252
122
1
601
-
82
108
6
395
36
-
526
302
122
2
426
-
71
98
21
332
51
-
340
134
122
3
353
-
66
84
3
233
42
-
492
129
112
4
359
180
68
79
7
266
28
-
232
122
102
5
320
165
78
73
10
338
33
-
366
119
93
6
465
108
57
58
9
373
25
-
379
119
83
7
845
113
58
59
14
243
-
152
77
102
81
8
1009
139
70
59
179
19
111
119
78
76
9
883
75
89
52
127
14
74
119
77
10
693
67
81
52
199
20
95
303
62
11
808
49
92
47
118
15
83
718
68
12
658
78
109
43
197
16
151
197
56
13
810
92
39
15
452
206
-
14
511
70
38
28
347
-
-
15
599
80
43
20
406
49
-
16
676
90
32
30
242
38
-
17
-
70
25
25
117
-
18
-
48
28
34
112
-
19
64
117
-
20
85
21
-
22
68
23
88
Average (All Years)
631
108
82
61
10
264
29
169
389
151
90
indicates that leachate collection data were not available for the site for that year
Shaded cells indicate future years (after 2016)
Figure 6-3 presents cell post-closure leachate collection rates as a percentage of the average annual
precipitation of the sites. The annual precipitation for each site was derived from NOAA data
(NOAA 2017) for the closest weather station to each site. The leachate collection rate for all the
study sites except Sites A, E, and G has been less than 5% of the annual rainfall. Excluding Site D
data and the collection rate for Site G4 at closure year, the precipitation-normalized leachate
71
-------
collection rate varies over an order of magnitude (0.5 GPAD/inch precipitation to 19.8
GPAD/inches) among the study sites excluding Site D. The smaller magnitude of the variability
in the precipitation-normalized leachate collection rate compared to the variability among the
leachate collection rate presented in Figure 6-1 suggests precipitation has a considerable impact
on leachate generation/collection as expected. Other factors, such as the effectiveness of the run-
on/run-off control system, size of the working face area that is exposed to the precipitation during
active operation, and waste composition, also impact the moisture influx into the landfill.
Site A
Site B
Site CI
Site C2
SiteD
SiteE
Site F
Site G3
Site G4
SiteH
Site I
Year After Closure
Figure 6-3. Annual Leachate Collection Rates as a Percentage of Average Annual Precipitation of
Study Cells after Closure
6.4.2. Modeled Leachate Collection Rates
A comparison of the measured and modeled data using the three approaches described above are
presented in Figures 6-4 through 6-13. In general, all three approaches provided comparable and
reasonable fits to the measured study site data for all the study sites except Site CI, D, and G3. It
should be noted that the y-axis scale and range used for Figures 6-4 through 6-13 are not the same
but vary based on the range of the measured and model flow rates. As discussed earlier, the Site D
measured leachate collection rate trend may not be representative of leachate generation due to
leachate flow meter calibration issues on several occasions and the draining of the leachate storage
pond before closure. It should also be noted that the Site F leachate collection rate diverges from
a decay-trend and begins increasing due to intrusion of stormwater into the LCS (as discussed
earlier). The leachate generation rate for Site B based on the HELP model shows a sharp decline
to an insignificant rate around the 22nd year after closure. This is due to the decline in the waste
moisture content below the field capacity used for simulation. As discussed earlier, a field capacity
of 0.073 (v/v) for MSW was used for the HELP simulations. HELP models the vertical percolation
of leachate only if the moisture content is above the field capacity. As discussed later, an initial
moisture content of 0.13 (v/v), which was very close to the assumed field capacity, was used for
MSW at Site B for the HELP model. As mentioned earlier, leachate generation modeling was not
performed for Site A due to the complexity of modeling leachate recirculation using these
72
-------
approaches. All of the leachate collected at Site A was recirculated for a majority of the period
after closure.
The HELP model-based leachate generation rate for the first year was much higher than the second
year for many of the sites. This is probably is due to drainable leachate associated with the assumed
moisture content of the drainage layer. For some of the sites, the HELP model leachate generation
rate sharply declined to an unreasonably low value and then increased gradually before starting a
gradual declining trend; further analysis of the cause of this unexpected fluctuation in the modeled
leachate generation rate was not conducted. These unreasonably low flow rates from the HELP
model were disregarded and not included in the figures for Site CI, D, F, G3, and I.
Unlike the unsaturated flow and first-order decay modeling approached described above, HELP
does not have a regression analysis feature that can automatically adjust various design inputs and
media properties that would yield the leachate collection rates that match most closely with the
measured data. The leachate collection/generation rates estimated using the HELP model are based
on a mix of default parameters and the best-fit results from the unsaturated flow model. The
estimates that best-fit the measured flow rates can be obtained using the HELP model by iteratively
running the model by varying the waste properties (e.g., wilting point, field capacity, porosity, and
hydraulic conductivity) values until the SSE between the modeled and measured data is
minimized. The HELP model was not iteratively executed in this study due to time constraints to
assess whether or not the HELP model estimates, which are presented in this section, represent the
best-fit to the measured leachate collection rate data. The magnitude of the difference between the
HELP model results and actual site data, therefore, should not be used to interpret that the first-
order model and the unsaturated flow models are either more or less accurate than the HELP model
for estimate leachate generation.
As mentioned earlier, among the approaches described above, only the HELP model provides the
capability to assess the impact of precipitation intrusion into waste and leakage through the liner
as the result of defects/damages to the final cover and bottom liner, respectively. The first-order
decay and unsaturated flow modeling approaches do not account for the impacts of moisture influx
through the final cover and leakage through the liner on the leachate collection rate.
73
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Year After Closure
j-4. Measured and Modeled Leachate Collection Rates from Site B
Site CI
Measured Rate
HELP Model
Unsaturated Flow Model
First-Order Decay Model
0
10
20
30
40
50
Year After Closure
Figure 6-5. Measured and Modeled Annual Leachate Collection Rates from Site CI
74
-------
Year After Closure
Figure 6-6. Measured and Modeled Annual Leachate Collection Rates from Site C2
0 10 20 30 40 50
Year After Closure
Figure 6-7. Measured and Modeled Annual Leachate Collection Rates from Site D
75
-------
Year After Closure
Figure 6-8. Measured and Modeled Annual Leachate Collection Rates from Site E
Year After Closure
Figure 6-9. Measured and Modeled Annual Leachate Collection Rates from Site F
76
-------
Year After Closure
Figure 6-10. Measured and Modeled Annual Leachate Collection Rates from Study Cell G3
Year After Closure
Figure 6-11. Measured and Modeled Annual Leachate Collection Rates from Study Cell G4
77
-------
Year After Closure
Figure 6-12. Measured and Modeled Annual Leachate Collection Rates from Site H
Year After Closure
Figure 6-13. Measured and Modeled Annual Leachate Collection Rates from Site I
78
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Table 6-4 presents the values of the input parameters corresponding to the best-fit based on the
first-order decay, unsaturated flow, and HELP models. The estimates based on the unsaturated
flow model fit the measured data with an r2 > 0.5 for seven of the ten study cells modeled. An r2
of 0.5 means that the modeling approach captures or explains 50% of the variations in the leachate
flow rate over time at a given site, and the rest 50% is due to factors (e.g., stormwater intrusion
into the landfill) that are not accounted for by the model. The first-order decay model also fits six
of the ten study cells with an r2 > 0.5. Sites CI, D, and G3 cells had the lowest r2 value for both
models; lower r2 for Site G3 is probably due to a spike in leachate collection rate ten years after
closure. Site H had five months of very high initial flow that rapidly declined, and Site E had
periods of leachate flow that appeared to deviate from the overall general trend in the data during
2009 and 2010.
Note that the unsaturated flow model provides an estimate of drainable porosity and moisture
content, which are different from the porosity and initial moisture content used by the HELP
model. Table 6-4 also shows the half-life (i.e., timeframe for leachate collection to half of the rate
at closure) corresponding to the estimated decay rate (ki) for the first-order decay model. The half-
life for all of the study cells (except CI, D, and G3) ranges from 2.3 to 14.5 years. Neither of the
models provided a reasonable fit to the actual data for Sites CI, D, and G3 dues to the lack of a
consistent trend in the leachate generation rate for these sites. The r2 values corresponding to the
HELP model flow rates were not calculated as the model was not iteratively run to minimize the
SSE between the modeled and the measured data.
The in-place waste saturated hydraulic conductivity values estimated based on the unsaturated
flow model for all sites with r2>0.5 ranged from 9.0><10"6 to 3.7><10"3 cm/s, which is within the
range reported in the literature (Townsend et al. 2015a). It should be noted that the default field
capacity value was used for the HELP model run, and the wilting point was calculated using pore-
size distribution (from the unsaturated flow model). The reasonableness of the estimates for the
estimated wilting point could not be assessed due to a lack of data reported in the literature for this
parameter.
Table 6-5 presents the leachate collection rate 30 years after closure estimated using different
modeling approaches. The estimates based on the first-order decay model are, in general,
substantially lower than those obtained using the HELP model and the unsaturated flow model.
The HELP model estimates were very close to those from the unsaturated flow model. All the
estimates except for the first-order decay model estimate for Sites CI and G3 are lower than the
most recent leachate generation rates.
79
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Table 6-4. Values Used for Different Modeling Approaches
Parameter
Site
B
CI
C2
D
E
F
G3
G4
H
I
First Order Decay Model
Total leachate
generation potential of
MSW, V0 (m3/Mg)
18
91
12
66
23
14
179
88
11
10
Decay Constant, kj
(year-1)
0.16
0.02
0.09
0.00
0.11
0.05
0.01
0.97
0.17
0.07
Half-Life (Years)
4.3
-
7.7
-
6.3
14.5
-
0.7
4.0
9.8
r2
0.84
0.09
0.95
<0
0.65
0.48
<0
0.51
0.75
0.98
Unsaturated Flow Model
Saturated hydraulic
Conductivity, Ks (xlO"4
cm/s)
9.8
0.004
3.8
2500
0.09
0.15
0.007
0.07
37
1
Drainable Porosity, (pd
(v/v)
0.50
0.62
0.51
0.59
0.54
0.53
0.62
0.53
0.58
0.46
Initial Drainable
Moisture Content, 0od
(v/v)
0.12
0.53
0.18
0.12
0.35
0.23
0.51
0.49
0.22
0.19
MSW Pore-size
Distribution Index (X)
0.68
0.70
0.44
0.26
0.57
0.48
0.70
0.52
0.31
0.46
r2
0.80
<0
0.98
<0
0.62
0.55
<0
0.82
0.76
0.96
HELP Model
Saturated Hydraulic
Conductivity, Ks (xlO-4
cm/s)
9.8
0.004
3.8
2500
0.09
0.15
0.007
0.07
37
1
Porosity, cp (v/v)
0.508
0.627
0.522
0.615
0.554
0.541
0.62
0.53
0.607
0.477
Initial Moisture
Content. 0O, (v/v)
0.131
0.450
0.175
0.1
0.359
0.2
0.51
0.49
0.23
0.185
Field Capacity. 9fc
(v/v)
0.073
0.073
0.073
0.073
0.073
0.073
0.073
0.073
0.073
0.073
Wilting Point. 9wp, (v/v)
0.012
0.012
0.027
0.043
0.018
0.024
0.011
0.020
0.039
0.025
Table 6-5. Estimated Leachate Collection Rate at the End of 30 Years after Closure Based on Modeling
Approaches Used
Most Recent
Estimated Leachate Collection Rate 30
Coefficient of Correlation
Leachate
Years after Closure (gallons per acre per
(r2) between the Measured
Site
Collection Rate
day)
and Modeled Flow Rates
(gallons per
HELP
Unsaturated
First-Order
Unsaturated
First-Order
acre per day)
Model
Flow Model
Decay Model
Flow Model
Decay Model
B
78
'—0
22
3
0.80
0.83
CI
48
39
36
58
<0
0.09
C2
28
20
19
8
0.98
0.95
D
14
1
7
10
<0
<0
E
197
75
65
16
0.62
0.65
F
64
12
12
9
0.55
0.48
G3
88
78
84
154
<0
<0
G4
38
37
38
<1
0.82
0.51
H
78
29
27
1
0.76
0.75
I
56
34
31
15
0.96
0.98
80
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6.4.3. An Evaluation of the Performance of the Primary Liner System
The bottom liner is designed to intercept leachate and prevent or minimize migration to the
underlying surficial aquifer. However, in reality, geomembranes are not entirely impervious as
these allow vapor transmission and may have manufacturing and construction defects. Six of the
study cells are lined with a double bottom liner system and have a secondary leachate collection
system or leak detection system (referred herein to as LDS). Figure 6-14 presents a temporal
variation of the leachate collection rate from LDS for the study cells with a double-liner system
and available LDS collection rate data; LDS collection rates were not available for study cells CI
and C2.
1000
Q
<
%
O
tu
03
Pi
£
o
'+->
o
tu
"o
U
cn
Q
h-l
100 -
10 4
¦%—
— Site G3
o
Site G4
T
— Site H
A
Site I
o
0.1 -
0.01
^ a-A a a a "A A
1 ir ~-§ ® ^ Q °
° o o
O O
I
10
I
12
I
14
I
16
18
I
20
I
22
24
Year After Closure
Figure 6-14. Temporal Variation of LDS Collection Rates at Study Cell(s) of Each Study Site After
Closure
As shown in Figure 6-14, the LDS collection rates of these study sites ranged from approximately
0.1 GPAD (for study cell G4) to 123 GPAD (for study cell G3). Over 50% and 85% of the values
across these sites are less than 1.5 GPAD and 8 GPAD, respectively. The LDS rates, in general,
show an overall declining trend over time after closure. As shown in Figure 6-14, the LDS
collection rate of G3, in general, was higher than the other sites. Bonaparte et al. (2002) reported
an average post-closure LDS rate from 14 closed cells at 4 MSWLFs (located in the northeast
region of the US) based on data collected over periods ranging from 4 to 64 months within ten
years after closure. The average leachate collection rate from these cells was reported to range
from 0 to 1,767 GPAD.
The aggregate primary liner efficiency was calculated by dividing the sum of annual LDS
collection rates by the sum of the corresponding annual LCS and LDS rates for each study site for
all years with available LCS and LDS collection rates. This analysis assumes that the leachate
leakage through the primary liner is the source of liquids in the LDS. As will be discussed later in
this section, in some cases, it appears that groundwater intrusion is a significant contributor to
liquids in the LDS. The assumption that 100% of LDS flow is the result of leachate leakage through
the primary liner provides the most conservative estimate of the primary liner performance. The
81
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analysis further assumes that the sum of the LCS and LDS flow rates represent the total leachate
generation rate. The analysis suggests that 2.8%, 0.4%, 0.4%, and 3.2% of leachate intercepted by
the primary liner leaked through the primary liner after closure for study cells G3, G4, H, and I,
respectively. In other words, the primary liner efficiency was estimated to be 97.2%, 99.6%,
99.6%, and 96.8% for study cells G3, G4, H, and I, respectively. The analysis assumes that the
leachate leakage through the secondary liner is insignificant. The primary liner efficiency would
be lower than the estimates presented above if the leakage through the secondary liner is
significant.
There are notable differences in the configuration of the primary liner system among these sites,
which may contribute to the differences in the LDS rates observed at these sites. One major
difference in the primary liner system among the sites was geomembrane material. The study cell
G3 was lined with a 36-mil Hypalon liner, while the primary geomembrane of G4, H, and I was
60-mil HDPE. The primary liner efficiency for G3 (97.2% as presented above) was in the range
estimated for the other sites with an HDPE liner. It should also be noted that a part of cell G3 does
not have a geomembrane underlaying its LDS, making it more susceptible to groundwater
intrusion than other sites.
The second notable difference is the nature of the layer underlying the primary geomembrane. It
should be noted that the part of the cell G3 bottom liner, which was on the slope of the adjacent,
is a composite liner, and the rest of the area was constructed in two phases with a double liner with
two varying configurations. The leachate collection rates were not available for these distinct areas
of G3. The primary liner of G3 (a part), G4, and Site I was constructed with a geomembrane
underlain by a low permeability GCL or compacted clay liner. The primary geomembrane of Site
H and a part of G3 was underlain by a high permeability layer (geonet for Site H and sand/granular
layer for part of G3); schematics depicting these liners configurations are presented in Attachments
G, H, and I.
The low permeability layer underlying the primary geomembrane of the Sites G3 (a part), G4, and
I is expected to impede leakage through any liner pinholes and defects, while a high permeability
layer underlying the primary geomembrane of Site H is expected to enhance leakage through the
geomembrane pinholes/defects. The leachate leakage rate through the primary geomembrane at
Site H is, therefore, expected to be greater than those at Sites G3 (part), G4, and I. However, for
the comparable primary leachate collection rates, the LDS collection rate at Site H was, in general,
lower than the other sites. The analysis suggests that other factors such as the head on the liner,
liner construction quality (e.g., frequency of construction defects), and groundwater intrusion
rates, which are typically unknown, may have a greater influence on the LDS collection rate than
the geomembrane material and/or the nature of layer underlying it.
The LDS collection rates at the study sites were substantially higher than those corresponding to
the equivalent hydraulic conductivity of geomembrane (e.g., ~10"13 cm/s for HDPE liner)
suggesting that the liquids in the LDS are primarily contributed by leachate leakage through
pinholes and construction defects in the primary liner and/or groundwater intrusion. The intrusion
of groundwater into the LDS has been reported to impact the LDS collection rate for Subtitle C
landfills (EPA 2017a). A review of groundwater well depth-to-liquid readings, site topography,
and bottom liner design for both Sites H and I show that the groundwater table was near and (at
least at certain times of the year) may have exceeded the bottom-most elevation of the liner. A
potential, therefore, existed for groundwater intrusion into the LDS and, ensuing dilution of the
LDS leachate. The available LDS leachate quality data were analyzed and compared with the LCS
82
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leachate quality and groundwater quality to assess the relative contributions of these sources to the
liquids collected from the LDS.
Leachate quality data were available for leachate collected from the LCS and LDS for Site I and
from the LDS, and LDS-LCS composite for Site H. Because more than 99.5% of the leachate
collected from Site H was primary leachate, it was assumed that composite leachate is
representative of LCS leachate quality. Several commonly detected parameters that have been
analyzed at both sites for primary leachate, secondary leachate, and groundwater were analyzed in
this study to assess the potential of groundwater intrusion into the LDS. The leachate quality data
for Site G were not analyzed as the available data represented the quality of leachate composited
from the closed and active cell(s) at the site.
Figure 6-15 presents a comparison of the distribution of chloride, the concentration of leachate
collected from LCS and LDS as well as groundwater at each of the two sites. It can be seen that
chloride concentration in LDS leachate is approximately an order of magnitude lower than that of
the respective LCS leachate for both the sites. The median chloride concentration in LDS leachate
is approximately two orders of magnitudes greater than the groundwater chloride concentration.
The leachate and groundwater quality data available after closure were used for the analysis;
chloride concentrations from five and three upgradient groundwater monitoring wells were used
for Site H and I, respectively.
Site I
Site H
10000
hJ
~5b
Ch
o
e
0)
o
S3
o
O
0)
T3
4=
O
1000 -
100 -
10 -
1 -
0.1
• •
* i ;
± 1
L
I : \
r
A.
%
'%¦ °*<
o.
M.
%
V
n
o
O,
M.
Figure 6-15. Chloride Concentration Distribution of Primary (or Composite) Leachate, Secondary
Leachate, and Background Groundwater Quality for at Site I and Site H
Figure 6-16 presents a comparison of the concentration distribution of several organic compounds
detected in more than 30% of the samples in both primary (or composite) and secondary leachate
at each of the two sites. The detection limit was used as a surrogate value for non-detect
measurements. Total organic carbon (TOC) was detected in all the leachate samples at Site H.
83
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Tert-butyl alcohol was below the detection limit in ten out of 15 secondary and two out of 15
primary leachate samples at Site I. Tetrahydrofuran was not detected in four out of 15 secondary
leachate samples for Site I. As shown in the figures below, the median LCS (or LCS-LDS
composite) concentration of each compound is approximately an order of magnitude greater than
the LDS concentration. This difference is similar to the difference in chloride concentrations
shown in Figure 6-15.
Considerable differences in the concentrations of parameters that are indicators of MSW leachate
(e.g., chloride, organic compounds) between leachate collected from the LCS versus the LDS
suggest that leachate leakage from the primary liner may not be the primary source of liquids
collected from the LDS. Assuming that the attenuation in chloride concentration is solely attributed
to dilution by groundwater (and not due to precipitation as chloride salts), the magnitude of the
difference in chloride concentration suggest that leachate leakage through the primary liner
represents 10% of the liquids collected from the LDS for Sites H and I. Therefore, the primary
liner leachate collection efficiency at these sites is expected to be higher than the estimates
presented earlier assuming that the leachate leakage, if any, through the secondary liner is
insignificant. A more detailed discussion of groundwater monitoring data for the study sites is
presented in Chapter 7.
Site I
Site I
SiteH
10000
c
o
O
-o
£
C3
1000
100
10
T
10000
G
O
G
o
O
o
-G
o
&
G
X)
1000
100
00
g
o
-O
S-H
O
o
H
10000
1000
100
Primary Secondary
Primary Secondary
Primary Secondary
Figure 6-16. Organic Compound Concentrations in Primary (or Composite) and Secondary
Leachate at Site H and I
6.5. Leachate Quality
6.5.1. Leachate Quality Data Available for the Study Sites
Leachate quality data were gathered from annual, quarterly, and/or monthly reports of the sites for
the LCS and, if available/applicable, for the LDS. The data were compiled from the closure year
through the most recently available data (as of 2016). Leachate quality data were available for all
the study sites as availability of these data was one of the site selection criteria. Although leachate
quality data were available for Site G, the data from this site represented the quality of leachate
composited from closed and active cell(s) of the site. The data for Site G were not included in the
analysis presented in this section as these were not truly representative of a closed MSWLF.
84
-------
Leachate quality data were analyzed for eight of the nine study sites. Out of these eight sites, only
LCS-LDS composite leachate quality data were available for one site (Site H). The chain-of-
custody suggests that the composite samples were collected at the on-site plant intake at Site H,
indicating that these samples contain the LCS and LDS liquids in the proportion of the LCS and
LDS collection rates. The LDS flow collected over the quality data evaluation period represented
approximately 0.4% of the total LCS and LDS liquids volume over this period; therefore, the
composite leachate quality data are representative of the LCS leachate quality. Site B leachate
quality data represented the quality of leachate collected from the site's closed lined cells, and
leachate collected from the perimeter toe drain of the site's one unlined closed cell. Leachate
quality data for study cells CI and C2 were available separately. However, because C2 was
piggybacked over CI and both cells were closed together, leachate quality data of both the cells
were analyzed as a single dataset to represent the leachate quality of study Site C.
The HHE impacts associated with leachate emissions after LCS termination would depend on the
nature of the contaminants present in leachate and the associated concentrations. The impact to
groundwater quality is expected to be the primary pathway of HHE impacts associated with the
LCS operation termination/modification and the resultant leachate emissions. The RCRA Subtitle
D regulations specify parameters for routine monitoring of groundwater quality at MSWLFs; App
I and II of §258 list these parameters (referred herein to as App I and App II parameters). Some
states recommend a comprehensive characterization of leachate with respect to groundwater
quality parameters (e.g., FDEP 2016, UDSHW 2012). The leachate composition data for
groundwater quality parameters would be an important input for contaminant fate and transport
modeling for estimating concentrations of COPCs in groundwater at the point of compliance or
the point of exposure resulting from potential leachate emissions from LCS termination or
alternative leachate management options. This section presents an evaluation of the
comprehensiveness of the available leachate quality data with respect to the App I and II
groundwater monitoring parameters.
Figure 6-17 shows the count of all the parameters and that of App I and App II parameters
measured for leachate quality at each study site at least once. App I and App II of §258 include 62
(15 metals and 47 organics) and 215 parameters (over 90% are organics), respectively. All of the
App I parameters are also included in the App II list. A few of the sites measured the concentration
of different congener compounds of App II parameters such as polychlorinated biphenyls (e.g.,
Aroclor-1016, Aroclor-1221) and chlordane (e.g., alpha-chlordane, beta-chlordane, and gamma-
chlordane) in the leachate. While counting the number of App II parameters analyzed at each site,
all congener compounds of a parameter measured at the site were grouped and counted as one App
II parameter. However, all the congeners measured were counted individually and included in
"other parameters."
As presented in Figure 6-17, the parameters monitored varied widely among the sites. Leachate
was most comprehensively monitored at Site F among all the study sites. All but three App II
parameters were analyzed for Site F leachate at least once after closure. In general, leachate at all
of the sites except Sites A, B, and C was monitored for most of App I parameters at least once after
closure. All the App I parameters were monitored for leachate at least once since closure at only
three of the eight sites (Sites D, F, and H). Only 16, 24, 40, 49, and 54 out of 62 App I parameters
were monitored for Sites A, B, C, E, and I, respectively. More than half of the parameters listed in
App II were analyzed for Site E. Ten or less App II parameters (not included in App I) were
monitored for Sites A, B, D, H, and I. At least one data point was available for 49 and 73 App II
85
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parameters (not included in App I) for leachate at Sites C and E, respectively, after closure.
Examples of the other parameters monitored at the study sites include field parameters (e.g., pH,
dissolved oxygen), ions (e.g., chloride, ammonia, sodium), and water quality parameters (e.g., total
dissolved solids (TDS), biological oxygen demand (BOD), chemical oxygen demand (COD),
TOC). The primary reason for such a wide variability in the available leachate characterization
data among the study sites is the lack of leachate monitoring requirements in the RCRA Subtitle
D regulations. The leachate quality monitoring is primarily driven by the state regulations and the
monitoring requirements of the WWTP(s) accepting the leachate for treatment and disposal.
The available leachate quality data suggest that sampling frequency varied widely among the study
site from once per month at Site C to twice per year at Site D. The leachate samples were collected
from a single location at all the study sites except Sites C and F. Leachate samples were collected
from three and four locations at Sites C and F, respectively. The analysis frequency, in general,
varied with parameters for all the study sites. Not all parameters were analyzed at every sampling
event. For examples, major ions (e.g., chloride, sulfate) and field parameters were analyzed twice
per year, metals were analyzed once in five years, and some organics (e.g., 1,4 Dioxane) were
analyzed only once after closure at Site F. Only one measurement was available for each of the
eight study sites for at least one organic compound after closure. The analysis frequency also
appears to vary with time. For example, arsenic data for Site A are available once a month for a
few years after closure, while only annual data are available for recent years.
T3
0)
N
^.
'cS
a
<
5-h
-------
data available for less than half of the App I parameters. More than half of the study sites had
leachate composition data available for ten or fewer App II parameters (excluding App I
parameters). The characterization frequency varied among the sites from once per month to twice
per year. It also varied with time and contaminants. Only a single measurement was available for
a few of the organic compounds for each site after closure. Apart from the lack of the data for a
large number of App I and II parameters, the number of measurements available for the
constituents measured at the study sites may also limit a reliable HHE impact assessment.
6.5.2. Impact of Leachate Quality on Groundwater
The impact of leachate on HHE can be evaluated in three sequential steps, as suggested by the
performance-based functional stability approach of HHE impacts evaluation (ITRC 2006). The
first step is to identify the COPCs by comparing the contaminants concentrations in leachate to
respective risk-based standards such as drinking water (40 CFR §141), surface water (40 CFR
§445), and groundwater standards (40 CFR §258.54) as well as any state limits. The contaminants
that have always been measured at concentrations below the respective risk-based thresholds are
not expected to pose a risk to HHE even without any attenuation. In the second step, contaminant
fate and transport modeling can be conducted to identify COPCs (identified in the first step) that
would be above the respective risk-based standards at monitoring wells or at other points of
compliance (e.g., surface water discharge outlet) in the event of leachate release associated with
the termination or scaling-back of LCS operation. For the contaminants that are evaluated to have
concentration above the respective risk-based standard at the point(s) of compliance, fate and
transport modeling can be conducted to assess the concentration of the COPCs at the state-
designated points of exposure in the event of leachate release from the MSWLF in the third step
(ITRC 2006). As discussed in Section 6.4.1, leachate is expected to accumulate above the liner
system after the LCS operation termination. The accumulated leachate would leak through the
liner defects and percolate to groundwater over time. Any decline in liner performance over time
or liner failure would increase the leakage rate.
The assessment presented in this chapter was limited to the first step as a detailed site-specific
HHE risk assessment was beyond the scope of this study. Available leachate quality data for eight
sites were analyzed to identify COPCs; the data from Site G were not analyzed as these represented
the quality of leachate composited from the closed and active cell(s) at the site. This screening
effort involves comparing contaminants concentration in leachate to relevant contaminant-specific
and exposure-pathways-specific risk-based standards. The leachate discharge to surface water
bodies (e.g., a wetland or creek near a landfill site) or to groundwater are two primary pathways
for the HHE impacts associated with LCS operation termination. For the analysis presented in this
report, groundwater was assumed to be the most likely water source affected by a leachate
discharge after the termination of LCS operation. Thus, only the groundwater quality thresholds
were used for comparison with the reported leachate constituents concentrations. This is one of the
limitations of the analysis presented in this section. The HHE impacts associated with leachate
releases to the surface water should also be evaluated for a comprehensive assessment.
Examples of risk-based thresholds relevant to groundwater include the Safe Drinking Water Act
drinking water standards (MCLs promulgated under 40 CFR §141 or SMCL listed under 40 CFR
§143), regional screening levels developed by EPA (EPA 2018), and the state-specified levels.
The RCRA Subtitle D landfill regulations use MCLs (for contaminants with an MCL) as protection
standards for groundwater. EPA (2017a) used MCL/SMCL to assess the leachate quality at the
closed Subtitle C landfills. MCL and SMCL were used for the screening analysis presented in this
87
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chapter. In most cases, these water quality thresholds are based on risk to human health upon
consumption of the water. However, for some chemical constituents, water quality thresholds are
based upon impacts to aesthetics (e.g., taste, odor) or to aquatic organisms. The study sites leachate
quality data were evaluated for the following parameters:
a) Parameters with an MCL and secondary maximum contaminant level (SMCL)
b) §258 App I and II parameters as these are the parameters that are monitored for
groundwater to assess impacts of MSWLFs to HHE
c) Parameters used by EPA (2017a) to assess the leachate quality of Subtitle C landfills
Overall, the leachate quality for a total of 272 parameters were evaluated for eight sites. A list of
these parameters is presented in Table J-l in Attachment J. Table J-2 in Attachment J shows a
complete list of all the App I and App II parameters that were measured at least once in the leachate
at each of the eight sites. The parameters are arranged in the table based on the monitoring
frequency across the sites - the most frequently analyzed parameters are at the top of the table.
Arsenic, cadmium, chromium, copper, nickel, selenium, and zinc were analyzed for leachate
quality at all eight sites. Fourteen parameters (i.e., lead, 1,1,1-trichloroethane, p-dichlorobenzene,
1,1-dichloroethane, cis-l,2-dichloroethylene, trichloroethylene, vinyl chloride, acetone, barium,
carbon disulfide, ethylbenzene, dichloromethane, silver, and toluene) were analyzed for leachate
at least once at seven sites.
The concentration of all the leachate quality parameters measured at each of the eight sites (i.e.,
excluding Site G among the nine sites) were compared to their respective MCL. Of the 272
leachate quality parameters selected for evaluation in this study, only 100 parameters have MCL
or SMCL, and among those 100 parameters, data were not available for 32 parameters (as listed
in Table J-3 of Attachment J) for any of the study sites. Table 6-6 lists all the parameters measured
above their respective MCL/SMCL at least once at the eight sites during the timeframe leachate
quality was evaluated. Table 6-6 also shows the MCL/SMCL, the total number of data points,
percent of samples that exceeded the MCL/SMCL, percent of samples that were measured above
their method detection limit (MDL), percent of samples that were measured below their MDL but
the MDL was greater than MCL/SMCL, and the number of sites where each parameter was
evaluated. The parameters that were never measured above their MCL or SMCL are not reasonably
expected to present a risk to HHE. The MDL was used as the concentration for the measurements
that were below their MDL. For non-detect measurements where the MDL was not available (0.3%
of all non-detects), the data point was evaluated as a non-detect without specifying any
concentration (i.e., considered as a blank in the statistical analysis).
Among all the 100 parameters that have an MCL/SMCL, a total of 56 parameters exceeded their
respective MCL/SMCL at least once; 43 and 30 of these parameters are App II and App I
parameters, respectively. However, for 11 parameters (all organic compounds) that were reported
as below detection for 100% of the samples, the reported detection limits were greater than the
MCL/SMCL; these parameters cannot be conclusively determined to be below MCL/SMCL. This
data quality issue represents a constraint for a reliable HHE impact analysis for a majority of the
organic compounds and a few metals (e.g., antimony, beryllium, and thallium). Of the remaining
44 parameters, 29 parameters were measured above their MDL (i.e., detected) in less than 50%
samples. Relatively low concentrations of a majority of contaminants, when compared to the
respective MCL/SMCL, may not necessarily be indicative of stabilized conditions but may be due
to the lack of exposure of landfilled waste to adequate moisture needed for
hydrolysis/solubilization of contaminants from solid to liquid phase. The parameters that were
88
-------
measured below their MDL in more than 50% of the samples were not further analyzed. For many
parameters, MDLs varied considerably among the sites and even among sampling events at the
same site.
The remaining 15 parameters (highlighted rows in Table 6-6) were measured above their MDL in
more than 50% of the samples. Of these 14 parameters, six parameters (i.e., arsenic, TDS, iron,
manganese, chloride, and color) were measured above their MCL/SMCL in more than 94% of the
samples. Although color data were available for only one site and had only three data points, it is
reasonable to assume that color would probably exceed its SMCL for MSWLF leachate. Turbidity
was measured above its detection limit in all the data points. As presented in Table 6-6, over 85%
of turbidity measurements were greater than 5 NTU (maximum allowable MCL for turbidity).
Among the other eight parameters, aluminum was measured above its SMCL range (0.05 to 0.2
mg/L) in approximately 56% of samples, sulfate and fluoride were respectively measured above
their SMCL and MCL in 25% and 13% of samples. Chromium, barium, copper, and toluene were
observed above their MCLs in 0.3% to 6% of the samples. pH was measured outside of its SMCL
range in 4% of all the data points. As presented in Table 6-6, there are a few parameters (e.g.,
mercury, 1,1,2-trichloroethane) for which all the exceedances were a result of the MDL being
greater than the respective MCL. It should be noted that the number of measurements for each
constituent are not evenly distributed among the study sites. For example, approximately 75% of
816 measurements for arsenic are for Site C (including CI and C2).
Among the 15 parameters, five parameters (i.e., arsenic, chromium, barium, copper, and toluene)
are also App I and App II parameters (Table J-2 in Attachment J) and were analyzed at seven of
the eight study sites at least once. The 15 parameters that were measured above their MDL in more
than 50% samples were further evaluated for the range of concentration, median, and arithmetic
mean values among the study sites, as shown in Table 6-7. Additional leachate quality indicator
parameters that do not have any MCL/SMCL (such as BOD, COD, ammonia-nitrogen, sodium,
etc.) were also evaluated; the concentrations of these parameters also varied over a wide range
among the sites. Table 6-7 also shows the total number of data points, percent of detected samples,
and the number of study sites that analyzed for the respective leachate quality parameter. The
MCL/SMCL for each parameter is also listed as a point of comparison with the results.
As shown in Table 6-7, the concentration of parameters generally varied over a wide range among
the sites. The median concentration of seven parameters (i.e., arsenic, turbidity, TDS, chloride,
iron, manganese, and color) was observed to be higher than their respective MCL/SMCL
(highlighted rows in Table 6-7). The median concentrations of the other eight parameters with an
MCL or SMCL shown in Table 6-7 were below their respective MCL or SMCL. It should be
noted that the analysis presented in this report was conducted solely based on the federal
MCL/SMCL. States may have state-specific risk-based standards, and landfill owners may be
required to use state-specific standards for an assessment of impacts to HHE.
Furthermore, the leachate quality is typically reflective of the decomposition status of the bottom-
most waste layer and does not necessarily represent the degree of stabilization of the entire landfill.
A well-decomposed waste layer above the LCS may attenuate the concentration of parameters
such as BOD and COD that are commonly used to assess leachate and waste stability (Kjeldsen et
al. 2002). The biodegradable organics in leachate from fresher waste in the above layers would be
consumed as it percolates through a well-decomposed carbon-limited waste layer above the LCS.
89
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Table 6-6. Summary of Leachate Parameters that Exceeded the MCL or SMCL at Least Once at
the Study Sites
Parameter
Units
MCL'1'/
SMCL<2)
Total
Data
Points
Samples (%)
Landfills
Reporting
Detected
Above
MCL*1'
or
SMCL<2)
Not-detected
and MDL*4' 5)>
MCL/SMCL
Contaminants with a Primary Drinking Water Standard
Arsenic
mg/L
0.01
816
97%
94%
2%
8
Barium
mg/L
2
275
92%
5%
1%
7
Chromium
mg/L
0.1
665
97%
6%
1%
8
Copper
mg/L
1.3131
759
54%
1%
0%
8
Fluoride
mg/L
4
140
54%
13%
2%
4
Toluene
Ug/L
1000
324
54%
0.3%
0%
7
Turbidity
NTU
5(6'
246
100%
86%
0%
4
Nitrogen, nitrate
mg/L
10
486
34%
6%
0%
7
Nitrogen, nitrite
mg/L
1
360
6%
56%
54%
3
Antimony
mg/L
0.006
256
38%
83%
57%
6
Beryllium
mg/L
0.004
254
6%
22%
21%
6
Cadmium
mg/L
0.005
567
17%
6%
4%
8
Lead
mg/L
0.015131
677
25%
6%
2%
7
Mercury (inorganic)
mg/L
0.002
424
3%
5%
5%
5
Selenium
mg/L
0.05
285
25%
16%
9%
8
Thallium
mg/L
0.002
196
6%
87%
85%
6
Cyanide (free cyanide)
mg/L
0.2
701
35%
0.1%
0%
5
1,1,2-trichloroethane
Ug/L
5
187
0%
30%
30%
5
1,1 -dichloroethy lene
Ug/L
7
187
0%
27%
27%
5
1,2,4-trichlorobenzene
Ug/L
70
49
0%
22%
22%
3
l,2-Dibromo-3-
chloropropane
Ug/L
0.2
175
1%
99%
99%
5
1,2-dichloroethane
Ug/L
5
230
17%
28%
28%
6
1,2-dichloropropane
Ug/L
5
187
2%
29%
29%
5
Benzene
Ug/L
5
324
38%
31%
27%
6
Benzo(a)pyrene
Ug/L
0.2
26
0%
88%
88%
2
Carbon tetrachloride
Ug/L
5
187
0%
28%
28%
5
Chlorobenzene
Ug/L
100
263
34%
11%
11%
6
Cis-1,2-
dichloroethylene
Ug/L
70
284
24%
11%
11%
7
Di(2-ethylhexyl)
phthalate
Ug/L
6
80
10%
78%
69%
3
Dichloromethane
Ug/L
5
312
11%
39%
34%
7
Endrin
Ug/L
2
28
4%
25%
25%
2
Ethylene dibromide
Ug/L
0.05
173
0%
99%
99%
5
Heptachlor
Ug/L
0.4
53
6%
15%
15%
2
Heptachlor epoxide
Ug/L
0.2
33
27%
24%
24%
2
Hexachlorobenzene
Ug/L
1
26
0%
73%
73%
2
Hexachlorocyclopenta
diene
Ug/L
50
26
0%
8%
8%
2
Notes: (1) Maximum contaminant level (MCL) per 40 CFR §141; (2) Secondary maximum contaminant level (SMCL)
per 40 CFR §143; (3) treatment-technique-specific action level; water systems are required to take additional steps if
10% of tap water samples exceed the action level; (4) Method Detection Limit; (5) Number of parameters that were
not detected and had their MDL greater than MCL/SMCL; (6) varies with treatment technique; maximum allowable
is 5 NTU. Highlighted rows are for the parameters that were detected in more than 50% of the samples.
90
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Table 6-6 (contd.). Summary of Leachate Parameters that Exceeded the MCL or SMCL at Least
Once at the Study Sites
Parameter
Units
MCL'1'/
SMCL<2)
Total
Data
Points
Samples (%)
Landfills
Reporting
Detected
Above
MCL*1'
or
SMCL<2)
Not-detected
and MDL<5' 6)>
MCL/SMCL
gamma-BHC
Ug/L
0.2
26
4%
31%
31%
2
p-dichlorobenzene
Ug/L
75
323
48%
10%
10%
7
Pentachlorophenol
Ug/L
1
26
0%
92%
92%
2
Polychlorinated
biphenyls
Ug/L
0.5
18
17%
61%
50%
2
Styrene
Ug/L
100
276
5%
9%
9%
6
Tetrachloroethylene
Ug/L
5
263
1%
31%
31%
6
Toxaphene
Ug/L
3
12
0%
67%
67%
1
Trans-1,2-
Dichloroethylene
Ug/L
100
225
0%
13%
13%
6
Trichloroethylene
Ug/L
5
275
7%
31%
30%
7
Vinyl chloride
Ug/L
2
286
28%
51%
47%
7
Contaminants with a Secondary Drinkin
> Water Standard
Total dissolved solids
mg/L
500121
723
100%
99%
0%
6
Sulfate
mg/L
250121
758
85%
25%
0%
8
Chloride
mg/L
250121
785
100%
97%
0%
7
Iron
mg/L
0.3121
626
99%
99%
0%
7
Manganese
mg/L
0.05121
599
99%
98%
0%
7
PH
S.U.
6.5-8.5121
1066
100%
4%
0%
8
Aluminum
mg/L
0.05-
0.2121
114
73%
56%
0%
4
Color
C.U.
15(2'
3
100%
100%
0%
1
Silver
mg/L
0.1(2'
259
16%
3%
2%
7
Zinc
mg/L
5(2'
799
42%
1%
1%
8
Notes: (1) Maximum contaminant level (MCL) per 40 CFR §141; (2) Secondary maximum contaminant level
(SMCL) per 40 CFR §143; (3) treatment-technique-specific action level; water systems are required to take
additional steps if 10% of tap water samples exceed the action level; (4) Method Detection Limit; (5) Number of
parameters that were not detected and had their MDL greater than MCL/SMCL Highlighted rows are for the
parameters that were detected in more than 50% of the samples.
91
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Table 6-7. Summary of Leachate Quality Parameters that were Detected in more than 50% of Samples and
Exceeded the Respective Drinking Water Standard at least Once at the Study Sites
Parameter
Units
MCL'1'/
SMCL<2)
Mini-
mum
Maxi-
mum
Median
Mean
Total
Data
Points
Samples
Detected
(%)
Landfills
Reporting
Contaminants with a Primary Drinking Water Standard
Arsenic
mg/L
0.01
0.0047
0.997
0.05
0.06
816
97%
8
Fluoride
mg/L
4
0.03
19.5
0.6
1.8
140
54%
4
Barium
mg/L
2
0.003
100
0.45
1.2
275
92%
7
Chromium
mg/L
0.1
0.00041
10
0.03
0.08
665
97%
8
Copper
mg/L
1.3
0.0007
50
0.01
0.1
759
54%
8
Toluene
Hg/L
1000
0.11
3,400
5
40
324
54%
7
Turbidity
NTU
5(3'
0.1
>1,000
27
95
246
100%
4
Contaminants with a Secondary Drinkin
5 Water Standard
Total
dissolved
solids
mg/L
500(2'
200
17,000
4,140
4,297
723
100%
6
Chloride
mg/L
250121
1.8
4,580
1,260
1,284
785
100%
7
Iron
mg/L
0.3121
0.1
12,400
5.1
27
626
99%
7
Manganese
mg/L
0.05121
0.0004
34.1
0.29
0.8
599
99%
7
Color
C.U.
15(2'
180
800
520
500
3
100%
1
PH
S.U.
6.5-8.5121
4.94
11.54
7.5
7.5
1,066
100%
8
Sulfate
mg/L
250121
0.07
1,430
70
141
758
85%
8
Aluminum
mg/L
0.05 to
0.2121
0.023
27.3
0.20
0.6
114
73%
4
Other Contaminants
Specific
Conductance
Hinlio
s/cm
3.08
21,750
8,395
8,983
1,066
100%
7
Alkalinity
mg/L
1.1
8,550
2630
2696
697
83%
6
Ammonia-N
mg/L
2.1
3,914
470
532
730
100%
5
BOD141
mg/L
3.7
4,620
76
153
787
98%
7
COD141
mg/L
51.6
7740
987
1,186
606
100%
7
TOC4'
mg/L
17
2700
238
276
480
100%
5
Calcium
mg/L
20
470
71
89
587
100%
5
Magnesium
mg/L
10.7
360
85
100
588
100%
5
Potassium
mg/L
0.5
1,040
364
376
589
100%
6
Sodium
mg/L
58
25,000
960
1,095
648
100%
7
Nickel
Hg/L
4.55
1,900
120
126
634
97%
8
Notes:
(1) Maximum contaminant level (MCL) per 40 CFR §141
(2) Secondary maximum contaminant level (SMCL) per 40 CFR § 143
(3) Varies with treatment technique; maximum allowable is 5 NTU
(4) COD = chemical oxygen demand, BOD = biochemical oxygen demand, TOC = total organic carbon.
-Highlighted rows are for the parameters that had a median concentration greater than their respective
MCL/SMCL
Significant figures used vary. The significant figures used for the minimum and maximum concentration are
the same as those corresponding to the respective concentration values. The number of significant figures used
for the mean and median for each parameter is the same as that of the concentration value with the least number
of significant figures.
92
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A common practice to assess the potential for waste-derived leachate to impact HHE is to compare
the leachate concentration to a regulatory water quality threshold (which are usually based on safe
water for human consumption). This analysis does not imply that waste-derived leachate would
ever be consumed; instead, this analysis allows screening of the chemicals that might pose a
concern if the waste-derived leachate mixed with a drinking water source and the degree of dilution
and attenuation that would be required to alleviate such concerns. Parameters that are consistently
measured above their respective MCL/SMCL may have the potential to impact groundwater
quality, while those below should not have an impact. The concentrations of contaminants detected
in more than one-third of the reported leachate samples, and at a concentration greater than
MCL/SMCL, were divided by their respective MCL/SMCL. Leachate constituents with a larger
ratio of concentration to MCL/SMCL would require more dilution and attenuation or treatment for
mitigating the impact on groundwater. This ratio can be thought of as the dilution and attenuation
factor (DAF) necessary to ensure that leachate mixed with groundwater fall below the MCL/SMCL
for a COPC.
Figure 6-18 presents the distribution of the MCL normalized concentration of the frequently
detected parameters with MCLs. The median MCL-normalized concentration is greatest for
arsenic, followed by benzene. The detection limit was used as the concentration for the samples
below detection. As presented in Table 6-6, benzene was detected in less than 40% of the samples
and its MDL was greater than the MCL for 27% of the sample analyzed. Benzene was below
detection in over 85% of the samples with MCL-normalized concentration values greater than one
(1). Similarly, chlorobenzene and p-dichlorobenzene were below detection in all of the samples
with MCL-normalized concentration values greater than one (1). Fluoride concentrations were
below the respective MCL in more than 85% of the samples. Barium, chromium, and copper were
below the respective MCL in more than 90% of the samples. Among all the constituents with MCL
that are measured at the study sites (except for turbidity), arsenic exhibited the greatest DAF.
Figure 6-19 presents the distribution of the SMCL normalized concentration of the frequently
detected parameters with SMCLs. The median SMCL-normalized concentration is greatest for
color, followed by iron. As mentioned earlier, color data were available only for three sampling
events at one site. Iron concentration is more than ten times its SMCL for over 75% of the
measurements from the study site. A DAF of approximately 17 would be needed for the median
iron concentration to decline below its SMCL. A DAF of less than ten would be needed for TDS,
chloride, and manganese concentrations to decline below their respective SMCL.
It should be noted that the SMCLs of 0.3 mg/L for iron is based on aesthetics (e.g., color, taste)
and technical considerations (e.g., impact to water treatment process) and are not based on human-
health risk considerations. As a point of comparison, the human-health-risk-based regional
screening level developed by EPA for tapwater ingestion for iron is 14 mg/L (for Hazard Index of
1.0) (more than 50 times the SMCL), which is greater than the median iron concentration of 5.1
mg/L measured at the study sites.
93
-------
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h-l
u
c
o
!-h
c
(L)
O
c
o
u
100 -
10 -
1 -
0.1 -
0.01 -
0.001 -
0.0001 -
0.00001 -
• •
n
A •
• T
if
T
T
.0^° 40^° ^ ^
c
-------
concentration of leachate. Dilution with the groundwater is expected to the primary mechanism
for the natural attenuation of TDS and chloride. The impact of these processes on concentrations
at the point of compliance or receptor wells can be assessed using contaminant fate and transport
models such as EPA's Composite Model for Leachate Migration with Transformation Products
(EPACMTP). The fate and transport modeling should be performed to estimate the concentrations
of the COPCs at the receptor wells (i.e., point of exposure) to assess the potential HHE impacts.
As mentioned earlier, the analysis presented in this chapter only identifies the COPCs associated
with leachate. A detailed fate and transport modeling was not performed to evaluate the
concentrations of these CPOCs at the point of compliance or at the receptor wells for a
comprehensive HHE impact assessment
6.5.* iporal Analysis of Leachate Quality
The observations made based on aggregate analysis conducted in the previous section may not
conclusively be extended to individual sites due to issues such as wide variation in the number of
data points from individual sites. For example, 75% of arsenic measurements were associated with
Site C while less than 1% measurements were from Site I. The objective of the assessment
presented in this section is to present an analysis of the data for CPOCs and other major leachate
indicator parameters for individual sites. Site-specific distribution and temporal trends in leachate
quality since closure until the most recent available data (as of 2016) were evaluated for five
parameters (i.e., arsenic, iron, manganese, TDS, and chloride) that were recorded above their MDL
in more than 50% of their measurements and which had a median concentration above their
MCL/SMCL. Three major leachate quality indicator parameters (i.e., BOD, COD, and ammonia-
N) that do not have any specific MCL/SMCL and pH were also evaluated. The leachate quality
data for all sampling locations at a site were compiled chronologically since the closure. For
example, Site C had three leachate sample collection points; all the leachate quality data for a
parameter (e.g., pH) collected from the three sampling locations were organized in a chronological
order starting the year of site closure. The first sampling event of the closure year is presented at
Year 0 in the figures presented below. A similar data organization was performed for Sites E and
F, which had three and four leachate sample collection points, respectively.
6.5.3.1, Arsenic
The distribution of arsenic (non-speciated) concentrations in leachate for the eight study sites
(starting from the respective closure year) is shown in Figure 6-20. Among these sites, the arsenic
concentration varied in the range of 0.0047 to 0.997 mg/L, with a median value of 0.050 mg/L.
Approximately 94% of the total 816 measured concentrations among the sites were greater than
the arsenic MCL of 0.01 mg/L. Townsend et al. (2015b) observed a median (of the mean for each
site) leachate arsenic concentration of 0.044 mg/L from a review of data from 54 MSWLFs in
Florida. Based on the median arsenic concentration, a DAF of approximately five would be needed
to lower the leachate concentration of arsenic below its MCL. The arsenic concentrations varied
among the sites over 1-2 orders of magnitude; for example, at Site C, arsenic concentrations were
generally within the range of 0.01 to 0.1 mg/L, whereas, at Site H, arsenic concentrations were
generally greater than 0.1 mg/L. The large variation in the observed arsenic concentrations among
the sites may be the result of the variable nature of the waste deposited and/or maybe related to the
type of cover soil used at the site.
95
-------
Site A Site B Site C Site D Site E Site F Site H Site I
Figure 6-20. Distribution of Leachate Arsenic Concentration at Eight Study Sites Since Closure
%
0.14
0.12
tJ
2P 0.10
0.08 -
S3
o
e8
n
-------
arid zone, does not appear to be significantly different than that of Site C, which is located in a
higher precipitation zone. Large differences in concentration can exist among the leachate
collected from different areas of the landfills. For example, arsenic concentrations in leachate
collected from two of the sumps of cell C2 at Site C were substantially different, as shown in
Figure 6-21. The median arsenic concentration for sump C2S leachate is more than double the
median for sump C2N. The wide variation in arsenic concentrations of leachate collected from a
single cell suggests that the nature and age of the deposited materials (MSW, daily cover soil) have
a considerable influence on leachate quality.
A temporal trend of arsenic concentration in Sites B, D, and H is shown in Figure 6-22. As shown
in Figure 6-22, the observed temporal trend of the arsenic concentration varied among the sites. In
general, the arsenic concentration in the leachate at Site B appears to be decreasing over time after
closure, whereas at Site D, the arsenic concentration showed an increasing trend. At Site H, arsenic
concentration exhibited a variably increasing trend since closure. Measurements were available
for only three and five sampling events for Sites F and I, respectively. Temporal trends were not
evaluated for these two sites due to the small number of measurements available. Overall, arsenic
showed a declining trend for Sites A, B, C2, and E. The four most recent arsenic measurements at
Site B were below the MCL, as shown in Figure 6-22. The most recent set of measurements at Site
A, C2, and E were above the MCL. Arsenic exhibited an increasing trend for Sites CI, D, and H.
00
Eh
o
5-h
"H
V
o
G
o
o
o
'3
GO
5-h
<
Year After Closure
Figure 6-22. Temporal Variability in Leachate Arsenic Concentration of Sites B, D, and H Since
Closure
The contaminant mass release rate, which would dictate the magnitude of impact to groundwater
quality, may reduce over time due to declining leachate generation rate even in the case where the
contaminant concentration is increasing. Figure 6-23 presents the arsenic release rate per unit area
(lbs per acre per day) for Site H, which exhibited a variably increasing trend for arsenic
concentration but a decreasing leachate generation rate trend. The arsenic release rates per unit
area were estimated by multiplying the arsenic concentration to the corresponding daily leachate
generation rate (estimated from the available month leachate generation data) and dividing by the
landfill area. The arsenic release rate from Site H exhibited a slightly declining trend over time,
97
-------
which is indicative of declining HHE impacts with respect to leachate arsenic emissions in the
event leachate migrate into the environment after the termination of LCS operation.
0.0008
0.0007
0.0006
U ^
"o3
P4 0.0005
y u
£ u
03 Q,
"3 S 0.0004
* £
.a n
Ph ^
u a 0.0003
2 £
< ~
0.0002
0.0001
0.0000
0 2 4 6 8 10
Years After Closure
Figure 6-23. Temporal Variability in Arsenic Release Rate of Site H Since Closure
6.5.3.2. Iron
Figure 6-24 shows the distribution of leachate iron concentrations in seven study sites since the
closure year. Among these sites, the iron concentration ranged from 0.1 to 12,400 mg/L, with a
median value of 5.1 mg/L. Excluding an outlier (one data point with an iron concentration of
12,400 mg/L) at one site, iron concentrations generally varied between 0.1 and 121 mg/L. Over
99% of over 600 measured concentrations among the sites were greater than the iron SMCL of 0.3
mg/L. Townsend et al. (2015b) observed a median (of the mean for each site) iron concentration
of 7.74 mg/L from a review of leachate iron concentration at 56 MSWLFs in Florida. Based on
the observed median concentration at seven of the study sites, a DAF of approximately 17 would
be needed to lower leachate iron concentrations to below the SMCL.
The temporal trend observed for leachate iron concentrations varied among the study sites. As an
example, as shown in Figure 6-25, iron appears to be slightly decreasing over time in the leachate
of Site I, while the iron concentration at Site H does not show a clear trend.
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105 1
104 -
a io3 -
10-2 J 1 1 1 1 1 1 1
Site A Site B Site C Site E Site F Site H Site I
Figure 6-24. Distribution of Leachate Iron Concentration at Seven Study Sites Since Closure
Year After Closure
Figure 6-25. Temporal Variability in Leachate Iron Concentration of Sites H and I Since Closure
6.5.3.3. Manganese
Figure 6-26 shows the distribution of leachate manganese concentrations at seven of the study sites
after closure. Among these sites, the manganese concentration varied in the range of 0.0004 to
34.1 mg/L, with a median value of 0.29 mg/L. Townsend et al. (2015b) observed a median (of the
mean for each site) manganese concentration of 0.19 mg/L in leachate quality data from 11
MSWLFs in Florida. More than 98% of approximately 600 measured concentrations among the
sites were greater than the manganese SMCL of 0.05 mg/L. Based on the observed median
manganese concentration from the seven sites, a DAF of approximately 5.7 would be needed to
lower median leachate manganese concentrations to below the SMCL.
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Leachate manganese concentrations did not appear to stabilize with time at any site. As an
example, Figure 6-27 shows the temporal variation in leachate manganese concentration at Site B
and H. Site B appears to show manganese concentration varying in a smaller range just after
closure followed by a decreasing trend, while manganese concentrations at Site H do not show an
increasing or decreasing trend.
102
10-4
Site A Site B Site C Site E Site F Site H Site I
Figure 6-26. Distribution of Leachate Manganese Concentration at Seven Study Sites Since Closure
101 -i
10"2 -I 1 1 1 1 1 1
0 2 4 6 8 10 12 14
Year After Closure
Figure 6-27. Temporal Variability in Leachate Manganese Concentration of Sites B and H Since
Closure
6.5.3.4. Total Dissolved Solids (TDS)
Figure 6-28 shows the distribution of leachate TDS for six of the study sites starting from the
closure year (Site I had only one TDS measurement in this duration). Among these sites, TDS
varied over a wide range from 200 to 17,000 mg/L, with a median of 4,140 mg/L. As a point of
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comparison, Townsend et al. (2015b) observed a median (of the mean for each site) TDS
concentration of 3,723 mg/L in leachate quality data collected from 56 MSWLFs in Florida. More
than 98% of over 700 measured TDS concentrations among these sites were greater than the SMCL
of TDS (i.e., 500 mg/L). Based on the median leachate TDS concentration at six study sites, a DAF
of 8.3 would be needed to lower leachate TDS concentrations to below the SMCL.
18000 -i
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§ •
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8 8000 -
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Site A Site B Site C Site F Site H Site I
Figure 6-28. Distribution of Leachate TDS Concentration at Six Study Sites Since Closure
In general, TDS at the sites appears to be stable to slightly decreasing with time. As an example,
Figure 6-29 shows the temporal trend of TDS concentrations in Site C leachate. The TDS
concentration appears to be slightly decreasing without stabilizing in more than 18 years since
closure.
Site A Site B Site C Site F Site H Site I
Distribution of Leachate TDS Concentration at Six Study Sites
101
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16000
14000
12000
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18
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Year After Closure
Figure 6-31. Temporal Variability in Leachate Chloride Concentration of Sites C and I Since
Closure
6.5.3.6. pH
Figure 6-32 shows the distribution of leachate pH at eight study sites since closure. Leachate pH
among the sites ranged from 4.9 to 11.5 s.u.; however, in general, the leachate pH remained within
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the SMCL range of 6.5 to 8.5 s.u. Leachate at the study sites was generally alkaline, with a median
value of 7.5. Leachate pH at all of the sites except Sites B and F ranged from 7.5 to 9 s.u., which
is typical of the methanogenic phase, as reported by Kjeldsenetal. (2002). Townsendetal. (2015b)
observed a median (of the mean for each site) pH of 7.28 s.u. based on leachate quality data
collected from 57 MSWLFs in Florida.
12 -
11 -
10 -
^ 9 -
8 -
6 -
5 -
4 -
Figure 6-32. Distribution of Leachate pH at Eight Study Sites Since Closure
The leachate pH at each site temporally appears to be stable; an evaluation of the buffering system
and its role on leachate pH was not conducted in this study. As an example, a temporal trend of
leachate pH at Sites A and F are shown in Figure 6-33. The pH at both sites generally varied in a
smaller range and was alkaline; pH at Site A was slightly higher than at Site F.
SMCL=8.5
SMCL=6.5
1
f f
I
Site A Site B Site C Site D Site E Site F Site H Site I
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9
SMCL=8.5
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° Site F
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10 12 14 16 18 20
Year After Closure
Figure 6-33. Temporal Variability in Leachate pH of Sites A and F Since Closure
6.5.3.7. Biochemical Oxygen Demand (BOD)
BOD is one of the key leachate quality indicator parameters that may be used to infer the extent of
waste degradation. Morris and Barlaz (2011) reported that a statistical evaluation of BOD
concentrations in leachate could be used as a primary measure of estimating leachate impact on
HHE. Figure 6-34 shows the distribution of leachate BOD in seven of the study sites starting in
the year of their closure. BOD data for Site F were not available, and very limited data were
available for two other sites (i.e., Site H and I). BOD varied over a wide range from 3.7 to 4,620
mg/L, with a median of 76 mg/L. Townsend et al. (2015b) reported a median (of the mean for each
site) BOD concentration of 84.5 mg/L based on data from 31 MSWLFs in Florida.
105
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10000
00
Q
O
PQ
1000 -
100 -
10 -
1
Site A Site B Site C Site D Site E Site H Site I
Figure 6-34. Distribution of Leachate BOD at Seven Study Sites Since Closure
In general, BOD appears to gradually decrease over time since closure for all the study sites where
BOD data were available. Figure 6-35 shows the temporal variation in leachate BOD at Site B and
Site C. A decrease in leachate BOD with landfill age is typical as the biodegradable carbon
compounds contents such as cellulose and hemicellulose decrease in the waste. Statom et al. (2004)
observed a similar decreasing trend of BOD in the leachate at a landfill located in Florida. It should
be noted that biodegradable organics in leachate from fresher waste in the above layers would be
consumed as it percolates through a well-decomposed carbon-limited waste layer above the LCS
and thus attenuating leachate BOD (Kjeldsen et al. 2002). The leachate BOD, therefore, is typically
reflective of the decomposition stage of the bottom-most waste layer and does not necessarily
represent the degree of stabilization of the entire landfill.
106
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00
Q
O
PQ
10000
1000 -
100 -
0
16
18
20
2 4 6 8 10 12 14
Year After Closure
Figure 6-35. Temporal Variability in Leachate BOD of Sites B and C Since Closure
6.5.3.8. Chemical Oxygen Demand (COD)
Figure 6-36 shows the distribution of leachate COD at seven of the study sites since closure. COD
data for one site were not available. COD ranged from 52 to 7,740 mg/L, with a median of 987
mg/L. Townsend et al. (2015b) reported a median (of the mean for each site) COD concentration
of 907.7 mg/L based on leachate quality data collected from 26 MSWLFs in Florida. The ratio of
the median values of BOD and COD after closure for all the study sites with the available BOD
and COD data except for Sites D and H was less than 0.1, which is a threshold that is commonly
used as an indicator of stabilized leachate. The ratio of median BOD to COD ratio for Sites D and
H after closure were approximately 0.28 and 0.12, respectively.
10000
Site A Site B Site C Site D Site E Site H Site I
Figure 6-36. Distribution of Leachate COD at Seven Study Sites Since Closure
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In general, leachate COD appears to be slightly decreasing after closure at all of the sites. Similar
to BOD, the COD data beyond ten years after closure were only available for two sites (i.e., Sites
B and C). Figure 6-37 shows the temporal variation in leachate COD at Site B and Site C. Site C
leachate COD appears to be slowly but consistently decreasing and does not appear to have reached
an asymptotic level in 18 years since closure. It should be noted that the leachate COD is typically
reflective of the decomposition stage of the bottom-most waste layer and does not necessarily
represent the degree of stabilization of the entire landfill. A well-decomposed waste layer above
the LCS may attenuate the COD concentration as leachate percolates through this waste layer
(Kjeldsen et al. 2002).
10000
Q
O
U
1000
100 -
0
18
2 4 6 8 10 12 14 16
Year After Closure
Figure 6-37. Temporal Variability in Leachate COD of Sites B and C Since Closure
6.5.3.9. Ammonia-Nitrogen
Although ammonia-nitrogen (hereafter referred to as 'ammonia-N') is not an App I or App II
parameter and does not have any MCL or SMCL, it is one of the critical long-term pollutants from
the HHE impact perspective(Kjeldsen et al. 2002). It has been reported as the primary cause of the
acute toxicity of MSWLFs leachate (Kjeldsen et al. 2002). Figure 6-38 shows the distribution of
leachate ammonia-N concentrations from five of the study sites after closure. Ammonia-N data
were not available from the other study sites. Ammonia-N concentration ranged from 2.1 to 3,914
mg/L, with a median of 470 mg/L. As a point of comparison, Townsend et al. (2015b) observed a
median (of the mean for each site) leachate ammonia-N concentration of 360 mg/L based on
leachate quality data collected from 57 MSWLFs in Florida, which was over 120 times the risk-
based standard for ammonia (2.8 mg/L) in Florida at the time of the study conducted by Townsend
et al. (2015b). The DAF needed for the median ammonia-N concentration of 470 mg/L for the
study sites to decline below this risk-based standard would be 168, which is the highest DAF
among all the parameters evaluated in this study.
Site A ammonia concentration appears to be substantially greater than the other sites. Considerably
higher ammonia concentration at Site A is potentially attributed to a higher degree of waste
decomposition at Site A and progressive accumulation of ammonia in leachate due to its continual
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recirculation into the landfill. Ammonia-N released from waste decomposition (such as proteins)
does not degrade in the anaerobic environment of a landfill and is released with leachate.
oo
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c
(L)
o
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s
<
5000
4000 -
3000 -
2000 -
1000 -
Site A
Site B Site C
Site E
Site H
Figure 6-38. Distribution of Leachate Ammonia-N Concentration at Five Study Sites Since Closure
In general, the leachate ammonia-N concentration trends appear to be stable to slightly decreasing
since closure. As an example, the temporal variation in leachate ammonia-N concentrations at Site
C and Site H are shown in Figure 6-39. Statom et al. (2004) observed a similar stable to a slightly
declining trend in leachate ammonia-N concentration in a landfill located in Florida. Based on an
evaluation of data from 50 landfills in Germany, Krumpelbeck and Ehrig (1999) reported no
significant decline in ammonia concentration even 30 years after closure (Kjeldsen et al. 2002).
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2500
0 2 4 6 8 10 12 14 16 18
Year After Closure
Figure 6-39. Temporal Variability of Leachate Ammonia-N Concentration of Sites C and H Since
Closure
6.6. Impacts of Leachate Collection System Failure and Subsequent
Leachate Leakage
Evaluation of primary liner leachate collection efficiency was performed and described in Section
6.4.3 using leachate quantity data from two of the nine sites. This evaluation suggests that only a
small fraction of the leachate that collects over the primary liner system leaks through the liner.
The sites/cells analyzed in this study have been closed for 8-19 years (as of 2016) and are expected
to still be within the service life of the liner. However, the integrity of the liner and its effectiveness
in intercepting leachate beyond its service life is a concern from a long-term HHE impact
perspective. In addition, deterioration in the performance of the LCS due to factors such as
drainage layer clogging would impact the ability to efficiently pump leachate out from the
collection system. These issues would result in an increase in leakage from defects in the liner due
to increased head on the liner. Additional liner deterioration and defects from liner aging may
further impact groundwater quality and subsequently may pose a risk to HHE. Leachate leakage
rate through the bottom liner defects can be estimated using the HELP model or published
analytical and empirical mathematical equations (e.g., Giroud et al. 1997).
The seepage of leachate through the final cover defects on landfill side slopes in the event of
leachate built-up in the landfill after LCS termination may subsequently impact surface water
quality. The impacts of leachate releases through all potential pathways should be evaluated to
assess the long-term impact of the termination of leachate collection and liner failure.
Leachate leakage is not expected to pose a risk to HHE if leachate analyte concentrations are below
the respective risk-based health and ecological protection standards developed for the media of
interest, such as groundwater or surface water (e.g., EREF 2006). However, it should be noted that
leachate contains elevated concentrations of several contaminants (e.g., ammonia) that do not have
a federal or state risk-based standard. Contaminant fate and transport modeling can be conducted
to assess the impact of contaminant transformation and dilution on concentrations at the point of
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compliance and receptor wells, which may be located well away from site groundwater monitoring
wells. Additional information on leachate quality, as observed from leachate samples analyzed
from the eight study sites evaluated in this study, is discussed in Section 6.4. Approaches for
evaluating the risk associated with leachate leakage are discussed in more detail in Section 6.7.
'' -Iisideration i • mi . sessimg and Mitigatirv f II oachat' In11 ts
on HUE
As described by EREF (2006) and ITRC (2006), there are three general approaches that may be
successively implemented to demonstrate that terminating LCS operations or reducing the
frequency of leachate monitoring/management would not have HHE impacts. The first and most
conservative approach is to show that the concentrations of all regulated leachate analytes are
below regulatory standards (i.e., state or federal) for groundwater and surface water. The second
is to demonstrate that groundwater or surface water concentrations cannot be reasonably expected
to exceed the regulatory standards at permitted points of compliance (e.g., groundwater monitoring
wells). The third is to demonstrate the concentrations would not exceed any regulatory or health-
based standard at the nearest point(s) of exposure.
As described previously, out of the 68 parameters with an MCL/SMCL that were measured at the
study sites, 15 parameters were detected in more than 50% of the samples. Of these 15 parameters,
ten were found to exceed their regulatory limit in 10% or more of the samples, and only three of
these ten parameters represent primary drinking water standards: arsenic, turbidity, and fluoride.
While it should be noted that there were a number of regulated and non-regulated contaminants
that were not analyzed, an analysis and comparison of historic leachate quality to regulatory
standards provides a valuable first step for screening the contaminants that represent a risk to HHE.
Typical points of compliance for MSWLFs (that may be impacted by damage or imperfections in
the bottom liner or final cover system) are groundwater monitoring wells downgradient of the
landfill and stormwater outfalls (typically located at/near the site property boundary). For those
leachate analytes with concentrations identified above the regulatory limit(s) during the initial step,
site owners/operators may proceed to the second step and choose to demonstrate that these
contaminants have never exceeded applicable standards at the site's permitted points of
compliance. If exceedances have been observed, the temporal trends for these parameters should
be analyzed to assess whether the concentration has consistently been decreasing to the point
below and can be reasonably expected to remain below the regulatory protection standard.
Examples of plots showing the temporal trend of arsenic and lead concentrations at groundwater
monitoring wells are presented in Chapter 7.
However, if temporal trends in groundwater/stormwater contaminant concentrations cannot be
used to demonstrate the absence of risk to HHE at the point of compliance, site owners/operators
may proceed to the third step and conduct the contaminant fate and transport modeling to assess
whether these would exceed the regulatory or health risk-based standards at the nearest point(s) of
exposure. The EPA Industrial Waste Management Evaluation Model (IWEM) is one screening-
level tool commonly used to assess contaminant fate and transport in groundwater. As described
by EPA (2017b), for version 3.1 of the tool, the user inputs site-specific parameters, climate, and
hydrogeological conditions to estimate the concentration of contaminants at specific downgradient
groundwater monitoring wells. Specific states may have their own recommended/required
modeling software programs. For example, the FDEP provides guidance and information on the
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selection of risk-based corrective action fate and transport models in Florida Administrative Code
(FAC) Chapter 62-780.100.
The analysis presented in this report included only the parameters that are required to be monitored
for groundwater at MSWLFs and have a federal MCL/SMCL. However, leachate may contain
constituents that are currently not required to be monitored for groundwater or for which regulatory
standards do not currently exist. These parameters may have the HHE impacts. Some examples of
these parameters include (but are not limited to) ammonia, pharmaceuticals, plasticizers, and
certain types of pesticides and flame retardants (Moody and Townsend 2017; Masoner et al. 2016;
Andrews et al. 2012; Musson and Townsend 2009; Barnes et al. 2004). A comprehensive risk
assessment should include an evaluation of these parameters as well.
Finally, the low contaminant concentrations observed during operating and post-closure phases
may not necessarily be indicative of stabilized conditions but may be due to partial
stabilization/mineralization because of the lack of exposure of landfilled waste to adequate
moisture. The considerations discussed above do not address the impacts of a potential increase in
the emission of contaminants with leachate after the LCS operation termination. Moisture intrusion
into the landfill due to compromises in the integrity of the final cover after PCC termination may
trigger decomposition of biodegradable waste, if any, and result in a release of contaminants with
leachate that poses a risk to HHE. The impacts of the potential future moisture exposure on
leachate emissions can be evaluated and mitigated by actively recirculating leachate or adding
other liquids sources (e.g., groundwater) during operating and post-closure phases when the site is
actively monitored by owners/operators and regulators. Adequate leachate/moisture volume
should be added to expose the waste mass deposited in the landfill to the elevated moisture
conditions.
6.8. Summary
6.8.1. Data Availability
Leachate collection rate and chemical quality data were available for all the study sites as the
availability of these data was a key site selection criterion. Leachate collection rates were available
for two distinct cells at two of the sites (C and G), thus along with the nine other single-celled
study sites, leachate collection rates were analyzed for 11 total landfill cells. Of the 11 study cells,
six are equipped with a double liner and LDS. Leachate quantity and quality data were available
for most of the period after closure for all of the sites except for Site I; these data were available
only for the first 12 of 19 years since closure for Site I. Two sites documented leachate recirculation
after closure and the volumes of leachate recirculated at these sites after closure were available.
LDS leachate quality data were available for two of the sites. The leachate quality data for Site G
were not included in the analysis as these were not truly representative of the closed MSWLF.
6.8.2. Measured Leachate Collection Rate
The available data were added for each year for each site and divided by the number of days data
were available and the area of the cell to calculate an annual average leachate collection rate per
unit area of the cell in GPAD. Since the closure, the annual leachate collection rate across all 11
cells varied from 3-2,070 GPAD, with a median of 92 GPAD and an average of 190 GPAD; 90%
of the annual collection rate measurements were less than 500 GPAD. Leachate generation
continues at all of the sites. Except for Sites A and E, which recirculated leachate since the closure,
the most-recent annual leachate collection rate for the landfills examined was below 100 GPAD,
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with the median annual leachate collection rate varying between approximately 6.6 and 159.2
GPAD among the non-leachate recirculating sites. The leachate collection rates for all sites except
for Sites A, CI, and F exhibited a general declining leachate production trend. Based on reported
chloride concentrations and observed spikes in leachate collection during the rainy season, the
recent spike in leachate collection rate at Site F appears to be a result of stormwater intrusion into
the LCS infrastructure. None of the landfill cells examined exhibited trends indicative of reaching
a steady state leachate collection/generation rate.
6.8.Z 1 «• chate Collection Rate Modeling Approai 1 , luation
Three approaches were used to estimate the future leachate collection rate from the study sites:
first-order decay modeling, unsaturated flow modeling, and HELP modeling. In general, all three
approaches provided comparable and reasonable fits to the measured study site data for all the
study sites except Sites CI, D, and G3. The estimates based on the unsaturated flow model fit the
measured data with an r2 > 0.5 for seven of the ten study cells modeled. The first-order decay
model also fit six of the ten study cells with an r2 > 0.5. Sites CI, D, and G3 cells exhibited the
lowest r2 value for both models. Unlike the first-order decay and unsaturated flow model, the
HELP model was not iteratively executed to obtain the best-fit to the measured leachate collection
rate data but was based on a mix of default parameters and the best-fit results from the unsaturated
flow model.
6.8,4. Hydraulic Efficiency of Prim, ler
Six of the study cells are lined with a double bottom liner system. The LDS collection rates at four
of these sites ranged from 0.1 GPAD (for study cell G4) to approximately 123 GPAD (for study
cell G3); LDS data were not available for the study cells CI and C2. Over 50% and 85% of the
values (across these sites) were less than 1.5 GPAD and 8 GPAD, respectively. The LDS rates, in
general, show an overall declining trend overtime after closure. The aggregate hydraulic efficiency
of the primary liner was calculated by dividing the sum of annual LDS collection rates by the sum
of the corresponding annual LCS and LDS rates for all years with available LCS and LDS
collection rates. The primary liner efficiency was calculated to be 97.2%, 99.6%, 99.6%, and
96.8%) for study Sites G3, G4, H, and I, respectively. A comparison of LCS and LDS leachate
quality data available for two of the sites suggests groundwater intrusion into the LDS might be a
significant source of liquids collected from LDS. Therefore, the primary liner efficiency at these
sites is expected to be higher than the estimates presented in Section 6.4.3.
6.8,5 Hat \chate Quality Data
The comprehensiveness of the available leachate quality data for eight study sites was evaluated
with respect to the parameters specified in the federal regulations for groundwater monitoring at
Subtitle D landfills (App I and App II of §258). A total of272 leachate constituents were monitored
at least once among the study sites (excluding G). The number of chemical constituents that must
be monitored for groundwater as part of App I and II of §258 are 62 and 215, respectively; all of
the App I parameters are also included in the App II list. The parameters monitored varied widely
among the sites. Only three sites reported leachate constituent data for every App I parameter, and
one of these sites reported at least one measurement for all but three of the App II parameters. Two
study sites had leachate characteristic data available for less than half of App I parameters. More
than half of the study sites reported leachate constituent concentration data for ten or less App II
parameters (excluding App I parameters). The characterization frequency varied among the sites
from once per month to twice per year, and further varied with time and contaminant. Only a single
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measurement was available for a few of the organic compounds for each site after closure. Apart
from the lack of data for a large number of App I and II parameters, the small number of
measurements available for some of the constituents measured at the study sites may limit a
reliable HHE impact assessment.
6.8.6. Contaminants of Potential Concern
A screening analysis was conducted to identify the contaminants that have been frequently
measured in leachate above the respective risk-based protection standards at the study sites after
closure to identify the COPCs. Parameters that were never measured above their MCL or SMCL
are not expected to present a risk to HHE. Drinking water standards were used as the risk thresholds
for this evaluation.
Of the 272 leachate quality parameters selected for evaluation in this study, only 100 parameters
have MCL or SMCL, and among those 100 parameters, data were not available for 32 parameters
(as listed in Table J-3 of Attachment J) for any of the study sites. Among all 100 of the parameters
with an MCL/SMCL, a total of 56 parameters exceeded their respective MCL/SMCL at least once.
However, 11 parameters (all organic compounds) reported as below detection, but the reported
detection limits were greater than the MCL/SMCL; these parameters cannot be conclusively
determined to be below MCL/SMCL. MDL was greater than the respective MCL for a majority
of organic compounds for which data were available for the study sites. This data quality issue
limits the HHE impacts assessment for these parameters.
Of the remaining 44 parameters, 29 parameters were measured above their MDL (i.e., detected) in
less than 50% samples. The remaining 15 parameters were measured at concentrations above their
MDL in more than 50% of the samples. Of these 15 parameters, six parameters (i.e., arsenic, TDS,
iron, manganese, chloride, and color) were measured above their MCL/SMCL in more than 94%
of the samples. Among all the constituents with MCLs that were measured at the study sites,
arsenic and turbidity were the only primary MCL parameters that consistently exceeded the MCL.
Among the secondary parameters, greatest dilution and attenuation would be needed for iron for
its concentration to decline below its SMCL of 0.3 mg/L. A majority of the iron measurements
were below the regional screening level developed by EPA for tapwater for the ingestion pathway
for iron (14 mg/L).
Relatively low exceedance frequency of a large number of contaminants above the respective
MCL/SMCL may be due to incomplete flushing out of contaminants associated with limited waste
stabilization because of the lack of exposure of landfilled waste to adequate moisture. Moisture
intrusion into the landfill due to compromises in the integrity of the final cover after PCC
termination may trigger further decomposition of biodegradable waste, if any, and result in a
release of contaminants with leachate at levels that pose a risk to HHE. The impacts of the potential
future moisture intrusion on leachate emissions can be evaluated and mitigated to some extent by
actively recirculating leachate or adding other liquids sources (e.g., groundwater) during operation
and after closure while the site is actively monitored by owners/operators and regulators.
iporal Trends of the Contaminants of Potential Concern
The temporal trend for most of the COPCs varied among the study sites. Arsenic showed a
declining trend for Sites A, B, C2, and E. Four of the most recent arsenic measurements at Site B
were below the MCL. The most recent set of measurements at Sites A, C2, and E were above the
MCL. Arsenic concentrations exhibited an increasing trend for Sites CI, D, and H. The arsenic
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mass release rate from Site H exhibited a slightly declining trend over time. TDS and ammonia at
the sites, in general, appeared to be stable to slightly decreasing with time. In general, BOD appears
to gradually decrease since closure for all the study sites for which BOD data were available.
Leachate COD appears to be slightly decreasing after closure at all of the sites.
tiioiis
The analysis presented in this chapter has the following limitations:
1. The primary liner performance evaluation analysis assumes that the leachate leakage
through the secondary liner is insignificant. The primary liner efficiency would be lower
than the estimates presented if the leakage through the secondary liner is significant.
2. Only limited leachate composition data were available for a few of the study sites. For
example, only five arsenic measurements collected during the 3-4 years after closure were
available for Site I. The contaminant concentration can vary substantially over time.
3. The MDL for a large number of organic compounds monitored for the study sites was
greater than the respective protection standard. These parameters cannot be conclusively
determined to be below MCL/SMCL.
4. The screening analysis conducted to identify the COPCs was based on the assumption that
groundwater was the water source affected by a leachate discharge after the termination of
LCS operation. Impacts on surface water may be a more probable pathway for some sites.
The use of surface water-specific risk-based thresholds would have resulted in a different
set of the COPCs than presented in this chapter.
5. An analysis of the historical leachate collection/emission rates and the leachate
characterization data are presented in this chapter. These estimates can be used for an
assessment of the HHE impacts. However, an HHE impact assessment was not an objective
of the study. Modeling approaches such as life cycle assessment and contaminant fate and
transport modeling coupled with risk assessment can be used to estimate the HHE impacts
of leachate emissions associated with the termination of LCS operation at the point of
compliance or at the point of exposure.
6. The screening analysis conducted to identify the COPCs was based on MCLs/SMCLs.
More than 60% of the parameters that were measured at least once do not have an
MCL/SMCL. Some of these parameters may be of HHE concern. As mentioned above, the
analysis only included groundwater contamination pathway. The surface water
contamination pathway was not evaluated.
7. It should be noted that the literature-reported leachate collection rate and composition data
and/or values observed at other sites cannot be used to reliably assess site-specific impacts
of terminating LCS operation on HHE due to the large variation reported in the literature.
Therefore, the data from the nine study sites should not be used as a proxy for conducting
a reliable site-specific impact assessment. The data analysis approaches presented in this
chapter can be used to estimate the site-specific leachate collection rates, which can be
used as inputs for a more reliable assessment of the HHE impacts.
8. Due to the small number of sites analyzed in this study, statistical evaluation was not
performed to estimate the representative leachate collection rates for different precipitation
zones of the US. Several states require mandatory routine reporting of leachate collection
rates and composition data. Future research should consider using these data for statistical
evaluation to assess the variation of leachate collection rates and composition with weather
conditions (e.g., precipitation).
115
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7. Groundwater Monitoring Data
rview
Groundwater monitoring data serves as one of the key criteria for assessing the MSWLF
containment system performance and the impacts on HHE. RCRA Subtitle D regulations require
the monitoring of groundwater quality per §258(e) for MSWLFs, which describes the requirements
for groundwater monitoring and corrective action, as one of the PCC requirements for MSWLF
sites. Groundwater monitoring systems for MSWLFs are typically designed to monitor
groundwater of the uppermost aquifer. The monitoring system consists of a network of background
(or upgradient) and downgradient wells. The background well(s) are located to monitor
groundwater quality that has not yet been, nor expected to be impacted by the landfill, while
downgradient wells are installed at locations to detect groundwater impacts from the landfill.
Groundwater monitoring events are required to be conducted on a semi-annual basis throughout
the operating and PCC period unless an alternative frequency is approved by the appropriate
regulatory agency.
Groundwater monitoring is conducted in two phases: detection monitoring and assessment
monitoring. As discussed in Chapter 2, detection monitoring includes the routine sampling of 62
parameters (App I) (though 15 inorganic parameters in this list may be modified by States with an
approved program). As required by §258.54(c), assessment monitoring is initiated if a statistically
significant increase over background concentrations is identified for one or more of the detection
monitoring parameters in any downgradient monitoring well. To assess whether a statistically
significant increase has occurred, §258.53(g) requires the selection of a statistical method to
analyze the concentrations of the various groundwater monitoring parameters. Additional
guidance on the use of these statistical methods is provided in EPA (2009).
Assessment monitoring includes the monitoring of 215 parameters (included as App II of 40 CFR
§258); over 90% of these parameters are organic compounds. All of the App I parameters are
included in the App II list. The owner may be required to implement corrective measures (e.g.,
groundwater remediation) if one or more of App II parameters are detected at statistically
significant levels above the respective groundwater protection standard. The MCL is the
groundwater protection standard for the parameters for which an MCL has been promulgated. For
constituents for which the background concentration is higher than the MCL, the background
concentration is used as the protection standard §258.55(h). The background concentration or a
state-specified human health risk-based standard should be used as the groundwater protection
standard for the parameters without an MCL. The first objective of the analysis presented in this
chapter is to identify parameters that were detected above the respective MCL.
Objective 1. Analyze groundwater quality data available for the study cells to identify the
parameters that were detected above the respective MCL.
Contaminant concentrations and the associated HHE impacts are expected to vary over time. The
screening analysis described above for identifying the contaminants that were detected above the
respective MCL at the study sites does not account for temporal variation of concentrations. The
contaminants that have been frequently measured above the respective MCL initially but have
declined below the respective MCL over time are not expected to be an HHE concern. Some state
guidance documents recommend a demonstration that the contaminants have not been detected
above the respective MCL or state standard for the past several years in order to reduce the
116
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groundwater monitoring requirements. The second objective of this chapter is to evaluate temporal
trends of contaminant concentrations that have been frequently detected above the respective MCL
to assess whether the concentration of any of these contaminants has declined below the respective
MCL over time after closure.
Objective 2. Assess whether the concentrations of the contaminants that were frequently detected
above the respective MCL have declined below the MCL over time after closure.
It should be noted that only 51 out of 215 App II parameters (including lead and copper) have an
MCL. The groundwater protection standard for the parameters without an MCL is the background
concentration established based on background well(s). The groundwater protection standards for
these parameters were not available for the study sites. Some of these parameters may have a state-
specific risk-based standard, which may be used to assess the HHE impacts of these parameters.
The third objective of the analysis presented in this chapter is to present examples of parameters
that were detected above the respective state-specific risk-based standard.
Objective 3. Analyze groundwater quality data available for the study cells to identify examples
of parameters that do not have an MCL and that were detected above the respective
state-specific risk-based standard/criteria.
This chapter also presents a discussion about the impact of biogeochemical changes in the
aquifer system, LCS failure, and subsequent leachate release on groundwater quality.
. Data Sources
Groundwater quality data are valuable in assessing both current and long-term groundwater
impacts at the site. The groundwater monitoring data from the nine study sites were analyzed to
assess the nature and frequency of groundwater issues observed after closure; only the data
collected after closure were analyzed in this report. As discussed in the individual site descriptions
(Attachments A-I), groundwater monitoring reports (including statistical analysis of groundwater
monitoring data) were not available from all nine sites-data from several sites were only available
in an unprocessed form (e.g., were downloaded from a tabulated state database, were received as
a spreadsheet from the site owner/operator).
Of the nine sites reviewed, groundwater/environmental monitoring reports were available for six
of the sites. Of these six sites, the most recent groundwater monitoring report suggests that three
of the sites appear to be in detection monitoring, and two are in assessment monitoring (or the
state-equivalent monitoring phase), and one is under the corrective action phase. The presence of
an unlined or a lined cell not constructed under §258 regulations near the study cell(s) at two of
the three sites under assessment monitoring and corrective action phase complicates a reliable
assessment of the groundwater impacts of the study cells. The following data and information were
compiled and analyzed for each of the study sites:
• Number of upgradient and downgradient groundwater monitoring wells
• Groundwater flow direction
• The date range of available groundwater monitoring data
• The available groundwater quality data. Only concentration data for parameters that were
historically detected in groundwater samples were available for some sites.
• Parameters observed to have exceedances of the respective MCLs after the closure
117
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• Examples of parameters observed to have exceedances of the respective and state-specific
risk-based standards after the closure
• Year(s) since closure when the exceedance(s) was observed
Due to data inconsistencies or unavailability at most of the sites, the following groundwater
information was not reviewed and summarized as part of the analysis:
• Groundwater data analysis with respect to the respective state groundwater protection
standards
• Groundwater flow gradients (change in total head over the associated horizontal distance)
• Groundwater well maintenance logs/procedures
• Adjacent or on-site land use activities that may be impacting groundwater quality or flow
direction around the study site
Table 7-1 provides a summary of the key features for the study cells of all nine sites pertaining to
the groundwater quality data analyzed in this chapter. The groundwater quality data were available
for the entire duration since closure for all the study sites. The groundwater sampling frequency at
the study sites is either semiannually or quarterly. The available data suggest that the laboratory
analysis frequency appears to be contaminant specific for a majority of these sites. For example,
some parameters appear to be analyzed on a quarterly basis, while a majority of the organics are
analyzed once every five years at Site F.
Table 7-1. Summary of Groundwater-Related Features for Study Cells of Each Studied Site
Feature
A
B
C
D
E
F
G
H
I
Years since Closure (as of 2016)
18
12
18
7
12
19
23 (G3),
16 (G4)
8
19
Years of Available Groundwater
Data
14
13
19
8
13
20
18
9
19
Piggybacked Over Closed, Non-
Subtitle D Cell
V
V
V
Adjacent to Active Cell
V
V
Consistent Groundwater Flow
Direction
V
V
V
V
V
V
V
# Upgradient Wells Analyzed
l1
4
4
2
11
4
9
10
3
# Downgradient Wells Analyzed
23
17
62
16
492
4
17
20
14
Typical Groundwater
Monitoring Frequency (#/year)
2
4
4
2
2
4
4
4
4
1 the Site had a variable groundwater flow direction.
2 includes all the wells without a known location relative to the site.
It should be noted that a typical groundwater data evaluation process, as required by §258.53(g),
involves the selection and implementation of a detailed statistical method to evaluate groundwater
quality impacts. The analysis of groundwater monitoring results presented in this section and
described below is not intended to meet or serve as a substitute to the requirements of §258.53(g)
but was used for screening groundwater quality to identify parameters with an elevated potential
to cause groundwater quality impacts.
118
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II i III1 riges to Isolating and Understate !ii i-, -i 1 III" ¦
< II¦ 11IIII ¦'! .('HI IIII111 ¦
Many of the study sites received waste before the promulgation of §258 regulations and contain
cells that are not required to meet requirements of §258 (e.g., unlined or lined cells that were closed
before the promulgation of these regulations). At these sites, the cells under the §258 regulations
are commonly constructed immediately adjacent to or piggybacked over an old cell (i.e., a cell
excluded from §258 regulations) for more efficient use of available space. In these cases, it
becomes challenging to conclusively assess whether groundwater impacts (if existent) are
associated with the cell(s) under §258 or cells that are not under §258 regulations.
Groundwater monitoring data were reviewed for all the monitoring wells around the study cells
for those sites where the studied cell(s) was physically separated from any unlined cell (e.g., the
studied cell was not piggybacked over an unlined cell) and/or could be analyzed independent of
all other non-Subtitle D cells and active cells. Only five studied sites (Site C, D, F, H and I) had
closed lined cell(s) that were completely isolated from other cells (i.e., active cells under §258
regulations as well as closed cells not under §258 regulations), appeared to have a consistent
groundwater flow direction, and had groundwater wells installed at both upgradient and
downgradient locations from the closed study cell(s) of interest. The data from all the monitoring
wells were reviewed for the other four sites (Sites A, B, E, and G). The closed lined cell at Site A
is located immediately adjacent to an unlined cell. Issues such as this complicate the evaluation of
groundwater data for an individual cell that is contiguous to other closed MSWLF cells not
constructed under §258 regulations. States (e.g., Florida) may allow the initiation of PCC for a
closed cell or group of cells only if these cells are maintained and/or monitored separately from
the rest of the landfill.
Change in groundwater flow direction is another factor that may complicate a reliable groundwater
impact assessment. For example, while the groundwater flow direction for Site A was fairly
consistent during the first portion of the PCC period, however, the groundwater reports suggested
a highly variable groundwater flow direction in the recent years. Changes in land use/groundwater
use around the site and long-term weather patterns (e.g., sequencing of dry and wet years) are some
of the factors that may result in localized changes in surficial aquifer groundwater flow direction.
MSWLF owners/operators attempting to make a demonstration in support of reducing the
frequency of or terminating PCC should consider an evaluation of these changes around the site
and the long-term impact of these changes on groundwater flow direction and quality.
1 ' ii i'iidwatn liiii| :r- i ih'3 Stuch-''! 'ii''
In this study, existing impacts on groundwater were identified at the sites for three sets of
parameters: parameters with MCLs, parameters with SMCLs, and example of parameters without
an MCL or SMCL. First, the data were analyzed only for the App I and II parameters with an
MCL. An MCL is a federal human-health risk-based standard. All the parameters that are
measured at concentrations above the respective MCLs at least once in a groundwater monitoring
well after closure were identified.
The second set of parameters that were evaluated are silver and zinc exceedances (i.e., with respect
to the SMCL); silver and zinc are the only parameters included in the App I and App II lists that
have an SMCL. SMCLs are based on aesthetics (e.g., color, taste), cosmetics (e.g., skin
119
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discoloration), and technical considerations (e.g., impact to water treatment process) and are not
based on human-health risk considerations.
A large number of App I and II parameters do not have an MCL/SMCL; only 32 and 51 of App I
and App II parameters, respectively, have an MCL, and only two parameters (i.e., silver and zinc)
have an SMCL. The groundwater impact with respect to these parameters without any
MCL/SMCL can be assessed by comparing the concentrations measured at downgradient wells to
those of upgradient well(s) or to a state standard. An analysis of a few parameters at a couple of
sites is presented to highlight the importance of analyzing the parameters without MCL/SMCL for
assessing the groundwater impacts.
Table 7-2 summarizes the MCL exceedances identified at each of the sites. No exceedances were
identified at Site F for the range of data available for review (i.e., post-closure Year 8 through Year
19). The exceedance counts presented in the table do not include groundwater monitoring data for
which the MDL was above the respective MCL. Several instances were found where the parameter
analysis detection limit was greater than the respective MCL. For example, antimony, beryllium,
thallium, and arsenic detection limits were higher than the respective MCL on several sampling
occasions at Site C. Several organics measured at Site F had detection limits greater than the
respective MCL.
As shown in the table, only occasional MCL exceedances were identified following cell closure
among the studied sites. A majority of the organic compounds measured at the study sites were
not detected in groundwater samples. Over 65% of the observed vinyl chloride exceedances at Site
E occurred at a well close to the unlined cell at the Site. The arsenic exceedances observed at Site
B are potentially associated with the unlined cells as these occurred at the downgradient wells that
are closer to the unlined cell than the study cell. Only Sites G, H, and I appear to have had
reoccurring groundwater exceedances of one or two constituents. More than 50% and 10% of
measurements at Site H had beryllium and thallium concentration above its respective MCL.
However, approximately 40% and 30% of the observed beryllium and thallium exceedances
occurred at nine and five upgradient wells, respectively. Due to the exceedances observed above
the respective MCL at several upgradient wells, the exceedances at the downgradient wells cannot
be conclusively attributed to lined cells at Site H.
Although 107 arsenic exceedances at Site G is the third-largest exceedance count among all the
study sites, these exceedances represent less than 15% of the arsenic measurements at the site;
arsenic was not detected in more than 80% of the groundwater samples collected at the site.
Approximately 11% and 52% of arsenic exceedances observed at Site G occurred at four
upgradient and five wells downgradient to the unlined cell, respectively, at the site. The number
of arsenic exceedances at the other wells at Site G ranged from two to eight, after the closure of
the study cell. It should be noted that the site had an active cell at the time of the study.
All the parameters, except for mercury, listed in Table 7-2 for Site G, also exceeded the respective
MCL at the upgradient well(s). Approximately 13% and 38% of the lead exceedances occurred at
upgradient wells and wells downgradient/close to an unlined cell, respectively, at the site. Less
than 20% of the lead exceedances at Site G have occurred in the last ten years. Due to the
exceedances observed above the respective MCL at several upgradient wells, the exceedances at
the downgradient wells cannot be conclusively attributed to unlined and/or lined cells at Site G.
Only one arsenic measurement for a single well was available for Site E after closure.
120
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Table 7-2. Site Summary of App I and II Parameters with Exceedances Above the Respective Maximum Contaminant Level After Closure
Total
Number of
Upgradient
Wells with
Exceedances
Number of
Post-
Closure
Duration
Data
Available
(Years)
Site
Parameter
number
of wells
data were
available
Total Number
of
Measurements
Number of
Detected
Measurements
Number of
Exceedances
Downgradient/
Side-gradient
Wells with
Exceedance
Year (after
Closure) of
Exceedance(s)
Arsenic
23
214
61
16
0
9
13
0,1,3,13
Cadmium
23
177
9
1
0
1
9
1
Chromium
23
229
37
1
0
1
14
11
A
Lead
23
178
29
3
0
2
10
0-2
Thallium
23
226
78
2
0
2
10
12,14
Bis(2-Ethylhexyl)
Phthalate
1
1
1
1
0
1
1
0
Vinyl Chloride
23
207
10
5
1
1
14
0,1,3,8
B
Arsenic
21
911
597
18
0
2
7
0-1, 7-11
Vinyl chloride
21
1042
183
1
0
1
8
1
Antimony
9
149
1
1
0
1
19
18
Arsenic
9
149
5
1
1
0
19
3
C
Beryllium
9
149
15
3
0
1
19
14, 15, 18
Cadmium
9
149
7
1
0
1
19
11
Thallium
9
149
1
1
0
1
19
15
Beryllium
18
288
118
2
2
0
8
0
D
Chromium
18
288
284
2
0
2
8
1, 3
Lead
18
288
183
7
1
3
8
0-2, 4, 5, 7
Thallium
18
288
115
5
0
5
8
0
Arsenic
1
1
1
1
1
10
E
T etrachloroethy lene
45
482
8
1
0
1
12
0
Vinyl chloride
45
482
39
15
0
3
12
1-11
F
8
12
Arsenic
24
799
128
107
3
18
17
5-13, 16-22
Cadmium
24
990
22
6
1
3
18
6, 7, 11, 12, 19
G
Lead
24
975
320
53
3
12
18
5-9, 12-14, 16,
17, 19-22
Selenium
22
234
32
7
1
6
16
6, 7
Antimony
22
239
13
13
3
10
16
7, 11, 19
121
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Table 7-2 (contd.). Site Summary of App I and II Parameters with Exceedances Above the Respective Maximum Contaminant Level After Closure
Site
Parameter
Total
number
of wells
data were
available
Total Number
of
Measurements
Number of
Detected
Measurements
Number of
Exceedances
Number of
Upgradient
Wells with
Exceedances
Number of
Downgradient/
Side-gradient
Wells with
Exceedance
Post-
Closure
Duration
Data
Available
(Years)
Year (after
Closure) of
Exceedance(s)
G
Thallium
22
235
21
19
3
7
16
7, 11-13
Mercury
22
241
4
2
0
2
16
9, 16
H
Arsenic
30
255
141
5
0
1
9
3-5, 7-8
Beryllium
30
255
232
149
9
14
9
0-8
Cadmium
30
255
129
1
0
1
9
2
Chromium
30
255
103
1
1
0
9
8
Copper
30
256
97
1
0
1
9
7
Lead
30
256
81
12
0
3
9
0-7
Thallium
30
255
27
27
5
13
9
1,2,4,5,8
I
Arsenic
16
297
114
114
0
7
19
0-18
Vinyl Chloride
14
141
1
1
0
1
1
0
Methylene Chloride
15
134
14
14
0
1
8
0-7
Constituents in the shaded rows were above MCL in upgradient well(s)
1 Data points only for the sampling events with at least one detected measurement.
122
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As shown in Table 7-2, elevated arsenic levels have been observed at seven different downgradient
groundwater monitoring wells located at Site I. Arsenic exceeded the MCL only once for two of
these wells after closure. The arsenic concentrations at the rest of the five wells were regularly
observed above the MCL (i.e., 0.010 mg/L), and one of these five wells has not been monitored
for the most recent five years. A temporal plot of arsenic concentrations at downgradient wells for
the remaining four wells is presented in Figure 7-1; the laboratory reporting limit of 0.010 mg/L
was used for the values below the reporting limit. Two data points that appeared to be outliers (one
point from Well A and one from Well C) were not included in the plot for clarity; these
concentrations were approximately an order of magnitude above the concentrations measured
during the sampling events immediately prior to and after the outlier measurement.
0.10 -
0.08
£
G
o
u-
+->
G
O
G
O
U
G
Xfl
u-
<
0.06 -
0.04 -
0.02 -
0.00
Well A
WellB
Well C
WellD
*
-w T
— T- -I" — — — X— X— ^"~T~TT »T TT~fr
MCL = 0.010 mg/L
XT X
A XX y
X ^ X
XX.
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19
Year After Closure
Figure 7-1. Temporal Variation of Arsenic Concentrations Observed Above the MCL at Four
Downgradient Wells at Site I After Closure
As shown in Figure 7-1, the arsenic concentrations observed at Well C have been consistently
below the reporting limit/MCL since year 13 after closure. Arsenic concentrations have been
consistently above the MCL at Wells A, B, and D. Arsenic concentrations exhibit an increasing
trend at Well D. Arsenic concentrations observed at Well B vary over a relatively wide range
compared to the other wells while the arsenic concentrations observed in Well A have been
relatively consistent since year 13. In summary, the variable arsenic concentration trends at these
wells (i.e., variable at one location, increasing and consistent at the other locations) suggest a need
for continued groundwater monitoring at the site to assess long-term impacts on groundwater
quality.
123
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A comparison of the leachate indicator parameter (e.g., total dissolved solids, chloride, and
ammonia) concentration in downgradient wells with the respective background concentration can
be used for assessing the magnitude of the groundwater impacts associated with leachate release.
Figure 7-2 presents chloride and ammonia distribution in leachate, an upgradient well, and four
downgradients wells that exhibited arsenic concentrations above MCL after closure and were
actively monitored at the time of the study. As expected, the groundwater chloride and ammonia
concentrations are 2-3 orders of magnitude lower than the respective concentration in leachate.
The chloride concentration in downgradient wells was much greater than that for the upgradient
well suggesting potential impacts from the study cell. The chloride and ammonia concentrations
in Well C were greatest among all the downgradient wells. However, as shown in Figure 7-1,
arsenic concentration observed at Well C was lowest among the downgradient wells suggesting
leachate release, if any, does not appear to be the primary cause of arsenic exceedances at this well.
The site monitoring reports documented that these elevated levels may be a result of subsurface
geochemical changes resulting from the development of the site, which included the elimination
of groundwater recharge (i.e., not due to leachate release into groundwater).
The availability of chloride and/or ammonia data for groundwater is requisite for this analysis. It
should be noted that neither chloride nor ammonia is an App II parameter. Chloride data were
available for all of the study sites. Chloride was measured only for upgradient well at Site D.
Ammonia data were available for all the study sites except for Sites D and F. These data were not
analyzed for the study sites as the analysis was primarily focused on App I and II parameters with
MCL/SMCL and examples of App I and II parameters with state standard but no MCL/SMCL.
Both chloride and ammonia should be collected and analyzed to assess potential leachate release
from MSWLFs and ensuing HHE impacts.
The analysis presented above was conducted only for parameters that have a federal MCL. In
addition to reviewing the groundwater parameters with an MCL, groundwater quality data were
also reviewed for App I and II list parameters with an SMCL-silver and zinc. These data were not
available for all of the sites. For example, zinc and silver measurements were not available for
Sites I and E, respectively. Only one site had an exceedance of one of these parameters; Site G had
a silver exceedance three years after closure at a single well. As described previously, only 32 out
of 62 App I and 51 out of 215 App II parameters have an MCL and two parameters in these lists
have an SMCL.
The parameters that do not have a federal MCL/SMCL would also need to be analyzed for a
comprehensive HHE impact assessment. The state may have a human-health risk-based standard
for parameters with no federal MCL. An analysis was conducted for App I and II parameters
without an MCL/SMCL for a study site to emphasize the importance of analyzing these parameters
for assessing the groundwater impacts. As of the time of this evaluation, all three of the sites under
assessment monitoring/corrective action had instances where parameters without an MCL
contributed to groundwater impacts. Examples of these parameters include vanadium (Site A), 1,1-
dichloroethane (Site F), and 1,4-dioxane (Site I). Establishment of a site-specific groundwater
protection standard for the App II (which includes all App I parameters) parameters that do not
have an MCL is required by §258. The site-specific groundwater protection standard may be based
on the background concentration (established based on background/upgradient well groundwater
quality) or the state-specified human-health risk-based standard.
124
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104
103 -
102 -
101 -
10° -
10-1
T
Leacliate Upgradient Well A Well B Well C
(a)
Well D
Leacliate Upgradient Well A Well B Well C Well D
(b)
Figure 7-2. Distribution of (a) Chloride, and (b) Ammonia for Leachate and Four Downgradient
Wells and an Upgradient Well at Site I
125
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In order to provide an example evaluation of a parameter with a state-mandated risk-based standard
(but no MCL), parameters at Site I without an MCL were reviewed to identify parameters with a
state-designated groundwater protection standard that appeared to be significantly elevated over
background quality. Parameter 1,4-dioxane was one that has never been detected in background
wells but was detected in three downgradient wells. Although there is no federal MCL/SMCL, a
state-specified risk-based groundwater protection standard of 3 |ig/L for 1,4-dioxane applies to
Site I. The concentrations of 1,4-dioxane in these three downgradient wells are presented in Figure
7-3.
It is interesting to note that two of the three downgradient wells with detected concentrations of
1,4-dioxane at Site I are the same wells (i.e., Wells B and C) that exhibited arsenic exceedances,
as presented in Figure 7-1. Well C, which exhibited the greatest 1,4-dioxane concentration, also
had the greatest median chloride and ammonia concentration among all downgradient wells, as
shown in Figure 7-2, suggesting potential impacts of the landfill on groundwater quality. Elevated
concentrations of 1,4-dioxane in the downgradient groundwater monitoring wells (particularly in
Well B and Well C), when compared to the concentrations observed at the background wells,
suggest impacts from the closed study cell at Site I. The exhibited exceedances and temporal trend
of concentrations suggests a need for continued groundwater monitoring of 1,4-dioxane for a
reliable long-term HHE impact assessment.
00
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'-S
aj
*
G
V
o
G
O
O
-------
of concentrations in groundwater and leachate could not be performed due to insufficient leachate
1,4-dioxane data. As presented in Attachment I, 1,4-dioxane concentration was available for only
one leachate sampling event after closure.
Groundwater monitoring reports documented groundwater impacts associated with elevated
vanadium concentrations at downgradient wells at Site A. Vanadium is included in both App I and
II and does not have an MCL. Figure 7-4 presents a comparison of the distribution of vanadium
concentrations at upgradient wells and downgradient wells after closure. Please note that only
downgradient wells that have shown vanadium concentrations over the reporting limit on at least
two sampling events are presented in the figure. As shown in Figure 7-4, the vanadium
concentration at three downgradient wells appears to be noticeably greater than at the upgradient
wells. In addition, there have been two occasions where a downgradient well exceeded the State's
risk-based standard - this well has generally shown vanadium concentrations approximately an
order of magnitude above background concentrations. Vanadium concentrations at this well have
exhibited a declining trend since these exceedances.
For all non-detect measurements, the detection limit was used as a surrogate for the concentration
value. Over 80% of the measurements at the upgradient wells were below the method detection
limit. Over 75% of the measurements at two downgradient wells (D-2 and D-3) were reported
below the detection limit.
100
State Standard = 49 Dg/L
10 -
1 -
0.1
U-l
U-2
D-l
D-2
D-3
Figure 7-4. Distribution of Vanadium Concentration at Upgradient (U) and Downgradient (D)
Groundwater Monitoring Wells of Site A After Closure
The groundwater impact assessment presented in the section is by no means a comprehensive
assessment and has several limitations. First, only App I and II parameters with MCL/SMCLs
were analyzed. Although an example of other parameters (e.g., 1,4-dioxane at Site I and vanadium
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at Site A) are presented to emphasize a need to examine these parameters for assessing
groundwater impacts, not all the groundwater monitoring data available for the site were analyzed.
Site owners should consider collecting and analyzing data for all of the App I and II parameters,
and leachate quality indicators (e.g., ammonia, chloride) as well as emerging contaminants such
as per- and polyfluoroalkyl substances and pharmaceuticals, which are not be currently included
in App I or II.
Second, the analysis was conducted based on federal MCL and SMCLs for App II parameters with
these limits. The HHE impact analysis would need to be conducted based on the state-specified
risk-based standards, which, if available, are expected to be lower than the respective federal MCL.
The outcome of the analysis based on state-specified standards may be different from that based
on federal limits. For example, Site B does not appear to pose a threat to HHE for arsenic and vinyl
chloride with respect to MCLs. However, the site is under corrective action for frequent
exceedances of several parameters, including arsenic and vinyl chloride. The state-specific vinyl
chloride and arsenic standard used for this site are an order of magnitude smaller than the
respective federal MCL. The state may require monitoring of additional parameters of HHE
concern. For example, exceedances of manganese, which is not an App II parameter, were
documented for Sites B and I.
Third, the available groundwater data had several inadequacies. Only limited data were available
for a few of the study sites. For example, only one mercury measurement was available for each
monitoring well after the closure of the study cell at Site C and only a single arsenic measurement
was available for Site E. Adequate numbers of data points are essential for reliable statistical
analysis and analyzing trends over time. The unavailability of data for review for this study does
not necessarily mean that these data were not collected. MSWLFs owners should consider
collecting and cataloging the groundwater quality data for easy retrieval for statistical and trends
analysis.
Fourth, the data for several parameters had data quality issues specifically pertaining to the method
detection limit used for the laboratory analysis. Several instances were found where the parameter
analysis detection limit was greater than the respective MCL. These data were excluded from the
analysis presented in this section. Finally, only data collected after closure were examined. All the
available data, including data collected before waste placement activities, should be analyzed for
a more comprehensive impact assessment. This analysis approach can be complemented with more
rigorous fate and contaminant transport modeling to assess the impacts of contaminant
transformation and dilution on groundwater quality at the point of compliance and the point of
exposure over time.
'• 11111 act of Biogeochemical Changes on Groundwat i ality
Biogeochemical changes in the aquifer system due to liner construction can also impact
groundwater quality. Several landfills throughout Florida have reported iron exceedances in
groundwater monitoring wells. The process commonly believed to be responsible for iron releases
at landfill sites is known as "reductive dissolution." In this process, reducing conditions develop
in the surficial aquifer at a landfill site and transform the oxidized solid-phase of naturally-
occurring iron to reduced, dissolved iron. This phenomenon has been observed and described at
landfill sites around the country. In many cases, the reductive dissolution of iron minerals triggers
exceedances with respect to other parameters (e.g., arsenic); as iron reduces to its soluble ferrous
form, arsenic sorbed to the iron is also released.
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Many of the sites where iron dissolution has occurred are older, unlined landfills. However,
growing evidence shows that similar problems can occur at lined landfill sites as well. The
presence of a liner system may result in reducing conditions by limiting oxygen transport from the
atmosphere to groundwater. Townsend et al. (2015c) observed an increase in groundwater iron
concentration following the construction of a liner system of a test cell at a landfill in Florida. The
test cell was constructed in an area away from other landfill cells at the site, and no waste was
placed in the cell.
Figure 7-5 presents the iron concentrations distribution in landfill leachate and at several
groundwater monitoring wells at a lined MSWLF-this site does not have an unlined cell (IWCS
2010). As can be seen from the figure, the iron concentrations recorded at groundwater monitoring
wells are much greater than that of leachate, suggesting that leachate release alone could not result
in the groundwater iron concentrations observed at this site.
J
Eh
o
oS
—
S3
CD
O
Eh
o
U
Eh
o
MW1 MW2 MW4 MW8 MW9 MW10MW11MW12 CW8 CW9 CW10 CW11 Leachate
Monitoring Wells
Figure 7-5. Distribution of Iron Measured at Various Groundwater Monitoring Wells
(identification labels with MW and CW prefixes are groundwater monitoring wells),
and Leachate at an Active Lined MSWLF in Florida (IWCS 2010)
Figure 7-6 shows the distribution of iron concentration at the four downgradient groundwater
monitoring wells at Site I that showed elevated arsenic concentrations after closure. Two of the
groundwater wells have median iron concentrations above that of leachate from the closed study
cell, suggesting leachate may not be the primary contributor to the elevated iron concentrations
observed at Site I groundwater.
Several indicator parameters can be used to assess whether leachate release and migration is the
cause of groundwater contamination at MSWLFs. Strong indicators of leachate contamination
include ions of soluble salts (such as chloride), ammonia, and organic chemicals. As discussed
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above, a comparison of chloride and ammonia concentrations in downgradient wells with the
respective background concentration (in upgradient wells) can be useful for assessing leachate
impacted on the groundwater quality. The contaminant and chloride concentration ratio in leachate
can be compared with that for groundwater to assess whether leachate release has resulted in the
groundwater impacts with respect to the contaminant. In the event a leachate release is the major
contributor to groundwater contamination, the chemical concentration ratio(s) for parameters such
as iron or arsenic to chloride in groundwater should be similar to or (due to attenuation) lower than
the ratio(s) for leachate.
00
Eh
_o
u-
£3
(D
O
Eh
O
U
Eh
o
Well A
Well B
Well C
Well D
Leachate
Figure 7-6. Distribution of Iron Concentrations Measured in Leachate and in Downgradient
Groundwater Monitoring Wells with Elevated Arsenic Concentrations at Site I after Closure
Figure 7-7 presents the distribution of arsenic: chloride and iron: chloride ratios for four monitoring
wells with consistent arsenic exceedances and for leachate from the Site I study cell. The
arsenic:chloride ratios for three of these monitoring wells are substantially more than that of
leachate, which also suggests that a leachate release, if any, is not the primary contributor to arsenic
exceedances. Similarly, the iron:chloride ratio for all monitoring wells is approximately 2-3 orders
of magnitude greater than for leachate, suggesting that leachate release, if any, is not the primary
contributor of elevated iron concentrations in groundwater. This analysis approach cannot be used
to conclude that leachate release has not contributed to the impact altogether as the magnitude of
the arsenic mobilized from reducing condition is not known. Leachate release, if any, can also
cause reducing conditions or further amplify prevailing reducing conditions and arsenic
mobilization.
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10-1
i
w
~=i=~
Leachate Well A
Well B
(a)
Well C
Well D
Leachate Well A
Well B
fb)
Well C
Well D
Figure 7-7. Distribution of (a) Arsenic-to-Chloride Ratio, and (b) Iron-to-Chloride Ratio for
Leachate and Four Downgradient Wells at Site I after Closure
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<' ii ii itic i i mi i ¦> ii oundwater Flow
As discussed above, MSWLF groundwater impacts are analyzed by statistically comparing the
groundwater quality observed at downgradient wells with that of upgradient wells, which represent
the background quality and are not expected to be impacted by landfill cell(s). The direction of
groundwater flow determines the classification of wells as upgradient or downgradient.
Groundwater flow direction and velocity are routinely estimated by plotting potentiometric maps
using water level measurements at the wells during groundwater sampling events. Temporal
changes in groundwater flow directions can complicate comparisons of upgradient and
downgradient water quality and may make identifying the true impacts of MSWLFs challenging
to assess. Of the nine study sites, periodic groundwater monitoring reports (which document
groundwater flow direction) were available from five sites (i.e., Site A, C, D, F, and H). Of these
sites, records for three sites (i.e., Site D, F, and H) suggest that there has been little/no variability
in groundwater flow direction after closure. Of the other two sites, reports suggest that
groundwater flow direction appears to be varying seasonally for one site (i.e., Site C) and may be
experiencing long-term changes in groundwater movement at the other site (i.e., Site A).
Changes/consistency in site groundwater flow directions at the study sites were not analyzed for
each sampling event.
Factors such as weather patterns (e.g., sequencing of dry and wet years), changes in land
use/zoning and associated changes in consumptive use of groundwater surrounding the sites (e.g.,
pumping from the municipal water supply or other well fields located in the vicinity of the landfill),
or shifting gradients resulting from seasonal variations or tidal influences can result in localized
changes in groundwater flow. The impact of factors like these should be considered while
assessing the long-term impact of PCC duration reduction/termination on HHE.
IiH",< -i iii , ; iii'1 II 11 ii ii in trice of Monit .1 mii.,« » ells
The ability to collect representative groundwater samples is paramount to the reliability of
groundwater data and the contingent analysis/decision making. Factors such as the clogging of
groundwater monitoring well screens with sediments and precipitates (USGS 2016, OH EPA 2009,
and EPA 1988) and accidental well damages (e.g., from equipment or vehicular traffic, surface
drainage) casing can impact the quality of the data collected. As groundwater well maintenance
records were not requested from the owners/operators of the study sites included in this study, an
evaluation of the nature and prevalence of issues that may impact the performance of groundwater
monitoring wells was not performed. The site owner should consider implementing a routine
groundwater monitoring well inspection and maintenance program to ensure the collection of
representative samples for an accurate and reliable assessment of groundwater water quality. The
site operator should consider maintaining and including these records as part of the demonstration
to request PCC termination or an alternative PCC duration.
7.8. Summary
The groundwater monitoring data for all the study sites were analyzed for the App II parameters
with a federal MCL or SMCL. The groundwater data were analyzed for all the wells at the study
site regardless if the study cell was contiguous to the other active or unlined cells (Sites A, B, E,
G). Only data from the wells around the study cells were analyzed for the site if the study cell was
not contiguous to active or unlined cells (Site C, F, H, and I). A majority of the organics measured
at the sites were not detected in groundwater. No exceedances were identified at Site F for the
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range of data available for review (i.e., post-closure Year 8 through Year 19). Only occasional
MCL exceedances were identified following cell closure among the studied sites. Only Sites G, H,
and I appear to have had reoccurring groundwater exceedances. More than 55% and 10% of 255
samples at Site H had beryllium and thallium concentration greater than the associated MCL,
respectively. Approximately 40% and 30% of beryllium and thallium exceedances observed at Site
H occurred at the upgradient well(s), respectively. Site G has 107 and 53 samples with arsenic and
lead concentrations greater than the associated MCL, respectively. More than 10% of arsenic and
lead exceedances observed at Site G occurred at the upgradient well(s). More than 50% of the
arsenic exceedances observed at Site G correspond to the wells that were downgradient or close to
the unlined cell at the site. The beryllium and thallium exceedances at Site H and arsenic and lead
exceedances at Site G cannot be conclusively attributed to lined cells at these sites due to the
exceedances observed above the respective MCL at several upgradient wells for these parameters
at these sites.
The arsenic exceedances at Site I have been consistently observed at several downgradient wells
after closure. An analysis of the arsenicxhloride ratio in groundwater and leachate at Site I suggest
that arsenic exceedances at the site are not likely associated with leachate release, if at all. The
change in the biogeochemical environment of the surficial aquifer resulting from landfill
construction can also result in mobilization of naturally-occurring constituents and impact to the
groundwater quality.
The available silver and zinc were analyzed to identify exceedance with respect to the SMCL.
These data were not available for all of the sites. For example, zinc and silver measurements were
not available for Sites I and E, respectively. Only one site with these data had an exceedance of
one of these parameters; Site G had a silver exceedance three years after closure at a single well.
Only 51 out of 215 App II parameters have an MCL and two parameters in these lists have an
SMCL. The parameters that do not have a federal MCL/SMCL may also need to be analyzed for
a comprehensive HHE impact assessment. The state may have a human-health risk-based standard,
which may be used for assessing the risk to HHE for these parameters. As of the time of this
evaluation, all three of the sites under assessment monitoring/corrective action had instances where
parameters without an MCL contributed to groundwater impacts. Examples of these parameters
include vanadium (Site A), 1,1-dichloroethane (Site F), and 1,4-dioxane (Site I). In addition,
leachate indicator parameters such as chloride and ammonia should be routinely monitored for
groundwater. These parameters are helpful for assessing whether or not the observed groundwater
impacts are associated with leachate emission from MSWLFs.
tiioiis
The analysis presented in this chapter has the following limitations:
1. Only limited groundwater data were available for a few of the study sites. For example,
only one mercury measurement was available for each monitoring well after the closure of
the study cell at Site C and only a single arsenic measurement was available for Site E. The
contaminant concentration can vary considerably over time.
2. Only the groundwater quality data available after closure of the study cell(s) was analyzed.
The analysis presented in this chapter does not include the data collected before the cell
closure.
3. The MDL for a large number of organic compounds and some metals monitored for the
study sites was greater than the respective protection standard. These parameters cannot be
133
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conclusively determined to be below MCL/SMCL These measurements were not included
in the analysis presented in this chapter.
4. The analysis was conducted for App II parameters with MCL/SMCL. More than 60% of
the parameters that were measured at least once do not have an MCL/SMCL. These
parameters, which may be of HHE concerns, were not comprehensively evaluated in this
study.
5. The analysis was conducted based on federal MCL and SMCLs for App II parameters with
these limits. The HHE impact analysis would need to be conducted based on the state-
specified risk-based standards, which, if available, are expected to be lower than the
respective federal MCL. The outcome of the analysis based on state-specified standards
may be different from that based on federal limits. For example, Site B does not appear to
pose a threat to HHE for arsenic and vinyl chloride with respect to MCLs. However, the
site is under corrective action for frequent exceedances of several parameters, including
arsenic and vinyl chloride. The state-specific vinyl chloride and arsenic standard used for
this site are an order of magnitude smaller than the respective federal MCL.
6. The analysis was conducted only for App II parameters (all App I parameters are included
in App II). The state may require monitoring of additional parameters of HHE concern. For
example, exceedances of manganese, which is not an App II parameter, were documented
for Sites B and I.
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8. Final Cover Performance
i nil I missions Monitoring
SEM data were available for review from six of the nine sites. However, a complete dataset
covering the entire period from the closure date to the time of this study was only available for two
of the sites (Sites D and H). Table 8-1 presents the number of locations with SEM exceedances (a
surface emissions concentration elevated over the 500-ppm standard) as a function of years since
closure. A indicates that SEM data were not available for the site for that year. Shaded cells
indicate future years. Please note that the values presented in the table only represent exceedances
identified over the cover system of the site's study cell(s).
Table 8-1. Number of Annual Surface Emissions Monitoring Exceedances Over Study Cells as a Function of
Time after Closure
Site
C1
D
E
F
H
I
Years of
Available
2014
Surface
2009-
2008-2015
2012-
2008-
2004-
Years
Since
Closur
Emission
2016
2016
2015
2016
2016
Monitoring
Data
Area
(Acres)
40.8
69
30 (Cell 2- 24 and Cell 3-
6)
38
60
51
e
Closure
Year
1998
2009
2004
1997
2008
1997
-
3
-
-
6
-
0
-
0
-
-
2
-
1
-
0
-
-
1
-
2
-
0
-
-
2
-
3
-
0
5
-
4
-
4
-
0
2
-
2
-
5
-
0
2
-
0
-
6
-
0
1
-
2
61
7
-
0
-
0
21
8
Exceedance
-
0
-
0
9
s per Year
-
-
0
10
-
0
-
0
11
-
-
0
12
-
-
0
13
-
-
0
14
-
0
0
15
n
0
0
16
iu
0
0
17
iu
0
0
18
-
0
19
1 Data were available for three quarters of year 16, 17, and 18 after the closure
Study cells at Sites C, E, and H had the most frequent exceedances based on the available data.
The highest number of exceedances was observed over two monitoring events at Site I. Well
penetrations through the final cover were documented as the primary factor contributing to these
exceedances. More than 80% of these exceedances were corrected within 30 days of the initial
SEM exceedance observation; details about measures taken to correct the exceedances were not
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available. As shown in the table above, the final SEM exceedance observed at five out of six sites
were within ten years since closure. SEM exceedances beyond the initial ten years of PCC were
only observed at one site (Site C), and exceedances were noted for all three years that data were
available (years 16-18). It should be noted the study cells at Site C are covered with an exposed
geomembrane cap, which lacks a soil cover layer where methane oxidation could occur. At Site E,
recent exceedances occurred near wellheads, which were repaired either by excavating the cover
soil and backfilling with bentonite or by installing an additional small well tied into the
geomembrane just outside the well casing. At Site H, two exceedances were observed near two
different gas wells during one of the monitoring events in 2015 (seven years after closure). The
PVC pipe collar between the gas well casing and the geosynthetic closure system was adjusted at
one exceedance location and additional soil filling was provided at the other location as a
corrective measure to address these exceedances. Exceedances were not observed in the following
10-day monitoring event at both Site H locations.
lement
With the exception of Site I (where annual settlement data were available), limited settlement data
were available from occasional surveys of the landfill surface for Sites B, C, E, and F. For Site C,
topographic surfaces generated from surveys conducted in the 0, 1st, 4th, 6th, and 18th year after
closure were available; however, settlement or elevation data at specific survey points were not
available. Similarly, for Site E, the topography of the final cover at the 0, 7th, and 8th year after
closure were available for review; however, the available data were insufficient to analyze
temporal trends in the rate of settlement. Site F topographic maps were available for PCC years 5,
9, and 15 as images; however, the topographic survey data for the closure year were not available;
therefore, an analysis of trends in the settlement rate for Site F was not conducted. Landfill
settlement measurements at ten settlement plates installed on the final cover at Site B were
available for 9th and 11th year after closure. The location details of these plates were not available.
Site I topographic survey data was available for PCC years 1 to 13, 15, and 17 for 20 discrete
survey points located on the study cell. Eleven (11) of these points were on the top deck.
Distribution of the annual settlement rate at the survey points located on the top deck of the cell is
shown in Figure 8-1. The overall settlement rate exhibits an overall declining trend after closure.
The median annual settlement decreased from 2.4 ft in the first year to 0.13 ft in the 17th year after
closure. The range of the settlements observed at the survey points appears to decrease over time
from 2.5 ft to approximately 0.5 ft during the first year and 17th year after closure, respectively.
The range of the measured settlement rate at a given point of time is indicative of differential
settlement. As discussed previously, the differential settlement is one of the primary contributing
factors that impact the integrity of the final cover. As shown in Figure 8-1, a few points exhibited
a negative annual settlement rate, which might be associated with regrading activities that are
typically conducted on an as-needed basis to fill depressions and maintain positive surface
drainage. Several approaches and models have been used for modeling waste settlement
(Townsend et al., 2015a). Some of these models account for mass/volume loss over time due to
the decomposition process (e.g., Hettiarachchi et al. 2007).
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4
-l H 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17
Year After Closure
Figure 8-1. Distribution of Yearly Point-to-Point Settlement Rate at Site I
Bottom grade construction drawings and topographic information were used to develop a
topographic surface representative of the base grades of the cell at Site I (i.e., using AutoCAD
Civil 3D 2013). The topographic surfaces were compared with bottom grade surfaces to estimate
in-place waste volume for each year, which were used to evaluate the annual volume loss rate. The
waste volume loss rate is expected to be driven by the waste decomposition rate, which typically
is modeled using a first-order decomposition rate equation (Hettiarachchi et al. 2007). Several
researchers have modeled MSW mass and volume loss rates (e.g., Sheridan 2003, Kim 2005).
Many of these either are empirical or do not resemble the first-order decay model used for landfill
gas modeling. In order to estimate a volume loss decay rate that is analogous to the waste decay
rate used for LFG modeling, Eqn. 8-1 was used to model the volume loss rate; this equation
resembles the equation (Eqn. 5-3) used for methane generation modeling for MSWLFs. This
approach is similar to the one used by Hettiarachchi et al. (2007) to model the volume loss rate
using the first-order decomposition rate equation.
Vt=kv-Vs(e~kvi) Eqn. 8-1
where,
Vt = waste volume loss rate at time t (yd3 per year)
Vs = waste volume loss over an infinite time horizon (yd3)
kv = first-order decay rate for volume loss (year)
t = time (years)
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An iterative regression analysis was conducted using the Excel Solver function in order to calculate
the decay rate constant and waste volume loss over an infinite time horizon that would best-fit the
model to the estimated volume loss based on the settlement data. Figure 8-2 shows the estimated
and the modeled volume loss rate data for Site I. Once calculated, the model parameter values (as
shown in the figure) can be used to estimate the annual volume loss rate at a site for future years.
The results of the regression analysis suggest a total volume loss of 670,000 yd3 over an infinite
time horizon and a first-order decay rate of 0.149 year"1. Approximately 526,000 yd3 of volume
loss (representing approximately 79% of the total estimated volume loss) had occurred within 17
years of closure. The model indicates that 99% of the volume loss is estimated to occur within 30
years of closure. Due to the lack of landfill gas data for Site I, the correlation between the estimated
volume loss rates and the methane collection rates, and that between the methane generation and
volume loss decay rates could not be assessed.
The settlement and volume loss estimates presented above assumed a constant topographic
condition for the landfill bottom and did not account for the foundation settlement at the site. The
settlement and volume loss estimates are regarded as overestimations due to unaccounted
settlement of subsurface soils below the landfill as a result of overburden pressure.
-------
9. ' 'I III nil li vl v I I" " ' li h" 'I vl| li
gs
BW
The MSWLFs owners are required to perform PCC for a period of 30 years after closure
(§258.61(a)), unless the time period is decreased or increased by the Director of an approved State
as necessary to protect HHE. Subpart F of §258 lists PCC-specific requirements for the MSWLFs.
Although the RCRA Subtitle D regulations allow modification of PCC duration, there is no federal
guidance or specific direction on approaches that can be used for making a demonstration
pertaining to HHE impacts for supporting the extension or termination of PCC activities. The EPA
collected, reviewed, and analyzed environmental monitoring data from the closed cell(s) of nine
MSWLFs located across the US to use as examples to:
(1) Assess the nature of the data available for MSWLFs that can be used for HHE impact
evaluation
(2) Present approaches that site owners/operators and engineers can use to evaluate
monitoring data for identifying the COPCs and estimate emission rates of these
contaminants
(3) Identify data gaps, and
(4) Present operating and monitoring considerations for MSWLFs owners to evaluate and
mitigate the long-term impacts of MSWLFs.
Five of the sites are located in the northeast, two are located in the southeast, one in the northwest,
and one in the southwest region of the US. Five of the study sites are publicly owned and operated,
and the rest four are privately owned and operated. The selected study sites each possessed at least
one cell that has been closed for five years or more and has maintained environmental monitoring
records (e.g., groundwater monitoring, landfill gas quantity and quality, leachate quantity and
quality) available for review. The area and capacity of the studied cell(s) at these study cells/sites
ranged from 6 to 69 acres and from 0.5 to 4.9 million tons of MSW, respectively.
' 1 ; 1 ¦ t-Closi • < we Cost
The available PCC cost data from the nine study sites were analyzed to evaluate the financial
impact of different MSWLF PCC activities. The available data were organized into the following
six major cost categories: (1) leachate management, (2) GCCS management, (3) final cover
maintenance, (4) groundwater and subsurface gas monitoring and maintenance, (5) engineering
support and administration, and (6) other miscellaneous expenditures.
Limited PCC cost data were available for the study sites. Actual PCC cost data were available only
for three sites (Sites B, D, and E); only estimated cost data were available for the other locations.
Additionally, only three sites (Sites G, H, and I) provided cost data exclusive to the study cells.
For the other sites, the available cost data included the cost of maintaining other site cells as well.
All the available PCC cost data were adjusted to 2017 dollars based on consumer price index
values.
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For the sites with available study cell-specific data, the annual PCC cost varied from approximately
$1,200 to $11,000 per acre of waste footprint, with an average of $5,300 per acre. For the sites
where PCC cost data represented the entire site, annual PCC cost ranged between $1,550 to
$37,000 per acre with an average of $6,450 per acre. Inconsistencies in cost categories used for
tracking PCC cost, and the necessity for occasional system upgrades (e.g., GCCS expansion at Site
E, and construction of a sewer connection for pumping leachate to the local WWTP) appears to be
one of the primary reasons for such a wide variation in annual PCC cost at the study sites.
Annual leachate management (including LCS operation and maintenance, hauling, onsite/off-site
treatment, and leachate sampling and analysis) cost represented the greatest category cost at six of
the nine sites with a range of approximately 3 % to 68% of PCC cost. The leachate management
ranged from $5.5 to $219 per 1,000 gallons of leachate collected with an average of approximately
$79 per 1,000 gallons of leachate collected. GCCS management costs ranged from approximately
11% to 44% of the annual average PCC cost. The annual GCCS management cost ranged from
approximately $2,700 to $593,000 per year, for seven of the nine sites based on available data.
The average annual groundwater and subsurface gas monitoring cost varied between
approximately 5% to 26% of the total average annual PCC cost. The annual monitoring cost among
the sites ranged from approximately $2,400 to $169,600. Annual final cover maintenance cost
among the sites (except Site H) ranged between approximately 0.2% to 14% of the total average
annual cost. The annual cover maintenance cost of two sites (Sites G and I), which had cell-specific
cover maintenance cost data available, ranged from approximately $56 to $600 per acre of the
waste footprint. The final cover maintenance cost at Site I was observed to decline with the amount
of differential settlement, which was observed to reduce at Site I over time. For remaining sites,
the annual cover maintenance cost ranged from approximately $24 per acre to $3,400 per acre.
The available data suggest that the cost varied over a wide range due to several factors, including
inconsistencies in cost categorization, mixed availability of actual cost and cost estimates, and the
necessity for occasional capital-intensive system upgrades. Because of these limitations, landfill
owners and engineers should consider tracking and using the site-specific cost data for evaluating
the financial impacts of PCC activities instead of using the data presented in the report as proxies.
1 1 : 1 dfill Gas
LFG emissions constitute one of the primary pathways for potential HHE impacts resulting from
MSWLFs. The available LFG collection and flow rates and subsurface probe monitoring data for
the study sites were analyzed to:
(1) Assess the frequency of subsurface methane migration
(2) Assess the timeframe needed for the annual NMOCs collection rate from study sites to
drop below the NSPS threshold of 50 (or 34) Mg per year and the timeframe for the LFG
flow rate to decline below 5% and 10% of the peak flow rates
(3) Assess whether MSWLFs would have sufficient NMOCs generation potential to generate
50 (or 34) Mg of NMOCs annually after the termination of GCCS operation, and
(4) Assess the remaining methane generation potential at 30 years after closure and assess the
timeframes needed for the remaining methane potential to drop below 25% and 10% of the
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total generation potential. Some states use the percent remaining methane potential as
criteria for assessing the PCC termination.
The subsurface methane monitoring probe data from seven of the study sites suggest exceedances
in the subsurface and structural methane monitoring is a relatively rare occurrence. Approximately
1.4% of all 7,598 methane measurements at the subsurface perimeter monitoring probes exceeded
the lower explosivity limit of methane. Approximately 90% of these exceedances were observed
at Sites B and E. The observed methane exceedances at these sites likely caused unlined cells or
cells lined with a compacted clay liner at these sites. All the exceedances occurred during the first
five years after closure for all the sites except for Site E. The presence of a geomembrane liner
and an active GCCS limit subsurface migration of landfill gas and are the likely primary reasons
for the relatively infrequent methane detection/exceedances. The GCCS at Site A was terminated
in the 13th year of closure and no exceedances were reported since. The structure methane
concentration data (2,105 measurements) were also available for Sites B, F, and H. None of the
methane measurements in structures at Sites B and H were above 25% of its LEL. Only 5 out to
the total 1,155 methane measurements in Site F structures were above 25% of its LEL.
A regression analysis was first conducted using data from all of the study sites (except Site I) to
estimate site-specific decay rates and methane collection potential of waste that provided the best-
fit of the first-order decay model to the available monthly methane flow rates data. LFG data
exclusive to the closed cell at Site I were not available. The estimated site-specific decay rates
suggest that the waste decomposition at all of the eight study sites (with GCCS data) occurred
more rapidly than the decay rates specified/recommended by NSPS, AP-42, and GHG Reporting
regulations. This suggests that the post-closure methane generation rate from MSWLFs are
expected to be lower than those estimated based on regulatory default or AP-42 recommended
decay rates.
The future annual methane and NMOCs generation rates were then estimated for the study sites
using the first-order waste decomposition rate equation based on site-specific decay rates and
methane collection potentials, disposal amounts, and NSPS-default NMOCs concentration of
4,000 parts per million by volume as hexane. The annual NMOCs collection rate at each study site
was estimated to decline to below 50 Mg per year within 30 years after closure. The annual
NMOCs collection rate was estimated to decline below 34 Mg per year within 30 years after
closure for all the study sites except one (Site F). The study sites are not representative of the size
of approximately 75% of MSWLFs in the US, as six of the study sites contain less than 4 million
metric tons of waste. In order to assess the post-closure timeframes for typical size MSWLFs to
achieve NMOCs reduction below the NSPS threshold, NMOCs and methane flow rates were
estimated for MSWLFs containing approximately 3.35 (small MSWLF), 7.85 (medium MSWLF),
and 19.1 (large MSWLF) million metric tons of MSW, which correspond to the 25th, 50th, and 75th
percentile of the capacity of MSWLFs in the US, respectively. In order to assess the impact of
decay rate on the annual NMOCs collection rate trend, first-order decay modeling was conducted
for each of the three hypothetical MSWLFs for the five decay rates ranging from 0.02 to 0.22 year
to represent waste decay conditions ranging from arid climate to bioreactor operation.
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The analysis suggests that the annual NMOC collection rate for small MSWLFs located in arid to
moderate precipitation areas are not expected to decline below 50 Mg per year within 30 years
after closure. The NMOCs collection rate for medium and large MSWLFs is not expected to
decline below 50 Mg per year within 30 years after closure irrespective of location. The annual
NMOCs collection rate of MSWLFs operated as a bioreactor is expected to decline below 50 Mg
per year within 20 years after closure. The LFG flow rate is expected to decline below 5% of the
peak flow rate after the annual NMOCs collection rate reduces below 50 Mg per year for all of the
typical-size MSWLFs scenarios modeled.
The analysis suggests that the annual NMOC generation rate can surge above the NSPS threshold
of 50 Mg per year after GCCS operation termination with an adequate increase in the decay rate.
The deterioration of the final cover, if any, after GCCS operation termination could allow moisture
into the landfill, which is expected to enhance waste decay rate, uncontrolled methane, and
NMOCs emissions. The final cover should be monitored and maintained until the remaining
generation potential and the leaching potential of the in-place waste have declined to the levels
which are unlikely to pose a risk to HHE. In addition, landfill owners and regulators should also
continue surface and subsurface emissions and odor monitoring to proactively identify signs of an
increase in LFG generation rate and have provisions in place to resume GCCS operation, if needed,
to control these issues.
Estimates of the remaining methane generation potential were predicted to decline to below 10%
of the total potential within 30 years after closure for all the sites except for Site F. The percent
remaining methane potential, however, is not an appropriate metric to assess the HHE impacts. A
smaller percent remaining methane generation potential at a large MSWLF may pose a greater risk
than a small MSWLF with relatively higher percent methane generation potential. A mass-based
threshold (e.g., NSPS threshold of 50 Mg/year for annual NMOCs generation rate) is a more
appropriate metric for HHE impact assessment than a percent-based criterion.
ctfill Leachate
Leachate collection rate and quality data available for the study sites were analyzed to:
(1) Determine whether any of the sites have stopped generating leachate or achieved a
historically low and stable leachate collection rate after the closure
(2) Assess approaches that can be used to estimate a site's long-term leachate collection rate
(3) Evaluate the comprehensiveness of the available leachate quality data with respect to the
parameters specified by the federal regulations for groundwater monitoring
(4) Conduct a screening analysis to identify the contaminants that have been frequently
measured in leachate above the respective human health risk-based protection standards at
the study sites after closure to identify the COPCs, and
(5) Assess whether the concentration of the COPCs identified based on the screening analysis
has declined below the respective protection standard with time after closure.
Leachate collection rates, quality, and quantity data were available for all study sites. Leachate
collection rates were also available for two separate and distinct cells at two of the sites (C and G).
The leachate collection rates were analyzed for 11 closed cells.
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The annual leachate collection rate across all 11 cells varied from 3 - 2,070 GPAD after closure,
with a median of 92 GPAD and an average of 190 GPAD; 90% of the annual collection rate
measurements were less than 500 GPAD. Leachate generation continued at all of the sites. The
leachate collection rates for all sites except for Sites A, CI, and F exhibited a generally declining
trend. Based on reported chloride concentrations and observed spikes in leachate collection during
the rainy season, the recent spike in leachate collection rate at Site F appears to be a result of
stormwater intrusion into the LCS infrastructure. None of the landfill cells examined exhibited
trends indicative of attaining a steady-state leachate collection/generation rate.
An estimate of the long-term leachate generation rate would be needed to reliably assess the HHE
impacts. Three approaches were used to estimate the future leachate generation rate from the study
sites: first-order decay modeling, unsaturated flow modeling, and HELP modeling. In general, all
three approaches provided comparable and reasonable fits to the measured study site data for all
the study sites except Sites CI, D, and G3. Sites CI, D, and G3 cells exhibited the lowest r2 value
for both the first-order decay and the unsaturated flow models. Unlike the first-order decay and
unsaturated flow model, the HELP model was not iteratively executed to obtain the best-fit to the
measured leachate collection rate data but was based on a mix of default parameters and the best-
fit parameters from the unsaturated flow model.
The comprehensiveness of the available leachate quality data at the study sites was analyzed with
respect to the parameters specified in the federal regulations for groundwater monitoring at Subtitle
D landfills (App I and App II of §258). The leachate quality data for Site G were not analyzed
since the data from this site represented the quality of leachate composited from closed and active
cell(s) of the site. The data for a total of 272 leachate constituents were evaluated to assess the
leachate quality of the study sites. The number of chemical constituents that must be monitored
for groundwater as part of App I and App II 40 CFR §258 are 62 and 215, respectively; all the
App I parameters are also included in the App II list. The parameters monitored varied widely
among the sites. Only three sites reported leachate constituent data for every App I parameter, and
one of these sites reported at least one measurement for all but three of the App II parameters. Two
study sites were encountered with leachate characteristic data available for less than half of App I
parameters. More than half of the study sites reported leachate constituent concentration data for
ten or fewer App II parameters (excluding App I parameters). The monitoring frequency varied
among the sites from once per month to twice per year, and further varied with time and
contaminant. Only a single measurement was available for a few of the organic compounds
measured at each site after closure. Apart from the lack of data for a large number of App I and II
parameters, the small number of measurements available for some constituents reported may limit
a reliable HHE impact assessment.
A screening analysis identified chemical constituents that have been frequently measured in
leachate above the respective risk-based protection standards at the study sites after closure to
identify the COPCs. Parameters that were never measured above the respective risk-based
standards are not expected to present a risk to HHE. The federal primary and secondary drinking
water standards were used as the thresholds for this evaluation. Fifteen parameters were measured
at concentrations above the respective MDL in more than 50% of the samples. Of these 15
parameters, six parameters (i.e., arsenic, TDS, iron, manganese, chloride, and color) were
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measured above their MCL/SMCL in more than 94% of the samples. Among all the constituents
with an MCL, arsenic and turbidity were the only primary MCL parameters that consistently
exceeded the MCL. Among the secondary parameters, greatest dilution and attenuation would be
needed for iron for its concentration to decline below its SMCL of 0.3 mg/L. A majority of the
iron measurements were below the regional screening level developed by the EPA for tapwater for
the ingestion pathway for iron (14 mg/L).
The temporal trend for most of the COPCs varied among the study sites. Arsenic showed a
declining trend for Sites A, B, C2, and E. Four of the most recent arsenic measurements at Site B
were below the MCL. The most recent set of measurements at Sites A, C2, and E were above the
MCL. Arsenic exhibited an increasing trend for Sites CI, D, and H. The arsenic release rate (lbs
per acre per day) from Site H exhibited a slightly declining trend over time. TDS and ammonia at
the sites, in general, appeared to be stable to slightly decreasing with time. In general, BOD
appeared to gradually decrease since closure for all the study sites for which BOD data were
available. Leachate COD appears to be slightly decreasing after closure at all of the sites. It should
be noted that the leachate quality is typically reflective of the decomposition status of the bottom-
most waste layer and does not necessarily represent the degree of stabilization of the entire landfill.
A well-decomposed waste layer above the LCS may attenuate the concentration of parameters
such as BOD and COD that are commonly used to assess leachate and waste stability.
The hydraulic efficiency of the primary liner was also evaluated for four of the study cells that are
lined with a double bottom liner system. The LDS rates, in general, show an overall declining trend
over time after closure. The aggregate hydraulic efficiency of the primary liner was calculated by
dividing the sum of annual LDS collection rates by the sum of the corresponding annual LCS and
LDS rates for all years with available LCS and LDS collection rates. The primary liner efficiency
was calculated to be 97.2%, 99.6%, 99.6%, and 96.8% for study cells G3, G4, H, and I,
respectively. However, a comparison of several leachate indicator parameters (chloride, and trace
organics or TOC) in primary and secondary leachate and groundwater suggests groundwater
intrusion into the LDS might be a significant source of liquids collected from LDS. The primary
liner efficiency at these sites is thus expected to be higher than the estimates presented above.
> - undwater Monitori » 1 ta
The groundwater monitoring data were analyzed for each of the study sites for App II parameters
to:
(1) Identify the parameters that were detected above the respective MCL or SMCL
(2) Assess whether the concentration of the contaminants that were frequently detected above
the respective MCL has declined below the MCL over time after closure, and
(3) Analyze groundwater quality data available for the study cells to identify examples of the
parameters that do not have an MCL and that were detected above the respective state-
specific risk-based standards.
It should be noted that the groundwater data were reviewed for all of the monitoring wells at the
study site for five sites (Sites A, B, E, and G). The data only from the wells around the study cells
at Sites C, D, F, H, and I were analyzed as these study cells were standalone cells and did not
adjoin an unlined or active cell.
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The impacts to groundwater were identified at the sites for three sets of parameters: parameters
with MCL, parameters with SMCL, and examples of parameters without MCL or SMCL. A large
number of App I and II parameters do not have an MCL/SMCL; only 32 and 51 of App I and App
II parameters, respectively, have an MCL, and only two parameters (i.e., silver and zinc) have an
SMCL. The groundwater impact with respect to these parameters without any MCL/SMCL can be
assessed by comparing the concentrations measured at downgradient wells to those of upgradient
well(s) or to a federal or state recommended risk-based standards.
Only occasional MCL exceedances were identified following cell closure among the studied sites.
No exceedances were identified at Site F for the range of data available for review (i.e., post-
closure Year 8 through Year 19). Only Sites G, H, and I were observed to exhibit reoccurring
groundwater exceedances. More than 55% and 10% of samples at Site H had beryllium and
thallium concentration greater than the associated MCL, respectively. Approximately 40% and
30%) of beryllium and thallium exceedances observed at Site H occurred at the upgradient well(s),
respectively. Arsenic and lead concentrations in 107 and 53 samples were greater than the
associated MCL at Site G, respectively. More than 10% of arsenic and lead exceedances observed
at Site G occurred at the upgradient well(s). More than 50% of the arsenic exceedances observed
at Site G correspond to the wells that were downgradient or close to the unlined cell at the site.
The beryllium and thallium exceedances at Site H and arsenic and lead exceedances at Site G
cannot be conclusively attributed to lined cells at these sites due to the exceedances observed above
the respective MCL at several upgradient wells for these parameters at these sites. The laboratory
detection limit used was greater than the respective MCLs for several reported measurements (e.g.,
several organics at Site F, several arsenic and thallium measurements at Site C).
The arsenic exceedances at Site I have been consistently observed at several downgradient wells
after closure. The site monitoring reports suggest that these elevated arsenic levels at Site I may
be a result of subsurface geochemical changes. An analysis of the arsenic:chloride ratio in
groundwater and leachate at Site I also suggest that leachate release, if any, is not the primary
contributor of arsenic exceedances at the site. The change in the biogeochemical environment of
the surficial aquifer resulting from landfill construction can also result in mobilization of naturally-
occurring constituents and impacts to groundwater quality.
The available silver and zinc data were analyzed to identify exceedance with respect to the SMCL.
These data were not available for all of the sites. For example, zinc and silver measurements were
not available for Sites I and E, respectively. Only one site with these data had an exceedance of
one of these parameters; Site G had a silver exceedance three years after closure at a single well.
It is recommended that the parameters that do not have a federal MCL/SMCL should also be
monitored for a comprehensive HHE impact assessment. The site-specific background
concentration and/or state-specified human-health risk-based standard can be used for assessing
the risk to HHE for these parameters. As of the time of this evaluation, all three of the sites under
assessment monitoring/corrective action had instances where parameters without an MCL
contributed to groundwater impacts. Examples of these parameters include vanadium (Site A), 1,1-
dichloroethane (Site F), and 1,4-dioxane (Site I).
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( :inal Covt 1 • ormance
The final cover performance at six of the nine study sites was evaluated by analyzing available
SEM and settlement data. However, a complete dataset covering the entire period from the closure
date to the time of this study was only available for two of the sites (Site D and Site H).
The available data suggest that the surface emissions exceedances frequently occurred at study
cells at Sites C, E, and H. The highest number of exceedances was observed over two monitoring
events at Site I; wells penetrations through the final cover were documented as the primary factor
contributing to these exceedances. Overall, site SEM data suggests a decreasing trend of SEM
exceedances over time. The final SEM exceedance observed at five out of six sites were within
ten years of closure. SEM exceedances beyond the initial ten years of PCC were only observed at
one site, (Site C), and exceedances at Site C were noted throughout the three years of available
data (years 16-18) potentially due to the lack of a soil layer where methane oxidation could occur
over the exposed geomembrane cap.
Limited settlement data were available for only a few study sites (Sites C, E, and F), with the
exception of Site I, which had annual settlement data available. The topographic data for the
closure year of Site F were not available; therefore, an analysis of trends in the settlement rate for
Site F was not conducted. The median annual settlement for Site I decreased from 2.4 ft in the first
year after closure to 0.13 ft in the 17th year after closure. The range of the settlements observed at
the survey points appears to decrease over time from 2.5 ft during the first year to approximately
0.5 ft in the 17th year after closure for Site I. The topographic surfaces at Site I were compared
with bottom grade surfaces to estimate in-place waste volume for each year, which were used to
estimate the annual volume loss rate. The volume loss rate was modeled using a first-order decay
equation. The results of the best fit analysis suggest a total volume loss of 670,000 yd3 over an
infinite time horizon at Site I. Approximately 99% of the total settlement is estimated to occur
within 30 years after closure. The settlement and volume loss estimates are regarded as
overestimations due to unaccounted settlement of subsurface soils below the landfill as a result of
overburden pressure.
p . i isideration i • mi sessing and Mitigatir I ng-tei in 111¦ | - ids
9.2.1. Operating Considerations
The risk to the human health and the environment is contingent on the contaminant(s) mass loading
rate, which is a combination of the flow rate (e.g., leachate generation rate or landfill gas flow rate)
and contaminant(s) concentration, into the environment (Morris and Barlaz 2011). Compromises
in the integrity of the final cover would potentially result in moisture intrusion into the landfill and
subsequent increase leachate collection/generation rates. The concentration of leachate
constituents and landfill gas flow rate and composition are expected to be contingent on the
biodegradability of the deposited waste and/or leaching potential when exposed to moisture. The
possibility of future final cover compromises and ensuing emissions is one of the key concerns of
PCC termination. These concerns can be alleviated by implementing operating strategies (e.g.,
bioreactor operation) that can stabilize waste and flush out or stabilize leachable contaminants
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before terminating PCC and consequently mitigate the importance of the containment system
performance in protecting the HHE after PCC termination (Morris and Barlaz 2011).
As bioreactor landfill operations enhance the waste decomposition rate and substantially reduce
the time period over which waste decomposition occurs, leachate contaminants (except for
contaminants such as ammonia and major ions such as chlorides) concentrations and landfill gas
generation rate are relatively high during the early phases of bioreactor operation and
comparatively low once the decomposition has occurred. The amount of in-place waste that can
potentially degrade in the future is expected to be lower with bioreactor operation than that
associated with the conventional dry tomb operation. In addition, the cap of a bioreactor landfill is
expected to be less susceptible to damage from the differential settlement than a dry tomb landfill
as a majority of differential settlement associated with waste decomposition has occurred before
closure or during the PCC period when the site is actively monitored. Bioreactor operation is also
expected to result in lower leachate management costs as some of the leachate recirculated into the
landfill would be absorbed by the waste. The bioreactor operation, however, also has design,
operating, and monitoring challenges including but not limited to unavailability of moisture source
especially in arid areas, the complexity of adding liquids to achieve uniform moisture distribution
in the landfill, flooding of gas collection devices, and a need to collect and manage excess leachate
at the end of bioreactor operation (Townsend et al. 2015).
9.2.2. Monitoring Considerations
The ability for the owner and/or operator of an MSWLF unit to evaluate the HHE impacts
following completion of PCC is dependent on having quality environmental monitoring data
specific to the MSWLF unit of interest. For several of the study sites evaluated in this report,
various categories of monitoring data were not available for analysis for this study. The following
monitoring data would be useful for assessing whether or not an MSWLF unit would be protective
of HHE in the event of termination or reduction of PCC:
Waste Tonnage and Composition - while waste-specific (e.g., MSW, CDD) tonnages are
typically well-documented and were available for a majority of the study sites, detailed
composition studies or records documenting specific incoming waste types were scarce. The waste
composition study data were not available for any of the study sites except Site D. In addition, the
amount and chemical characterization data of different non-MSW materials such as industrial
waste, combustion residuals, and contaminated soils deposited in the landfill are valuable for HHE
impact assessment.
GCCS - Typically, GCCS atMSWLFs (e.g., Site C) with active cells are progressively expanded
with time to collect LFG from newly filled areas, and the LFG from several closed and active cells
are routed to a blower/flare system. Although LFG flow rate and composition data are collected
from each well, these are not typically tracked for individual cells. Study cell-specific LFG flow
rate and composition data were not available at several sites (e.g., Sites B, C, E, and G). From a
PCC monitoring standpoint, MSWLF owners/operators would benefit by installing devices that
allow frequent monitoring of the flow rate and quality of LFG collected from standalone closed
MSWLF units as these data would allow estimation of decay rate and the remaining methane
potential of these cells.
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LCS - The leachate quality (e.g., Site G) and quantity (e.g., Site B, Site E) at some of the study
sites were not independently tracked for the study cell. Depending on the timing of the closure of
various cells and the PCC goals, the owner/operator may consider PCC termination of standalone
cells or cluster of cells while operating the other cells at the site. However, the lack of monitoring
data from standalone individual or clusters of cells may preclude early PCC termination of these
cells. Like LFG, leachate is typically routed to a single location (e.g., for leachate storage/pre-
treatment) before transporting/pumping off-site for treatment and disposal. The study site data
suggest that leachate generation/collection rate and quality may vary substantially among various
collection points. MSWLFs owners/operators should consider monitoring leachate collection rates
and quality from various collection points. These data would allow identification of individual
cells (e.g., with elevated leachate collection rates and/or contaminant levels) that may need to be
monitored and operated (e.g., cell-specific leachate recirculation) more rigorously than the other
cells/areas. These data over time would also be helpful in identifying localized stormwater
intrusion, if any, into LCS/leachate sump.
As leachate quality tracking at individual leachate collection points can be cost-intensive,
contaminant-specific monitoring frequency should be considered. For example, field parameters
(e.g., pH and specific conductivity) can be monitored more frequently than the laboratory
parameters. The laboratory parameters that are frequently measured above the respective
groundwater protection standards can be monitored more frequently than the parameters that are
consistently undetected, thus reducing the cost burden while collecting adequate leachate quality
data for a robust HHE impact assessment.
MSWLFs owners and operators should consider routine monitoring of leachate quality even
though it not required by RCRA Subtitle D regulations. For a more comprehensive HHE impact
assessment, MSWLF owners/operators should consider harmonizing the list of monitoring
parameters for groundwater and leachate and ensuring that the laboratory reporting limits of the
monitored parameters are lower than the respective groundwater protection standard. Leachate
quality should be monitored for the parameters that are required for groundwater monitoring as
well as for constituents that occur at elevated levels in leachate (e.g., chloride, ammonia) and those
that can be used for assessing the waste stability (e.g., BOD and COD).
Groundwater Monitoring System - A specific challenge was identified when attempting to
analyze groundwater impacts associated with the study cells at the sites with the presence of
adjoining unlined or active cells. While isolating the groundwater impacts of adjoining cells may
not be practically feasible, independent groundwater monitoring of standalone closed cells would
allow identification of the sources of groundwater impacts, if any. A periodic review of changes
such as surrounding land use/zoning changes that can impact the groundwater flux and flow
direction should be considered while assessing the long-term impacts of modifying or terminating
PCC. Monitoring of groundwater quality with respect to leachate indicator parameters such as
chloride, ammonia, BOD, and COD would help to assess the contribution of leachate release, if
any, on groundwater quality impacts.
Settlement - Differential settlement of the landfill surface represents one of the most probable
risks to the integrity of the final cover. Routine settlement monitoring data can be used to estimate
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the future settlement rate. In addition, settlement data, when used in conjunction with a temporal
analysis of LFG collection and leachate quality, can provide an indication of waste stabilization.
Settlement data were not available for several sites, including Sites A, D, and G. Routine
topographic surveys of the final cover during PCC, including a survey of the final cover
immediately following the closure, provide an opportunity to evaluate the rate of settlement and
waste stabilization.
Monitoring Records - Some monitoring data (e.g., subsurface probe monitoring and SEM),
which are typically required to be routinely monitored for MSWLFs, were not available for several
monitoring events for review/analysis for this study. For example, SEM data were not available
for review for several years for Sites C, E, and I. The state regulators and MSWLFs
owners/operators should consider implementing systems for cataloging monitoring data for
prompt retrieval and analysis.
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