EPA600/R-21/162 | July 2021 | www.epa.gov/research
Life Cycle Inventories of End-of-Life
Pathways of Construction and
Demolition Materials
&EPA
, „ „ . „ United States
Office of Research and Development Environmental
Center for Environmental Solutions and Emergency Response Protection Agency
Land Remediation and Technology Division
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E P A/600/R-21 /162
July 2021
Life Cycle Inventories of End-of-Life
Pathways of Construction and
Demolition Materials
Briana Niblick1, Danny Hage2, Sarah Cashman2, Justin Smith3, Pradeep
Jain3, Ashley Edelen4, Timothy Townsend3, Wesley Ingwersen1
Environmental Decision Analytics Branch
Land Remediation and Technology Division
Center for Environmental Solutions and Emergency Response
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH 45268
1 U.S. EnvironmentalProtection Agency
2 Eastern Research Group, Inc.
3 Innovative Waste Consulting Services, LLC
4 Oak Ridge Institute for Science and Education
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Acknowledgements and Notice
The U.S. Environmental Protection Agency (US EPA), through its Office of
Research and Development (ORD), funded and managed the research described
herein under contractnumberEP-C-15-012 to Computer Sciences Corporation (CSC)
and under contract number EP-C-16-015 to Eastern Research Group, Inc. (ERG).
This report has been subjected to review by the Office of Research and Development
and approved for publication. Approval does not signify that the contents reflect the
views of the Agency, nor does mention of trade names or commercial products
constitute endorsement or recommendation for use.
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Foreword
The U.S. 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-
d riven research and provides r espo nsi v e t ech n i cal 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, Ph.D., Director
Center for Environmental Solutions and Emergency Response
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Executive Summary
The Sustainable and Healthy Communities Program has a mission to develop
data and tools that enable community leaders to integrate environmental, societal,
and economic factors into their decision-making processes and thus foster
community sustainability. This report examines one key area of community
sustainability interest, the management of materials from the construction and
demolition of buildings, roads, and other structures at their end of life (EOL).
Life Cycle Assessment (LCA) is an approach frequently used to examine the
environmental implications of the EOL management of materials, and while
much LCA research has focused on materials from household and community
activities (e.g., municipal solid waste), very little effort has focused on
construction and demolition debris (CDD). Though CDD constitutes a substantial
volume of material, the impact that these materials have on human and ecological
health has not been recognized in the same manner as other wastes, and thus they
have been less studied.
A meaningful LCA requires a robust database of information (e.g., material
composition and magnitude, energy consumption, and environmental emissions)
from throughout a material or product's life cycle. Compilations of such data-a
life cycle inventory (LCI) - provide the backbone for conducting an LCA to
examine different materials management strategies. The primary objective ofthe
work presented here was to extensively assess the body of knowledge regarding
CDD life-cycle data and to compile US-specific LCIs for distinct CDD material
categories from publicly available sources. These LCI datasets are intended to
complement background databases, which include LCIs for a variety of processes
and services such as natural resource extraction, manufacturing, energy
production, and transportation. An additional objective of this research was to
identify data gaps pertaining to CDD LCIs and thus identify future areas of focus.
LCIs were developed for the EOL management perspective of the following CDD
materials: asp halt pavement, asphalt shingles, gyp sum dry wall, CDD wood, land
clearing debris (LCD), Portland cement concrete, polyvinyl chloride (PVC),
fiberglass insulation, carpet and padding, clay bricks, copper wire, and vinyl
composition tile (VCT). Current EOL management practices were identified
based on published industry, government, and peer-reviewed literature.
Although the CDD LCIs in this report represent the most comprehensive datasets
currently available on this material stream, they are still limited due to the relative
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scarcity of US-specific data since CDD has not been examined to the same degree
as other EOL materials. Major data gaps include, but are not limited to, the
following:
i. State-specific tracking data for quantities of CDD materials processed or
recycled. States tend to track the quantities landfilled, which is only one
potential route for final disposition of CDD materials.
ii. Some materials, such as copper wire, are lost to activities outside of
tracking capabilities, e.g. removed from the building site by outside
parties before entering the processing stream. There need to be better
estimates for these non-tracked quantities.
iii. Long-term liquid and gaseous emissions from CDD and MSW landfills,
especially as the mix of different CDD materials may present a different
environmental profile than an individual material alone.
iv. More precise emissions modeling for material incineration scenarios.
Many of the air emissions due to incineration are based on laboratory tests
of standard or proxy materials, which can vary from what actually occurs
in practice.
v. Transport distances between management processes for discarded and
processed CDD materials. Distances for modeling were typically set at a
fixed, assumed quantity since these transport data were not available in the
literature, nor were sufficient data available to enable an accurate
geospatial analysis.
To complement the data gap analysis, several recommendations for material-
specific data collection were identified and are described in each chapter of the
report.
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Table of Contents
Table of Contents vii
List of Figures xi
List of Tables xiii
List of Abbreviations, Acronyms, and Initialisms xvi
1 Introduction 1-1
1.1 Background 1-1
1.2 Scope of Work and Objectives 1-2
1.2.1 CDD Materials in this Report 1-2
1.3 Report Organization 1-6
1.4 References 1-6
2 Approaches to Life-Cycle Data Management in the EOL Phase 2-7
2.1 The EOL Phase B ound ary 2-7
2.2 Life Cycle Inventory 2-7
2.3 Organization of LCI Datasets 2-9
2.4 Common Technosphere Inputs 2-9
2.4.1 Transportation 2-9
2.4.2 Equipment Use and Fuel Combustion 2-10
2.4.3 Water Consumption 2-10
2.4.4 Stormwater Management 2-11
2.4.5 Aggregates and Soil 2-11
2.4.6 Particulate Emissions 2-12
2.5 Mixed CDD Processing 2-12
2.6 Landfilling 2-16
2.6.1 Background 2-16
2.6.2 Transportation 2-16
2.6.3 Equipment Use and Equipment-Specific Fuel Consumption 2-18
2.6.4 Construction Phase 2-19
2.6.5 Operations Phase 2-20
2.6.6 Closure and Post-Closure Phase 2-22
2.6.7 Landfill Leachate Emissions 2-27
2.6.8 Landfill Gas Emis sions 2-27
2.6.9 Landfill Gas and Leachate Collection and Treatment 2-30
2.6.10 Data Gap Analysis of Landfill Gas and Landfill Leachate Collection and
Treatment 2-31
2.7 References 2-32
3 Asphalt Pavement 3-1
3.1 Introduction 3-1
3.2 LCI Sources 3-2
3.3 LCI Related to Disposal 3-2
3.4 LCI Related to Recycling 3-4
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3.4.1 RAP Processing 3-4
3.4.2 RAP Use as Aggregate 3-4
3.5 Data Gap Analysis and Opportunities for Additional LCI Data 3-6
3.6 References 3-8
4 Asphalt Shingles 4-1
4.1 Introduction 4-1
4.2 EOL Management 4-2
4.3 LCI Sources 4-3
4.4 LCI Related to Disposal 4-3
4.5 LCI Related to Recycling 4-5
4.5.1 Shingle Processing 4-5
4.5.2 HMA/WMA Production with Asphalt Shingles 4-8
4.6 Data Gaps and Future Opportunities 4-9
4.7 References 4-12
5 Gypsum Drywall 5-1
5.1 Introduction 5-1
5.2 EOL Management 5-1
5.3 LCI Sources 5-3
5.4 LCI Related to Disposal 5-4
5.5 LCI Related to Recycling 5-6
5.5.1 Recovered Drywall Processing 5-6
5.6 Data Gap Analysis and Opportunities for Additional LCI Data 5-8
5.7 References 5-9
6 Wood 6-1
6.1 Introduction 6-1
6.2 EOL Management 6-2
6.3 LCI Sources 6-4
6.4 LCI Related to Disposal 6-5
6.5 LCI Related to Recycling 6-9
6.5.1 Processing of CDD Wood 6-9
6.5.2 Land Application of Mulched CDD Wood 6-10
6.5.3 Wood Ash 6-11
6.6 Data Gap Analysis and Opportunities for Additional LCI Data 6-12
6.7 References 6-14
7 Land Clearing Debris 7-1
7.1 Introduction 7-1
7.2 EOL Management 7-1
7.3 LCI Sources 7-2
7.4 LCI Related to On-Site Burning 7-3
7.5 LCI Related to Landfill Disposal 7-5
7.6 LCI Related to Recycling 7-6
7.6.1 LCD Used as Mulch 7-6
7.7 Data Gap Analysis and Opportunities for Additional LCI Data 7-10
7.8 References 7-12
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8 Portland Cement Concrete 8-1
8.1 Introduction 8-1
8.2 EOL Management 8-2
8.3 LCI Sources 8-4
8.4 LCI Related to Removal/Demolition 8-5
8.5 LCI Related to Disposal 8-6
8.6 LCI Related to Recycling 8-7
8.6.1 Concrete Processing 8-7
8.6.2 RCA Use as Aggregate 8-8
8.6.3 Demolished Concrete Use as Soil Fill Replacement 8-9
8.7 Data Gap Analysis and Opportunities for Additional LCI Data 8-10
8.8 References 8-12
9 Polyvinyl Chloride 9-1
9.1 Introduction 9-1
9.2 EOL Management 9-1
9.3 LCI Sources 9-3
9.4 LCI Related to Disposal 9-5
9.5 LCI Related to Incineration 9-5
9.6 Data Gap Analysis and Opportunities for Additional LCI Data 9-7
9.7 References 9-8
10 Fiberglass Insulation 10-1
10.1 Introduction 10-1
10.2 EOL Management 10-2
10.3 LCI Sources 10-2
10.4 LCI Related to Disposal 10-3
10.5 Data Gap Analysis and Opportunities for Additional LCI Data 10-5
10.6 References 10-5
11 Carpet 11-1
11.1 Introduction 11-1
11.2 EOL Management 11-3
11.3 LCI Sources 11-5
11.4 LCI Related to Disposal 11-6
11.5 LCI Related to Recycling 11-7
11.6 LCI Related to Incineration 11-9
11.7 LCI Related to Cement Kiln Combustion 11-12
11.8 Data Gap Analysis and Opportunities for Additional LCI Data 11-13
11.9 References 11-15
12 Clay Bricks 12-17
12.1 Introduction 12-17
12.2 EOL Management 12-17
12.3 LCI Sources 12-18
12.4 LCI Related to Disposal 12-19
12.5 LCI Related to Recycling 12-19
12.5.1 Clay Brick Demolition 12-19
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12.5.2 Clay Brick Sorting 12-19
12.5.3 Clay Brick Use as Aggregate 12-20
12.6 Data Gap Analysis and Opportunities for Additional LCI Data 12-20
12.7 References 12-20
13 Copper Wire 13-22
13.1 Introduction 13 -22
13.2 EOL Management 13-22
13.3 LCI Related to Disposal 13-23
13.4 LCI Related to Incineration 13-23
13.5 Data Gap Analysis and Opportunities for Additional LCI Data 13-24
13.6 References 13-24
14 Vinyl Composition Tile 14-26
14.1 Introduction 14-26
14.2 EOL Management 14-26
14.3 LCI Related to Disposal 14-27
14.4 LCI Related to Incineration 14-27
14.5 Data Gap Analysis and Opportunities for Additional LCI Data 14-30
14.6 References 14-30
15 Summary and Future Research Needs 15-1
15.1 Summary 15-1
15.2 Data Gaps and Future Research Opportunity 15-3
15.3 References 15-5
16 Acknowledgements 16-1
17 Appendix A: LCI Review Template 17-1
18 Appendix B: Bridge Process Creation 18-1
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List of Figures
Figure 1-1. Composition by mass of CDD generated in the United States (US EPA 2020).
Due to differences in estimation procedures, land clearing debris, fines, cardboard,
and plastics have been excluded from the total mass 1-3
Figure 2-1. Typical composition by mass of CDD materials processed by material recovery
facilities in the United States (CDRA 2015) 2-13
Figure 2-2. Unit process flow diagram for a typical mixed CDD material recovery facility
(MRF), also known as a CDD processing plant 2-13
Figure 2-3. Unit process flow diagram for construction, operation, closure, and post-closure
care of a CDD landfill receiving biodegradable CDD materials, such as wood and
paper 2-24
Figure 2-4. Unit process flow diagram for construction, operation, closure, and post-closure
care of a CDD landfill receiving gypsum drywall 2-24
Figure 2-5. Unit process flow diagram for construction, operation, closure, and post-closure
care of a CDD landfill receiving non-biodegradable CDD materials 2-25
Figure 2-6. Unit process flow diagram for construction, operation, closure, and post-closure
care of an MSW landfill receiving biodegradable CDD materials, such as wood and
paper 2-25
Figure 2-7. Unit process flow diagram for construction, operation, closure, and post-closure
care of an MSW landfill receiving gypsum drywall 2-26
Figure 2-8. Unit process flow diagram for construction, operation, closure, and post-closure
care of an MSW landfill receiving non-biodegradable CDD materials 2-26
Figure 3-1. Process flow diagram for end-of-life management of asphalt pavement 3-1
Figure 3-2. Unit process flow diagram for the use of reclaimed asphalt pavement as general
fill 3-6
Figure 4-1. Process flow diagram for end-of-life management of asphalt shingles 4-2
Figure 4-2. Unit process flow diagram for asphalt shingles at the processing facility 4-8
Figure 5-1. Process flow diagram for end-of-life management of gypsum drywall. MRF =
material recovery facility 5-1
Figure 5-2. Sources of gypsum drywall discarded for processing or final disposal in
California (CIWMB 2007). California data were used due to data availability and a
growing CDD sector 5-2
Figure 5-3. Unit process flow diagram for gypsum recovered from drywall processing 5-8
Figure 6-1. Process flow diagram for end-of-life management of treated and untreated
wood. Four types of wood treatments are included: ACQ, CBA, CCA, and DOT.
Only untreated wood is processed for mulch. ACQ = alkaline copper quaternary;
CBA= copper azole; CCA= chromate copper arsenate; DOT = disodium octaborate
tetrahydrate; MRF = material recovery facility 6-2
Figure 6-2. Unit process flow diagram for application of CDD wood mulch to land 6-11
Figure 7-1. Process flow diagram for end-of-life management of land clearing debris
(LCD) 7-1
Figure 7-2. Unit process flow diagram for land clearing debris at the site of open-burning
combustion 7-5
Figure 7-3. Unit process flow diagram for land clearing debris at the air curtain incinerator.
7-5
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Figure 7-4. Unit process flow diagram for mulched land clearing debris at the processing
facility 7-7
Figure 7-5. Unit process flow diagram for the land application of mulched land clearing
debris 7-10
Figure 8-1. Process flow diagram for the end-of-life management of Portland cement
concrete 8-2
Figure 8-2. Distribution of recycled concrete applications in the US (Deal 1997) 8-3
Figure 8-3. Unit process flow diagram of recycled concrete aggregate used as general site
fill 8-9
Figure 8-4. Unit process flow diagram for concrete debris used as general site fill 8-10
Figure 9-1. Process flow diagram for EOL management of PVC products without
plasticizers 9-1
Figure 9-2. Approximation of 2011 CDD PVC material EOL management pathways.
Calculated from US EPA 2015a and vanHaaren et al. 2010 9-2
Figure 9-3. Unit process flow diagram for incineration of PVC products without
plasticizers at waste-to-energy (WTE) facility 9-7
Figure 10-1. Process flow diagram for EOL management of fiberglass insulation products.
10-2
Figure 11-1. Fraction of carpet managed through different EOL pathways (CARE 2017) 11-1
Figure 11-2. Mass composition of carpet waste (Ucar and Wang 2011) 11-2
Figure 11-3. Residential fiber mix (Realff 2011 as cited in US EPA 2019) 11-2
Figure 11-4. Process flow diagram for end-of-life management of carpet and carpet
padding. MRF = material recovery facility; PU = polyurethane 11-3
Figure 11-5. Unit process flow diagram for carpet recycling 11-8
Figure 11-6. Unit process flow diagram for carpet padding recycling 11-9
Figure 11-7. Unit process flow diagram for carpet incineration at an MSW waste-to-energy
facility 11-10
Figure 11-8. Unit process flow diagram for carpet padding incineration at an MSW waste-
to-energy facility 11-12
Figure 11-9. Unit process flow diagram for carpet (21% by mass) and coal (79% by mass)
combustion in a cement kiln 11-13
Figure 12-1. Process flow diagram for end-of-life management of clay bricks 12-17
Figure 13-1. Process flow diagram for end-of-life management of copper wire 13-22
Figure 14-1. Process flow diagram for end-of-life management of vinyl composition tile
(VCT) 14-26
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List of Tables
Table 1-1. Generation by mass of CDD materials in the United States in 2018 (US EPA
2020) 1-4
Table 1 -2. CDD processes by NAICS category 1-4
Table 2-1. Standard LC A flow types, descriptions of flow types, and examples of product
systems 2-8
Table 2-2. Terms commonly used in EOL CDD management, their respective descriptions,
and examples of each term 2-8
Table 2-3. Estimated transport distances for aggregate materials (developed from USCB
2010) 2-12
Table 2-4. Percentages of CDD materials transported to CDD landfills and MSW landfills
by highway, rail, and waterway 2-17
Table 2-5. Mass percentages and transport distances of CDD materials to MSW and CDD
landfills, by transport mode 2-18
Table 2-6. Landfill gas production properties forthree CDD materials: lumber, cardboard,
and gypsum drywall 2-28
Table 2-7. Comparison of methane and carbon dioxide emissions resulting from disposal
of CDD materials in CDD and MSW landfills 2-30
Table 3 -1. End-of-life management processes for reclaimed asphalt pavement 3-1
Table 3 -2. Life cycle inventory (LCI) data sources for reclaimed asphalt pavement 3-2
Table 3-3. Life cycle inventory and associated data sources for the unit process: "asphalt
pavement, at processing plant" 3-4
Table 3-4. Life cycle inventory data sources for reclaimed asphalt pavement. "P" indicates
partial coverage of the data, "X" indicates full coverage, and indicates no data 3-8
Table 4-1. Life cycle inventory (LCI) data sources for asphalt shingles 4-3
Table 4-2. Life cycle inventory data sources for asphalt shingles. "P" indicates partial
coverage of the data, "X" indicates full coverage, and indicates no data 4-11
Table 5-1. End-of-life management processes for gypsum drywall 5-3
Table 5-2. Overview of life cycle inventory data sources associated with gypsum drywall 5-4
Table 5-3. Overview of life cycle inventory data sources associated with gypsum drywall.
"P" represents partial coverage of the required data, while "X" represents full
coverage. Section 5.6 describes the levels of data coverage in more detail 5-9
Table 6-1. End-of-life management processes for treated and untreated wood 6-3
Table 6-2. Overview of life cycle inventory data sources associated with treated and
untreated wood 6-5
Table 6-3. Overview of life cycle inventory data sources associated with treated and
untreated wood. "P" represents partial coverage of the required data, while "X"
represents full coverage 6-14
Table 7-1. End-of-life management processes for land clearing debris (LCD) 7-2
Table 7-2. Overview of life cycle inventory data sources associated with land clearing
debris (LCD) 7-3
Table 7-3. Overview of life cycle inventory data sources associated with land clearing
debris. "P" represents partial coverage of the required data, while "X" represents full
coverage 7-11
Table 8-1. End-of-life management processes for Portland cement concrete 8-4
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Table 8-2. Overview of life cycle inventory data sources associated with Portland cement
concrete 8-5
Table 8-3. Overview of life cycle inventory data sources associated with Portland cement
concrete. "P" represents partial coverage of the required data, while "X" represents
full coverage 8-11
Table 9-1. End-of-life management processes for PVC 9-3
Table 9-2. Overview of life cycle inventory data sources associated with PVC 9-3
Table 9-3. Overview of US-based LCI data available. "P" represents partial coverage of
the ideal data 9-8
Table 10-1. Approximate mass composition of non-faced fiberglass insulation (US EPA
2015) 10-1
Table 10-2. End-of-life management processes for fiberglass 10-2
Table 10-3. Overview of life cycle inventory data sources associated with fiberglass
insulation 10-3
Table 11-1. End-of-life management processes for carpet 11-4
Table 11 -2. Overview of life cycle inventory data sources associated with carpet 11-5
Table 11-4. Overview of US LCI data utilized for dataset development. "P" represents
partial coverage of the required data, while "X" represents full coverage 11-15
Table 12-1. End-of-life management processes for clay brick 12-18
Table 12-2. Overview of life cycle inventory data sources associated with clay bricks 12-19
Table 13-1. End-of-life management processes for copper wire 13-22
Table 14-1. End-of-life management processes for vinyl composition tile (VCT) 14-26
Table 15-1. Summary of CDD life cycle inventory processes included in this report. Refer
to AppendixB for the creation of bridge processes for additional data and emissions.
15-1
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List of Abbreviations, Acronyms, and Initialisms
ACQ Alkaline Copper Quaternary
ADC Alternative Daily Cover
AP-42 Compilation of Air Pollutant Emission Factors
A SMI Athena Sustainable Materials Institute
BDL Below-Detection Limit
BOD Biochemical Oxygen Demand
BTEX Benzene Toluene Ethylene Xylene
BTS Bureau of Transportation Statistics
Btu British Thermal Units
CDD Construction & Demolition Debris
CB A Copper Azole
CCA Chromated Copper Arsenate
CDRA Construction and Demolition Recycling Association
CFCs Chlorofluorocarbons
CKD Cement Kiln Dust
CLT Cross-Laminated Timber
COD Chemical Oxygen Demand
CORRIM Consortium for Research on Renewable Industrial Materials
DCA Dense Concrete Aggregate
DOT Disodium Octaborate Tetrahydrate
EAPA European Asphalt Pavement Association
EIA Energy Information Administration
EOL End of Life
ERG Eastern Research Group, Inc.
FDEP Florida Department of Environmental Protection
FGD Flue Gas Desulfurization
FHWA Federal Highway Administration
FML Flexible Membrane Liner
ft3 Cubic Foot
GCCS Gas Collection and Control Systems
GHG Greenhouse Gas
GREET Greenhouse Gases, Regulated Emissions and Energy Use in
Transportation
HCFCs Hydrochlorofluorocarbons
HDPE High Density Polyethylene
HHD Heavy Heavy-Duty
HMA Hot Mix Asphalt
hr Hour
IE Impact Estimator
ILCD International Reference Life Cycle Data
IPCC Intergovernmental Panel on Climate Change
IVL Swedish Environmental Institute
IWCS Innovative Waste Consulting Services, LLC
kg Kilogram
kJ Kilojoule
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km
Kilometers
kWh
Kilowatt-Hours
kWh/kg
Kilowatt-Hours per Kilogram
L
Liter
lb
Pound
LCA
Life Cycle Assessment
LCD
Land Clearing Debris
LCI
Life Cycle Inventory
LCS
Leachate Collection System
LCRS
Leachate Collection and Removal System
LEAF
Leaching Environmental Assessment Framework
LEED
Leadership in Energy & Environmental Design
LFG
Landfill Gas
LMOP
Landfill Methane Outreach Program
m
Meter
m2
Square Meter
m3
Cubic Meter
MDF
Medium-Density Fiberboard
Hg
Micrograms
mg
Milligrams
MJ/m3
Mega Joules per Cubic Meter
MMT
Million Metric Tons
MOVES
Motor Vehicle Emissions Simulator
MSW
Municipal Solid Waste
MSW-DST
Municipal Solid Waste-Decision Support Tool
MT
Metric Tons
n.d.
No Date
ng
Nanogram
n.r.
No Record
NAICS
North American Industry Classification System
NAPA
National Asphalt Pavement Association
NCSU
North Carolina State University
ND
Not Detected
NMOC
Non-Methane Organic Carbon
NMVOC
Non-Methane Volatile Organic Compound
NO A A
National Oceanic and Atmospheric Administration
NOx
Nitrogen Oxides
NRC
National Resources Canada
NREL
National Renewable Energy Laboratory
NRMRL
National Risk Management Research Laboratory
O&M
Operation(s) and Maintenance
ORD
Office of Research and Development
OSB
Oriented Strand Board
PAH
Polycyclic Aromatic Hydrocarbons
PAS
Publicly Available Specification
PCC
Portland Cement Concrete
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PE
Polyethylene
PM
Particulate Matter
PVC
Polyvinyl Chloride
RAP
Reclaimed Asphalt Pavement
RCA
Recycled Concrete Aggregate
RCRA
Resource Conservation & Recovery Act
SHC
Sustainable and Healthy Communities
SOx
Sulfur Oxides
SPLP
Synthetic Precipitation Leaching Procedure
t*km
Tonne-Kilometer
TCLP
Toxicity Characteristic Leaching Procedure
IDS
Total Dissolved Solids
TK
Transfer Coefficient
TNT
Trinitrotoluene
TOC
Total Organic Carbon
Tonne
Metric Ton
TRPH
Total Recoverable Petroleum Hydrocarbon
TSS
Total Suspended Solids
US
United States
US EPA
United States Environmental Protection Agency
USCB
United States Census Bureau
USD A
United States Department of Agriculture
USGS
United States Geological Survey
VOC
Volatile Organic Compound
WARM
Waste Reduction Model (US EPA Tool)
WRATE
Waste and Resources Assessment Tool for the Environment
WWTP
Wastewater Treatment Plant
XRF
X-ray Fluorescence
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1 Introduction
1.1 Background
Decision-making processes used to evaluate environmental, economic, and social implications in
a resource-constrained world are often poorly characterized in terms of potential impacts to human
health, ecosystem services, economic vitality, and social equity. The need for a decision-making
approach that accounts for all three pillars of sustainability (environment, economy, and society)
is widely recognized as a key component for transitioning to a more sustainable world (USGS
1998, US EPA 2009, Pereira 2012). The US EPA's Sustainable and Healthy Communities
Research Program (SHC) strives to provide tools for community decision-makers to incorporate
human health, socio-economic, environmental, and ecological factors into their decisions to
promote effective community sustainability. Decisions pertaining to waste and materials
management have been identified as one of the highest priorities for implementing sustainable
practices (US EPA 2012). Life cycle assessment (LCA), a scientific method that evaluates the
environmental impacts throughout the life cycle of a material or process, can be used as a tool to
comprehensively assess environmental and human health implications of material management
options. One of the critical underlying components of LCA is the life cycle inventory (LCI), which
consists of quantitative input and output flows associated with the management of a material (e.g.,
energy, material properties, and associated emissions and transformations). As part of the SHC
research program, the US EPA is developing tools and data that can be used by communities to
conduct LC As to better understand and manage their material and energy waste streams.
Although computer-based LCA tools have been developed to analyze waste materials and
processes, the overwhelming focus has been on municipal solid waste (MSW). Construction and
demolition debris (CDD) has largely been excluded from previously developed models, partly
owing to the perceived nature of CDD as chemically inert and lack of available data since CDD is
often regulated less stringently than MSW in the United States. CDD is generated from the
construction, renovation, repair, and demolition of structures such as residential and commercial
buildings, roads, and bridges. Wood, asphalt pavement, Portland cement concrete (PCC), masonry,
shingles, and drywall represent the dominant fractions of CDD materials. CDD materials can also
include lesser amounts of such materials as metals, plastics, insulation, cardboard, and soil. Trace
quantities of chemical products such as paints, solvents, and adhesives may also be present.
In 2018, 600 million US tons of CDD were generated in the United States (US EPA 2020).
Approximately 75 percent of this CDD was reused, the vast majority of which was used for
concrete aggregate for paving. In light of this volume and potential for environmental impacts and
benefits, the US EPA identified the collection of data and information regarding CDD management
practices as a priority area for LCA research.
Analysis of CDD materials typically occurs during the end-of-life (EOL) phase of an LCA. The
EOL phase in the CDD sector includes energy and material inputs from the unit processes
associated with managing the material as well as the associated emissions from that material. The
phase extends from management, through disposition, and after final disposition. For example, an
LCA pertaining to the recycling of gypsum drywall recovered from building demolition might
include the materials (e.g., steel, lubricants) and fuel (e.g., diesel) used by the equipment to grind
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and screen the drywall. The LCA should also include the emissions associated with the production
and use of these materials and energy, the emissions associated with the recycling process itself
(e.g., dust emissions), as well as the emissions associated with processing equipment and facility
decommission. The LCI data for this example would then include the accounting of energy and
raw material inputs and emissions to various environmental compartments (i.e., air, water, and
soil) over the life cycle of the process (i.e., construction, operation, and decommissioning).
1.2 Scope of Work and Objectives
The objective of this research was to assess the body of knowledge regarding CDD life-cycle data
in the US and compile LCIs using publicly available EOL data of selected CDD materials and
management processes. Peer-reviewed literature, government and private industry publications,
and LCA modeling tools were reviewed to identify current management practices and associated
LCI datasets. If LCI data were not available, process descriptions and documentation (e.g.,
emission categories, background data used to compile the dataset, geographic location, and time
period of the data) were reviewed to evaluate the completeness of the dataset. Primary sources
were used to the greatest extent possible in development of the datasets.
LCIs were not compiled for processes where significant data were lacking or where a CDD
management practice was determined to be uncommon. If publicly available information for a
given unit process was unavailable (e.g., liquids emissions from the disposal of CDD), proxy
information was reviewed and included as applicable (e.g., CDD leaching data). In cases where
U.S. datagaps existed, LCIsdeveloped for non-U.S. conditions were reviewed tobetterunderstand
the inputs and approaches used. Unit process data unavailable for the US were identified as data
gaps in need of further research.
A final objective of this research was to make the compiled LCI data available online in an open-
source format for public use. The LCIsdeveloped in this project were formatted in the US Federal
Life Cycle Inventory Unit Process Template (Cooper et al. 2015). Once in the template, the
datasets were imported into an openLCA database, and further revised and formatted with
metadata. The datasets are then published and made available in standard LCA dataformats on the
Federal LCA Commons atwww.lcacommons.gov.
1.2.1 CDD Materials in this Report
Based on US EPA (2020) estimates, approximately 600 million U.S. tons of CDD were generated
in the U.S. in 2018. Figure 1-1 illustrates the percent composition by material type of the majority
of these materials. Due to differences in estimation procedures, land clearing debris, fines,
cardboard, and plastics have been excluded. Still, the waste types in Figure 1-1 comprise the vast
majority of CDD waste.
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Drywall, 2.5%
AsphaltShingle
2.5%
Brickanc
Tile, 2.i
Figure 1-1. Composition by mass of CDD generated in the United States (US EPA 2020). Due to
differences in estimation procedures, land clearing debris, fines, cardboard, and plastics have been
excluded from the total mass.
This report presents a compilation of life cycle inventories (LCI) of the unit processes needed to
conduct an LCA of EOL management of the following major CDD constituents:
1. Asphalt Pavement
2. Asphalt Shingles
3. Gypsum Drywall
4. Wood Products
5. Land Clearing Debris (LCD)
6. Portland Cement Concrete (PCC)
7. Polyvinyl Chloride (PVC)
8. Fiberglass Insulation
9. Carpet
10. Clay Bricks
11. Copper Wire
12. Vinyl Composition Tile (VCT)
Table 1-1 lists the respective quantities of selected major CDD components that underwent EOL
management in 2018. The quantities of stockpiled material (i.e., for asphalt pavement and asphalt
shingles) were excluded from the table. For this report, temporary stockpiling is not considered
EOL management since it is part of a larger process. Table 1-2 lists the CDD and associated
industrial equipment LCI processes developed as a result of this study by North American Industry
Classification System (NAICS) category.
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Table 1-1. Generation by mass of CDD materials in the United States in 2018 (US EPA 2020),
Material
CDD generation (US tons)
Portland Cement Concrete
405
Asphalt Pavement
107
Wood Products
40.8
Asphalt Shingles
15.1
Gypsum Drywall
15.2
Land Clearing Debris
N/A
Total
583
Table 1-2. CDD processes by NAICS category.
Process
NAICS Category
Processing of reclaimed asphalt pavement; at processing
2379
Use of asphalt shingles; ground; as general fill; at fill site
2379
Use of concrete debris; as general fill; at fill site
2379
Use of reclaimed asphalt pavement; as general fill; at fill site
2379
Use of recycled concrete aggregate; as general fill; at fill site
2379
Dispensing of compressed natural gas
2389
Dispensing of diesel; at pump
2389
Dispensing of gasoline; at pump
2389
Dispensing of liquefied petroleum gas
2389
Operation of compressed natural gas equipment; industry average; > 19 kW and
< 56 kW
2389
Operation of compressed natural gas equipment; industry average; > 56 kW and
< 560 kW
2389
Operation of diesel equipment; industry average; < 19 kW
2389
Operation of diesel equipment; industry average; >19 kW and <56 kW
2389
Operation of diesel equipment; industry average; >56 kW and <560 kW
2389
Operation of diesel equipment; industry average; 560 kW and < 900 kW
2389
Operation of diesel equipment; industry average; > 900 kW
2389
Operation of gasoline equipment; 2-stroke; industry average; < 19 kW
2389
Operation of gasoline equipment; 4-stroke; industry average; < 19 kW
2389
Operation of gasoline equipment; industry average; > 19 kW and<56 kW
2389
Operation of gasoline equipment; industry average; >56 kW and <560 kW
2389
Operation of liquefied petroleum gas equipment; industry average; > 19 kW
and <56 kW
2389
Operation of liquefied petroleum gas equipment; industry average; > 56 kW
and <560 kW
2389
Processing of gypsum; milled; at drywall processing facility
3274
Processing of asphalt shingles; ground; at processing plant
3279
Waste transport; at local road; unpaved
4842
Stockpiling of material; at drop site
4931
Recovery of construction and demolition debris; segregated; at mixed CDD
processing facility
5621
Combustion of land clearing debris; at air curtain incineration
5622
Combustion of land clearing debris; at open burning
5622
Crushing of clay brick; as soil fill
5622
Crushing of clay brick; at crushing facility
5622
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Process
NAICS Category
Incineration of carpet padding; at MSW waste-to-energy facility
5622
Incineration of carpet; at MSW waste-to-energy facility
5622
Incineration of copper wire; 12-2; with ground; at MSW waste-to-energy
facility
5622
Incineration of polyvinyl chloride, PVC; at MSW waste-to-energy facility
5622
Incineration of vinyl composition tile, VCT incineration; at MSW waste-to-
energy facility
5622
Land application of CDD wood mulch; at application
5622
Land application of land clearing debris mulch; at application
5622
MSW landfilling of ACQ-treated wood; at MSW landfill
5622
MSW landfilling of CBA-treated wood; at MSW landfill
5622
MSW landfilling of DOT-treated wood; at MSW landfill
5622
MSW landfilling of carpet padding; at MSW landfill
5622
MSW landfilling of carpet; at MSW landfill
5622
MSW landfilling of clay brick; at MSW landfill
5622
MSW landfilling of copper wire; 12-2; with ground; at MSW landfill
5622
MSW landfilling of demolished concrete; at MSW landfill
5622
MSW landfilling of fiberglass insulation; at MSW landfill
5622
MSW landfilling of gypsum drywall; at MSW landfill
5622
MSW landfilling of land clearing debris; at MSW landfill
5622
MSW landfilling of polyvinyl chloride, PVC; at MSW landfill
5622
MSW landfilling of untreated wood; at MSW landfill
5622
MSW landfilling of vinyl composition tile, VCT; at MSW landfill
5622
MSW landfilling of wood ash; at MSW landfill; CCA 4 kg/m3
5622
MSW landfilling of wood ash; at MSW landfill; CCA 40 kg/m3
5622
MSW landfilling of wood ash; at MSW landfill; CCA 40 kg/m3
5622
MSW landfilling of wood ash; untreated wood; at MSW landfill
5622
Processing and cement kiln combustion of carpet; at cement kiln; 21 % carpet,
79% bituminous coal (mass)
5622
Processing of copper wire scrap; chopped, insulation removed; at processing
5622
Processing of land clearing debris; mulched; at processing facility
5622
Recycling of carpet padding; at rebond polyurethane foam production facility
5622
Recycling of carpet, residential; at recycling
5622
Reuse of clay brick; at construction
5622
Unlined CDD landfilling of ACQ-treated wood; at unlined CDD landfill
5622
Unlined CDD landfilling of CBA-treated wood; at unlined CDD landfill
5622
Unlined CDD landfilling of CCA-treatedwood; at unlined CDD landfill
5622
Unlined CDD landfilling of DOT-treated wood; at unlined CDD landfill
5622
Unlined CDD landfilling of carpet padding; at unlined CDD landfill
5622
Unlined CDD landfilling of carpet; at unlined CDD landfill
5622
Unlined CDD landfilling of clay brick; at unlined CDD landfill
5622
Unlined CDD landfilling of concrete; at unlined CDD landfill
5622
Unlined CDD landfilling of copper wire; 12-2; with ground; at unlined CDD
landfill
5622
Unlined CDD landfilling of fiberglass insulation; at unlined CDD landfill
5622
Unlined CDD landfilling of gypsum drywall; at unlined CDD landfill
5622
Unlined CDD landfilling of land clearing debris; at unlined CDD landfill
5622
Unlined CDD landfilling of polyvinyl chloride. PVC; at unlined CDD landfill
5622
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Process
NAICS Category
Unlined CDD landfilling of untreated wood; atunlinedCDD landfill
5622
Unlined CDD landfilling of vinyl composition tile, VCT; at unlined CDD
landfill
5622
Unlined CDD landfilling of wood ash; untreated wood; at unlined CDD
landfill
5622
Use of clav brick; as soil fill
5622
1.3 Report Organization
This report is organized into nine chapters. Following Chapter 1, Chapter 2 summarizes important
details of the materials investigated in this report and then presents LCI information on unit
processes common to the management of multiple CDD materials targeted in this report, including
landfilling, equipment use, transportation, and environmental emissions (e.g., leachate and gas).
Chapters 3 through 14 present material-specific details, including current EOL management
practices, an estimate of the quantity of the material managed at the EOL phase (if available), LCI
needs and sources reviewed, LCIs for the different EOL management processes, and data gaps
indicating additional research needs. The materials examined in these chapters include asphalt
pavement, asphalt shingles, gypsum drywall, wood, land clearing debris, Portland cement
concrete, polyvinyl chloride (PVC), fiberglass insulation, carpet, clay bricks, copper wire, and
vinyl composition tile. Finally, Chapter 15 summarizes the data gaps identified for the various
CDD materials and presents opportunities for future research.
1.4 References
Cooper, J., Kahn, E., and Ingwersen, W. (2015). US Federal LCA Commons Life Cycle Inventory
Unit Process Template (Version 1.0): USDA Agricultural Research Service. DOI:
10.15482/USDA.ADC/1178137
Pereira, E.G., daSilva, J.N., de Oliveira, J.L., and Machado, C.S. (2012). Sustainable Energy: A
Review of Gasification Technologies. Renewable and Sustainable Energy Reviews, 16,
4753-4762.
US EPA (2009). Sustainable Materials Management: The Road Ahead. EPA530-R-09-009, June
2009.
US EPA (2019). Waste Reduction Model (WARM) Tool: User's Guide. Version 15, May 2019.
Available: https://www.epa.gov/warm/documentation-chapters-greenhouse-gas-emission-
energy-and-economic-factors-used-waste-reduction.
US EPA (2020). Advancing Sustainable Materials Management: 2018 Fact Sheet. Office of
Resource Conservation and Recovery. EPA 530-F-20-009. Washington, DC.
USGS (1998). Materials Flow and Sustainability. USGSFact Sheet FS-068-98, June 1998.
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2 Approaches to Life-Cycle Data Management in the EOL Phase
2.1 The EOL Phase Boundary
The EOL phase begins when a material is discarded after the use phase. Many LCA studies
reported in the literature and from institutional efforts (e.g., LEED green building efforts) have
focused on building materials (among other CDD)from inception to the point of sale (i.e., cradle
to gate), or on the life cycle of the service phase (i.e., while in active use). The EOL phase is often
neglected in LCA studies of building materials due to the lack of LCIs representing management
practices in the US. The LCIs provided in this report begin to address this long-term data gap.
Upstream material manufacturing and associated processes (e.g., raw material extraction and
processing) were considered in LCI data collection when the material was recycled in a closed
loop (e.g., reclaimed asphalt pavement (RAP) used in the production of new asphalt pavement).
For materials not typically recycled through a closed loop, upstream processes were not further
explored since these processes would not impact environmental emissions within the EOL
boundary. Upstream processes were considered in a few open-loop recycling cases. One such
example is dimensional lumber recycled into particle board. Portland cement concrete recycled as
road-base aggregate can also include consideration of upstream processes associated with primary
wood and aggregate production.
It is important to consider all life-cycle phases of facilities and equipment used in EOL
management processes. While one may expect the bulk of the environmental burdens from a CDD
recycling facility to occur during the operational phase, the emissions associated with construction
and decommissioning/demolition of the facility must also be taken into consideration to
understand the full life-cycle impacts.
2.2 Life Cycle Inventory
An LCA is only as accurate and comprehensive as the underlying LCI data. LCI datasets should
include input and output flows (materials, energy, emissions, etc.) for all processes within the
system boundary. Section 2.2 presents terminology frequently used with the LCI datasets.
A fundamental aspect of a product or process LCI is the stream of substances, each of which may
be categorized as an elementary, product, or waste flow (see Table 2-1). Flows are quantified by
at least one property, such as volume, area, mass, time, or energy, and are contextualized as inputs
or outputs to the environment (elementary flows) or to the technosphere (as a product of one or
more other processes). Table 2-1 summarizes and provides examples of the flow types.
Once a material reaches the EOL phase, the environmental burdens associated with a particular
management process are quantified by evaluating the inflows and outflows of energy, materials,
and process emissions. Broadly speaking, emissions may be fuel-related (e.g., emissions from fuel
extraction, processing, transport, combustion) or non-fuel-related (e.g. dust, leachate). Emissions
may be generated during facility construction, operation, or decommissioning, and may be the
result of production and use of materials that are ancillary to the process (i.e., operation and
maintenance consumables) as presented in Table 2-2. Table 2-2 lists and defines the terms used to
describe the categories of emissions and materials included in the LCI datasets.
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Table 2-1. Standard LCAflow types, descriptions of flow types, and examples of product systems.
Flow Type
Description
Product System Example
Elementary
Material or energy entering or exiting
the system without prior or subsequent
treatment, respectively.
Input: Water
Output: Particulate matter emissions from
heavy equipment operation
Product
The usable output of a system, or in the
case of intermediate products, the
desired input to another process.
Input: Gasoline for combustion in heavy
equipment
Output: Softwood sawed and planed lumber
Waste
Material leaving the product system
Residuals from CDD processing
Table 2-2. Terms commonly used in EOL CDD management, their respective descriptions, and
examples of each term.
Term
Description
Example
Pre-combustion
Emissions
All emissions released as a result of
fuel or electricity production
Air emissions from crude oil
extraction, transport and processing
for diesel or gasoline fuel
production
Manufacturing/
Construction
Emissions
Emissions released during the
manufacture of a product or piece of
equipment/construction of a facility
Dust emissions from land clearing
activities for concrete plant
construction.
Non-fuel Emissions
Emissions which are not associated
with fuel combustion. These
emissions are those emitted during a
processing step, not as a direct result
of energy use or fuel consumption
Emissions from stormwater run-off
or landfill leachate to surface or
groundwater
Operation and
Maintenance
(O&M)
Consumables
Those materials which are used by a
process but are not incorporated into
the product of interest.
Lubricating oils for process
equipment
Decommissioning/
Demolition
Emissions
Emissions released as a result of
removing or disposing of process
facilities or equipment.
Particulates released as part of
material recovery facility demolition
Primary Material
Material produced from virgin
resources
Asphalt produced from petroleum
refining
Recycled Material
Materials produced from processing
of discarded products
Aggregate produced from crushing
discarded Portland cement concrete
One of the major emissions considered in LCA and in this life cycle inventory is carbon dioxide,
as it is one of the major greenhouse gases (GHGs) that contributes to climate change. Carbon
dioxide emissions can be considered fossil or biogenic. Biogenic carbon dioxide is released due to
transformation of biologically active carbon (e.g., biological decomposition or combustion of
biomass), whereas fossil carbon dioxide is usually released from the combustion of carbon
compounds from a fossil origin (e.g., diesel fuel or plastics). Many models and datasets do not
consider biogenic carbon dioxide emissions for quantification of emissions associated with human
activity since biogenic carbon dioxide emissions would occur regardless of human activities (RTI
International 2003, US EPA 2012a). However, any anthropogenic increase in greenhouse gas
emissions from biologically active carbon above and beyond what would have naturally occurred
is typically considered in LCA models. For example, the landfill disposal of biomass results in
methane emissions which have a GHG potency over 25 times greater than those of carbon dioxide.
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Because it is unlikely that these methane emissions would have naturally occurred without human
activity, their impact on the environment would typically be included in LC A models.
LCI data on emissions to water vary based on assumed environmental controls (e.g., wastewater
treatment) and the associated pollutant removal efficiency, which may be quantified on a
constituent-specific basis. For example, the US-specific Municipal Solid Waste-Decision Support
Tool (MSW-DST) assumes removal rates of 97% and 21.6% for aqueous biochemical oxygen
demand (BOD)and phosphate, respectively (NC State and ERG 2011). Leachate-related emissions
for landfill disposal of materials are an example of waterborne emissions considered for EOL
material management. LCI data are often available for chemicals in leachate that are legally
required to be monitored on a routine basis (e.g., BOD, arsenic). Ash from material combustion,
wastewater treatment plant sludge, and solids collected in air-pollution-control devices are some
examples of solid wastes that should be accounted for in the EOL management of CDD materials.
Environmental emissions from these materials and processes are not yet typically included in LCA
models and thus present an opportunity for research (ASMI 2014).
2.3 Organization of LCI Datasets
USDAandUS EPA collaborated to create a Microsoft Excel-based template called theUS Federal
LCA Commons Life Cycle Inventory Unit Process Template (Cooper et al. 2015) for creating and
documenting life cycle inventory data, which gives dataset developers the ability to document the
sources, data, and calculations behind each compiled life cycle inventory. The template file was
designed to allow the integration of the datasets into OpenLCA, an open-source LCA program.
All LCIs developed in this report were prepared in the template. The flows included in the LCI
datasets are quantified in terms of a reference flow, typically in the amount of 1 of the appropriate
unit (e.g., kg, metric ton). All inputs and outputs included in an LCI are scaled with respect to the
reference flow. The reference flow is always included under the output flows, but OpenLCA users
can switch reference flows for processes that produce multiple products of interest.
All unit process LCIs were assessed for data quality as described in Edelen and Ingwersen (2016).
Additional information on the organization of the datasets can be found in the completed template
files, which are available on the EPA's Environmental Dataset Gateway (https://edg.epa.gov).
2.4 Common Technosphere Inputs
2.4.1 Transportation
Emissions associated with transportation are often normalized with respect to the amount of
material multiplied by the shipment distance, typically expressed as ton-miles (or tonne-km). Ton-
miles provide the best single measure of the overall demand for freight transportation services;
this measure in turn reflects the overall level of industrial activity in the economy (Dennis 2005).
Ton-miles have historically been used by the U.S. Census Bureau (USCB) to analyze the
magnitude and modes of freight transportation at a national level for different commodities (2007
Commodity Flow Survey by USCB (2010)).
For quantifying the emissions associated with the general transport of materials discussed in this
report, the process "Truck transport, class 8, heavy heavy-duty (HHD), diesel, short-haul, load
factor 0.75" was used to simulate the one-way transport of materials between individual EOL
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material management locations for distances less than or equal to 322 km, or 200 miles, where 200
miles is the upper boundary for short-haul distances in the United States Life Cycle Inventory (US
LCI) database (US LCI 2012). A similar process was used to model long-haul transport for
distances greater than 322 km. A default distance of 20 km was assumed when the actual distance
was unknown.
2.4.2 Equipment Use and Fuel Combustion
Similar to electricity use, several CDD management processes require the use of heavy equipment
for a variety of tasks (e.g., material loading, unloading, sorting, on-site transport). The US EPA
(2014) MOVES database was used to quantify equipment use as well as allocate diesel
consumption to specific heavy equipment types per unit reference flow. The MOVES database
includes information on the operational lifespans of different pieces of equipment based on fuel
consumption over the equipment's lifetime. The database also includes specific emissions
information based on heavy equipment category (e.g., diesel excavators, 4-stroke rubber tire
loaders) and engine power rating range (e.g., 50 < hp < 75, 75 < hp < 100). Total facility/process
volumetric diesel consumption, equipment-specific volumetric diesel consumption rates (from
equipment manufacturer or vend or literature), and a diesel density of 0.845 kg/liter (i.e., No. 2 fuel
oil at 7.05 lb/gal as reported in US EPA (1998)) were used to estimate equipment use per unit
reference flow. If the equipment-specific diesel consumption for equipment used at EOL
management facilities or processes were not found in the literature, facility-wide diesel
consumption was allocated to specific types of heavy equipment by one of the following two
methods:
1. Total facility or process-specific diesel consumption was allocated to specific equipment
types by using manufacturer/vendor-listed hourly diesel consumption rates (assuming
simultaneous operation) for the case where all pieces of equipment would potentially be
used simultaneously.
2. Total diesel consumption was allocated to specific equipment types based on the
nationwide diesel consumption rates reported for equipment types as listed in the US EPA
(2014) MOVES database.
Due to an absence of readily available data, operation and maintenance consumables (e.g., oil use,
tire/track use, filters, drive belts) were not accounted for in the datasets.
2.4.3 Water Consumption
2.4.3.1 Airborne Dust Suppression
Water consumption for the wet suppression control of particulate emissions from a variety of CDD
material processing operations was estimated using the midpoint of a range of water consumption
values provided for an industry-popular dust emissions control device. For water-balance
purposes, it was assumed that all water sprayed into the air for dust suppression purposes is
released to the atmosphere as water vapor.
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2.4.3.2 Unpaved Road Dust Control
Particulate emissions from unpaved roads are accounted for as detailed in the Particulate Emissions
section. Therefore, while a zero default value is included in each of the datasets, additional water
consumption associated with unpaved road dust emissions control was parameterized with a range
including a maximum value (i.e., 0.125 gallons per square yard applied every 20 minutes) for near
100% dust emissions control efficiency, based on US EPA (1985). For water balance purposes,
and considering the hot, dry conditions which typically warrant the use of water for unpaved road
dust control, it was assumed that all water applied to unpaved road surfaces evaporates and is
released to the atmosphere as water vapor.
2.4.4 Stormwater Management
CDD EOL management facilities commonly collect and impound rainfall runoff onsite in
stormwater ponds. A fraction of the impounded stormwater is released to the atmosphere as water
vapor, and a fraction infiltrates into the ground. The volumetric fraction of stormwater released to
each part of the environment is dependent on a number of site-specific factors (climate, soil
permeability, etc.). Similarly, stormwater quality is dependent on site-specific factors such as the
types, sizes, and geometries of uncovered material stockpiles at the site. For water balance
accounting purposes, the datasets include a flow to surface water (based on an average nationwide
annual precipitation volume derived from NOAA2016) leaving the system boundary, since values
for pond water loss through groundwater infiltration and evaporation are unknown and stormwater
quality data were unavailable.
2.4.5 Aggregates and Soil
2.4.5.1 Primary Aggregate Transport
Several CDD materials may contain primary aggregates (asphalt pavement, asphalt shingles, PCC)
or may be used as a substitute for the production and transport of primary aggregates as an EOL
management option. As a result, it is necessary to develop an estimate of the average nationwide
modal (i.e., truck, rail, or water) distance that primary aggregates typically travel from production
to end-use. These estimates are summarized in Table 2-3.
The USCB Commodity Flow Survey provides the total amount, distance-amount, and average
miles per shipment for gravel and crushed stone in the US. However, the distances and quantities
provided are not organized by transport path or end users (e.g., hot mix asphalt (HMA) plants or
PCC plants). Due to lack of end-user data, the average distances for the commodity "Gravel and
crushed stone" presented in USCB (2010) are assumed to represent the average US-wide aggregate
transport distance from production sites to multiple primary (e.g., HMA Plants, PCC plants) end
uses for various transport modes. The average distance per mode was calculated by dividing the
total ton-mile data for the mode by the total amount (tons) for all modes combined. Only single-
mode transport data provided by the survey were used in the analyses presented in this report.
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Table 2-3. Estimated transport distances for aggregate materials (developed from USCB 2010).
Material
transport
Source and
representative
commodity
Total amount
transported by
single mode
(million tons)
Mode
Ton-miles
(millions)
Average
transport
distance
(miles)
Average
transport
distance
(km)
Aggregate
transport:
mine to
paving
mix plant
USCB (2010)
gravel and
crushed stone
1,930
Truck
80,600
41.7
67.1
Rail
23,400
12.1
19.5
Water
15,500
8.01
12.9
2.4.6 Particulate Emissions
2.4.6.1 Material Loading and Unloading
US EPA (2006b) AP-42 (Chapter 13, Section 2.4) was used to develop a datasetto estimate the
particulate matter emissions from aggregate loading and unloading operations. This process
dataset simulates the particulate matter emissions resulting from drop operations associated with
material handling and storage piles. A process default of two drops was assumed. The dataset user
should specify the number of material drops which occur at the site, and if known, can specify
wind speed values and a material-specific moisture content from the ranges presented in US EPA
(2006b). Average and median values of moisture content and wind speed were selected from the
ranges for each of these parameters as provided in US EPA (2006b). In addition to these modeling
parameters, it should be noted that the Occupational Safety and Health Administration (OSHA)
mandates that workers in the construction industry be exposed to concentrations of airborne
crystalline silica no greater than 50 |ig/m3 per 8-hour workday (OSHA 2016).
2.4.6.2 Transport over Unpaved Roads
AP-42 (US EPA 2006a) was also used to develop a dataset which estimates the particulate matter
emissions from vehicular movement on unpaved road surfaces. The dataset user should specify
the length of the unpaved road, a representative silt content using the range of silt contents provided
for different materials listed in US EPA (2006a), the loaded weight, and the capacity of the haul
vehicle. A default assumption was made for an unpaved road round-trip distance of 2 km (1 km
each way). The average gross vehicle weight and capacities of three articulated truck models were
selected for estimating unpaved road emissions, as listed in CAT (2006). The median silt content
of the range presented US EPA (2006a) was selected for the default silt content value included in
this process.
2.5 Mixed CDD Processing
The CDD material recovery facility (MRF) process dataset is an agglomeration of data from a
nationwide survey of CDD MRFs conducted by the Construction and Demolition Recycling
Association (CDRA) and includes detailed infrastructure and equipment inventories from an in-
depth study of two large-scale mixed CDD MRFs located in the southeastern US. Sixteen
respondents to the CDRA (2015) provided information on mixed CDD MRF throughput, diversion
rates, facility size, electricity consumption, and fuel consumption. Figure 2-1 illustrates the
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composition of materials typically recovered or discharged to CDD MRFs. Figure 2-2 illustrates
the flows included in the dataset for this process.
Recovered Screen
Residual 20%
Fuel Product
AsphaltShingles
Cardboard
Bricks and
Metal
Green Waste
Other
Concrete 33%
in the Figure 2-1. Typical composition by mass of CDD materials processed by material recovery
facilities United States (CDRA2015).
T
Water
t <
Surface Solid
water waste
-Mixed CDD; at generation
Transport, single unit truck;
short haul; diesel powered
- - -Concrete; ready-mix, truck mix; at plant- -
Hot mix asphalt, HMA;
at production facility
Surface dried lumber;
at planer mill
Gypsum wallboard product; regular;
0.5 inch, 12.7 mm
-Heavy equipment -
Equipment support infrastructure and
other equipment; unspecified
Recovery of
construction and
demolition debris,
segregated; at
mixed CDD
processing facility
Land use Electricity Water Diesel
Concrete; demolished; at mixed CDD MRF *
--Fines; recovered screen material; at mixed CDD MRF-h
Fuel product; unspecified; at mixed CDD MRF 1
Green waste; unspecified; at mixed CDD MRF 1
Metal; unspecified scrap; at mixed CDD MRF -»
Bricks and masonry; at mixed CDD MRF 1
Cardboard; unspecified scrap; at mixed CDD MRF—i
Asphalt shingles; tear-off; at mixed CDD MRF -»
--Gypsum drywall; demolished; at mixed CDD MRF »
-Glass; broken; at mixed CDD MRF *
-Reclaimed asphalt pavement; at mixed CDD MRF— >
-Top soil; at mixed CDD MRF 1
Plastic; unspecified scrap; at mixed CDD MRF
Carpet or carpet padding; unspecified scrap;
at mixed CDD MRF
Wood; unspecified scrap; at mixed CDD MRF
Paper; unspecified scrap; at mixed CDD MRF
Legend
Elementary Flow
Technosphere Flow_
Hi N°t Included
Included
System Boundary
Figure 2-2. Unit process flow diagram for a typical mixed CDD material recovery facility (MRF),
also known as a CDD processing plant.
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Diesel consumption on a per-kilogram-material-processed basis was estimated as 0.00171 liters
by dividing the total amount of diesel used by the CDD processing facilities by the total mass of
material received by all the mixed CDD processing facilities that responded to the CDRA (2015)
survey. The following equation was used for calculations:
1"= i Dj
F=^=1Mt
Where,
Df = diesel consumption per kilogram of CDD processed (L/kg)
Di = total diesel consumption at the ith CDD processing facility for 2013 (L)
Mi = total mass of material received at the ith CDD processing facility for 2013 (kg)
Electricity data were calculated in a similar fashion as 0.00308 kWh per kilogram of material
processed. Equipment inventory and water consumption data were developed from a detailed
inspection of two large-scale mixed CDD MRFS in the southeastern US (IWCS 2015). Additional
flows represent average values taken from 16 individual mixed CDD MRFs from across the US.
Mixed CDD MRFs typically utilize automated and manual processes to separate a mixed incoming
CDD stream into individual material categories and then market the segregated materials to
different end users. Some CDD MRFs include additional processing steps (e.g., concrete crushing,
woody debris grinding). This datasetis intended to represent the average nationwide CDDMRF
employing average recovery technologies and so additional processing steps have been excluded.
A primary assumption of this dataset is that all incoming CDD loads arrive at theMRF as mixed
material. In actual practice, some loads already represent individual materials and are a product of
specific demolition activities, such as road demolition, land clearing, and roofing projects. These
loads may not (and commonly do not) receive the same degree of processing as mixed loads.
Because the CDRA-reported data do not provide details on the fraction of loads arriving in mixed
vs. material-specific loads, this dataset presents all loads arriving as mixed CDD. Another primary
assumption is that the equipment inventory compiled from a detailed inspection of two large-scale
mixed CDD MRFs (IWCS 2015) is reflective of a representative inventory of the average
nationwide facility.
This dataset assumes that particulate matter (PM) emissions resulting from paved surfaces,
processing, and stockpile management are negligible. The extent of wet suppression techniques
utilized at the sites suggests a very high degree of dust control.
Equipment operation/utilization and materials used in facility construction were estimated for two
mixed CDD MRFs in the southeastern US. Construction materials were estimated by taking onsite
dimensional measurement of facility infrastructure. The total amount of each of these materials
was then quantified on a mass basis using typical density values for each material type. The total
tonnage of materials accepted at each of the two facilities was found for the years 2013-2015. An
average annual tonnage was then calculated using these tonnage amounts. Service lives for the
different materials were estimated using values presented in Townsend and Cochran (2010). The
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total tonnage of each material was then divided by its service life to approximate the quantity of
that material necessary each year. These quantities were then divided by the average annual facility
throughput to approximate the mass of each construction material needed per mass of throughput.
The equipment manufacturer-listed horsepower ratings of each piece of equipment inventoried at
the two sites were used in conjunction with the US EPA (2014) Motor Vehicle Emissions
Simulator (MOVES) database to quantify equipment usage according to fuel consumption. The
US EPA (2014) MOVES database was also used to estimate equipment utilization based on the
nationwide average total fuel consumption values listed for specific equipment types.
Utilization of electric-powered equipment (e.g., trommels, finger screens), equipment-support
infrastructure, staircases and platforms, and steel wall partitions was estimated on a mass basis,
assuming a 20-year functional/serviceable life. The total mass of each equipment/equipment -
support infrastructure type was divided by the average annual throughput reported by the facility
from 2013-2015 (i.e., quantities reported to the state's permitting authority).
Land occupation was estimated by dividing the total site areas reported by all 16 respondents and
dividing this value by the total throughput reported by the corresponding throughput, assuming a
typical facility lifespan of 20 years. A 12' x 56' standard office trailer was assumed for use for
administrative purposes at mixed CDD MRFs.
Mixed CDD MRFs commonly use onsite wells as a source of water for wet dust suppression;
therefore, water consumption is typically not tracked at these facilities. Total facility water
consumption was estimated from three specific uses: water use by misting cannons, water use by
the sprinkler system located at the top of the perimeter wall, and water use for paved surface dust
control. Manufacturer specifications of misting cannon equipment present at one of the two sites
and operating time period assumptions were used to estimate water consumption for these devices.
The number and style of sprinkler heads, sprinkler nozzle specifications, and the same operating
time period assumptions were used to estimate sprinkler water consumption. The total area of
paved surfaces, and a water application rate and frequency to achieve nearly 70% dust control
(with brooming), as described in US EPA (1985), was used to estimate the total amount of water
necessary for paved surface dust suppression.
A water balance was conducted assuming that all water consumption for the purposes of dust
control is returned to the atmosphere as water vapor, whereas all water resulting from rainfall onto
paved surfaces is collected and managed as stormwater. Average nationwide precipitation was
estimated from NOAA (2016). The average paved fraction of the total site area was generated via
Google Earth aerial imagery, given the four CDR A-respondent sites. This fraction was multiplied
by the land occupation value and the average nationwide precipitation. Due to a lack of data
availability, specific emissions associated with stormwater discharge to surface water from these
facilities have not been included in this dataset. Therefore, the surface water emissions flow is only
considered partially accounted for in this dataset.
Several of the construction and equipment manufacturing input flows are approximated using
surrogates due to flow availability in the existing US LCI database. Steel consumption uses "steel,
billets, at plant" to approximate the environmental burdens associated with structural steel
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production. "Surfaced dried lumber, at planer mill, US SE" was used to represent all onsite
dimensional lumber use (e.g., fence posts and slats, 2"x4"s).
No nationwide average transport distance was found for mixed CDD movement between the point
of generation and a mixed CDD processing facility; therefore, a distance of 20 km was assumed.
Diesel fuel consumption occurs at CDD processing facilities from heavy equipment used to
unload, sort, transport, and load waste at facility tipping floors. The Doka (2009) diesel fuel
consumption estimate for fuel usage was also based on literature-derived values corresponding to
the fuel demand of a skid-steer loader; values ranged from 2.95 - 5.9 MJ/m3 of sorted building
waste with an average value of 4.4 MJ/m3.
2.6 Landfilling
2.6.1 Background
Although recycling rates have risen steadily based on increased consumer awareness and grants
and other incentives to encourage recycling, landfilling is still the predominant EOL management
method for discarded materials in the US. While CDD disposal typically occurs at unlined CDD
landfills, a small percentage of CDD materials are still disposed of at MSW landfills.
Material disposal LCIs should include the materials, energy inputs and emissions associated with
landfill construction, waste placement and compaction, and closure and post-closure-care activities
along with the long-term liquids and gaseous emissions pertaining to biogeochemical
decomposition of deposited materials. The domain of disposal-related inputs and outputs
considered varies significantly among different LCA models and LCI databases. For example, US
EPA's WARM only considers GHG emission from materials transport, waste placement and
compaction activities. Another tool, the Waste and Resources Assessment Tool for the
Environment (WRATE), includes materials and other inputs and emissions associated with landfill
construction, operation and closure. Several other LCA models account for a subset of these
emissions. This section describes datasets which account for landfill construction, operation,
closure and post-closure activities.
2.6.2 Transportation
Transport mode, distance traveled and the associated waste mass for each mode are the primary
data pieces used to specify transport flows for the landfill process LCI datasets. The Waste
Business Journal (WBJ) maintains a database of over 8,000 solid waste facilities in the US. The
database includes two categories for landfills: "demolition landfill" and "landfill." The database
also includes types of waste accepted. All "demolition landfills" are listed as receiving CDD, and
all "landfills" are listed as receiving MSW. The categories "demolition landfill" and "landfill" are
therefore assumed to correspond to "CDD landfill" and "MSW landfill", respectively. The WJB
database also includes facility waste acceptance rates, and the transportation modes/access types
by which waste may be received at the facility. The transportation modes for each CDD and MSW
landfill in the database include one or a combination of these modes: highway, rail, and waterway.
It should be noted that every CDD Landfill and MSW landfill included in the WBJ database has
highway access. However, as expected, none of these facilities are exclusively accessed by rail or
waterway.
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Waste acceptance rates specific to landfills with specific access types were summed from WBJ
(2014) to find the access-type specific fraction of waste received at each landfill type (i.e., MSW
landfill, CDD landfill). In the absence of nationwide, mode-specific CDD transport data, several
engineering, best-judgement assumptions were made regarding the overall fraction of CDD
transported to landfills with different modal access types: It was assumed that a facility with
highway access and rail access would receive 90% of its waste via rail. It was assumed that a
facility with highway and waterway access would receive 90% of its waste by waterway. It was
assumed that a facility with highway, rail, and waterway access would receive its waste 90% via
rail, 5% via highway, and 5% via waterway. Table 2-4 shows the estimated waste fraction
transported by each of three modes for CDD and MSW landfills.
Table 2-4. Percentages of CDD materials transported to CDD landfills and MSW landfills by
Transport mode
CDD Landfill
MSW Landfill
Highway
91.4%
74.6%
Rail
8.63%
25.4%
Waterway
0.01%
0.04%
Due to the insignificant waste fraction transported by waterway, waterway transport was not
included in the landfill LCIs dataset. The material fraction associated with waterway transport was
neglected since it falls below the precision of the data included in the LCI dataset.
Publicly available records for 3 operating CDD landfills in Florida (Costal Landfill Disposal of
Florida, LLC, 2010; Environmental Resources, LLC, 2010; and Orange Blossom Disposal
Facility, 2010) and 3 operating MSW landfills in Ohio (Sunny Farms Landfill, LLC, 2016; Apex
Environmental, LLC, 2016; and American Landfill, LLC, 2016) were used to estimate waste
transport distance to MSW and/or CDD landfill sites. The 3 Ohio MSW landfill sites also accepted
CDD. Because records detailing transport information for rail-transported CDD to a CDD landfill
sites were not found, CDD transport data reported by Ohio MSW landfills was assumed to reflect
the transport and modes of CDD transported to a CDD landfill. The records included information
on the tonnage of each waste type (e.g., MSW, CDD) county of origin (i.e., county, state), and
mode of transport (i.e., rail, truck). Transport distances, however, are not provided, so they were
approximated using a GIS analysis tool. For the county where the landfill was located, the distance
from the center of the county to a visually representative edge of the county was used to estimate
the haul distance. For the other counties, the straight-line distance between the landfill site and the
county centroid was used to estimate the haul distance. Because of the straight-line approximation,
it should be noted that the distances represent underestimates of the actual distances. Each
shipment was then classified as short (<322 kilometers) or long haul (>322 kilometers); this
distance is the cutoff between short-haul and long-haul transport.
The total tonnage shipped through each transportation mode (short-haul truck, long haul truck, and
rail) was then calculated for both MSW and CDD landfills.
The average mode-specific shipment distance calculation is shown in the following equation:
„ n. , l?(TonsMixDistMi)
Av9Dlst« = TonsTotalM
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Where,
AvgDistM = The average mode-specific kilometers traveled by waste arriving at a specific
landfill type (i.e., MSW, CDD)
M = the specific mode of transport (i.e., long-haul, short-haul, rail)
N = the total number of counties of origin
TonsMi = the tonnage of the ith waste origin hauled by a specific transport mode
DistMi = the kilometer distance of the ith waste origin hauled by a specific transport mode
TonsTotalM = the total tonnage transported by a specific transport mode
Table 2-5 summarizes the results of this calculation:
Table 2-5. Mass percentages and transport distances of CDD materials to MSW and CDD landfills,
Mode
MSW Landfill
CDD Landfill
Mass
Percent
Distance
(km)
Mass
Percent
Distance
(km)
Short-Haul
72.4%
26.5
91.3%
97.1
Long-Haul
2.16%
619
0.0136%
420
Rail
25.4%
469
8.63%
807
Due to the very low fraction of waste received at CDD landfills through long-haul truck transport,
this transport mode was considered negligible for a CDD landfill.
2.6.3 Equipment Use and Equipment-Specific Fuel Consumption
EREF (1999) presented equipment-specific fuel consumption for the operation of landfill sites that
do and do not have a daily cover soil requirement. Heavy equipment use and associated diesel
consumption at sites that have a daily cover soil requirement were assumed to be reflective of
operations at MSW landfills, while operation at sites which do not have a daily cover requirement
was assumed to be reflective of CDD landfills, which commonly have a weekly or less-frequent
cover frequency requirement. EREF (1999) provides equipment fractional diesel consumption for
general equipment categories (e.g., scraper, loader). Landfill operation plans from three Florida
MSW landfills (i.e., New River Regional Landfill, Orange County Landfill, and Bay County
Steelfield Road Landfill) and three CDD landfills (i.e., Costal Landfill Disposal of Florida, LLC;
Lake Environmental Resources LLC, Florida; Lordstown Landfill, Ohio) were reviewed and
detailed equipment inventories (including equipment model numbers) for the equipment used at
these sites were compiled. It is assumed that similar equipment models are used for both CDD and
MSW landfills. The horsepower rating of each piece of equipment was obtained from equipment
manufacturer or vendor websites. In the event the specifications for the equipment model listed in
the operation plans were not available, the specification for a similar/the latest available model
was used. The horsepower rating for several equipment types (e.g., bulldozers) varied across the
sites. For these equipment types, the average horsepower was calculated. It should be noted that
two of the equipment types, drum roller and grader, as listed in EREF (1999) were not found in
the referenced operation plans; therefore, specifications of a drum roller and a grader used in
landfill operations were selected from the Caterpillar handbook (Caterpillar 2006).
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EREF (1999) presents equipment diesel consumption for all MSW landfill phases; however, EREF
(1999) only provides CDD operations phase diesel consumption information (i.e., for landfill sites
that do not have a daily cover soil requirement). Because of the relatively common practice of
using old quarry sites as unlined CDD landfills, it was assumed that the environmental burdens
associated with unlined CDD landfill construction were negligible.
Several states require a soil closure cap for CDD landfills (e.g., Florida, Virginia). A 0.61-m thick
soil closure cap was assumed forthis dataset.
2.6.4 Construction Phase
Landfill construction requires a variety of material and energy inputs. Landfills are built as
containment systems with the goal of minimizing direct (e.g., waste-related) emissions to the
surrounding environment. The bottom liner of landfill cells is generally constructed via a
combination of low-permeability (typically <10"7 cm/sec) compacted earthen material and
geosynthetic materials (typically 60-mil-thick HDPE). The purpose of the bottom liner is to
contain and collect leachate and remove it out of the landfill cell, generally using porous drainage
media (e.g., gravel), piping, and mechanical pumps to prevent build-up of liquid on the liner
system. Emissions resulting from landfill cell construction occur during liner material
manufacturing, transport and use of heavy equipment for on-site soil excavation and liner
installation.
The fuel consumption and material resources required for landfill construction would depend on
the level of environmental controls installed at the site. Composite liner systems are frequently
installed at MSW landfills, which often include multiple layers of different geosynthetic materials
for both leachate collection and leak detection purposes. Geosynthetics may be placed in contact
with underlying low-permeability earthen material, which commonly will require compaction
prior to the placement of the geosynthetics. In addition to the energy consumption of equipment
needed to transport, place, compact, and weld liner components, the environmental burdens
resulting from geosynthetic manufacturing should be considered.
Ecobalance developed an LCI for US MSW landfills for the Environmental Research and
Education Foundation in 1999 based on a survey of more than 100 MSW landfills across the US
and Europe. Part of this survey included compiling the average characteristics and fuel demand
necessary for the construction of the liner system and other support infrastructure. Ecobalance
(1999) provides the average thickness and the density of each of the liner components and, based
on survey results, presents the average airspace use per MSW landfill footprint area. Ecobalance
data were used as a primary input for developing landfill construction LCIs forMSW-DST and
EASETECH.
Literature-reported density ranges of CDD materials as presented by CCG (2006), Huang et al.
(1992) and Angus Hankin et al (1995) were used in this analysis to estimate the MSW landfill
footprint required per mass of CDD material accepted at MSW landfills. This allowed an estimate
of the mass of individual materials needed per mass of CDD material accepted at an average MSW
landfill. It should be noted that except for land clearing debris (where a compacted value was
reported) all the literature-reported densities represent uncompacted material; therefore, the
environmental burdens associated with land filling these CDD materials may be overestimated.
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Ecobalance (1999) also developed an estimate of the amount of diesel fuel required forMSW
landfill construction on a mass-acceptance basis. The report details the average transport distances
for each construction material; these distances were multiplied by their respective masses and the
resulting mass-distance amounts were organized and summed by whether materials were
transported more or less than 322 km. Material transport greater than 322 kilometers was modeled
as long-haul transport and all other transport was modeled as short-haul.
EREF (1999) reported equipment used and the associated fuel consumption for MSW landfills for
site construction, operation, and closure. The equipment inventory presented in EREF (1999) was
classified according to the detailed equipment power classifications used in the US EPA (2014)
Motor Vehicle Emission Simulator (MOVES) program. Detailed equipment classifications were
made using several landfill equipment inventories posted on publicly-accessible state
environmental agency solid waste department websites from Florida and Ohio. Total fuel
(quantified as diesel) consumption by each piece of equipment was then estimated based on the
equipment-specific fuel consumption percentages presented in EREF (1999).
CDD landfills are typically unlined since no federal requirements for liners andleachate collection
systems currently exist for CDD landfills, though some states require liner construction for CDD
landfills. Data that detail the energy and material burdens associated with the construction of an
unlined CDD landfill were not found. Therefore, MSW landfill construction LCI data were used
as a proxy for CDD landfill data. No construction LCI are developed for unlined CDD landfill
construction due to lack of data.
2.6.5 Operations Phase
Waste placement and compaction begins following construction of the liner/leachate collection
system. Waste is filled in designated cells and lifts in a sequenced filling plan. Landfill operations
generally include placing and compacting waste materials as well as periodically applying cover
soil to the exposed waste surface. MSW landfills will commonly install a daily cover over the
active waste face while CDD sites may install a weekly cover or no cover at all. Besides the diesel
energy necessary to place and compact incoming waste, electricity is necessary to power numerous
site facilities and buildings (e.g., scale house, workshop, offices, lighting). Cover soil is assumed
to represent 10% of the volume of waste material placed at MSW landfills. However, cover soil is
assumed to only represent 1.43% of the volume of waste material placed at CDD landfills (i.e., it
is only placed once on a weekly basis or l/7th of the daily cover amount). Literature-reported CDD
material densities were used to estimate the corresponding mass of cover soil necessary for
material placement at either a CDD or MSW landfill site, assuming a soil density of 1,330 kg/m3
(USD A 2013). These densities were necessary in order to translate cover soil requirements from a
volumetric to a gravimetric basis.
The volume of surface water impounded at an MSW landfill site was estimated based on
precipitation amount and leachate. The quantity of precipitation estimated for a US MSW landfill
sites in EREF (1999) was 89 cm per year. Assuming a landfill operational timeframe of 40 years
and assuming 228,000 m3 of waste-occupied airspace per hectare as calculated from EREF (1999),
the total volume of precipitation for each m3 of landfilled waste is approximately 1570 1/m3 and
accounting for the total volume of leachate generated (34.0 1/m3), a total of 1530 liters of surface
water is generated per cubic meter of waste. Surface water impounded at CDD landfills was
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calculated in a similar fashion but assumed a landfill operating life of 20 years; the total quantity
of surface water generated is half that of the amount generated at MSW landfills.
EREF (1999) presented individual equipment and the associated fuel consumption for
construction, operation, and closure and post-closure care phase of MSW landfills, but only the
operations phase for landfills without daily cover. The data presented for landfills without a daily
cover requirement were used to estimate operations phase fuel consumption for CDD landfills.
EREF (1999) provides fractional diesel consumption data for general equipment categories (e.g.,
scraper, loader). Landfill operation plans from three Florida MSW landfills (New River Regional
Landfill, Orange County Landfill, and Bay County Steel Road Landfill) and three CDD landfills
(Costal Landfill Disposal of Florida, LLC; Lake Environmental Resources LLC, Florida;
Lordstown Landfill, Ohio) were reviewed. Detailed site equipment inventories, including
equipment model numbers, were compiled. It is assumed that similar equipment models would be
used across all phases of an MSW landfill.
Once the detailed equipment lists were compiled, the horsepower rating of each piece of equipment
was obtained from equipment manufacturer or vendor websites. Based on information provided
by several MSW landfill owners in Florida, water is primarily used for dust control purposes at
MSW landfills (IWCS 2016). As described in US EPA (1985), near 100 percent dust control can
be achieved with an application of 0.125 gallon per square yard of road every 20 minutes.
Assuming 1 km of road with a 6-m width, 150 10-hour operating days per year, an MSW landfill
site would use approximately 15.3 million liters of water per year for controlling dust emissions
from road surfaces. Using the average annual MSW landfill site disposal rates averaged from WBJ
(2014), it is estimated that approximately 0.116 liters of wateris utilized for each kilogram of CDD
landfilled at an MSW landfill site. Assuming a 500-meter unpaved road length for CDD landfills
and the average annual CDD landfill site disposal rates from WBJ (2014), it is estimated that 0.351
liters of water is utilized for each kilogram of CDD landfilled at an unlined CDD landfill site. In
the absence of additional data, it is assumed that all water applied to road surfaces for the purpose
of dust control evaporates.
US EPA (2006a) AP-42, Chapter 13, Sections 2.2 and 2.4 were used to estimate particulate matter
emissions from unpaved road surfaces and from aggregate loading/unloading operations,
respectively. Particulate emissions from these sources are separately estimated by independent
datasets, as described in the Particulate Emissions section in Chapter 2 of this report.
The total quantity of nonmethane organic compounds (NMOC) produced during MSW
decomposition was assumed as a surrogate for volatile organic compound (VOC) emissions from
landfilling. The US EPA (2005) Landfill Gas Emissions Model (LandGEM) default gas generation
parameters were used to estimate the total mass of VOCs produced per kilogram of methane
generated. The calculated methane-to-VOC (i.e., NMOC) mass ratio was then multiplied by the
net quantity of methane released per kg of waste (i.e., after accounting for the methane attenuated
through cover soil or through destruction by flaring or combustion in a gas-to-energy project).
Operation-phase gasoline consumption was estimated from data obtained through IWCS (2016)
personal communication with a Florida MSW landfill site and publicly-available waste quantity
reports corresponding to site operations from 2007-2012. Approximately 1.35E-4 liters of gasoline
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is required per kg of material landfilled at an MSW landfill. Operations-phase gasoline
consumption for unlined CDD landfilling was estimated assuming the operation of a single 2010
Ford, F-150, 2-wheel drive, 6-speed automatic transmission. Fuel consumption was calculated
using default operating assumptions provided by US DOE (2016), resulting in an estimate of
approximately 1.53E-5 liters of gasoline used per kg of material landfilled at an unlined CDD
landfill.
2.6.6 Closure and Post-Closure Phase
Once a landfill has reached permitted capacity, it will undergo a closure process that usually
involves installing a low-permeability or impervious final cover system to minimize precipitation
infiltration and landfill gas emissions. A gas collection and control system (GCCS) is typically
installed at an MSW landfill before closure to control fugitive LFG emission. Unlike MSW
landfills, CDD landfills do not typically have an active GCCS. LCIs associated with closure
include cap installation (material and energy usage), construction of other site infrastructure (such
as roads), and continued operation (if applicable) of leachate and gas collection and management
systems, environmental monitoring, and post-closure care activities.
Ecobalance (1999) summarizes the quantities of individual cap materials necessary to close an
MSW landfill, based on a "typical final closure cover profile," which includes layers of soil,
geotextile, sand, clay, and HDPE. Consumption of soil and clay materials is aggregated by
Ecobalance (1999) into a single "soil" material category to quantify fuel consumption for soil
production- and transport-related emissions. The materials necessary for installing a GCCS and
gas monitoring system are also provided for the closure phase of the MSW landfill and are
organized into the consumption of HDPE and PVC.
During the post-closure-care period (assumed 30 years), Ecobalance (1999) assumes that 10% of
the cap will need to be replaced due to erosive wear. The LCI includes the soil and fuel (diesel)
requirement for replacing 10% of the cap over the 30-year post-closure-care period. The soil and
fuel amounts provided by Ecobalance (1999) for closure were increased by 10% to account for
this additional soil needed over the post-closure care period. The fuel consumption resulting from
site inspections (eight inspections were assumed to occur annually) and site mowing was also
estimated on an annual basis. These emission factors were multiplied by the 30-year post-closure-
care period and are included as "Gasoline combustion, in industrial equipment."
From an LCI perspective, constructing and operating a GCCS entails emissions from producing
and transporting system components and energy demands from GCCS construction and
installation. GCCS commonly include a flare or other destruction device (e.g., an internal
combustion engine) to oxidize methane and other chemicals of concern to carbon dioxide.
However, the MSW landfilling LCIs do not include materials and energy input for constructing a
blower/flare station.
Several states require a soil closure cap for CDD landfills (e.g., Florida, Virginia). A 0.61-m thick
soil closure cap was assumed for this dataset. The post-closure monitoring requirements and
duration varies among states. Some states, similar to the requirements for MSW landfills, require
30-year post-closure care monitoring of CDD landfills whereas others require post-closure care
for a much shorter duration than that of MSW landfills (e.g., Florida require post-closure care
monitoring for five years for CDD landfills). A post-closure care period of 10 years was assumed
2-22
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for this dataset. One-third of the post-closure fuel usage reported by EREF (1999) for MSW
landfills was used to estimate fuel consumption for post-closure care of CDD landfills (i.e. the
CDD landfill post-closure care period is one-third of the MSW landfill post-closure care period).
2-23
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Biodegradable materials;
end of life; at collection
Transport, truck; short-haul —
-Transport, truck; long-haul
--Transport, train; diesel powered-
Waste transport; at local road;
unpaved
Heavy equipment
- Stockpiled material; at drop site -
t t t i t
Fuel combustion Surface Ground Water _
Leachate CO
products water water , vapor ,
_j i i ! i L
t t t
CH4 VOCs
Cover soil; from unlined CDD
landfill stockpile
Unlined CDD landfilling of
biodegradable CDD materials;
at unlined CDD landfill
Legend
Elementary Flow
Technosphere Fjow
¦ Partially Included
| Included
System Boundary
Land occupation; j i
construction and demolition Electricity Gasoline Diesel
debris landfill I | i
Water
I ill
Figure 2-3. Unit process flow diagram for construction, operation, closure, and post-closure care of a
CDD landfill receiving biodegradable CDD materials, such as wood and paper.
1
Surface
water
Ground
water
t
Water
vapor
t
C02
t
CH4
t
VOCs
t
h2s
Gypsum drywall;
end of life; at collection
Transport, single unit truck;
short-haul; diesel powered
Transport, single unit truck;
long-haul; diesel powered
Transport, train; diesel powered-
Waste transport; at local road;
unpaved
Heavy equipment «
Stockpiled material; at drop site*
Cover soil; from offsite source-
Unlined CDD landfilling of
gypsum drywall; at unlined CDD
landfill
Legend
Elementary Flow
Technosphere Flow^
| Included
H Not Included
System Boundary
Land use
Electricity Gasoline
*
i
i
Diesel
i
i
Water
Figure 2-4. Unit process flow diagram for construction, operation, closure, and post-closure care of a
CDD landfill receiving gypsum drywall.
2-24
-------
Non-biodegradable materials;
end of life; at collection
_-Transport, single unit truck;
short-haul; diesel powered
Transport, single unit truck;
long-haul; diesel powered
-Transport, train; diesel powered
Waste transport; at local road;
unpaved
-Pipe, polyvinyl chloride; at plant*
Heavy equipment
Stockpiled material; at drop site-*
Cover soil; from offsite source-
Office trailer
Fuel combustion
products
*
Surface
*
Ground
water water
I
Leachate
+
Water
vapor
Unlined CDD landfilling of non-
biodegradable materials; at
unlined CDD landfill
Legend
Elementary Flow
_ _Technosjphej;(^F[qw_
Partially Included
Included
System Boundary
Land occupation;
construction and demolition
debris landfill
T~
Electricity Gasoline Diesel Water
Figure 2-5. Unit process flow diagram for construction, operation, closure, and post-closure care of a
CDD landfill receiving non-biodegradable CDD materials.
t
Fuel
t
T T , !
Leachate LFG; captured water
products
_ _ Biodegradable materials;
end of life; at collection
-Transport, single unit truck; short-haul; diesel powered*
-Transport, single unit truck; long-haul; diesel powered*
— "Transport, train; diesel powered *
Waste transport; at local road; unpaved *¦
Geomembrane; HDPE; at plant ~
Pipe; HDPE; at plant- >
Pipe; PVC; at plant ~
-Steel billets; at plant *-
Clay; at mine »•
Hot mix asphalt, HMA, at production facility—
Concrete; at plant ~
Stockpiled material; at drop site ~
Cover soil; from MSW landfill stockpile ~
-Geotextile; at plant
Heavy equipment
Soil; at excavation
Sand; at mine
Fugitive
Surface leachate
combustion wate roundwate (toWWTP) for energy vapor
1 1 production 1
I
rF t i
' FugitiveVOCs
CHa
MSW landfilling of
biodegradable materials; at
MSW landfill
Legend
Elementary Flow
Technosphere Flow
Partially Included
H Included
System Boundary
Land occupation;
municipal solid waste landfill
t
t
Electricity Gasoline Diesel
t
Water
Figure 2-6. Unit process flow diagram for construction, operation, closure, and post-closure care of
an MSW landfill receiving biodegradable CDD materials, such as wood and paper.
2-25
-------
t
Fuel
combustior"
products
Surface
I 1 t
Fugitive Leachate LFG; captured
leachate 't0 WWTP) for energy
(to ground water) production
t
Water
r
Fugitive
CH4
t
h2s
Gypsum drywall;
end of life; at collection
Transport, single unit truck;
short-haul; diesel powered
Transport, single unit truck;
long-haul; diesel powered
Transoort. train, diesel Dowered--
¦Waste transport; at local road; unpaved -
Geomembrane, HDPE, at plant
Pipe, HDPE, at plant
Pipe, PVC, at plant
Steel, billets, at plant
Clay; at mine
__Hot mix asphalt (HMA),__
at production facility
-Concrete; at plant
Stockpiled material; at drop site
_ Cover soil; from MSW
landfill stockpile
Geotextile; at plant
Heavy equipment
Soil; at excavation
Sand; at mine
MSW landfilling of gypsum
drywall; at MSW landfill
Legend
Elementary Flow
Tech nosphere_Flow ^
¦ Not Included
¦ Partially Included
H Included
System Boundary
Land use
Electricity
Gasoline
Diesel
Water
Figure 2-7. Unit process flow diagram for construction, operation, closure, and post-closure care of
an MSW landfill receiving gypsum drywall.
1
Fuel
combustion
products
Non-biodegradable CDD materials;
end of life, at collection
- Transport, single unit truck; short-haul; diesel powered
--Transport, single unit truck; long-haul; diesel powered-
————Transport, train; diesel powered
Waste transport; at local road; unpaved -
Geomembrane; HDPE; at plant
Pipe; HDPE; at plant-
Pipe; PVC; at plant
Steel, billets, at plant
Clay; at mine
Hot mix asphalt HMA,
at production facility
Concrete; at plant
- Stockpiled material; at drop site
Cover soil; from MSW landfill stockpile
Geotextile; at plant ~
Heavy equipment
Soil; at excavation
Sand; at mine
f ~
Fugitive
Surface leachate
water (to ground
water)
1 t
Leachate Water
(to WWTP) vapor
MSW landfilling of non-
biodegradable materials; at
MSW landfill
Legend
Elementary Flow
Technosphere Flow^
| Partially Included
Included
System Boundary
~T { T f ]
Land occupation; i
municipal solid Electricity Gasoline Diesel Water
waste landfill ) i i i
I 1,1
i I i
Figure 2-8. Unit process flow diagram for construction, operation, closure, and post-closure care of
an MSW landfill receiving non-biodegradable CDD materials.
2-26
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2.6.7 Landfill Leach ate Emissions
Landfill leachate is generated as precipitation or waste-entrained moisture percolates through the
waste material and dissolves and retains various compounds. Unlined landfills or landfills with
damaged bottom liners have the potential to release landfill leachate to underlying soils and
groundwater. In LCI databases and models, leachate emissions are generally modeled as emissions
to water, though leachate may contain dissolved gaseous species that are ultimately released to the
atmosphere. As mentioned previously, CDD landfills do not carry a federal requirement for liners
and leachate collection systems like MSW landfills. Leachate emissions are caused by the release
of compounds/elements in the waste materials themselves, resulting in direct (i.e., waste-specific)
emissions. Models and/or databases often do not handle leachate emissions on a waste-specific
basis but rather on assumptions of leachate composition from mixed waste streams (e.g., MSW
leachate), due to the relative lack of data on emissions from individual waste components,
particularly over large spans of time. The timeframe over which leachate emission continues to
occur after the waste placement complicates estimations of long-term leachate emissions.
Therefore, for the purposes of developing a landfill disposal management LCI for the different
CDD materials presented in this report, literature was reviewed to identify sources of material-
specific Synthetic Precipitation Leaching Procedure (SPLP) and Toxicity Characteristic Leaching
Procedure (TCLP) data. SPLP data was selected as representative of leaching which occurs as a
result of precipitation infiltration through CDD materials in either a CDD landfill or a beneficial
use land application (e.g., use as a fill material). TCLP data was used to estimate leachability of
CDD materials in an MSW landfill. The majority of the CDD materials were only modeled with
respect to disposal in an unlined CDD landfill or land-applied beneficial use application.
2.6.8 Landfill Gas Emissions
Landfill gas emissions result from the decay of landfilled organic materials and depend on the
landfilled waste composition. In an anaerobic MSW landfill environment, LFG tends to be
comprised of approximately 55% methane and 45% carbon dioxide with trace amounts of other
gases for the majority of the landfill's active and post-closure life. The CDD stream typically
contains smaller quantities of readily biodegradable wastes (generally the largest biodegradable
component of CDD is wood and paper); thus a lower total bulk gas production is observed (Doka
2003a). However, at CDD landfill sites, H2S, a malodorous compound produced typically from
decay of sulfur-containing wastes (e.g., gypsum drywall), can be produced. Since gas-production
rates are expected to be low at CDD landfills, there is no federal requirement for an active GCCS
at a CDD landfill, and employing combustion-based treatment systems can be challenging at CDD
sites as a result of the small amount of gas produced.
Gas production at MSWs is typically assumed to follow a first-order decay relationship, and the
production rates can be estimated using computer modeling tools (e.g., the US EPA's Landfill Gas
Emissions Model) (US EPA 2005). Since a GCCSis rarely used at CDD landfill sites, thetemporal
modeling or estimation of gaseous emissions from CDD landfills (either controlled or
uncontrolled) may be challenging because of the lack of data collected on gas composition and
quantities at operating facilities. Several researchers have used field surface emissions monitoring
and laboratory columns to quantify sulfur gas release from CDD landfills (Lee et al. 2006, Eun et
al. 2007, Xuetal. 2014).
2-27
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The generation, composition, and controlled/uncontrolled emissions of LFG are based on the gas-
production properties of the landfilled material (e.g., decay rate, total gas production potential),
GCCS operation and coverage area, and cover soil and cap installation and characteristics. Gas
treatment or inhibition mechanisms (e.g., co-disposal of drywall with ash or lime) employed will
also impact environmental emissions (Plaza et al. 2007, Panza and Belgiorno 2010, Xu et al. 2010,
Sungthong and Reinhart 2011).
The methodology used by the US EPA (2019) for WARM landfilling process was selected for
estimating the gas emissions from the landfill disposal of CDD materials. US EPA (2012b)
provides LFG production properties for numerous waste materials placed in landfills. While the
specific focus of the documentation is MSW materials, several materials are also part of the CDD
waste stream including dimensional lumber (selected to represent CDD wood), branches (selected
to represent LCD), cardboard and gypsum drywall. The data presented in US EPA (2012b) used
for estimating the LFG emissions for CDD and MSW landfill disposal of the different organic
CDD materials discussed in this report is presented below in Table 2-6.
Table 2-6. Landfill gas production properties for three CDD materials: lumber, cardboard, and
gypsum drywall.
Material
Initial mass
fraction of
carbon
(Barlaz 1998)
Fraction of
methane carbon in
initial carbon
(Barlaz 1998)
Ratio of
dry mass
to wet
mass
MSW landfill gas collection
efficiency, assuming typical
landfill moisture conditions
(Barlaz et al. 2009)
50Dimensional
Lumber/Branches1
0.49
0.12
0.9
0.9
Cardboard
0.47
0.22
0.95
0.89
Gypsum Drywall
0.05
0.18
0.94
0.872
1 US EPA (2012b) used experimental gas production results from branches as a proxy for gas production from dimensional lumber.
In the current report, the authors used gas production from branches as a proxy for gas production from LCD. A moisture content
of 50% was used for LCD instead of the 10% used in US EPA (2012b) for branches. The methane and carbon dioxide emissions
estimates for branches were adjusted for this difference in moisture content.
2 A gas collection efficiency for LFG produced from the decomposition of gypsum drywall was not provided, however, the gas
collection efficiency for waste paper was assumed since this is the organic portion of the drywall which will produce
methane/carbon dioxide.
The total amount of methane emitted from a landfill after the placement of any of the materials
listed in Table 2-6 in a CDD landfill was calculated using the following equation:
16
MGi = CDi xMcixWtx —
Where,
Moi = methane generated from landfill disposal of 1 kilogram of the ith material (kg)
CDi = initial carbon mass fraction of the dry ith material (%)
Mci = methane carbon as a fraction of the initial carbon (%)
Wi = ratio of dry to wet mass
16/12 = conversion factor methane to carbon mass
2-28
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Assuming thatLFG is not collected at CDD landfills, the total amount of methane emitted from
CDD landfills was assumed to be 90% of the amount generated due to an assumed cover soil
oxidation rate of 10% (as discussed previously in the WARM landfill gas emission section). The
amount of degraded carbon is equally allocated to methane and carbon dioxide (based on results
from Barlaz et al. 1989). Therefore, the amount of carbon dioxide generated from the placement
of one of the CDD materials in a CDD landfill prior to cover soil oxidation may be approximated
according to the following equation:
44
^Gi ~ CDi X MCi XWtX —
Where,
Coi = carbon dioxide generated from landfill disposal of 1 kilogram of the ith material (kg)
44/12 = conversion factor of the molecular ratio of carbon dioxide to carbon
The carbon dioxide emission estimated from equation above was multiplied with 1.1 to account
for the carbon dioxide from the oxidation of methane in the cover soil.
US EPA (2012b) also presents information which allows an estimate of the amount of methane
and carbon dioxide emitted from the MSW landfill disposal of the CDD materials presented in
Table 2-6. US EPA (2011) reported that approximately 72% of all landfill-produced methane is
generated at landfills with a GCCS. Methane emissions would result from fugitive LFG at MSW
landfills without as well as with GCCS. The methane emissions from the disposal of the CDD
materials at an average nationwide MSW landfill was estimated using the following equation:
MEMswi = MGi x 0.9 x (Lgccs x (1 — I]C() + (1 — Lgcc5))
Where,
Memswi = mass of methane emitted from the disposal of 1 kilogram of the ith material at
an MSW landfill (kg)
Lgccs = percentage of methane from MSW landfills with a GCCS (i.e. 72%)
Hci = gas collection efficiency for the /th material (see Table 2-6)
0.9 = factor accounting for cover soil oxidation of uncollected methane. Other variables as
defined above.
In addition to the carbon dioxide generated from organic material decomposition, carbon dioxide
would also result from the combustion of methane collected in an MSW landfill GCCS and from
the oxidation of uncollected methane emitted through the landfill cover soil. Carbon dioxide
emissions from the disposal of the CDD materials at an average nationwide MSW landfill may be
estimated according to the following equation:
44 / \
^EMSWi = ^Gi + MGi X — X (^Gccsnc;) + 0-1 X ((^GCCs)(l ~~ Hci) + (1 ~~ ^GCCs)))
Where,
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-------
CEMswi = mass of carbon dioxide emitted from the disposal of 1 kilogram of the ith material
at an MSW landfill (kg)
0.1 = factor accounting for the carbon dioxide produced from cover soil oxidation of
uncollected methane
44/16 = conversion factor of the molecular ratio of carbon dioxide to methane. Other
variables as defined above.
Table 2-7 presents a summary of the estimated methane and carbon dioxide emissions for the CDD
and MSW landfill disposal of the organic CDD materials discussed in this report based on the
calculation methodology presented above.
Table 2-7. Comparison of methane and carbon dioxide emissions resulting from disposal of CDD
materials in CDD and MSW landfills.
Material
CDD Landfill Emissions
MSW Landfill Emissions
Methane
(kg/kg)
Carbon
Dioxide (kg/kg)
Methane
(kg/kg)
Carbon
Dioxide (kg/kg)
Dimensional Lumber and
Engineered Wood
0.064
0.21
0.022
0.33
Cardboard
0.12
0.40
0.042
0.60
Gypsum Drywall
0.010
0.034
0.0038
0.052
LCD
0.036
0.12
0.012
0.183
2.6.9 Landfill Gas and Leach ate Collection and Treatment
Because themajority ofUS CDD landfills do not have a GCCS, an LCI process dataset that models
the environmental burdens associated with gas collection and management at a CDD landfill was
not developed.
As of 2012, 17 states required that CDD landfills have liners and leachate collection systems
(IWCS 2012). Since this is fewer than half the states, it was assumed that the typical US CDD
landfill is unlined. The emissions resulting from installing, operating, and decommissioning a
leachate collection and removal system (LCRS) at an MSW landfill still need to be considered for
CDD materials placed in MSW landfills. Two subcategories of emissions result from leachate
collection and treatment. One subcategory is emissions based on the total volume of leachate
handled (e.g., emissions resulting from electricity used to pump leachate). The other subcategory
is emissions based on the treatment of specific leachate constituents (e.g., amount of energy
necessary for aeration of leachate to reduce the concentration of BOD below regulatory limits).
While Ecobalance (1999) provides an estimate of the average leachate collected per ton of MSW
deposited in MSW landfills at different intervals after placement (i.e., 20, 100, and 500 years), it
does not estimate the average amount of electricity necessary to collect and transport leachate. The
only energy consumption estimate provided in Ecobalance (1999) is for the WWTP treatment of
BOD, given as 0.001 kWh/g BOD removed. While WWTP removal efficiencies are provided for
seven leachate constituents/parameter categories (i.e., COD, BOD, ammonia, phosphate, total
suspended solids, heavy metals, and trace organics), the specific energy required for the removal
of each of these is not provided. While the emission path for each parameter is specified (e.g.,
BOD treatment will release carbon dioxide to air, emitbiomass sludge), an additional complicating
2-30
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factor is the potential for the sludge from the WWTP to be disposed of at the MSW landfill from
which the leachate came. Furthermore, it does not appear that the removal efficiencies are
temporally weighted to account for the average LCRS operational period (estimated as 40 years).
To estimate the mass of a particular contaminant (or treatment byproducts of a particular
contaminant) released to surface water or to air as a result of WWTP treatment, it would be
necessary to estimate the mass emissions of the contaminant through MSW leachate over time.
Because of these limitations and the data gaps identified in the next section, it was not possible to
create anLCIdatasetforthe collection and treatment of leachate produced from disposing of CDD
materials in an MSW landfill.
2.6.10 Data Gap Analysis of Landfill Gas and Landfill Leachate Collection and
T reatment
To develop an LCI dataset that simulates emissions resulting from the use of an MSW landfill
GCCS to collect gas produced from CDD materials, the following data gaps need to be addressed:
1. Average energy required to construct and manufacture the components of a GCCS.
Significant quantities of steel are used in the manufacturing of flare stations used for an MSW
landfill GCCS; the quantity of steel and other components used is important for developing a
representative LCI for GCCS construction. Furthermore, determining the average energy
associated with gas well drilling/trenching and GCCS pipe welding would allow for a more
accurate estimate of the emissions associated with GCCS construction and installation.
2. Average energy required to collect a unit volume of LFG. Electricity is necessary to power
GCCS equipment; therefore, emissions associated with powering GCCS equipment should be
allocated to organic CDD materials on a gas-collected-per-mass-disposed-of basis.
3. Data on common practices for decommissioning GCCS equipment. Information on the
EOL management of GCCS equipment is necessary to estimate the complete environmental
burdens associated with its serviceable life. Depending on whether the flare station or other
GCCS components are recovered or simply disposed of will have an impact on the total
emissions resulting from the MSW landfill disposal process dataset.
Besides the leachate data limitations presented for each of the materials in its respective chapter,
the following additional information would be necessary to develop a representative dataset for
leachate emissions resulting from the placement of CDD in an MSW landfill:
1. Measured Volume of leachate collected over the serviceable life of an MSW landfill
LCRS. The CDD material-specific MSW landfill datasets developed as part of this project
estimated leachate volume by assuming a percent of the total precipitation infiltrates into the
waste and is collected, based on EREF (1999). Total leachate volume is necessary to estimate
the average emissions associated with collecting and transporting MSW leachate, irrespective
of leachate quality.
2. Mass fraction of different CDD-material-produced leachate constituents, which may be
expected to be released in leachate during the service life of the LCRS. Currently, only
limited leachate concentration information is available for a relatively small number of
2-31
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parameters for a relatively narrow range of CDD materials. It is not possible to estimate the
total mass fraction of many leachate parameters that would be collected and treated at a WWTP
from the landfilling of CDD. For example, it is expected that calcium would leach from the
placement of demolished concrete in an MSW landfill environment. There currently appears
to be no information that specifically predicts the total amount of calcium that will leach out
from concrete placed in an MSW landfill over the operational life of an LCRS. Like LFG
production, it is likely that individual leachate contaminants will be released at different rates
depending on the specific CDD material. Long-term leaching data specific to each CDD
material will likely be necessary before an LCI dataset representing the emissions resulting
from treating leachate from MSW-landfill-placed CDD can be created.
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3 Asphalt Pavement
3.1 Introduction
Asphalt pavement is constructed in multiple layers: top surface, intermediate, and base. The top
two layers typically consist of 95% aggregate and 5% asphalt (FHWA 2011). Asphalt (also
referred to as bitumen), which is a product of petroleum refining and includes the denser fraction
of crude oil hydrocarbons, is used as an aggregate binder. Aggregates used for asphalt pavement
production may include gravel, sand, and crushed stone; crushed stone may include various rock
types such as limestone, dolomite, and granite. While there are numerous processes for paving mix
production, HMA production is the most commonly used process; approximately 94% of US roads
are paved with HMA (US EPA 2012, Kelly 2011). NAPA (2020) estimated that approximately
382 million metric tons (MMT) of asphalt pavement were produced and used for road construction,
rehabilitation, restoration, and resurfacing in the US in 2019.
Asphalt pavements are routinely rehabilitated, resurfaced, and reconstructed due to surface
wearing over time. Asphalt pavement is removed by either road milling or demolition through
excavation. Milling involves grinding the road surface using a machine which has a toothed rotary
drum that can be lowered or raised to adjust the milling depth. Road pulverization and excavation
may be used in instances where the road base is compromised and no longer provides sufficient
structural support for the overlying layers or for instances where milling is not feasible or
economically justified (e.g., parking lots, small road stretches).
Figure 3-1. Process flow diagram for end-of-life management of asphalt pavement.
Table 3-1 lists the processes that should be considered to conduct an LCA of EOL management
of asphalt pavement. The emissions associated with energy and materials requirements and
process non-energy emissions (e.g., fugitive dust, liquid emissions associated with disposal of
RAP in a landfill) were taken into account in compiling the different LCI datasets.
Table 3-1. End-of-life management processes for reclaimed asphalt pavement.
Process
LCA considerations
Processing of reclaimed
asphalt pavement; at
processing
RAP processing may include additional crushing and fractionation (i.e.,
sorting into different size categories). The extent of RAP processing is
dependent on many factors such as end use, amount of RAP used
relative to primary materials for HMA, production method, and the
duration of RAP storage in a stockpile.
3-1
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Process
LCA considerations
Use of reclaimed asphalt
pavement; as general fill;
at fill site
This process models the placement and use of reclaimed asphalt
pavement chunks for general fill in the environment.
3.2 LCI Sources
Peer-reviewed literature, government and private industry publications, and various LCA
modeling tools were reviewed to identify available LCI datasets pertaining to asphalt pavement
EOL management processes. Table 3-2 lists data sources reviewed to develop the LCI presented
in this chapter. If LCI data were not available, process metadata and documentation (e.g., included
emission categories, background data used to compile the dataset, geographic location and time
period of the data) were reviewed to evaluate the completeness of the dataset. If available, the
primary sources of information used to develop the LCI datasets and information were reviewed.
Table 3-2. Life cycle inventory (LCI) data sources for reclaimed asphalt pavement.
Source
Description
Wilburn and
Goonan
(1998)
The authors provide energy requirements associated with crushing/sorting stone, sand
and gravel, and RAP. These data were taken from the Portland Cement Association
and an energy audit of a recycling facility in Denver, Colorado.
US EPA
(2019)-
WARM
EPA's Waste Reduction Model presents data on GHG emissions associated with
source reduction, transport, recycling, and landfilling (i. e., collection and placement)
of asphalt pavement.
RSMeans
(2011)
Construction cost data which includes estimates for types and productivity of heavy
equipment typically used to manage RAP in EOL processes (e.g., use as a soil fill).
3.3 LCI Related to Disposal
The annual survey on recycled asphalt materials conducted by the National Asphalt Pavement
Association (NAPA), which in 2019 included 212 asphalt mix production companies from 48
states representing 1,101 production plants, consistently finds that less than 1% of all RAP is
disposed of in a landfill (NAPA 2020). The vast majority is used for other purposes, such as
aggregate or HMA.
Air emissions from landfill disposal of asphalt pavement result from the operation of landfill
equipment during material and cover soil compaction and placement, including both fuel-related
and pre-combustion emissions. RAP exposure to precipitation or other liquids (e.g., landfill
leachate) results in leached emissions. In addition, asphalt mixtures can contain many different
additives that, while small in number, may prove to be environmentally signfiicant.
Leachable emissions from RAP were estimated using the SPLP (batch test) and leaching column
data reported by Townsend and Brantley (1998). Townsend and Brantley (1998) conducted batch
and column leaching tests on asphalt pavement samples collected from six sites. Batch test data
were used for contaminants except heavy metals. Batch test concentrations were multiplied by the
total solution volume and divided by the sample mass to estimate leachability on a per-kilogram-
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asphalt pavement basis. Contaminant emissions were not estimated for parameters below the
detection limit in more than half (i.e., three) of the samples; bromide, sodium, and potassium were
excluded. The detection limit was used as the concentration for measurements below detection for
samples in which the contaminant emission was quantified. Nitrate and sulfate data reported by
the study were not used since the SPLP extraction fluid contains these anions. The emissions of
non-purgeable organic compounds were not estimated as this compound is not included as an
elementary flow in US LCI (2012). Total dissolved solid (TDS) data reported by the study were
also not included to avoid double counting of emissions as some of the contaminants are included
in the TDS measurement.
None of the measured heavy metals and organic compounds leached above the detection limits in
batch leaching tests. Therefore, column test data were used to estimate heavy metals and organics
emissions. All the organic compounds (VOCs and PAHs) and heavy metals, except lead, targeted
in this study were below detection in the column tests. The total lead leached in saturated column
test was greater than that of the unsaturated column test; the lead emission was estimated based on
the saturated column test data, which corresponds to a liquid-to-solid (L:S) ratio of 0.257. Lead
emission was estimated for each of the six samples by dividing the total amount of lead leached
from a given RAP sample by the total RAP amount used for the test. The average lead leached
from six samples was estimated to be 6.58 |ig/kg of RAP.
Azah (2011) conducted batch and (saturated and unsaturated SPLP) column leaching tests on RAP
samples collected from five Florida locations to assess PAH leaching. Leaching data were selected
from column tests for PAHs, which were detected in over half of the samples. Four PAHs were
detected in over half of the samples during unsaturated column testing (i.e., fluoranthene, pyrene,
benzo(k)fluoranthene, benzo(a)pyrene) while three PAHs were detected in over half of the samples
during saturated column testing (i.e., pyrene, benzo(g,h,i)perylene, benzo(b)fluoranthene).
Because the final L:S was higher for saturated column testing compared to unsaturated column
testing (i.e., approximately 2 versus 1.2, respectively) saturated column leaching results were
selected over unsaturated column testing results for pyrene since this compound was detected in
over half of the samples for both column testing conditions. Data were used to estimate liquids
emissions from the disposal of RAP in CDD material landfills. Column leaching data were used
to estimate the total leachable amount of PAHs - all below-detection-limit (BDL) measurements
subsequent to detected concentrations were excluded from the analysis. The other BDL
measurements were included at the detection limit concentration. Temporal column PAH
concentrations (ng/L) were multiplied by the L:S ratio (L/kg) of the solution and values were
summed for each PAH to estimate the PAH's total leachability on a per-kilogram-asphalt-
pavement basis.
The energy use and the associated emissions from landfill operation (e.g., waste placement,
compaction) include diesel use in heavy equipment and electricity use in landfill buildings (e.g.,
administrative buildings, workshop). In the absence of additional data, it was assumed that asphalt
pavement would be transported 20 km for landfill disposal. Diesel consumption from landfill
operations and electricity consumption from landfill administrative offices and workshop areas
were estimated from EREF (1999) and IWCS (2014), respectively, and these flows are included
as the "CDD landfill operations" input flow, as detailed in Chapter 2. Details on how cover soil
was assigned for the placement of demolished asphalt pavement in a CDD landfill is also included
in Chapter 2 and is based on the bulk density of asphalt pavement as provided in CCG (2006).
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RAP leaching data, energy consumption data from landfill operations, and the assumed transport
distance were used to develop an LCI process dataset for the disposal of RAP at an MSW and an
unlined CDD landfill, as discussed in Section 2.5.10.
3.4 LCI Related to Recycling
3.4.1 RAP Processing
RAP may need to be processed prior to recycling as an aggregate or for the production of new
paving mix. The extent of RAP processing would be dependent on a number of factors, such as
the pavement removal method (i.e., milling, demolition and excavation), duration it is stockpiled,
and end-use specifications. For those instances where RAP is produced as a result of road
demolition and excavation, RAP must be size-reduced before it can be recycled into new HMA or
for most aggregate applications. Additional processing is also likely to be required when RAP
constitutes greater than 20% of the new paving mix for the mix to meet specified fractionation
requirements. According to Brock and Richmond (2007), many continuously fed HMA plants will
have a closed-circuit crushing system on the front end of the aggregate feeder where oversized
milled material and aggregate is screened away from the feed, crushed, and then returned to the
screen. Emissions from crushing and screening RAP millings could in this case be considered part
of the total emissions released from the HMA plant. However, it is possible that some RAP
produced as a result of pavement demolition and excavation may be crushed off site from an HMA
plant.
Table 3-3 summarizes the LCI dataset for RAP processing. Similar to aggregate crushing and
sorting during production, it is expected that particulates would be a source of emissions released
from RAP crushing and sorting operations. These fugitive dust emissions are of such small relative
quantity that they are considered negligible and are therefore not included in Table 3-3. If
recovered pavement is not processed prior to use, the diesel and electricity consumption should be
excluded from the LCA. In the absence of national average data on the distance from the road
demolition site to RAP processing, a distance of 20 km was assumed with transport by a single-
unit, short-haul, diesel-powered truck.
Table 3-3. Life cycle inventory and associated data sources for the unit process: "asphalt pavement,
Input Flow
Source
Unit
Amount
Asphalt pavement, from road demolition
N/A
kg
1.00
Diesel, combusted in industrial equipment
Wilburn and Goonan
(1998)
L
0.000213
Electricity, at industrial user
Wilburn and Goonan
(1998)
kWh
0.00229
Truck transport, class 8, heavy heavy-duty
(HHD), diesel, short-haul, load factor0.75
USCB (2010)
t*km
0.02
Output Flow
Source
Unit
Amount
Reclaimed asphalt pavement, at processing
N/A
kg
1.00
3.4.2 RAP Use as Aggregate
RAP produced from road demolition and excavation activities will likely need to be crushed prior
to use in structural fill applications; however, it is possible that millings may be used without
3-4
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additional processing. The primary emissions pathway resulting from the use of RAP as a fill
material include leaching to ground water (Townsend and Brantley 1998, Azah 2011). The use of
RAP as aggregate would avoid the emissions resulting from the production and transport of
primary aggregates as presented in Chapter 2, as well as use as binder replacement for the asphalt.
This process models the placement and use of reclaimed asphalt pavement chunks for general fill
in the environment. This dataset accounts for material transport to the site, particulate emissions
associated with material placement/stockpiling, equipment utilization, equipment fuel combustion
emissions relating to material placement, and leaching emissions after placement. Specific
activities included in this dataset which contribute to the overall environmental burdens associated
with the process include the placement of the material and leaching of the material once it is put
in place. Figure 3-2 shows the flows included in the dataset.
A primary assumption of this dataset is that all precipitation that comes in contact with the
stockpiled material is discharged to groundwater, though regional stormwater management
practices may vary. For leaching data, the SPLP test data are assumed to be reflective of the long-
term leachate emissions from crushed demolished asphalt pavement used as a general fill in the
environment. The fuel combustion and equipment used to place the RAP is taken from RSMeans
(2011) and Caterpillar (2006) data on the placement of riprap.
For non-metal inorganics and lead, contaminant volumetric concentrations were converted to mass
concentrations by multiplying the volumetric concentration with the L:S ratio used in the SPLP
test. Parameters with concentrations below detection limit (BDL) for more than 50% of the
samples were not included in the dataset. For purposes of determining an average leachability
value, the detection limit concentration was used to replace the BDL value for the rest of the
parameters.
For PAH concentrations, column leaching data were used to estimate the total leachable mass of
PAHs-all BDL concentration measurements subsequent to detected concentrations were excluded
from the analysis. Other BDL measurements were included at the detection limit concentration.
Temporal column PAH concentrations (ng/L) were multiplied by the L:S ratio (L/kg) of the
solution and all values were summed for each PAH to estimate the PAH's leachability on a per-
kilogram-asphalt-pavement basis.
This dataset calls in a flow from an additional dataset to simulate the release of PM as a result of
material stockpiling. These PM emissions are accounted for by the "stockpiling of material; at drop
site" inflow. The dataset user should modify the upstream processes to account for 1 material drop
(i.e., discharge). Additional process-specific parameters (e.g., wind speed, moisture content)
should be updated for this process if information is available. Data on the fines fraction of asphalt
pavement chunks was not found in literature, therefore, it was assumed that asphalt pavement
chunks only have 10% of the fines present in the aggregates that were tested in the development
of the AP-42 emission factor methodology used to simulate material stockpiling.
Information on equipment utilization and equipment fuel use to place RAP was taken from
RSMeans (2011) item 31 37 13.10 0300, "Riprap, dumped, 50 lb average". This data includes
bulldozer use and productivity (tons per day). Fuel use for the bulldozer from RSMeans (2011)
was estimated using fuel use data from five bulldozers given in Caterpillar (2006). Since all the
engines in Caterpillar (2006) were larger than the engine size given in RSMeans (2011), the engine
3-5
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size data from Caterpillar (2006) with the hourly fuel use per engine size (L per hour per
horsepower). The fuel use estimate for the equipment, combined with the equipment productivity
from RSMeans (2011), was used to estimate the total fuel use per functional unit. The lifetime of
the nonroad equipment (i.e., for equipment utilization per functional unit) was estimated using the
equipment-specific lifetime fuel consumption values presented in the US EPA (2014) Motor
Vehicle Emissions Simulator (MOVES) database. The MOVES database presents estimates of the
average annual fuel consumption for specific equipment categories.
t
Fuel combustion
Ground
Leachate
byproducts water i
1 1 1
Transport, single unit truck;
short-haul; diesel powered
Heavy equipment ~
Use of reclaimed
Reclaimed asphalt pavement;
from road demolition; ~
at demolition site
asphalt pavement; as
general fill; at fill site
Stockpiled material;
at drop site
* i r
Diesel Water Oxygen
! 1 1
1
Legend
Elementary Flow
Technosphere Flow^
| Not Included
| Included
System Boundary
Figure 3-2. Unit process flow diagram for the use of reclaimed asphalt pavement as general fill.
3.5 Data Gap Analysis and Opportunities for Additional LCI Data
In the review of US-based asphalt pavement management LCI data. Three models and three
publications were identified that contained at least partial emissions data for some portion of
asphalt management. With the exception of WARM and NREL (WARM uses information from a
limestone mining process included in the 2009 NREL database), each of these data sources is
independent of each other, (i.e., the same primary data are not being used in multiple sources). As
shown in Table 3-4, WARM process data are partial because WARM only analyzes GHG
emissions. Similar to WARM, AP-42 only includes partial data because the datasetsonly include
process air emissions. Wilburn and Goonan (1998) only provide energy requirements for RAP
processing, without distinguishing what fuels are used to provide this energy. Townsend and
Brantley (1998) used SPLP batch and column testing to estimate the leachability of numerous
organic, inorganic, and metal parameters. However, the leaching data are only considered partial
because testing occurred in 1997-1998, and it appears that some laboratory detection limits have
significantly decreased since that time. For example, Azah (2011) reports the leachable amount of
PAHs from RAP, where the detection limit for these tests ranged between 0.0001 and 3.5 |ig/L
while the detection limit for Townsend and Brantley (1998) ranged between 0.5 and 5 |ig/L. Also,
Townsend and Brantley (1998) column testing occurred over 40 days while column testing by
Azah (2011) occurred over 35 days. Additional extended runs would be necessary to observe
concentration trends to estimate the total leachable concentrations from RAP placed in a general
3-6
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fill or unlined landfill. Datafrom Azah (2011) are partial because Azah only presents the leachable
amounts of PAHs. All identified models and publications that include environmental burdens of
some portion of asphalt pavement management are presented in Table 3-4. The table shows that
LCI data are better documented for primary material extraction and HMA production processes
than for pavement removal and EOL management.
Based on a review of these sources, the following needs for additional US-specific LCI data with
respect to the EOL management of asphalt pavement were identified:
1. Milling/Excavation equipment fuel consumption, operation, and manufacturing
emissions. The US EPANONROAD model simulates air emissions from non-road equipment,
including milling and excavation equipment, on a per-horsepower-operating-hour basis.
Published equipment-specific loading capacities (e.g., from equipment manufacturer
performance handbooks) could be used to estimate a conservative loading rate to project
emissions on a mass-emission-per-mass-pavement-removed basis by using these emission
factors as provided in NONROAD. This is a similar approach to that taken by the Washington
State Department of Transportation in an LCA they undertook to compare PCC road
rehabilitation alternatives (Weiland and Muench 2010). However, while this information
would allow for estimating fuel-related air emissions due to equipment operation, no estimates
of particulate matter emissions from road abrasion or environmental burdens associated with
milling/excavation equipment manufacturing were found in the literature. Also, while it
appears that water spray is used to suppress particulate emissions from milling, no estimates
were found on the quantity or quality of the water released from spraying.
2. Leachable emissions from the landfill disposal of RAP as aggregate. Although an estimate
of liquid emissions from RAP disposal in unlined CDD landfills and for use as aggregate is
included in the LCI datasets, these estimates are likely lower than the actual emission due to
partial leaching of contaminants attributed to the L:S ratio of the batch and column RAP
leaching test data they were based on. Future research should consider estimating the long-
term leaching of RAP.
3-7
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Table 3-4. Life cycle inventory data sources for reclaimed asphalt pavement. "P" indicates partial
Process
US EPA
(2019)-
WARM
GaBi
Athena
Wilburn
and
Goonan
(1998)
US LCI
(2012)
Townsend
and
Brantley
(1998)
Azah
(2011)
Crude Oil
Extraction
P
X
X
-
X
-
-
Asphalt
Production
P
X
X
-
X
-
-
Transport
P
-
X
P
-
-
-
Landfill
Disposal
P
-
-
-
-
P
P
RAP
Processing
-
-
-
P
-
-
-
RAP Use as
Aggregate
-
-
-
-
-
P
P
3.6 References
Athena (2001). A Life Cycle Inventory for Road and Roofing Asphalt. Prepared by Franklin
Associates, March 2001. http://www.athenasmi.org/wp-
content/uploads/2011/10/5_Road_And_Roofing_Asphalt.pdf. Accessed 19 February
2014.
Aurangzeb, Q., Al-Qadi, I.L., Ozer, H., and Yang, R. (2014). Hybrid Life Cycle Assessment for
Asphalt Mixtures with High RAP Content. Resources, Conservation and Recycling, 83,
77-86.
Azah, E.M. (2011). The Impact of Polycyclic Aromatic Hydrocarbons (PAHs) on Beneficial Use
of Waste Materials. Ph.D. Dissertation. University of Florida, Gainesville, FL, USA.
Bolen, W.P. (2013). 2011 Minerals Yearbook - Sand and Gravel, Construction [AdvanceRelease],
Published by the United States Geological Survey, MEPMay 2013.
http://on.doi.gov/lzpWK2z. Accessed 12 March 2014.
Brock, J.D. and Richmond, J.L. (2007). Technical Paper T-127: Milling and Recycling. Asteclnc.,
Chattanooga, Tennessee.
Caterpillar (2006). Caterpillar Performance Handbook. Edition 36. Cat publication. Caterpillar
Inc., Peoria IL.
CCG (2006). Targeted Statewide Waste Characterization Study: Waste Disposal and Diversion
Findings for Selected Industry Groups. A Report Prepared by Cascadia Consulting Group
for the California Integrated Waste Management Board, June 2006.
COLAS (2003). The Environmental Road of the Future: Life Cycle Analysis. Energy
Consumption & Greenhouse Gas Emissions. September 2003.
https://www.colas.com/sites/default/files/publications/route-future-english_l.pdf.
Accessed 4 April 2014.
Copeland, A. (2011). High Reclaimed Asphalt Pavement Use. FHWA Publication No.: FHWA-
HRT-11-057, McLean, VA,USA. September 2011.
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Ecoinvent (2014). Swiss Center for Life Cycle Inventories: Ecoinvent Centre. Dataset Information
(UPR): Treatment of Waste Asphalt, Sanitary Landfill, CH.
Eurobitume (2012). Life Cycle Inventory: Bitumen. 2nd Edition, The European Bitumen
Association, Brussels, Belgium, July 2012. https://www.eurobitume.eu/fileadmin/pdf-
downloads/LCI%20Report-Website-2ndEdition-20120726.pdf. Accessed 20 February
2014.
FHWA (2011). Reclaimed Asphalt Pavement in Asphalt Mixtures: State of the Practice,
Publication No. FHWA-HRT-11-021. The Federal Highway Administration, US
Department of Transportation at the Research, Development, and Technology Turner -
Fairbank Highway Research Center, McLean, VA, USA, April 2011.
FHWA (2012). Warm Mix Asphalt (WMA). Federal Highway Administration, US Department of
Transportation, January 2012. http://Lusa.gov/lqKulQH. Accessed 11 June 2014.
IWCS (2014). Personal communication with confidential client and Pradeep Jain, Innovative
Waste Consulting Services.
Jullien, A., Moneron, P., Quaranta, G., and Gaillard, D. (2006). Air Emissions from Pavement
Layers Composed of Varying Rates of Reclaimed Asphalt. Resources, Conservation and
Recycling, 47, 356-374.
Kelly, W.M. (2011). Mineral Industry of the State of New York: 2007-2010. New York State
Museum Record 3. In: Ruffer, R. and Gardner, K. Report on the Economic Impact of the
New York State Mining and Construction Materials Industry.
http://www.nysm.nysed.gov/common/nysm/files/nysmrecord-vol3_0.pdf.
Lee, W., Chao, W., Shih, M., Tsai, C., Chen, T.J., and Tsai, P. (2004). Emissions of Polycyclic
Aromatic Hydrocarbons from Batch Hot Mix Asphalt Plants. Environmental Science &
Technology, 38 (20), 6274-6280.
Legret, A., Odie, L., Demare, D., and Jullien, A. (2005). Leaching of Heavy Metals and Polycyclic
Aromatic Hydrocarbons from Reclaimed Asphalt Pavement. Water Research, 39, 3675-
3685.
NAPA (2020). Asphalt Pavement Industry Survey on Recycled Materials and Warm-Mix Asphalt
Usage: 2019, Information Series 138. National Asphalt Pavement Association, Greenbelt,
Maryland, September 2020.
NAPA and EAPA (2011). The Asphalt Paving Industry: A Global Perspective, 2nd Edition.
European Asphalt Pavement Association and the National Asphalt Pavement Association,
February 2011.
Natural Resources Canada (2005). Road Rehabilitation Energy Reduction Guide for Canadian
Road Builders. Canadian Industry Program for Energy Conservation, Ottawa, Ontario.
https://www.nrcan.gc.ca/sites/www.nrcan.gc.ca/files/oee/pdf/industrial/technical-
info/benchmarking/roadrehab/Roadhab_eng_web.pdf. Accessed 19 February 2014.
NIST (2018). Building for Environmental and Economic Sustainability (BEES) Online 2.0
Technical Manual. NIST Technical Note 2032, National Institute of Standards and
Technology, December 2018.
https://nvlpubs.nist.gov/nistpubs/TechnicalNotes/NIST.TN.2032.pdf.
PE International (n.d.). GaBi Software. Search GaBi Databases.
https://gabi.sphera.com/america/databases/gabi-data-search. Accessed May 2014.
RSMeans, (2011). Building Construction Cost Data. 69th Ed. Reed Construction Data, Norwell,
MA.
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Stripple, H. (2001). Life Cycle Assessment of Road - A Pilot Study for Inventory Analysis, 2nd
Revised Edition. A Report Prepared by the IVL Swedish Environmental Research Institute
for the Swedish National Road Administration, March 2001.
https://trid.trb.org/view/689935. Accessed 20 February 2014.
Townsend, T.G. and Brantley, A.S. (1998). Leaching Characteristics of Asphalt Road Waste.
Report #98-2. State University System of Florida, Florida Center for Solid and Hazardous
Waste Management.
TRB (2011). NCHRP Report 691: Mix Design Practices for Warm Mix Asphalt. A Report
Prepared by Ramon Bonaquist of Advanced Asphalt Technologies, for the National
Cooperative Highway Research Program, Transportation Research Board of the National
Academies, Washington, D C., USA.
US EIA (2013). 2010 Manufacturing Energy Consumption Survey Data. US Energy Information
Administration, Washington, D.C., USA. http://Lusa.gov/U4guIv. Accessed 6 June 2014.
US EIA (2014). Petroleum and Other Liquids, Supply and Disposition: Asphalt and Road Oil. US
Energy Information Administration, Washington, D.C., USA. http://Lusa.gov/lnePIsv.
Accessed 24 March 2014.
US EIA, (n.d.). International Energy Statistics - Units. US Energy Information Administration.
Accessed May 5th, 2016. http://www.eia.gov/cfapps/ipdbproject/docs/unitswithpetro.cfm
US EPA (1995). Compilation of Air Pollutant Emission Factors -5th Edition, Volume I: Stationary
Point and Area Sources. Office of Air Quality Planning and Standards, Office of Air and
Radiation, US http://Lusa.gov/lmgxh6jUS.
US EPA (2000). Hot Mix Asphalt Plants Emission Assessment Report. A Report Prepared by
Emissions Monitoring and Analysis Division of the US Environmental Protection Agency
for Midwest Research Institute and Eastern Research Group, Inc., December 2000.
US EPA (2004). Section 11.19.2 - Crushed Stone Processing and Pulverized Mineral Processing.
AP 42, Fifth Edition, Volume I, Chapter 11: Mineral Products Industry, US Environmental
Protection Agency. http://Lusa.gov/ljerheN. Accessed 4 April 2014.
US EPA (2006). Section 7.1 - Organic Liquid Storage Tanks AP 42, Fifth Edition, Volume I,
Chapter 7: Liquid Storage Tanks, US Environmental Protection Agency.
http://Lusa.gov/lqkteVT. Accessed 2 June 2014.
US EPA (2012). Asphalt Concrete. US EPA for the Waste Reduction Model, Version 12, US
Environmental Protection Agency, February 2012. http://Lusa.gov/VLedD3.Accessed 19
February 2014.
US LCI (2012). US Life Cycle Inventory Database. National Renewable Energy Laboratory.
http://www.nrel.gov/lci/. Accessed 20 February 2014.
USCB (2010). 2007 Economic Census, Transportation, Commodity Flow Survey, EC07TCF-US.
United States Department of Transportation, Research and Innovative Technology
Administration, Bureau of Transportation Statistics and the United States Department of
Commerce, Economics and Statistics Administration, United States Census Bureau. April
2010.
Walker, D. and Davis, J. (2008). An Overview of Storage and Handling of Asphalt. Asphalt: The
Magazine of the Asphalt Institute. 13 August 2008. http://asphaltmagazine.com. Accessed
10 March 2014.
Weiland, C.D. and Muench, S.T. (2010). Life Cycle Assessment of Portland Cement Concrete
Interstate Highway Rehabilitation and Replacement, WA-RD 744.4. WSDOT Research
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Report Published by the Washington State Department of Transportation, Office of
Research and Library Services, February 2010.
Wilburn, D.R. and Goonan, T.G. (1998). Aggregates from Natural and Recycled Sources:
Economic Assessments for Construction Applications - A Materials Flow Analysis. US
Geological Survey Circular 1176, US Geological Survey and US Department of the
Interior.
Willett, J.C. (2013). 2011 Minerals Yearbook - Stone, Crushed [AdvanceRelease], US
Geological Survey, March 2013. http://on.doi.gov/lxTlZZo. Accessed 12 March 2014.
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4 AsphaltShingles
4.1 Introduction
Asphalt shingles are more commonly used over other roofing alternatives (e.g., wood, tile, slate,
and metal) due to their relatively lower material and installation cost and their durability (ARMA
2014). Four out of five homes in the US are covered with asphalt shingles (ARMA 2014). More
than 12 MMT (12.5 billion ft2) of shingles are manufactured in the US annually; approximately
65% are used for re-roofing projects and 35% are for new roofs (Brock 2007). The sources of
discarded shingles are post-manufacturing (i.e., pre-consumer) and post-consumer (i.e., from
construction, renovation, and demolition activities). Post-consumer shingles generated from
construction, renovation, and demolition activities are commonly referred to as "tear-offs".
Although old shingles may be overlain by new shingles during re-roofing, most building codes
limit maintenance of one re-roof without removing the existing shingles. The shingles, therefore,
are removed at some point after their service life, which typically is 20 years (NCHRP 2013).
Approximately 15 MMT of asphalt shingles are discarded annually in the United States (US EPA
2020). In 2018, approximately 13 MMT were landfilled, whereas approximately 2 MMT were
recovered for further use (US EPA 2020). Roughly 90% of recovered asphalt shingles are
comprised of tear-off shingle scrap; 10% come from pre-consumer scrap (VANR 1999, Sengoz
and Topal 2005). Pre-consumer scrap tendstobe more uniform compared to tear-offs and typically
consists of shingles and packaging material (e.g., paper or plastic). However, tear-off shingles
contain other roofing debris (e.g., wood, paper, metal) and may experience oxidation given the
material composition and exposure outdoors.
The discarded shingles are transported either to a landfill for disposal or to a processing facility
and eventually used for asphalt pavement production as depicted in Figure 4-1. While not an
established practice in the US, the use of discarded shingles as a fuel source in an industrial
application (e.g., cement manufacturing) has been explored on a limited scale (OCC 2008, Lee
2011). Figure 4-1 identifies the flow of materials and processes that should be considered for LCA
of asphalt shingles EOL management.
4-1
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End-of-life
Product
Removal
MSW
Landfilling
CDD MRF
Combustion
for Cement
Kiln Fuel
Shingle
Recovery,
Processing
Asphalt
Pavement
Production
Figure 4-1. Process flow diagram for end-of-life management of asphalt shingles.
4.2 EOL Management
Several regional composition studies suggest that composition roofing, which consists of asphalt
shingles and attached roofing tar and tar paper, comprises 2.7 to 18.3% (by weight) of CDD
materials received at landfills (CCG2006, CCG2008, CCG2009, CDM2009, CDM2010, and
RWB et al. 2010). Given that 129 MMT of CDD materials were sent to landfills in 2018, this
suggests that approximately 3.5 to 23 MMT of composition roofing were sent to landfills (US EPA
2020).
As identified in Figure 4-1, three potential EOL management pathways for asphalt shingles are
HMA production, combustion in a cement kiln, and disposal. Closed-loop recycling of asphalt
shingles into new asphalt shingles is not a viable recycling option at present due to challenges in
meeting stringent manufacturer feedstock specifications (Snyder 2001, OCC2008). Consequently,
applicable closed-loop recycling option and the associated LCI data (i.e., raw material extraction
and product manufacturing data for asphalt shingles) are not discussed in this report.
The use of asphalt shingles in asphalt paving mix production has stayed relatively steady in recent
years: shifting only from 0.85 MMT in 2017 to 0.83 MMT in 2019 (NAPA 2020). Shingles also
provide aggregate needed for paving mix production. In 2019, paving mix producers recycled
approximately 0.016 MMT of asphalt shingles as aggregates (NAPA 2020). Another small but
growing use of asphalt shingle aggregate is for dust control on rural roads as a replacement for
calcium carbonate. This use is not quantified here, but could be considered if the practice becomes
more widespread.
Asphalt shingles, due to their significant energy content, present a potential opportunity for energy
recovery, including use as a supplemental fuel in cement kilns. Combustion of discarded asphalt
shingles for energy is still under development in the US (OCC 2008, Lee 2011). The US EPA
(2012) used emissions from the combustion of oil and lubricants as a proxy for the emissions from
shingles combustion due to a lack of shingles-specific combustion emissions data and assumed
that shingles combustion in cement kilns would offset emissions associated with refinery fuel gas
combustion for assessing emission factors for the use of shingles in a cement kiln for the WARM
model. As shingles combustion in a manufacturing application is not widely practiced, dataneeded
to develop LCI for this process are lacking and shingles combustion is not further discussed in this
report.
4-2
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Using paving industry recycling estimates and the approximate amount of asphalt shingle waste
produced annually, it is estimated that more than 80% of waste asphalt shingles are disposed of in
landfills, typically within CDD landfills (Sengoz and Topal 2005, CIWMB 2007, CMRA 2007a,
and NAPA 2020). Some landfill facilities may separate incoming loads of asphalt shingles foruse
as road base material for temporary access roads or for truck pads.
4.3 LCI Sources
Peer-reviewed literature, government and private industry publications, and LCA modeling tools
were reviewed to identify available LCI datasets pertaining to asphalt shingles EOL management
processes. If LCI data were not available, process metadata and documentation were reviewed to
evaluate the completeness of the dataset (e.g., emissions categories were included, background
data were used to compile the dataset, and the geographic location and time period of the data were
considered). The primary sources of information used to develop the LCI datasets and information
identified, if available, were reviewed. Table 4-1 presents the data sources reviewed to compile
LCI for shingle EOL management options.
Table 4-1. Life cycle inventory (LCI) data sources for asphalt shingles.
Source
Description
Cochran
(2006)
Cochran (2006) presented diesel energy requirements for asphalt shingles
processing.
US EPA
(2012)
EPA's Waste Reduction Model presents data on GHG emissions associated
with source reduction, transport, recycling, and landfilling (i.e., collection and
placement) of asphalt shingles.
AP-42 (US
EPA 1995)
Provides emission factors of filterable particulate matter (PM), total organic
carbon (TOC), and carbon monoxide (CO) for asphalt roofing manufacturing
processes.
NIST
(2018)
BEES
The National Institute of Standards and Technology Building for
Environmental and Economic Sustainability model enables an economic and
environmental impact comparison across various building materials.
Documentation suggests that the model uses data from SimaPro, US LCI
(2012) and Trumbore etal. (2005) for estimating LCI for asphalt shingles
manufacturing. No LCI specific to EOL management of shingles are provided.
4.4 LCI Related to Disposal
Emissions associated with shingles disposal in a landfill include air emissions from equipment
used for placing shingles in the landfill, emissions associated with landfill construction and
operation, and liquids and gaseous emissions from material decomposition in the landfill
environment. Asphalt shingles are not expected to significantly degrade biologically and,
therefore, are not expected to produce gaseous emissions (US EPA 2012). There are various
sources which provide landfilling emission factors related to equipment use and landfill
construction and operation; however, none are specific to asphalt shingle landfilling. Generalized
landfill construction and operations LCI data were presented in Chapter 2.
4-3
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Asphalt shingles are typically disposed of with other discarded CDD materials and not in a single-
material landfill, or "monofill". Field-scale leachate quality data specific to asphalt shingles
disposal in a landfill are not available. The available laboratory-scale shingles leaching studies
data were reviewed for developing an estimate of liquids emission from shingles disposal in
landfills. Kriech et al. (2002) measured the total and leachable (leached using a TCLP solution)
concentration of 29 PAHs in four primary roofing asphalt samples from a commercial source of
roofing asphalt for built-up roofs. The total PAH results indicated concentrations in the roofing
shingles ranging from 4.0 to 23 mg/kg. None of the 29 PAHs analyzed were detected in TCLP
leachate, suggesting that PAHs do not readily leach from different asphalt materials.
Commercial Recycling Systems (Scarborough, Maine) reported leaching (TCLP) results for
ground shingles (CDRA 2010). The results indicated that the VOCs, Semi-VOCs, PAHs, and
metals were not readily leachable. Low concentrations of some metals (eight RCRA metals) were
reported. Some constituents of ground shingles, most notably PAHs, were reported to exceed the
concentration standards for state de minimis risk levels (Appendix A of Chapter 418 Maine Solid
Waste Rules [MDEP 2012]). The data were not available for further evaluation.
Azah (2011) conducted batch and column leaching tests (SPLP) on a shingle sample collected
from a recycling facility in Florida to assess PAHs leaching from shingles. These data were used
to estimate PAHs emissions from disposal of shingles in CDD materials in landfills. Batch test
data were used for PAHs that were measured above the method detection limits. Batch test
concentrations were multiplied by the total solution volume and divided by the sample mass to
estimate leachability on a per-kilogram-asphalt-pavement basis. The column leaching test (under
saturated conditions) data were used to estimate leaching emission of PAHs (acenaphthene,
phenanthrene, anthracene, and dibenzo(a,h)anthracene) that were not detected in batch leaching
tests. The column leaching test was conducted under saturated conditions until aL:S ratio of 3.08
(L of liquids per kg of shingles) was achieved. PAHs concentrations were measured at liquids-to-
solid ratios (0.62, 1.23, 1.85, 2.47, and 3.08 L of liquid per kg of shingles). The contaminant mass
released between two sampling events was estimated by multiplying the L:S ratio increment from
the previous sampling event to measured PAH concentration. The cumulative leaching amount
was estimated by adding the leaching amount from each sampling interval.
Jang (2000) conducted batch and column leaching tests of several individual CDD materials,
including asphalt shingles using the SPLP extraction fluid to assess leaching of conventional
water-quality parameters. These data were used to estimate liquids emissions of calcium, chloride,
and sodium with asphalt shingles disposal in an unlined CDD landfill; the parameters measured
below detection were not used. The concentrations were multiplied by the total solution volume
and divided by the sample mass to estimate leachable mass per kilogram of shingles. The results
for nitrate and sulfate were not included due to the presence of nitric acid and sulfuric acid in the
SPLP leaching solution, the influences of which are not entirely known.
The asphalt shingles leaching data and energy consumption data from landfill operations were
used to develop an LCI process dataset for disposal of asphalt shingles at an unlined CDD landfill.
Emissions are provided per kilogram "Asphalt shingles, at unlined CDD landfill" flow. Although
the actual liquids emissions are expected to be greater than the estimated liquids emission, as the
material would be subjected to leaching a higher L:S ratio than that used by Azah (2011) and Jang
(2000) for thebatch leaching tests, using these emissions for LCAuntil the total emission estimates
4-4
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become available would be more accurate than excluding liquids emission altogether. Details on
diesel and electricity consumption as a result of landfill operations (included in the"CDD landfill
operations" flow) and details on the calculation of cover soil requirements for placement of asphalt
shingles at an unlined CDD landfill are provided in Chapter 2. The bulk density of asphalt shingles
provided by CCG(2006) was used to estimate cover soil requirements. In the absence of average
nationwide distance data, the site of asphalt shingle removal was assumed to be 20 km from the
CDD landfill disposal site.
4.5 LCI Related to Recycling
Post-consumer shingles generated from construction, renovation, and demolition activities are
commonly referred to as tear-offs. Although old shingles may be overlain by new shingles during
re-roofing, most building codes limit maintenance of one re-roof without removing the existing
shingles. The shingles, therefore, are removed at some point after their service life, which typically
is 20 years (NCHRP2013).
4.5.1 Shingle Processing
Asphalt shingle processing includes removing contaminants (e.g., nails, metal flashing, plywood)
from discarded asphalt shingles and reducing their size. The emissions and energy requirements
for processing equipment should be considered for asphalt shingle processing LCI. Asphalt
shingles, once removed, may be segregated from other CDD materials at the construction site or
they may be commingled with other demolition waste. While some asphalt shingle processing
facilities receive only source-separated asphalt shingles (e.g., from manufacturer waste or roofing
waste), others may not require source separation and sort materials at the facility instead. This
sorting often includes removing non-shingle materials from the discarded shingles.
This process dataset accounts for the emissions and other environmental burdens associated with
the operation of an asphalt shingles processing facility per kilogram of processed material. It
accounts for diesel and heavy equipment utilization; PM emissions associated with shingle
grinding, stockpiling, and unpaved road transport; land occupation; and water consumption (for
PM control from grinding). Grinder diesel consumption is based on a review of grinding equipment
listed in permit documents from 7 sites across 4 states; the average grinder horsepower from these
facilities was used to select a representative piece of grinding equipment and associated diesel
consumption rate. Diesel consumption and equipment utilization for mobile heavy equipment are
based on a detailed equipment inventory performed for a confidential concrete processing
operation (IWCS 2016) and using equipment manufacturer literature and lifetime diesel
consumption values reported for various heavy equipment types in the MOVES database by US
EPA (2014). This process includes all equipment used for loading, grinding, conveying, and
stockpiling asphalt shingles received and handled at asphalt shingle processing facilities.
As detailed below, this dataset assumes that the average nationwide asphalt shingles processing
facility uses diesel to power their grinding equipment; however, amount of fuel consumption will
depend on the quality and type of shingles being processed. This dataset also assumes that the
average nationwide asphalt shingle processing operation uses a grinder rated for processing
approximately 59 tons of shingles per hour.
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The degree and methods of processing asphalt shingles depend on the quality of materials received
and facility design; post-consumer shingles require more intensive sorting due to the higher level
of contamination compared with pre-consumer scrap shingles; post-consumer asphalt shingles
may also have variable properties due to variation in degree of weathering and age among loads
of asphalt shingles (NAHB 1998). Asphalt shingles that arrive at a mixed CDD facility are
typically manually picked from the CDD debris and stockpiled until an appreciable amount of
material has been acquired for processing. Once the shingles have been sorted and all the
undesirable materials removed, the shingles are size-reduced using various grinding and screening
methods to obtain the size necessary for the intended recycling application (CMRA 2007b).
Various grinding and screening methods have been used to grind shingles for recycling, including
shredders, hammer mills, and different screen arrangements. A grinder typically consists of a
loading hopper, feeding drum, grinding chamber, size screen, and an exit conveyor. Water is
sometimes added during shredding to keep the grinder and shingles cool and to control dust
(CMRA 2007b, TRB 2013). The ground-up shingles are typically screened. The fraction that does
not pass through the screen may be used for a process with larger size specifications or they may
be fed back into the grinder forfurther size reduction (Marks and Gerald 1997, VANR 1999, and
Grodinsky et al. 2002). Screenings that are greater than 3/4 of an inch can typically be used as an
aggregate; in most HMA applications the shingles must be reduced to a size smaller than V2 an
inch (CMRA 2007b). Sand may be added to the ground shingles to prevent agglomeration of the
materials during storage (IWCS 2010, TRB 2013), though this process was not accounted for in
the developed dataset.
Publicly available permit documents from state environmental agencies were reviewed for 7
asphalt shingle processing facilities (across 4 states). Information on the facilities permitted
throughput, grinding equipment (i.e., make, model, horsepower), and location was compiled for
the development of this dataset. Grinder fuel consumption was found by first averaging the
horsepower listed for grinders listed at all the facilities. This average horsepower was used to select
an asphalt shingle grinder with approximately the same power rating. Grinder engine fuel
consumption for the identified grinder was utilized from Lakeside Industries, Inc. (2014), which
matched fuel consumption values obtained from an operating facility (IWCS 2016). Six out of the
seven permitted facilities used diesel-powered equipment, and five out of the seven had screening
equipment attached to the grinding equipment and operated by the same engine; this was assumed
to be the typical operational set-up. A mobile equipment inventory and total mobile and screening
equipment fuel consumption values were obtained from an operating facility in Minnesota (IWCS
2016). The fuel consumption for specified engines were obtained from specifications sheets
(Bobcat 2007, Caterpillar 2006). The equipment manufacturer-listed horsepower ratings of each
piece of equipment were used in conjunction with the US EPA (2014) Motor Vehicle Emissions
Simulator (MOVES) database to quantify equipment usage according to fuel consumption.
Water use is assumed to consist of unpaved road dust control based on US EPA (1985), and water
use for dust control equipment (assumed to be the water consumption associated with a single dust
control unit). While a defaultvalue of zero is included in this dataset, additional water consumption
associated with unpaved road dust emissions control was parameterized in the US EPA template
with a range including a maximum value for near 100% dust emissions control efficiency, based
on US EPA (1985).
4-6
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Land occupation is based on a review of aerial imagery of the seven shingles recycling facilities
evaluated in the development of this dataset. Concrete pavement thickness is based on personal
communication with an operating facility in Minnesota (IWCS 2016).
PM emissions from the asphalt shingle grinding unit are based on emission limits for primary rock
crushing from AP-42 (Table 11.19.2-2), as was assumed in a facility technical support document
for an asphalt shingle processing facility in Washington State (Lakeside Industries, Inc. 2014).
This dataset calls inflows from additional data sets to simulate the release of PM as a result of
materials transport over unpaved roads, and as a result of material stockpiling operations. These
PM emissions are accounted for by the "transport; unpaved road" and "stockpiled material; at drop
site" inflows, respectively. The dataset user should modify these upstream processes to account
for a 51 m, one-way length of unpaved road, and a total of 2 material drops (i.e., discharge from
conveyor, loading into haul trucks). Unpaved road length is estimated assuming that the average
unpaved road length at a facility is equal to the one-way distance from the edge to the center of
the site. It was assumed that incoming loads of asphalt shingles provide a negligible contribution
to PM emissions during stockpiling operations; only ground shingles were assumed to result in
PM emissions during stockpiling. These and/or additional process-specific parameters should be
updated for these upstream processes as and whenmore reliable databecome available.
In the absence of additional information, all incoming material was assumed to be transported 20
km to the processing facility. Industry experience suggests that 12' x 56' standard office trailers
are commonly used for administrative purposes at various construction and demolition debris
processing facilities. Office trailer utilization was accounted for in this dataset.
A water balance was conducted assuming that all water consumption for the purposes of dust
control is returned to the atmosphere as water vapor, whereas all water resulting from rainfall onto
paved surfaces is collected and managed as stormwater. Average nationwide precipitation was
estimated from NOAA (2016). Dust control values are calculated using water consumption values
from Dust Control Technology (2016) for a DB-60 dust control device. The calculated value is
comparable to the water consumption value of one active processing facility (IWCS 2016).
Due to an absence of information, specific emissions associated with stormwater discharge to
surface water from these facilities was not included in this dataset. Therefore, the surface water
emissions flow is only considered partially accounted for in this dataset. Liquids emissions to
surface and groundwater from short-term storage of unprocessed and processed shingles in
stockpile should be considered; particulate emissions are unavailable.
4-7
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Transport, single unit truck;
short-haul; diesel powered
Stockpiled material;
at drop site
-Waste transport; at local road; unpaved»
Asphalt shingles;
demolished; at collection
Heavy equipment —
Office trailer ~
Paved area; 4" thick »
PM
Fuel combustion
products
~
Surface
water
~
Water
vapor
Processing of asphalt shingles;
ground; at plant
Land use
Water Diesel Electricity
Wood; recovered from shingles
processing facility; ~
at shingles processing
Metal;
recovered from shingles processing;--*
at shingles processing
Recycled asphalt shingles;
ground; at processing facility
Metal;
-recovered from shingles processing; ferous;n
at shingles processing
Waste residual; to landfill;
at concrete processing facility
Legend
Elementary Flow
Technosphere Flow ^
Not Included
Partially Included
Included
System Boundary
Figure 4-2. Unit process flow diagram for asphalt shingles at the processing facility.
4.5.2 HMA/WMA Production with Asphalt Shingles
Use of asphalt shingles for producing paving mix has significantly increased in recent years due
to the shingles' asphalt content and the increase in asphalt prices. The benefits of using asphalt
shingles for paving mix production include reduced demand for primary asphalt cement and
aggregate, reduced paving mix production cost, and improved resistance to pavement cracking and
rutting due to the reinforcement provided by fibers contained in shingles (Brock and Shaw 1989,
Krygowski 1993, Ali et al. 1995, Button et al. 1995, NAHB 1998, Foo et al. 1999, Mallick 2000,
and Sengoz and Topal 2005). Several state departments of transportation (e.g., Georgia,
Minnesota, Montana, and South Carolina) have specifications allowing from 3 to 8% of HMA to
be replaced by tear-off shingles; other laboratory experiments have acknowledged that up to 7%
of HMA material can be replaced by tear-off shingles without adverse effects (Mallick 2000, OCC
2008).
Another potential environmental concern with the recycling of asphalt shingles is the emission of
PAHs during the production of HMA. Since asphalt is a mixture of paraffinic and aromatic
hydrocarbons, heating of asphalt can result in the emission of PAHs (ARMA 1998; US EPA 2000;
Lee et al. 2004). PAHs are one of the major classes of air pollutants emitted from HMA facilities
(US EPA 2000). While the quantity of PAH emissions from HMA facilities has been fairly well
documented, the impact of using recycled asphalt shingles on PAH emissions is not well
understood. Currently, there are no available data to suggest that emissions of PAHs during HMA
production with recycled asphalt shingles would be different from emissions from production
without asphalt shingles. TRB (2013) reported that HMA plants may require more frequent
cleaning and adjustments may need to be made to temperature settings to melt the more hardened
shingle asphalt, suggesting greater energy demand for HMA production using asphalt shingles.
However, data are not currently available to estimate the additional energy and material demand
associated with using shingles for HMA production.
4-8
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Asphalt and aggregates constitute approximately 19-36% and 20-38% (by weight) of shingles,
respectively (NAHB 1998, CIWMB 2007). Therefore, asphalt shingles supplement the use of
primary asphalt and aggregates in paving mix production. The US EPA (2012) assumed that
asphalt shingles provide 22% and 38% (by weight of shingles) of asphalt and aggregate,
respectively, for estimating emission offsets associated with avoiding primary asphalt and
aggregate production. The US EPA (2012) also assumed a loss of 7.2% of the product during
recycling. Based on the procedure used by the US EPA (2012), recycling 1 kg of asphalt shingles
forHMA production is estimated to replace 0.2 kg and 0.35 kg of primary asphalt and aggregates,
respectively. It is assumed that there are no differences in emissions associated with use of asphalt
shingles in HMA production (e.g., particulate matter, liquid, and the energy requirements) as
compared to those from the use of primary materials in the production process. It is assumed that
there would be additional avoided emissions resulting from the prevention of the transport of
primary aggregates and asphalt to the HMA plant; additional details on the transport distances for
these primary materials are provided in Chapters 2 and 3.
4.6 Data Gaps and Future Opportunities
Table 4-2 summarizes the type of data presented by various sources reviewed for compilation of
asphalt shingle EOL management LCI. Only WARM (US EPA 2012), Athena (2001), AP-42 (US
EPA 1995), MSW-DST, Azah (2011), Wilburn and Goonan (1998), and GaBi provide information
with respect to US-based processes. Some sources incorporate data from other sources listed in
Table 4-2. For example, the US EPA (2012) used data from NREL database and Cochran (2006).
As shown in the table, many sources present only part of the data/information needed for compiling
LCI. For example, WARM uses only GHG emissions associated with equipment fuel consumption
to estimate landfill emission. Similarly, Ecoinvent only has partial landfill leachate emissions data
because leachate from CDD landfills is not considered.
A majority of LCI information available on asphalt shingles pertains to the manufacturing aspects
of the life cycle. Only limited EOL-specific LCI are available. Based on a review of the available
information, the following datagaps were identified for compilation of a more comprehensive LCI
dataset for asphalt shingles EOL management:
1. Data pertaining to shingles use in pavement mix production. The use of asphalt shingles in
HMA production has shown a significant increase in recent years. Similar to the US EPA
(2012), the LCI presented in the report for shingles use for HMA production are based on the
assumption that shingles do not impact emissions associated with HMA production. Data are
not available to estimate the additional energy and material demand and emissions associated
with using shingles for HMA production. Future research should consider assessing the impact
of shingles on energy and materials input and emissions associated with HMA production. The
impact of using shingles in HMA on the quality and service life of the pavement should also
be assessed. The discussion presented in this chapter focused on recycling asphalt shingles in
HMA. However, there has been significant growth in the use of WMA in recent years. There
are insufficient data pertaining to the use of shingles in WMA.
2. Data pertaining to asphalt shingles use as fuel source in industrial applications. The LCI
data for asphalt shingles uses as a supplemental fuel in cement kiln are lacking due to the rarity
of this practice in the US (OCC 2008). The cement industry has experimented with the use of
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asphalt shingles in cement kilns (e.g., Lafarge cement plant, Brookfield,Nova Scotia, and St.
Mary's Cement, Charlevoix, Michigan (Lee 2011)); however, there is very little research data
available on the subject. A Department of Energy-sponsored project conducted in 2007
investigated the feasibility of using asphalt shingles (pre-consumer and post-consumer) in the
manufacture of cement and in circulating fluidized bed boilers (OCC 2008). OCC (2008)
presented emissions from the combustion of the shingles and the potential impacts on the
quality of the products and the cement kiln dust (CKD); however, the data were interpreted as
being rudimentary. None of the existing US-specific data sources listed in Table 4-2, except
WARM (US EPA 2012), included emissions associated with shingles in cement kiln
manufacturing. The WARM model does consider GHGs associated with shingle combustion
in a cement kiln; however, GHGs were estimated using lubricants as a proxy. Future research
into improved combustion data, as well as potential recycling applications, is recommended.
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Table 4-2. Life cycle inventory data sources for asphalt shingles. "P" indicates partial coverage of the data, "X" indicates full coverage, and
indicates no data.
Process
WARM
MSW-
DST
Ecoinvent
GaBi
Athena
(2001)
AIM 2
US LCI
(2012)
Wilburn and
Goonan
(1998)
Jang
(2000)
Azah
(2011)
Cochran
(2006)
Aggregate
Production
P
-
X
X
-
P
X
P
-
-
-
Asphalt Production
P
-
-
X
X
P
X
-
-
-
-
Transport
P
-
-
-
X
-
-
P
-
-
-
Landfill Construction
& Operation
P
X
-
X
-
-
-
-
-
-
-
Landfill Leachate
Emissions
-
-
P
-
-
-
-
-
P
P
-
Shingle Processing
P
-
-
-
-
-
-
-
-
-
P
HMA Production
P
-
-
-
-
-
-
-
-
-
-
4-11
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4.7 References
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Asphalt Concrete Mixtures Containing Reclaimed Roofing Materials. In: Proceedings of
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ARMA (1998). Polyaromatic Hydrocarbon Emissions from Asphalt Processing and Roofing
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Azah, E. M. (2011). The Impact of Polycyclic Aromatic Hydrocarbons (PAHs) on Beneficial Use
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CDRA (2010). Case Study of Successful Asphalt Shingle Recycling: Commercial Recycling
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TRB (2013). Recycled Materials and Byproducts in Highway Applications. Volume 8:
Manufacturing and Construction Byproducts. NCHRP Synthesis 435, National
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Owens, R. (2005). Emissions Factors for Asphalt-Related Emissions in Roofing
Manufacturing. Environmental Progress 24 (3), 268-278.
US EPA (1985). Dust Control at Hazardous Waste Sites. Hazardous Waste Engineering Research
Laboratory, Cincinnati, OH. EPA 540 2-85 003.
US EPA (1995). AP-42, Fifth Edition, Volume I, Chapter 11: Mineral Products Industry, Section
11.2 - Asphalt Roofing. US Environmental Protection Agency. http://Lusa.gov/lr8s2WR.
Accessed 8 July 2014.
US EPA (2000). Hot Mix Asphalt Plants Emission Assessment Report. EPA-454/R-00-019,
Emissions Monitoring and Analysis Division, Office of Air Quality Planning and
Standards, US Environmental Protection Agency, Research Triangle Park, NC, USA.
December 2000.
US EPA (2019). Waste Reduction Model (WARM) Tool: User's Guide. Version 15, May 2019.
Available: https://www.epa.gov/warm/documentation-chapters-greenhouse-gas-emission-
energy-and-economic-factors-used-waste-reduction.
US EPA (2013). Analysis of Recycling of Asphalt Shingles in Pavement Mixes from a Life Cycle
Perspective. EPW07020, 24 July 2013.
US EPA (2014). Motor Vehicle Emission Simulator (MOVES): User Guide for MOVES 2014.
EPA-420-B-14-055. July 2014.
US EPA (2020). Advancing Sustainable Materials Management: 2018 Fact Sheet. Office of
Resource Conservation and Recovery. EPA 530-F-20-009. Washington, DC.
US LCI (2012). US Life Cycle Inventory Database. National Renewable Energy Laboratory.
http://www.nrel.gov/lci/. Accessed 20 February 2014.
VANR (1999). Recycled Asphalt Shingles in Road Application: An Overview of the State of
Practice. Vermont Agency of Natural Resources, September 1999.
https://dec.vermont.gov. Accessed 24 March 2014.
Wess, J.A., Olsen, L.D., and Sweeney, M.H. (2004). Asphalt (Bitumen). Concise International
Chemical Assessment Document 59, A Report Prepared by Joann Wess, Dr. Larry Olsen,
and Dr. Marie Sweeney for the National Institute for Occupational Safety and Health,
World Health Organization, Geneva, Switzerland.
Wilburn, D.R. and Goonan, T.G. (1998). Aggregates from Natural and Recycled Sources:
Economic Assessments for Construction Applications - A Materials Flow Analysis. US
Geological Survey Circular 1176, US Geological Survey and US Department of the
Interior.
Willett, J.C. (2013). 2011 Minerals Yearbook - Stone, Crushed [AdvanceRelease], US Geological
Survey, March 2013. http://on.doi.gov/lxTlZZo. Accessed 12 March 2014.
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5 Gypsum Drywall
5.1 Introduction
Gypsum drywall (also referred to as gypsum board, wallboard, or plasterboard), typically
manufactured and sold as 4-ft-by-8-ft or 10-ft-by-12-ft sheets or panels, is widely used as an
interior wall and ceiling finishing in residential, commercial, and institutional structures.
Although there are a variety of drywall products, including many varieties of fire-resistant and
water-resistant products, ^-inch-thick regular and Type X (fire-resistant) gypsum boards
combined constitute over 80% (by weight) of the prefabricated gypsum products (Crangle 2014).
In2013, approximately 19.5 billion ft2 of gypsum drywall products were sold in theUS (Crangle
CDD MRF
Grinding,
Screening
End-of-life
Product
Removal
MSW
Landfilling
2014).
CDD
Landfilling
Figure 5-1. Process flow diagram for end-of-life management of gypsum drywall. MRF = material
recovery facility, depicts the various gypsum drywall EOL processes.
Figure 5-1. Process flow diagram for end-of-life management of gypsum drywall. MRF = material
recovery facility.
5.2 EOL Management
Approximately 13.8 MMT of waste gypsum drywall was processed in the United States in 2018,
of which nearly 12 MMT was disposed of in landfills and 1.9 MMT was recovered through CDD
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processing facilities (US EPA 2020). The California Waste Management Board (2007) reports
fractions of (California) waste drywall which are produced from new construction, demolition,
manufacturing, and renovation activities (Figure 5-2). Distinguishing between these waste drywall
sources is important because the drywall quality differs substantially: manufacturing waste and
post-construction drywall cutouts are relatively clean, while demolition and renovation drywall
typically has contaminants such as paint, joint compounds, nails, and any other material that may
have been applied to the drywall during installation and use. Drywall which is recovered for
recycling is typically taken to drywall processing facilities where processing removes
contaminants and separates gypsum from the paper backing (WRAP 2008).
Figure 5-2. Sources of gypsum drywall discarded for processing or final disposal in California
(CIWMB 2007). California data were used due to data availability and a growing CDD sector.
Due to its relative high quality (e.g., absence of nails, paint), the drywall discarded from drywall
manufacturing and new construction accounts for the bulk of the drywall waste stream recycled
(Venta 1997, WRAP 2008, US EPA 2012). Gypsum drywall that is not recycled (representing the
majority of waste drywall) is disposed of in landfills. Drywall is estimated to account for
approximately 8% of landfilled CDD materials (Barnes 2000, Sandler 2003, CCG 2006, CDM
2009, RWB et al. 2010). MSW landfills may limit the amount of drywall accepted, or the drywall
may be banned from disposal altogether, often due to potential generation and release of hydrogen
sulfide gas. While specific practices vary, most demolition drywall is still landfilled.
Table 5-1 presents the processes included in EOL management of gypsum drywall. For all
applications in which discarded drywall is processed to produce gypsum (referred to herein as
recycled gypsum) and used as a substitute for primary gypsum, production of primary gypsum was
identified as an end boundary for LCI compilation. The production and use of recycled gypsum
avoids primary gypsum production and thus avoids its associated emissions. Emissions
downstream of the primary gypsum production (e.g., those associated with drywall manufacturing
from primary gypsum) were assumed to occur regardless of whether downstream processes used
primary or recycled gypsum unless the available data indicated otherwise.
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Table 5-1. End-of-life management processes for gypsum drywall.
Process
LCA considerations
Unlined CDD
landfilling of gypsum
drywall; at unlined
CDD landfill
This process entails placing and compacting drywall and long-term
physical, chemical, and biological decomposition of drywall in a
landfill.
MSW landfilling of
gypsum drywall; at
MSW landfill
This process models materials and energy inputs from disposal of
gypsum drywall in an MSW landfill environment. It also includes the
environmental burdens associated MSW landfill construction, operation
and closure/post-closure care
Processing of
gypsum; milled; at
drywall processing
facility
Processing includes grinding waste drywall and screening paper and
other contaminants that may be present in the drywall waste to produce
ground-up gypsum.
5.3 LCI Sources
Peer-reviewed literature, government and private industry publications, and various LCA
modeling tools were reviewed to identify available LCI datasets pertaining to gypsum drywall
EOL management processes. Table 5-2 lists data sources reviewed to compile LCI presented in
this chapter. If LCI data were not available, process metadata and documents were reviewed to
evaluate the completeness of the dataset (e.g., emissions categories, background data, and
geographic location and time period of the data). Primary sources were identified wherever
possible.
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Table 5-2. Overview of life cycle inventory data sources associated with gypsum drywall.
Source
Description
Athena
(2011)
Athena Institute developed this report for the Gypsum Association. The report
presents cradle-to-gate LCI of 1/2-inch regular and 5/8-inch Type X gypsum
drywall. It also presents US-specific LCI for raw gypsum extraction, gypsum paper
manufacture, and drywall assembly based on the data from seven quarries/mining
sites, three paper plants, and 17 drywall board plants to develop the LCI datasets
with the intent of uploading them to the US LCI (2012) database.
Venta
(1997)
The report presents cradle-to-gate LCI of different types of gypsum boards and
associated finishing products. These LCI appear to be used by the Athena Impact
Estimator for Buildings, an LCA model which evaluates the environmental impacts
of structures created from a variety of building materials. The data used for LCI
development were primarily specific to Canada. US-specific data such as
distribution of different type of boards produced were used for developing the LCI.
Cochran
(2006)
Cochran provided an estimate of the energy requirements of equipment for
processing gypsum drywall for recycling (in MJ/hour) based on a survey of
equipment manufacturers.
5.4 LCI Related to Disposal
The emissions from gypsum drywall landfill disposal result from operating landfill equipment
during material and cover soil compaction and placement, including both fuel-related and pre-
combustion emissions, as well as emissions associated with the physical, chemical, and biological
decomposition of gypsum drywall in landfill. The exposure to precipitation or other liquids (e.g.,
landfill leachate) is expected to result in leaching emissions. The liquids and gaseous emissions to
the environment are expected to depend on the biogeochemical environment of the landfill (e.g.,
MSW landfill, CDD landfill), as well as in-place environmental controls employed at the landfill.
The disposal emissions of drywall are included in the landfilling datasets for CDD materials
discussed in Chapter 2. Figure 2-4 and Figure 2-7 show the flows included for landfilling of
gypsum drywall in CDD and MSW landfills, respectively.
The EPA Waste Reduction Model (WARM) estimates landfilling emission factors related to
drywall, which include GHG emissions from transportation (including heavy equipment) and the
operation of the landfill, carbon sequestration, and methane generation from biological
decomposition of the facing and backing paper (US EPA 2019). WARM does not consider liquid
emissions from landfills. It does include fugitive methane emissions associated with drywall paper
decomposition from drywall disposed of in a CDD landfill site without gas collection and control
systems (GCCS). WARM uses methane generation potential reported by Staley and Barlaz (2009)
to assess methane generation from drywall. Staley and Barlaz (2009) estimated the methane
generation potential of drywall by multiplying the methane generation potential of old corrugated
cardboard (OCC) and Kraft bags reported by Eleazer et al. (1997) by 0.1 to account for the relative
mass of paper (approximately 10% of drywall by mass based on the data reported by National
Gypsum Company (2008)) and adjusted for the paper-specific moisture content (6% by mass
reported by Tchobanoglous et al. (1993)). The drywall methane generation potential was estimated
to be 15.2 m3 per dry MT of drywall. WARM assumes a methane oxidation rate of 10% in the
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landfill cover based on work by Czepiel et al. (1996) to estimate fugitive emissions from methane
generation. Additional details on the information used to estimate methane and carbon dioxide
emissions from the disposal of drywall in a CDD and MSW landfill are presented in Section 2.6.8.
The CDD landfill methane and carbon dioxide emissions are estimated to be 0.010 kg and 0.034
kg, respectively.
WARM considers only methane emissions from drywall decomposition in anaerobic landfill
environments for estimating GHG impacts. The production of hydrogen sulfide from the biological
decomposition of organic matter in anaerobic conditions in the presence of dissolved sulfate
(primarily from gypsum) has been reported to be a major environmental concern associated with
gypsum drywall disposal in landfills (Jang 2000, Xu 2005). Several factors, including moisture
content, organic content, pH, and temperature, may contribute to the production of hydrogen
sulfide in landfills (Elsgaard et al. 1994, Knoblauch and Jorgensen 1999, Koschorreck 2008).
Several studies have indicated that the amount of organic matter present in CDD landfills, although
significantly lower than in MSW landfills, is not a limiting factor for hydrogen sulfide production;
the paper backing on drywall is sufficient to sustain a viable microbial community that produces
hydrogen sulfide (Hardy Associates 1978, Townsend 2002, New Hampshire Department of
Environmental Services, 2004). Tolaymat et al. (2013) has reported a decay rate constant for
drywall decomposition and the associated hydrogen sulfide production. Xu (2005) has assessed
the impact of different cover materials in reducing hydrogen sulfide emissions from CDD landfills
and reported the hydrogen sulfide concentration in gaseous emissions from drywall decomposition
based on laboratory experiments.
Anderson et al. (2010) evaluated hydrogen sulfide emissions and sulfur content of CDD fines from
nine landfills in the United States. The rate of emission was determined to be 5,360 ft3 of hydrogen
sulfide per ton of sulfur. From stoichiometry, sulfur represents 18.6% of the gypsum (by weight).
Assuming that drywall is comprised of 92% gypsum (Marvin 2000), this equates to 0.041 kg of
hydrogen sulfide released per kg of drywall disposal (using a density of about 1.42 g/L of hydrogen
sulfide at 20° C and 1 atm pressure). Plaza et al. (2007) estimated the attenuation of hydrogen
sulfide by different landfill cover materials, including clay and sandy soils; the average attenuation
of these two soils was 47.5%. Using the same assumptions listed in Chapter 9, and in the absence
of other US data, the average estimated hydrogen sulfide release rate based on these studies
(including average cover soil removal efficiency) is 0.021 kg hydrogen sulfide per kilogram of
drywall disposed of in a CDD landfill.
Jang (2000) conducted batch and column leaching tests of several individual CDD materials,
including new gypsum drywall, using Synthetic Precipitation Leaching Procedure (SPLP)
extraction fluid (and US EPA SW-846 Method 1312). Drywall material was cut into square pieces
approximately 5 cm on each side. Results from these experiments were used to estimate liquids
emissions that would result from disposing of gypsum drywall in an unlined landfill. Unlined
landfill disposal may occur due to the classification of CDD type wastes as more chemically inert
than MSW components; in some areas, the term inert material landfill is used for CDD sites (Doka
2003). Batch test data were used for constituents with concentrations measuring above detection
limits due to a greater L:S ratio (20:1 vs. 5.3:1 for column tests); higher L:S ratios are generally
capable of leaching a higher quantity of the total constituent mass from the solid material (i.e.,
gypsum drywall). Batch test concentrations were multiplied by the total solution volume, and then
divided by the sample mass to estimate leachability on a per-kilogram-drywall basis.
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Column leaching test data were used to estimate the leaching emissions of potassium and
magnesium, as these were not detected in batch leaching tests. The column leaching test entailed
percolation of approximately 160 L of SPLP extraction fluid through a 30-cm-diameter polyvinyl
chloride column containing 30 kg of drywall over a 3-month period; an L:S ratio of 5.3 was
achieved in this test. The amount of potassium and magnesium leached between two sampling
events was estimated by multiplying the leachate volume collected since the previous sampling
event with the measured concentration. The cumulative leaching amount was estimated by adding
the leaching amount for each sampling interval.
The gypsum drywall leaching data and energy consumption data from landfill operations were
used to develop an LCI process dataset for disposal of drywall at an unlined CDD landfill.
Emissions are provided per kilogram "Gypsum drywall, at unlined CDD landfill" flow. Although
the actual emissions are expected to be greater than the estimated liquids emission as the material
would be subjected to leaching a higher L:S ratio than used by Jang (2000) for the batch leaching
test, using these emissions for LC A until the total emission estimates become available would be
more accurate than excluding liquids emission altogether. The methane emission rate estimate
presented earlier in the section is included in the dataset. Details on the diesel and electricity
consumption included in the "CDD landfill operations flow" and on how the cover soil
requirement was determined are provided in Chapter 2. The density of bulk gypsum drywall for
the cover soil requirement estimation was provided by CCG (2006). In the absence of average
nationwide distance data, the site of gypsum drywall removal was assumed to be located 20 km
from the CDD landfill disposal site. While actual distances may vary, this assumption was used to
represent a typical distance between the drywall removal site and the nearest landfill.
5.5 LCI Related to Recycling
Drywall recycling has been gaining momentum as landfills place more restrictions on drywall
disposal due to actual or potential odor issues. Discarded drywall sources include off-spec drywall
generated at the gypsum board plant, as well as scraps from new construction or renovation, or
material produced from structural demolition. Due to quality issues, the majority of waste gypsum
used for recycling comes from the plant's off-spec material and scraps from construction and
renovation projects. At drywall manufacturing plants, scrap drywall generated from off-spec
boards is recycled back into the manufacturing process for new drywall material. The drywall
generated from demolition activities is not recycled as frequently as drywall from other sources
due to possible contamination from other CDD materials such as nails, paint, and joint compound
(Venta 1997, Cochran 2006).
5.5.1 Recovered Drywall Processing
This process accounts for the emissions and other environmental burdens associated with the
operation of a drywall processing facility per kilogram of processed material. It includes
electricity, diesel, and heavy equipment utilization; particulate matter (PM) emissions associated
with shingle grinding, stockpiling, and unpaved road transport; land occupation; and water
consumption (for PM control from grinding). Much of the data used to construct this dataset is
based on personal communication with a drywall processing operation (IWCS 2016), which
provided data on facility material throughput, equipment inventory, electricity, diesel, and water
consumption, as well as facility layout including paved area, and covered building area. A primary
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assumption of this dataset is that this facility is representative of the average nationwide drywall
processing facility.
In addition to the equipment inventory, equipment use was reported using equipment manufacturer
lifetime diesel consumption values of various heavy equipment types in the MOVES database (US
EPA 2014). This includes all equipment necessary for loading, milling, conveying, and stockpiling
gypsum received and handled at a drywall processing facility. Electric-powered stationary
equipment is assumed to be the primary equipment used, with infrequent use of stationary diesel
equipment included as operating 50% of the time. Figure 5-3 shows the flows that are included in
this dataset.
Land use figures were based on average property footprints derived from aerial images of seven
facilities representing three states. The footprint was divided by the capacity and assumed facility
lifespan to determine land use per kg of drywall processed.
All incoming material was assumed to be transported 20 km to the processing facility. While actual
distances may vary, this assumption was used to represent a typical transport distance. Industry
experience suggests that 12' x 56' standard office trailers are commonly used for administrative
purposes at various construction and demolition debris processing facilities. It is assumed that
processing occurs at a covered facility.
A water balance was conducted assuming that all water consumption for the purposes of dust
control is returned to the atmosphere as water vapor, whereas all water resulting from rainfall onto
paved surfaces is collected and managed as stormwater. Average nationwide precipitation was
estimated from NOAA (2016) to estimate stormwater volumes. Due to lack of data, emissions
associated with stormwater discharge to surface waterfrom these facilities was not included in this
dataset. Therefore, the surface water emissions flow is only considered partially accounted for in
this dataset.
Diesel consumption for diesel equipment is based on total diesel consumption (IWCS, 2016) and
is allocated to each piece of equipment using the US EPA (2014) Motor Vehicle Emissions
Simulator (MOVES) database to quantify equipment usage according to fuel consumption. The
loaders are reduced by a use factor since only one loader is operated at a time. Likewise, the diesel
trommel equipment flows are reduced by a use factor since the example facility did not employ it
as often as the electric trommels (IWCS 2016).
This process is assumed to take place at a facility with a covered area for processing which requires
a foundation and would likely include paved access to the covered area. Based on experience with
operating recovery facilities, abundant spraying of water for dust control is assumed to be
sufficient to prevent dust emissions (IWCS 2016).
PM emissions for processing equipment is based on PM emission limitations of a mill at a plaster
processing facility in Wales, UK due to a lack of available information on US facilities
(Environment Agency 2010). Total PM emissions are distributed based on category 4 mineral
grinding PM distribution values given in AP-42 (US EPA 1996).
This dataset calls inflows from additional data sets to simulate the release of PM as a result of
material stockpiling operations. These PM emissions are accounted for by the "stockpiled material;
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at drop site" inflows. The dataset user should modify these upstream processes to account for a
total of 3 material drops (i.e., unloading at facility, discharge from conveyor, loading into haul
trucks). These process-specific parameters should be updated to include upstream processes as
more reliable data become available.
t
PM
Transport, single-unit truck;
short-haul; diesel powered
Stockpiled material;
at drop site
-Waste transport; at local road; unpaved
Gypsum drywall; demolished; _
at collection
Heavy equipment
Office trailer
Fuel combustion Surface
products
water
i
Water
vapor
L_
Processing of gypsum; milled;
at drywall processing facility
Paper; unspecified scrap;
at drywall processing facility
Gypsum; milled;
at drywall processing facility
Metal; ferrous scrap;
at drywall processing facility
Solid waste; to landfill;
at drywall processing facility
Land use
Water Diesel Electricity
Legend
Technosphere Flow
| Not Included
| Partially Included
I Included
System Boundary
Figure 5-3. Unit process flow diagram for gypsum recovered from drywall processing.
5.6 Data Gap Analysis and Opportunities for Additional LCI Data
Table 5-3 summarizes the type of data presented by various sources reviewed for compilation of
drywall EOL management LCI.
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Table 5-3. Overview of life cycle inventory data sources associated with gypsum drywall. "P"
represents partial coverage of the required data, while "X" represents full coverage, and
indicates no data. Section 5.6 describes the levels of data coverage in more detail.
Process
Venta
(1997)
Athena
(2011)
US EPA (2019) -
WARM
Cochran
(2006)
Gypsum Mining
X
X
P
-
Paper Backing
Production
X
X
-
-
Drywall
Manufacturing
X
X
X
-
Landfilling
-
-
P
-
Drywall Grinding and
Paper Screening
-
-
X
P
Transportation
-
X
X
-
A majority of LCI information available on drywall pertains to the manufacturing aspects of the
life cycle. Only limited EOL-specific LCI are currently available, though newer inventories are
being reviewed, such as Athena (2020). Based on currently available sources, the following data
gaps were identified for compiling a more comprehensive LCI dataset for drywall EOL
management:
1. Long-term leachable emissions from drywall placed in a landfill. As described earlier, the
liquid emissions presented in this study are based on SPLP tests, which simulate leaching from
land application or disposal in an unlined CDD landfill. The batch leaching data used for
estimating liquid emissions correspond to an L:S ratio of 20 and, therefore, do not represent
complete liquid emission. As gypsum drywall is typically disposed of with other discarded
materials and not disposed of in a monofill, field-scale leachate data specific to gypsum drywall
disposal in a landfill were not available. The liquid emissions from gypsum drywall placement
in CDD and MSW landfills would therefore need to be based on laboratory-scale studies
simulating long-term liquids emissions.
2. Long-term gaseous emissions from drywall placement in landfills. The disposal of gypsum
drywall in CDD and MSW landfills is expected to produce methane and hydrogen sulfide.
Staley and Barlaz (2009) used the methane generation potential of OCC and Kraft paper as a
proxy for the methane generation potential of gypsum paper while WARM used the methane
generation potential estimate provided by Staley and Barlaz (2009). Further studies would be
needed to improve the modeling of long-term gaseous emissions for inclusion in these EOL
datasets.
5.7 References
Anderson, R., Jambeck, J.R., and McCarron, G.P. (2010) Modeling of Hydrogen Sulfide
Generation from Landfills Beneficially Utilizing Processed Construction and Demolition
Materials. A Report Prepared for the Environmental Research and Education Foundation,
Alexandria, VA. February 2010.
Athena Sustainable Materials Institute (2011). A Cradle-to-Gate LCA of 1/2" Regular and 5/8"
Type X Gypsum Drywall. December 2011.
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Athena Sustainable Materials Institute (2020). A Cradle-to-Gate Life Cycle Assessment of 1/2"
Lightweight and 5/8" Type X Conventional Gypsum Board for the USA and Canadian
Markets. April 2020. Available: https://gypsum.org/download/14369.
Balazik, R. (1995). Gypsum. USGS Minerals Yearbook 1995. United States Geological Survey,
US Department of the Interior.
Barnes, A.H. (2000). Feasibility of Recycling Scrap Gypsum Drywall from New Construction
Activities in Florida. Master's Thesis, University of Florida, Gainesville, FL, USA. May
2000.
CCG (2006). Targeted Statewide Characterization Study: Detailed Characterization of
Construction and Demolition Waste. A Report Prepared by Cascadia Consulting Group for
the California Environmental Protection Agency Integrated Waste Management Board.
June 2006.
CDM (2009). Illinois Commodity/Waste Generation and Characterization Study. A Report
Prepared by Camp, Dresser, and McKee (CDM) for the Illinois Department of Commerce
and Economic Opportunity and contracted by the Illinois Recycling Association. 22 May
2009.
California Integrated Waste Management Board (CIWMB) (2007). Construction and Demolition
Recycling: Wallboard (Drywall) Recycling.
http://www.calrecycle.ca.gov/conDemo/Wallboard. 26 July 2007.
Cochran, K.M. (2006). Construction and Demolition Debris Recycling: Methods, Markets, and
Policy. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Crangle, R.D. (2013). Gypsum. Produced by the US Geological Survey, United States Department
of the Interior. US Geological Survey Minerals Yearbook-2011. June 2013.
Crangle, R. (2014). Mineral Commodity Summaries 2014. Produced by the US Geological Survey,
United States Department of the Interior. Mineral Commodities Profile, Gypsum Industry
Profile. February 2014.
Czepiel, P.M., Mosher, B., and Harriss, R.C. (1996) Quantifying the Effect of Oxidation on
Landfill Methane Emissions. Journal of Geophysical Research, 101, 16, 721-16, 729.
Doka, G. (2003). Life Cycle Inventories of Waste Treatment Services, Part III: Landfills,
Underground Deposits, Landfarming. Ecoinvent Report No. 13, Swiss Centre for Life
Cycle Inventories, Diibendorf, December 2003
Doka, G. (2009). Life Cycle Inventories for Waste Treatment Services, Part V: Building Material
Disposal. Ecoinvent Report No. 13, Swiss Centre for Life Cycle Inventories, Dubendorf.
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Eleazer, W.E., Odle, W.S., Wang, Y.S., and Barlaz, M.A. (1997). Biodegradibility of Municipal
Solid Waste Components in Laboratory-Scale Landfills. Environmental Science &
Technology, 31, 911-917.
Elsgaard, L., Prieur, D., Mukwaya, G.M., and Jorgensen, B.B. (1994). Thermophilic Sulfate
Reduction in Hydrothermal Sediment of Land Tanganyika, East Africa. Applied
Environmental Microbiology, 60 (5), 1473-1480.
Environment Agency (2010). Notice of Variation and Consolidation with Introductory Notes. For
Siniat Limited, Bristol BS20 0NF.
https://www.gov.uk/government/uploads/system/uploads/attachment_data/file/339413/Va
riation_and_consolidation_notice.pdf Accessed May 27th, 2016.
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Golder Associates. (2007). Life Cycle Assessment of Plasterboard, PBD014, Landfill Emissions
Inventory. A Report Prepared by Golder Associates (UK) Ltd. for Environmental
Resources Management Ltd., October 2007.
Hardy Associates Ltd. (1978). Investigation of means to control sulphide production in drywall
landfill disposal operations. Gypsum Subcommittee of the Lower Mainland Refuse Project,
19 December 1984.
Isaac, C., Morris, A. (2012). Evaluation of Metal Leaching from Contaminated Soils, FGD, and
oth Coal Combustion Byproducts in Ruse Scenarios. Mined Gypsum LEAF Methods 1313
and 1316. Draft Report.
IWCS (2016). Personal communication with confidential client and Innovative Waste Consulting
Services.
Jang, Y.C. and Townsend, T.G. (2003) Effect of Waste Depth on Leachate Quality from
Laboratory Construction and Demolition Debris Landfills. Environmental Engineering
Science, 20 (3), 183-196.
Jang, Y. (2000). A Study of Construction and Demolition Waste Leachate from Laboratory
Landfill Simulators. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Kellenberger, D., Althaus, H.-J., Jungbluth, N., and Kiinniger, T. (2004). Life Cycle Inventories
of Building Products. Final Report ecoinvent 2000 Published by Swiss Centre for LCI,
EMPA-DU, Diibendorf, CH.
Kellenberger, D., Althaus, H., Kiinniger, T., and Lehmann, M. (2007). Life Cycle Inventories of
Building Products. Ecoinvent Report No. 7, Swiss Centre for Life Cycle Inventories,
Diibendorf. December 2007.
Knoblauch, C. and Jorgensen, B.B. (1999). Effect of Temperature on Sulphate Reduction, Growth
Rate and Growth Yield in Five Psychrophillic Sulphate-Reducing Bacteria from Arctic
Sediments. Environmental Microbiology, 1 (5), 457-467.
Koschorreck, M. (2008). Microbial sulphate reduction at a low pH. FEMS Microbial Ecology, 64,
329-342.
Marvin, E. (2000). Gypsum Wallboard Recycling and Reuse Opportunities in the State of
Vermont. Waste Management Division, Vermont Agency of Natural Resources. August
2000.
National Gypsum Company (2008). National Gypsum Company: Drywall, cement board, gypsum
board.www.nationalgypsum.com. Accessed January 9, 2008.
National Gypsum (2013). Gypsum Board Frequently Asked Questions.
https://www.nationalgypsum.com. Accessed April 2014.
New Hampshire Department of Environmental Services. (2004). Interim best management
practices Approved for Co-Disposal of Construction and Demolition Process Fines with
Municipal Solid Waste. November 22, 2004.
NOAA (2016). National Centers for Environmental Information, State of the Climate: National
Overview for Annual 2015, published online January 2016.
http://www.ncdc.noaa.gov/sotc/national/201513.
Plaza, C., Xu, Q., Townsend, T.G., Bitton, G., and Booth, M. (2007). Evaluation of Alternative
Landfill Cover Soils for Attenuating Hydrogen Sulfide from Construction and Demolition
Debris Landfills. Journal of Environmental Management, 84 (3), 314-322.
R.W. Beck, Inc., CCG, and IWCS (2010). Statewide Construction and Demolition Debris
Characterization Study. A Report Prepared by R.W. Beck, Inc., Cascadia Consulting
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Group, Innovative Waste Consulting Services, LLC, for Georgia Department of Natural
Resources, Sustainability Division. June 2010.
Sandler, K. (2003). Analyzing What's Recyclable in C&D Debris. Biocycle, 44 (11), 51-54.
Staley, B.F. and Barlaz, M.A. (2009). Composition of Solid Waste in the United States and
Implications for Carbon Sequestration and Methane Yield. Journal of Environmental
Engineering, ASCE, 135, 901-909.
Tchobanoglous, G., Theisen, H., and Vigil. S. (1993). Integrated Solid Ewaste Management:
Engineering Principles and Management Issues. McGraw Hill. ISBN 0-07-063237-5
Tolaymat, T.M., El Badawy, A. M., and Carson, D. A. (2013). Estimate of the Decay Rate Constant
of Hydrogen Sulfide from Drywall in a Simulated Bench-Scale Study. Journal of
Environmental Engineering, ASCE, 139, 538-544.
Townsend, T.G. (2002). Gypsum Drywall Impact on Odor Production at Landfills: Science and
Control Strategies. Florida Center for Solid and Hazardous Waste Management, University
of Florida, Gainesville, FL.
US EPA (1996) AP 42, Fifth edition, Volume I - Appendix B.2 General particle size distributions.
https://www3.epa.gov/ttn/chief/ap42/appendix/appb-2.pdf Accessed May 27th, 2016.
US EPA (2019). Waste Reduction Model (WARM) Tool: User's Guide. Version 15, May 2019.
Available: https://www.epa.gov/warm/documentation-chapters-greenhouse-gas-emission-
energy-and-economic-factors-used-waste-reduction.
US LCI (2012). US Life Cycle Inventory Database. National Renewable Energy Laboratory.
http://www.nrel.gov/lci/. Accessed 20 February 2014.
Venta, G.J. (1997). Life Cycle Analysis of Gypsum Board and Associated Finishing Products.
Prepared by George J. Venta of Venta, Glaser & Associates for The Athena Sustainable
Materials Institute. March 1997.
WRAP (2008). Life Cycle Assessment of Plasterboard: Quantifying the environmental impacts
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Xu, Q. (2005). Hydrogen Sulfide Emissions and Control Strategies at Construction and Demolition
Debris Landfills. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
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6 Wood
6.1 Introduction
Wood is the third most widely used construction material in the US after asphalt concrete and
Portland cement concrete (PCC) (Cochran and Townsend 2010). Softwood and hardwood lumber
constituted approximately 58% (by weight) of 60 MMT of the solid wood products manufactured
in the US in 2011 (Howard and Westby 2013). Approximately 46% and 14% of the total solid
wood products consumed in the US in 2009 were used for residential (single family, multi-family,
and mobile homes) and non-residential building construction and renovation, respectively
(McKeever and Howard 2011). Light wood framing utilizing dimensional lumber and engineered
wood is employed heavily in residential construction (Wacker 2010, McKeever and Howard
2011). This section covers the following solid wood product wastes, which commonly appear in
CDD, including dimensional lumber and engineered wood products:
• plywood and oriented strand board (OSB),
• particleboard,
• medium-density fiberboard (MDF),
• structural laminated veneer lumber,
• glue laminated timber, and
• wood I-joists
Woody wastes from land clearing debris (LCD) activities are covered in Chapter 7 of this report.
Paper products, although representing the largest wood-derived product stream, represent a small
fraction of discarded CDD materials by weight and therefore are not included in the scope of CDD
materials investigated in this report. A large amount of wood waste (residue) is also generated
during manufacturing, and more than 98% of this waste is used by the wood products
manufacturing industry as fuel or as a feedstock for other products such as particle board. In an
LCA context, the management of these residues and the associated environmental impacts are
typically attributed to the product manufacturing phase and therefore are outside the scope of this
report (Puettmann and Wilson 2005).
Figure 6-1 illustrates the end-of-life processes associated with wood products. The wood products
used for construction are discarded at the end of their service life, which ranges from 50 to 100
years (Cochran and Townsend 2010). As depicted in Figure 6-1, theEOL management options of
discarded wood products include landfill disposal (either in MSW or CDD landfills), combustion,
and recycling. Discarded wood products can also be used in the production of new wood products.
However, this is not a common practice and therefore is not included in the LCI dataset.
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End-of-life
Product
Removal
CDD MRF
Chipping,
Removal of
Nails
MSW
Landfilling
of Wood
CDD
Landfilling
of Wood
Figure 6-1. Process flow diagram for end-of-life management of treated and untreated wood. Four
types of wood treatments are included: ACQ, CBA, CCA, and DOT. Only untreated wood is
processed for mulch. ACQ = alkaline copper quaternary; CBA = copper azole; CCA = chromate
copper arsenate; DOT = disodium octaboratetetrahydrate; MRF = material recovery facility.
6.2 EOL Management
In 2018, wood products contributed to approximately 37 MMT of CDD waste generated in the
United States. Approximately 25% of CDD wood (equivalent to more than 9 MMT) is recycled as
mulch or biomass fuel (US EPA 2020). A small fraction, approximately 4%, of CDD wood waste
is used for the production of pulp chips, color mulch, pressed fire logs, and fuel pallets (Wiltsee
1998).
The type of modification to the built environment (e.g., construction, renovation, or demolition)
impacts the relative fraction of wood waste present in the waste stream, which in turn impacts the
viable EOL management processes. The fraction of wood in CDD waste generated from
construction, renovation, and demolition of residential structures has been reported to be higher
than for non-residential structures (Cochran et al. 2007). Cochran et al. (2007) also reported a
greater wood fraction in the waste stream from renovation than from demolition of residential and
non-residential buildings.
The material handling at the point of generation (e.g., segregation from other CDD materials)
impacts the quality, and in turn, the EOL management options. For example, the economic
viability of recovering and recyling wood commingled with other CDD materials would depend
on a variety of factors, such as processing required, market demand for the end use, and other
situational considerations. In addition to contamination by other materials, the presence of treated
wood may also dictate the final EOL management options.
Landfilling of wood wastes is the most common EOL management strategy employed in the US.
Combustion of discarded CDD wood is practiced on a limited scale and is often considered a form
of recycling (MGE 1997, Falk and McKeever 2004, Cochran 2006). Soil application may also
have limited viability if the wood was chemically treated.
Discarded CDD wood is not commonly reused in the United States. The deconstruction of
buildings, rather than demolition, has been proposed as an alternative and practiced on a very small
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scale to enhance material recovery and reuse (NAHB 1997, Denhart 2010). Closed-loop recycling
of wood, unlike many other waste materials, is severely limited (MGE 1997). Although engineered
wood products can be manufactured using discarded dimensional lumber, dimensional wood
cannot be manufactured using engineered wood products due to the manufacturing processes
involved. Recycling dimensional lumber may also involve processing into smaller pieces of
lumber or size reduction (i.e., chipping) to produce engineered wood products (Merrild and
Christensen 2009). This downcycling has been referred to as a wood cascade chain (Sathre and
Gustavsson 2006, Hoglmeier et al. 2013) and entails energy and carbon balances including land
use, primary material substitution, transit, and manufacturing.
Table 6-1 lists and describes the LCA processes included in this report's associated datasets for
EOL management of wood products.
Table 6-1. End-of-life management processes for treated and untreated wood.
Process
LCA considerations
Unlined CDD
landfilling of DOT-
treated wood; at unlined
CDD landfill
This process includes materials (e.g., equipment, soil, water) and
energy (fuel, electricity) inputs for placement and compaction of CDD
wood in an unlined CDD landfill, along with non-energy emissions
(e.g., particulate emissions from equipment operation, gas emissions
from the decomposition of wood, and liquids emissions associated
with biogeochemical degradation of CDD wood in a landfill). Energy
recovery from the collection and combustion of landfill gas is also
included.
MSW landfilling of
DOT-treated wood; at
MSW landfill
This process includes material and energy inputs as well as gas and
leachate emissions from disposal of DOT-treated wood products in an
MSW landfill. It also includes landfill construction, operation, and
closure/post-closure care.
Unlined CDD
landfilling of CCA-
treated wood; at unlined
CDD landfill
This process includes material and energy inputs and emissions from
the degradation of Chromated Copper Arsenate (CCA)-treated wood
in an unlined CDD landfill. It also includes the placement of cover soil
and landfill operation, closure, and post-closure care.
MSW landfilling of
CBA-treated wood; at
MSW landfill
This process includes material and energy inputs as well as gas and
leachate emissions from disposal of CBA-treated wood products in an
MSW landfill. It also includes landfill construction, operation, and
closure/post-closure care.
Unlined CDD
landfilling of CBA-
treated wood; at unlined
CDD landfill
This process includes material and energy inputs and emissions from
the degradation of Copper Azle (CBA)-treated wood in an unlined
CDD landfill. It also includes the placement of cover soil and landfill
operation, closure, and post-closure care.
MSW landfilling of
ACQ-treated wood; at
MSW landfill
This process includes material and energy inputs as well as gas and
leachate emissions from disposal of ACQ-treated wood products in an
MSW landfill. It also includes landfill construction, operation, and
closure/post-closure care.
Unlined CDD
landfilling of ACQ-
This process includes material and energy inputs and emissions from
the degradation of Alkaline Copper Quaternary (ACQ)-treated wood
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Process
LCA considerations
treated wood; at unlined
CDD landfill
in an unlined CDD landfill. It also includes the placement of cover soil
and landfill operation, closure, and post-closure care.
MSW landfilling of
untreated wood; at
MSW landfill
This process includes material and energy inputs from disposal of
untreated wood in an MSW landfill. It also includes the landfill
construction, operation, and closure/post-closure care.
Unlined CDD
landfilling of untreated
wood; at unlined CDD
landfill
This process includes material and energy inputs and emissions from
the degradation of untreated wood in a CDD landfill. It also includes
the placement of cover soil and landfill operation, closure, and post-
closure care.
MSW landfilling of
wood ash; untreated
wood; at MSW landfill
This process includes material and energy inputs from disposal of
wood ash resulting from the combustion of untreated wood in an MSW
landfill. It also includes landfill construction, operation, and
closure/post-closure care.
Unlined CDD
landfilling of wood ash;
untreated wood; at
unlined CDD landfill
This process includes material and energy inputs and emissions from
the placement of wood ash from untreated wood combustion in an
unlined CDD landfill. It also includes the placement of cover soil and
landfill operation, closure, and post-closure care.
MSW landfilling of
wood ash; at MSW
landfill; CCA 9.6 kg/m3
This process includes material and energy inputs as well as leachate
emissions from disposal of wood ash resulting from the combustion of
CC A-treated wood (9.6 kg/m3) in an MSW landfill. It also includes the
MSW landfill construction, operation, and closure/post-closure care.
MSW landfilling of
wood ash; at MSW
landfill; CCA 40 kg/m3
This process includes material and energy inputs as well as leachate
emissions from disposal of wood ash resulting from the combustion of
CCA-treated wood (40 kg/m3) in an MSW landfill. It also includes
landfill construction, operation, and closure/post-closure care.
MSW landfilling of
wood ash; at MSW
landfill; CCA 4 kg/m3
This process includes material and energy inputs as well as leachate
emissions from disposal of wood ash resulting from the combustion of
CCA-treated wood (4 kg/m3) in an MSW landfill. It also includes
landfill construction, operation, and closure/post-closure care.
Land application of
CDD wood mulch; at
application
This process includes emissions associated with the land application
of mulched CDD wood at the application site, including transportation
and the spreading of mulch.
6.3 LCI Sources
Peer-reviewed literature, government and private industry publications, and various LCA
modeling tools were reviewed to identify available LCI datasets pertaining to the EOL
management of CDD wood. Table 6-2 lists the data sources reviewed to compile the LCIs
presented in this chapter. If LCI data were not available, process metadata and documents were
reviewed to evaluate the completeness of the dataset (e.g., emissions categories included,
background data used to compile the dataset, geographic location and time period of the data).
Although LCI from many information sources listed in Table 6-2 may not pertain specifically to
the US, these sources are presented and discussed for better understanding of the inputs used to
develop these LCI and the LCI information available globally.
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Table 6-2. Overview of life cycle inventory data sources associated with treated and untreated wood.
Source
Description
AP-42 (US
EPA 1995a)
US EPA (1995a) provides air emissions factors for plywood manufacturing,
reconstituted wood products (OSB, particleboard, medium-density fiberboaid
(MDF), hardboard, and fiberboard) manufacturing, wood preservation, and
manufacture of engineered wood (including glulam, laminated veneer lumber, and
others).
Cochran
(2006)
Cochran (2006) compiled energy requirements for recycling waste wood into mulch,
which has been identified as the primary EOL management option employed in the
US.
Dubey et al.
(2010)
Dubey et al. (2010) presents leaching data for four types of pressure-treated wood,
simulating contaminant leaching under a variety of recycling and disposal scenarios.
Jambeck
(2004)
Jambeck (2004) includes batch tests conducted on CCA-treated wood at different
L:S ratios withdeionized water as the leaching fluid; metals leached were quantified.
Jang (2000)
Jang (2000) reported anions, cations, and metals leaching from individual CDD
materials subjected to synthetic precipitation.
Townsend
et al. (1999,
2004)
These studies reported column and batch leaching test data on new and weathered
untreated and treated wood. Batch leaching tests utilized multiple set-up protocols to
evaluate the impact on metal leachability.
US EPA
(2019)
US EPA (2019) presents GHG emissions factors pertaining to source reduction,
recycling, combustion, and disposal of dimensional lumber, MDF, and hardwood
flooring for the WARM model.
Athena
The Athena Impact Estimator (IE) for Buildings life-cycle model includes the energy
requirements for demolishing wood-framed structures. The Athena Sustainable
Materials Institute (ASMI)has developed cradle-to-gate LCI for the following wood
products: cross-laminated timber (CLT), Glulam (glue laminated timber), wood I-
joists, laminated veneer lumber, MDF, OSB, particle board, softwood plywood
sheathing, and softwood lumber.
RSMeans
(2011)
Data on the cost of construction which includes estimates for types and productivity
of heavy equipment. This includes equipment necessary to land-apply end-of-life
wood products, e.g. mulch.
6.4 LCI Related to Disposal
Disposal of CDD wood product waste in landfills is the most commonly encountered management
practice in the US (Falk and McKeever 2012). Due to lower tipping fees, CDD wood disposal is
more common in CDD landfills than in MSW landfills. The potential forleachate and landfill gas
(LFG) release to the environment depends on the biogeochemical environment of the landfill and
the environmental controls, as discussed in Chapter 2. The emissions associated with production
and use of different material and energy inputs for landfill construction, operation, and closure, as
well as those associated with leachate and gas, are included in the LCIs for wood disposal. The
details of leachate and gas emissions are presented in this section.
Due to its organic nature, wood falls under the category of biodegradable waste. Landfilling of
biodegradable waste in CDD and MSW landfills are discussed in Chapter 2, with flows shown in
Figure 2-3 and Figure 2-6, respectively. The decay of wood waste in an anaerobic environment
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produces methane, which may be collected by a GCCS and converted to biogenic carbon dioxide
via flaring or energy-conversion technology. Most of the wasteLCA models (e.g., WARM, MSW-
DST, EASETECH) account for methane as the only gaseous emission associated with wood
decomposition in a landfill. WARM and MSW-DST use material-specific methane yield and decay
rates reported by several sources (i.e., Eleazer et al. 1997, Barlaz 1998, Staley and Barlaz 2009,
De la Cruz et al. 2010).
The MSW-DST model contains LCI data for leachate emissions from the landfill disposal of an
array of waste materials contained in MSW. CDD wood product is not included as a material
category in the model. The WARM model used methane yield of branches (reported on dry weight
of branches) as a proxy for developing emission factors for dimensional lumber, medium-density
fiberboard, and wood flooring. The methane yield was adjusted for moisture content to estimate
methane emission per unit wet weight of the material. The moisture content used for dimensional
lumber and medium-density fiberboard was the same as that of branches. The moisture content
used for wood flooring was greater than that of branches resulting in approximately 15% lower
methane emission factors for wood flooring than for the other wood products. WRATE provides
wood-specific emissions of approximately 30 gaseous constituents. The material-specific gaseous
emissions of various constituents are not based on actual material-specific measurements, but
rather on a theoretical allocation of the total emissions to individual waste components (Golder
Associates 2005).
The LFG production properties of branches were used as a proxy for estimating gas generation as
a result of the landfill disposal of CDD wood. Additional information on LFG emissions of CDD
wood is provided in Section 2.6.8. The methane and carbon dioxide emissions from the landfill
disposal of CDD wood is estimated as 0.064 and 0.21 kg, respectively, for placement in a CDD
landfill.
A significant fraction of methane is captured and combusted to carbon dioxide at landfills with
GCCS (e.g., MSW landfills). The methane and biogenic carbon dioxide emission from landfills
with GCCS would be lower and higher, respectively, than from landfills with no GCCS. Using the
average national statistics for the percentage of landfills that have GCCS and assuming a 90%
average gas collection efficiency as provided for dimensional lumber in WARM, the methane and
carbon dioxide emissions from the landfill placement of one kilogram of wood in an MSW landfill
are estimated to be 0.022 kg, and 0.33 kg, respectively. Wood treatment chemicals were assumed
not to have an impact on gas generation.
Wood products are often treated with preservative chemicals for protection from the weathering
elements and biota. Chemicals may either be applied to the wood's surface and/or impregnated via
pressure treatment into the wood itself (Haverty and Micales-Glaeser 2004, US EPA 1999). CCA
was, at one time, the most extensively used chemical in the treatment of lumber and other wood
products, datingback to the 1940s. However, CCA was phased out of production in the early 2000s
due to toxicity-related health concerns of arsenic exposure. Prior to the phase-out, three types of
CCA-treated wood were widely available, with Type C CCA being the most common (Jambeck
2004). Other chemical treatments, such as copper-based alkaline copper quaternary (ACQ), copper
azole (CBA), and disodium octaborate tetrahydrate (DOT) are now being used in greater
quantities. Jambeck et al. (2007) estimated that the peak quantity of CCA wood in the waste stream
would occur in 2008, when approximately 9.7 million m3 would be disposed of. The presence of
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wood treatment chemicals and other chemicals in paints (which may contain lead), stains, etc. has
a significant impact on the quality of liquid emissions (Lebow et al. 2004, Townsend et al. 2004)
and complicates and often impedes the EOL management via disposal or reuse/recycling.
Lebow et al. (2004) conducted an extensive review of the published leaching data from treated
wood and reported that several factors such as particle size, wood species, leaching water
characteristics, and surface finishes (e.g., paint) have significant impact on preservative leaching
from pressure-treated wood. For example, Lebow et al. (2004) reported that red oak leached
approximately 15% of the total arsenic, while yellow poplar leached only approximately 1% of
arsenic for the same treatment type (Type C CCA). Townsend et al. (2004) reported that for the
same treated wood, a 100-g block leached at levels approximately a quarter of the arsenic leached
by sawdust particles.
Wood is typically disposed of together with a mix of other CDD materials. The actual
measurements of long-term wood-specific liquids emissions from full-scale landfills are not
available and are not expected to be available in the future. The laboratory-scale leachate quality
data published by various sources (Townsend et al. 1999, Jang 2000, Lebow et al. 2002, Lebow et
al. 2004, Jambeck 2004, Dubey 2005, Jambeck et al. 2006, Dubey et al. 2007, Mitsuhashi et al.
2007, Dubey et al. 2010, Hasan et al. 2010, Clausen et al. 2010, Tao et al. 2013, Tao 2014) were
reviewed to estimate liquids emission from the disposal of untreated and treated wood in landfills.
The following criteria were used to select datafor estimating liquids emission from wood products
disposal in landfills:
1. Sample size. Wood products are not expected to undergo a significant size reduction
during waste placement and compaction in a landfill. The data from leaching tests
conducted on larger particle sizes (e.g., wood blocks) were preferred over data from tests
on aggressively size-reduced wood samples (e.g., sawdust) (e.g. Townsend et al. 2004, and
Townsend etal. 2005) for liquids emission estimates.
2. L:S ratio. The cumulative amount of chemicals leached from treated wood has been
reported to be a function of the amount of liquid the wood is exposed to (Jambeck 2004,
Tao et al. 2013). Many studies assessed leaching from sample columns exposed to natural
or synthetic precipitation (Jang 2000, Jambeck 2004, Tao et al. 2013, and Tao 2014) for a
limited timeframe. None of these studies reported 100% leaching of the preservatives. The
L:S ratio for these studies was either not reported or significantly lower than the L:S ratio
of the standardized leaching tests such as SPLP and TCLP. The datafrom tests with greater
L:S ratios were preferred over those from lower L:S ratio tests.
3. Leaching fluid. As landfill leaching is expected to occur under slightly acidic conditions,
leaching data associated with neutral or basic fluids such as deionized water (e.g., Jambeck
2004, Jambeck et al. 2006, Dubey etal. 2007) were not used for estimating liquid emissions
associated with disposal of untreated/treated wood in landfills. Data from tests using SPLP
and TCLP extraction fluids were used for estimating liquid emissions from an unlined CDD
landfill and MSW landfill, respectively.
Based on these criteria, results from the SPLP batch leaching test (L:S=20) data reported by Jang
(2000) were used for estimating liquid emissions (for COD, chloride, potassium, calcium, arsenic,
chromium, copper, and manganese) from untreated wood and CCA wood disposal in an unlined
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CDD landfill. Jang (2000) conducted batch leaching tests using SPLP extraction fluid on
individual CDD materials, including untreated and CCA-treated wood. The column leaching data
from Townsend et al. (1999) were used for parameters that were measured below the detection
limit by Jang (2000). Townsend et al. (1999) conducted column leaching tests on individual CDD
materials, including wood (untreated, new southern pine lumber) with SPLP extraction fluid. The
overall L:S ratio for the column experiment was approximately 5:1.
Distinct data are included in untreated wood disposal in an unlined CDD materials landfill. The
bulk density of wood waste presented in CCG (2006) was used to estimate cover soil requirements
for placement of untreated wood waste at a CDD landfill. Additional details on the diesel and
electricity requirements included in the "CDD landfill operations" flow and details on the cover
soil estimate are provided in Chapter 2. An average distance of 20 km was assumed between the
wood removal site and CDD landfills.
Leaching of chemicals from wood treated with CCA and other chemicals have been investigated
by several authors (e.g., Jang 2000, Townsend et al. 2004, Jambeck 2004, Dubey et al. 2010, Hasan
et al. 2010). Based on the criteria discussed above, data presented by Jang (2000) were selected to
estimate liquid emissions associated with CCA-wood (Type C, chemical retention rate of 4.0
kg/m3) disposal in unlined CDD materials landfill. Data from SPLP tests conducted by Dubey et
al. (2010) on sawdust of CCA-treated wood (Type C, retention rate of 6.4 kg/m3) were used for
the parameters not measured by Jang (2000). The CDD landfill dataset includes data on CCA wood
disposal in an unlined CDD materials landfill. COD, chloride, calcium, manganese, and potassium
emission from CCA-treated wood are at similar levels to those from untreated wood.
Three datasets were compiled for disposal of ACQ-, CBA-, and DOT-treated wood in unlined
CDD landfills. The liquids emissions for these treated wood types were estimated based on SPLP
data presented by Dubey etal. (2010). Dubey etal. (2010) conducted tests on sawdust from various
treated wood using a variety of extraction fluids, including SPLP, TCLP, and leachates fromMSW
landfills. The same quantity of cover soil, diesel and electricity consumption for landfill
operations, and transport distance between the site of wood removal and the CDD landfill were
assumed; these same parameters were also assumed for the CDD landfill disposal of untreated
wood.
Leachate-related emissions from wood products in an MSW landfill environment were estimated
based on TCLP data reported in literature. The MSW landfilling dataset includes emissions data
for four types of treated wood based on the TCLP test data reported by Dubey et al. (2010). The
cover soil requirements, as well as the material and energy flows included for MSW landfill
construction, operation, closure, and post-closure care are detailed in Chapter 2 of this report. The
bulk density of wood products presented in CCG (2006) was used to estimate cover soil
requirements.
Several factors should be considered when using the proposed liquids emissions for MSW
landfills. First, the emissions presented in the dataset should be considered a partial representation,
as these are based on batch leaching tests with an L:S ratio of 20. In reality, wood placed in a
landfill would result in leaching at a much greater L:S ratio (assuming that the landfilled waste
will never be reclaimed). Second, leachate from an MSW landfill is typically collected and treated
before the effluent is discharged into the environment during active disposal, closure, and post-
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closure care. The wastewater treatment process partitions contaminants from the liquid phase into
the treated effluent and solid residues (sludge, biosolids). The contaminant quantities released into
the environment with effluent discharge depends on the treatment plant's contaminant-removal
efficiency, which in turn, depends on the contaminant type (NCSU and ERG 2011). A treatment
efficiency of 85% for heavy metals is used by MSW-DST. The sludge from a wastewatertreatment
plant is commonly managed by either land-application or disposal at MSW landfills. Based on a
nationwide survey NEBRA (2007) estimated that approximately 49% of the wastewatertreatment
plant (WWTP) sludge generated in the US is applied to soil. The inorganic contaminants contained
in sludge could potentially leach and be released into the environment (groundwater, surface
water).
Additional pathways by which leachate releases contaminants into the environment include
fugitive leachate emissions through imperfections in the bottom liner (e.g., geomembrane pin holes
that occur during construction), leaching of chemicals from land-applied sludge, and the cyclic
process of contaminant release from WWTP sludge deposited in landfills, leachate treatment, and
sludge disposal at landfills. Following the post-closure care period, it would be expected that all
leachate would eventually discharge into the environment. For example, MSW-DST and
EASETECH account for leachate collection and treatment for a default period of 100 years.
6.5 LCI Related to Recycling
Approximately 25% of CDD wood (equivalent to more than 9 MMT) is recycled as mulch or
biomass fuel (US EPA 2020). A lack of demand for end products (mulch, biomass) coupled with
competition from wood product manufacturing industries are potential challenges for CDD wood
recovery and recycling. The US mulch demand in 2005 of approximately 3 MMT (estimated by
Cochran 2006) was small compared to the amount of residue produced by wood product
manufacturing (177 MMT in 2002 as reported by McKeever (2004)). Moreover, CDD wood is
often considered to be a less desirable feedstock for mulch production due to aesthetics.
6.5.1 Processing of CDD Wood
Energy inputs and emissions associated with CDD wood waste processing to produce mulch
include those related to manufacturing and the use of sorting, grinding, and screening equipment
(Cochran 2006). Based on the data reported by Morbark (2006), Diamond Z (2006), and Bandi
(2006) for a horizontal grinder and manufacturer equipment specifications for an excavator and
loader, the diesel equipment requires 29.5 MJ per MT of wood waste processed in a mixed CDD
material recovery facility (MRF). This is equivalent to a fuel consumption of approximately 0.755
L of diesel per MT of wood. Fuel consumption data reported by Cochran (2006) is used here.
However, there exists a need for more accurate and publicly available energy and material input
data from built facilities so that more reliable LCIs can be developed.
The non-energy-related emissions from wood grinding include particulate matter emissions and
liquid emissions from wood/wood chip stockpiles. AP-42 presents air emission factors for log
chipping as part ofMDF manufacturing. These data can be used as a proxy for CDD wood grinding
until measurements from operating facilities become available. Unlike CDD wood processing
facilities, however, engineering controls such as cyclone and/or fabric filter collection are
implemented to control particulate matter emissions from chipping operations at MDF
6-9
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manufacturing facilities. Using log chipping air emission as a proxy would therefore result in
underestimating particulate matter emissions from CDD wood processing facilities. As the wood
decomposition in this scenario would occur under aerobic conditions, gas emissions from the land
application of mulch were estimated by assuming that 100% of carbon content will decompose to
produce carbon dioxide. Using the biogenic carbon content of branches (published by Barlaz 1998)
as a proxy for wood products, approximately 1.63 kg of carbon dioxide (biogenic) would be
produced from aerobic decomposition of 1 kg of wood products. This estimate is based on 0.494
g of carbon content as C per dry kg of wood product and 0.9 kg of dry wood per kg of wet wood
product.
6.5.2 Land Application of Mulched CDD Wood
This process models the land application of mulched CDD wood and includes the emissions
associated with the transportation and spreading of CDD wood mulch. This process is from gate
to application site, and does not include the production of mulch. It does account forthe byproducts
from aerobic mulch degradation. Figure 6-2 shows the flows included in the LCI for this process.
Emissions resulting from land application assume aerobic conditions and leachable concentrations
of constituents equivalent to those released from SPLP testing.
Fuel use is based on equipment required to spread mulch (RSMeans 2011) and themain equipment
fuel use (Caterpillar 2006). Gas emissions from the land application of mulch were estimated by
assuming that 80% of carbon content of CDD wood will decompose aerobically to produce carbon
dioxide. The carbon content of yard waste branches reported by Barlaz (1998), the fraction of
decomposable carbon reported by Beck-Friis et al. (2000), and the moisture content of oven-dried
wood reported by Briggs (1994) were used for this estimation. The liquid emissions from land
application of mulch are expected to be the same as those from wood disposal in a CDD landfill
as leaching is primarily influenced by natural precipitation; leachate emissions are estimated from
SPLP measurements by Jang (2000) and Townsend et al. (1999). Decomposition products of
aerobic decomposition, such as water and residual organic carbon in soil, are not included, nor is
the oxygen input required to form carbon dioxide emissions. In the absence of additional
information, it was assumed that the CDD wood mulch would be transported 20 km to the mulch
end user.
Fuel use and lifetimes in non-road equipment were estimated using the MOVES database (US
EPA 2014). The MOVES model runs fleet-average uses of non-road equipment in the construction
and industrial sectors, and provides total fuel usage for all equipment estimated to be in operation
in the US by type and horsepower class. Types are defined by MOVES-specific implementation
of EPA Source Classification Codes. Total fuel use is divided by reported population to estimate
fuel use per unit equipment peryear. Using primary datafor non-road equipment used on site, each
is matched to a MOVES type and horsepower class. Median lifetimes in years are provided in
MOVES by equipment classification and engine horsepower. The amount of a vehicle used is
based on the proportion of fuel used divided by the MOVES estimated fuel use over the
equipment's lifetime.
6-10
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Transport, single unit truck;
short-haul; diesel powered
Heavy equipment
CDD wood; mulched;
at processing facility
Stockpiled material;
at drop site
+
Ground
water
t
Leachate
Land application of CDD
mulch; at application
Gasoline
Water
Oxygen
Legend
Elementary Flow ^
Technosphere Flow
H Not Included
| Included
System Boundary
Figure 6-2. Unit process flow diagram for application of CDD wood mulch to land.
6.5.3 Wood Ash
Wood ash is generated at a rate of approximately 2.7 MMT annually in the US (Risse 2010). In
the US, ash is managed by two major pathways: landfilling (approx. 65%) and land application
(9%), while 25% goes towards "other" undefined uses (Spokas 2010). When used as a soil
amendment, wood ash is capable of providing valuable nutrients (e.g., potassium, phosphorus,
magnesium), as well as acting as a liming agent, raising the pH and thus assisting in the retention
of nutrients (Kahl et al. 1996, NEWMOA2001, ASTSWMO 2007, ODEQ 2011). Land application
is practiced more frequently in the Northeastern US (at a rate of approximately 80% generation);
in contrast, the Southeastern US practices land application at only about 10%, and the Midwest at
about 33% (Vance 1996, Risse 2010).
The presence of CCA-treated wood has been recognized as a major issue with CDD wood waste
combustion (Cochran 2006). Incinerated CCA-treated wood can produce ash with heavy metal
concentrations that exceed toxicity characteristic hazardous waste limits (Solo-Gabriele et al.
2002). In batch leaching tests (TCLP, SPLP) performed by Solo-Gabriele et al. (2002), wood ash
produced from mixed wood waste with only 5% CCA-treated wood (by mass) caused consistent
exceedance of toxicity limits for arsenic and intermittent exceedances for chromium. For ash
resulting from the combustion of CCA-treated wood retaining high levels of preservative, heavy
metals represented up to 36% (by weight) of the resulting ash. Regulations related to beneficial
use of wood ash reflect concerns over the presence of treated wood combustion ash. For example,
Florida allows the land application of wood ash provided it was not produced from combustion of
treated or painted wood (FDEP 2002). Preliminary results from Mitchell et al. (2020) indicate that
it may be possible to stabilize the metals and reduce environmental impacts from leaching by
producing a cementitious aggregate when applied to soil.
Wood ash composition and characteristics (e.g., leaching behavior, reactivity) can vary
significantly depending on the temperature of the system and the characteristics and degree of
contaminants in the fuel wood (Jenkins et al. 1998). Several studies have examined the chemical
characteristics and the wood ash effects on plant growth as well as total metals extractable under
variably acidic conditions (Zhan et al. 1994, Demeyer et al. 2001, Norstrom et al. 2012). Other
researchers in the US have examined ash leaching using deionized water, ammonium citrate, and
6-11
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humic and fulvic acids as extraction fluids (Erich 1991, Clapham and Zibilske 1992, Chirenje et
al. 2002). Except for the limited testing performed by Solo-Gabriele et al. (2002), leaching data
from wood ash in conditions simulating precipitation exposure or a landfill environment were not
found in the literature.
The datasets for untreated or treated wood (provided at different treatment levels) ash in both CDD
and MSW landfill environments are based on the data reported by Solo-Gabriele et al. (2002). The
SPLP results (used to simulate wood ash placed in a CDD landfill) and TCLP results (used to
simulate wood ash placed in an MSW landfill) were derived from leaching tests performed on
seven ash samples. It was assumed that treated-wood-derived ash would be disposed of in a lined
MSW landfill and not in an unlined CDD landfill. These samples included ash from untreated
wood (one sample, southern yellow pine), ash from pre-consumer CCA-treated wood (three
samples, each with a different treatment level), and ash produced from recycled wood recovered
from CDD processing facilities (three samples, each recovered from a different facility). The SPLP
and TCLP leaching results from the ash samples produced from the combustion of the recycled
wood recovered from the three CDD processing facilities were each averaged to provide the
leaching LCI data. Only copper, chromium, and arsenic leaching results were provided in this
study. The expected leachable concentrations of non-metal organics and inorganics are unknown;
however, the total and leachable ash concentrations of these parameters as published in other
studies (as listed above) suggest that detectable concentrations of numerous other parameters (e.g.
calcium, manganese, iron) may be encountered if analyzed. Wood ash derived from untreated
CDD wood did not show detectable leached concentrations of any of the three metals. The only
environmental burdens included in the LCI which represents the placement of untreated wood at
a CDD or MSW landfill would be those associated with landfill construction, operations, and
closure/post-closure care. The liquids emissions presented for MSW landfill disposal represent the
emissions with untreated leachate. Leachate from lined landfills is typically collected and treated
during active landfill operation, closure, and post-closure care prior to discharge into the
environment.
The datasets for placing ash in a CDD landfill may be used to simulate the land application of
wood ash if the flows associated with landfilling and cover soil are removed. Except for untreated
wood ash, all contaminants measuring below the detection limit were analyzed at the detection
limit for dataset development purposes. The average distance between wood combustion and
CDD/MSW landfills was assumed to be 20 km. For estimating the amount of landfill cover soil
required, a bulk wood ash density of 702 kg/m3 was assumed (Huang et al. 1992).
6.6 Data Gap Analysis and Opportunities for Additional LCI Data
Table 6-3 summarizes the type of data presented by various sources reviewed for the compilation
of wood products EOL management LCI. Several LCA models (e.g., WARM, MSW-DST) and
LCI databases provide data on US-based processes. As shown in the table, there is a tendency for
each source to present only part of the information needed for LCI compilation. Therefore, sources
must be combined to achieve a more robust LCI dataset.
A majority of LCI information available on wood products pertains to the manufacturing aspects
of the life cycle. Based on a review of the available information, the following data gaps were
6-12
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identified that, if collected, would allow for a more comprehensive LCI dataset for wood product
EOL management:
1. Long-term teachable emissions from wood products placed in a landfill. The liquid
emissions presented in this study are based on SPLP and TCLP tests, which simulate leaching
from synthetic rainwater or an aggressive MSW landfill environment, respectively. The batch
leaching data used for estimating liquid emissions correspond to an L:S ratio of 20 and,
therefore, do not represent complete liquid emissions. Leaching data of only certain chemical
constituents are reported in literature. Furthermore, the standardized leaching tests simulate
leaching associated with physical and chemical mechanisms and do not simulate leaching
associated with biological decomposition of wood due to the short duration of these tests (18
hours). As wood products are typically disposed of with other discarded materials and not
disposed of in a monofill, field-scale leachate quality data specific to wood products disposal
in landfills are rare. The liquid emissions from the placement of wood products in landfills
would, therefore, need to be based on laboratory-scale studies. Future research should assess
the leaching of a larger suite of chemicals over a greater L:S ratio and those associated with
biological decomposition.
2. Long-term gaseous emissions from wood biodegradation in a landfill. The data reported
for branches were used as a proxy for estimating gaseous emissions from anaerobic
biodegradation of wood production disposed of in landfills due to lack of CDD wood-specific
data. Future research should consider quantification of a larger suite of gaseous emissions from
dimensional and engineered wood products to assess the impact of resins and treatments on
gaseous emissions.
3. Wood combustion ash. Numerous studies characterized wood ash to assess its benefits as a
soil amendment (e.g., as a lime substitute). Leaching data (SPLP, TCLP), however, are lacking
to be able to fully assess liquid emission for land application of untreated wood ash or for
disposal scenarios. Limited leaching data are available for CCA-treated wood ash. Future
research should quantify leaching emissions of a wider suite of chemicals and for higher L:S
ratios for wood ash.
4. Environmental impact of emerging wood preservation chemicals. Although certain
chemical treatments (e.g., CCA) have been phased out, future research efforts should assess
the impacts of emerging treatment preservatives such as nano-zinc oxide (Clausen et al. 2010).
6-13
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Table 6-3. Overview of life cycle inventory data sources associated with treated and untreated wood.
"P" represents partial coverage of the required data, while "X" represents full coverage, and
indicates no data.
Process
AP-42
Cochran
(2006)
us
EPA
(2019)-
WARM
MSW-
DST
NREL
Jang
(2000)
Dubey
et al.
(2010)
Solo-
Gabriele
et al.
(2001)
Dimensional Lumber
manufacturing
-
-
-
-
X
-
-
-
WPM- Engineered Wood
Production
X
-
-
-
X
-
-
-
WPM- Ancillary Materials
(e.g., CCA)
-
-
-
-
P
-
-
-
Landfill Construction &
Operation
-
-
P
X
-
-
-
-
Landfill Leachate Emissions
-
-
-
-
-
X
X
-
Landfill Gas Emissions
-
X
-
X
-
-
-
-
Wood Processing
-
X
-
-
-
-
-
-
Mulch-Liquids Emission from
Land Application
-
-
-
-
-
X
-
-
Untreated Wood Combustion-
Air Emission
X
-
P
-
-
-
-
-
Treated Wood Combustion
Ash- Liquids Emission
-
-
-
-
-
-
-
X
6.7 References
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Major Forest Landowner Types. Forest Policy and Economics, 9, 237-248.
Athena (2012). LCA in Construction: Status, Impact, and Limitations. A Report by ASMI and PE
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Barlaz, M.A. (1998). Carbon Storage During Biodegradation of Municipal Solid Waste
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Clausen, C.A., Green III, F., and Kartal, S.N. (2010). Weatherability and Leach Resistance of
Wood Impregnated withNano-Zinc Oxide. Nanoscale Research Letters, 5, 1464-1467
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Cochran, K.M. and Townsend, T.G. (2010). Estimating Construction and Demolition Debris
Generation Using a Materials Flow Analysis Approach. Waste Management, 30, 2247-
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Cochran, K.M. (2006). Construction and Demolition Debris Recycling: Methods, Markets, and
Policy. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Cochran, K., Townsend, T., Reinhart, D., and Heck, H. (2007). Estimation of Regional Building-
Related C&D Debris Generation and Composition: Case Study for Florida, US. Waste
Management, 27, 921-931.
De La Cruz, F.B. and Barlaz, M.A. (2010). Estimation of Waste Component-Specific Landfill
Decay Rates Using Laboratory-Scale Decomposition Data. Environmental Science &
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Denhart, H. (2010). Deconstructing Disaster: Economic and Environmental Impacts of
Deconstruction in Post-Katrina New Orleans. Resources, Conservation, and Recycling, 54,
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Dubey, B.K. (2005). Comparison of Environmental Impacts of Wood Treated with Chromated
Copper Arsenate (CCA) and Three Diffeent Arsenic-Free Preservatives. Ph.D.
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Dubey, B., Townsend, T., Solo-Gabriele, H., and Bitton, G. (2007). Impact of Surface Water
Conditions on Preservative Leaching and Aquatic Toxicity from Treated Wood Products.
Environmental Science & Technology, 41, 3781-3786.
Dubey, B., Townsend, T., and Solo-Gabriele, H. (2010). Metal Loss from Treated Wood Products
in Contact with Municipal Solid Waste Landfill Leachate. Journal of Hazardous Materials,
175, 558-568.
Eleazer, W.E., Odlelll, W.S., Wang, Y.S., and Barlaz, M.A. (1997). Biodegradibility of Municipal
Solid Waste Components in Laboratory-Scale Landfills. Environmental Science &
Technology, 31, 911-917.
Falk, R.H. andMcKeever, D.B. (2012). Generation and Recovery of Solid WoodWaste in theUS.
Biocycle August 2012, 30-32.
Golder Associates Ltd. (2005). Life Cycle Assessment Landfill Emissions. A Report Prepared by
Gold er Associates (UK) Ltd. For Environment Agency, Bristol, UK. 16 November 2005.
Hasan, A.R.,Hu, L., Solo-Gabriele, H.M.,Fieber, L., Cai, Y., and Townsend, T.G. (2010). Field-
Scale Leaching of Arsenic, Chromium, and Copper from Weathered Treated Wood.
Environmental Pollution, 158, 1479-1486.
Haverty, M. and Micales-Glaeser, J. (2004). Wood Destroying Organisms in the New Millennium:
Where Have We Gone Since Bend 1989?. Proceedings of the Fifth Joint Meeting of the
Western Forest Insect Work Conference and Western International Forest Disease Work
Conference. 26- 30 April 2004. San Diego, CA. 72-74.
Hoglmeier, K., Weber-Blaschke, G., and Richter, K. (2013). Potentials for Cascading of
Recovered Wood from Building Deconstruction- A Case Study for Southeast Germany.
Resources, Conservation and Recycling, 78, 81-91.
Howard, J.L. and Westby, R.M. (2013). US Timber Production, Trade, Consumption and Price
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Huang, H., Campbell, A.G., Folk, R., and Mahler, R.L. (1992). Wood Ash as a Soil Additive and
Liming Agent for Wheat: Field Studies. Communications in Soil Science and Plant
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Jambeck, J.R. (2004). The Disposal of CCA-Treated Wood in Simulated Landfills: Potential
Impacts. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Jambeck, J.R., Townsend, T., and Solo-Gabriele, H. (2006). Leaching of Chromated Copper
Arsenate (CCA)-Treated Wood in a Simulated Monofill and its Potential Impacts to
Landfill Leachate. Journal of Hazardous Materials, A13 5,21-31.
Jambeck, J., Weitz, K., Solo-Gabriele, H., Townsend, T., and Thorneloe, S. (2007). CCA-Treated
Wood Disposed in Landfills and Life-Cycle Trade-Offs with Waste-to-Energy and MSW
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Jang, Y. (2000). A Study of Construction and Demolition Waste Leachate From Laboratory
Landfill Simulators. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Lebow, S., Brooks, K., and Simonsen, J. (2002). Environmental Impact of Treated Wood in
Service. Proceedings of Forest Products Society Conference. 11-13 February 2002.
Kissimme, FL. 205-215.
Lebow, S., Cooper, P., and Lebow, P. (2004). Variability in Evaluating Environmental Impacts of
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McKeever, D. (2004). Woody Residues and Solid Waste Wood Available for Recovery in the
United States, 2002. USDA Forest Service, Forest Products Laboratory.
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chips, 2" deep, Skid Steer Loader
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7 Land Clearing Debris
7.1 Introduction
Land clearing debris (LCD) is comprised of treetops, branches, and stumps, and can also include
materials such as soil, rocks, and shrubs resulting from vegetation removal for building and
infrastructure construction and land development (US EPA 2011). The moisture content of LCD
is typically greater than that of other CDD wood sources (e.g., dimensional lumber, plywood),
estimated at approximately 50% (wet basis) (Maker 1994, US EPA 2003, Tumuluru et al. 2011).
Estimates of LCD production are difficult because this category of materials is often excluded
from regulation as a solid waste. Wiltsee (1998) indicated difficulty with estimating LCD
quantities because the major management approaches for LCD (which includes chipping on-site
and burning without energy recovery) often do not involve any mass or volume estimates.
Figure 7-1 identifies the flow of materials and processes included in this report's associated
datasets for LCD EOL management.
End-of-life
Product
Removal
Air Curtain
Incineration
Figure 7-1. Process flow diagram for end-of-life management of land clearing debris (LCD).
7.2 EOL Management
Based on a survey of LCD contractors, Wiltsee (1998) found that the most common method of
LCD EOL management is on-site burning of LCD materials. Air emissions and liquids emission
from combustion ash are the primary environmental concerns with burning LCD. Depending on
state regulations, LCD may be disposed of in a designated LCD, inert waste, CDD, or MSW
landfill that accepts LCD. LCD can be used for mulch production, compost production, and
biomass fuel. These management options typically require reducing the size of the debris with a
chipper or grinder. Based on a survey of 180 wood collection and processing facilities in 14
counties in Michigan, Nzokou et al. (2011) reported that LCD constituted 61% of the total amount
of material processed by these facilities. In addition, mulch, wood chips, firewood, and industrial
fuel were the top four types of recycling products produced at the processing facilities; these
materials made up approximately 94% of the recycled products with mulch and chips comprising
7-1
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42% and 38.6% of all the recycled products, respectively. No national estimate of the quantities of
LCD managed with these EOL options is currently available.
Table 7-1 lists and describes the LCA processes included in this report's associated datasets for
EOL management of LCD.
Table 7-1. End-of-life management
Process
LCA considerations
Processing of land
clearing debris;
mulched; at processing
facility
This process includes energy and non-energy emissions associated
with LCD and compost processing and handling, as well as long-
term liquid emissions from land-applied compost.
Land application of
land clearing debris
mulch; at application
This process includes emissions associated with grinding and land
application, as well as those from leaching due to exposure to
precipitation.
Unlined CDD
landfilling of land
clearing debris; at
unlined CDD landfill
This process includes material and energy inputs for placement
and compaction of LCD in a CDD landfill along with non-energy
emissions (e.g., dust emissions from equipment operation and
liquid emissions associated with biogeochemical degradation of
LCD in a landfill).
MSW landfilling of
land clearing debris; at
MSW landfill
This process includes material and energy inputs from LCD
disposal in an MSW landfill. It also includes landfill construction,
operation, and closure/post-closure care
Combustion of land
clearing debris; at open
burning
This process includes emissions from LCD processing and
combustion, as well as those associated with management of
combustion ash.
Combustion of land
clearing debris; at air
curtain incineration
This process includes the preparation of LCD for combustion, air
emissions from combustion, and the long-term liquids emission
from combustion ash.
7.3 LCI Sources
Peer-reviewed literature, government and private industry publications, and various LCA
modeling tools were reviewed to identify available LCI datasets pertaining to EOL management
processes of LCD. Table 7-2 presents a list of data sources reviewed to develop the LCI presented
in this chapter. If LCI data were not available, process metadata and documentation (e.g., included
emission categories, background data used to compile the dataset, geographic location, and time
period of the data) were reviewed to evaluate the completeness of the dataset. If available, the
primary sources of information used to develop the LCI datasets were reviewed.
7-2
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Table 7-2. Overview of life cycle inventory data sources associated with land clearing debris (LCD).
Source
Description
Cochran
(2006)
Cochran (2006) presented diesel energy requirements for CDD wood
grinding equipment based on a survey of equipment manufacturers.
AP-42 (US
EPA 1996,
2002, 2003)
AP-42 provides air emissions data for trench air curtain burning and wood
chipping.
Lutes and
Kariher (1997)
Lutes and Kariher (1997) present VOC, SVOC, PAH, and criteria air
pollutant emission data from burning LCD with and without an air blower.
RSMeans
(2011)
RSMeans (2011) provides information on the cost of construction,
including estimates for types and productivity of heavy equipment. This
includes equipment necessary to apply LCD end products, such as mulch,
to land.
7.4 LCI Related to On-Site Burning
An estimate of the amount of LCD disposed of in the US is not available. However, it has been
reported that most LCD wood is disposed of at the site of generation. Open burning is one method
of LCD disposal; the debris is typically heaped in piles or placed in pits and burned in the absence
of emission control devices. As open burning is typically regulated at the state or local level in the
US, emission control requirements and resulting emissions vary by geographic location (ERGI
2001). Figure 7-2 shows the flows included in the LCI for open burning of LCD.
Combusting LCD using an air curtain incinerator (ACI) is another EOL management approach for
LCD at the point of generation. In this case a blower blows air into the burning debris to enhance
combustion, speeding up the combustion process and, ideally, reducing emissions by achieving
more complete combustion. There are two major types of ACIs: 1) a trench ACI which is
comprised of a mobile air blowing system that is placed in a constructed trench that contains debris
and blows air into the trench; and 2) a self-contained firebox unit, where the debris is placed in the
firebox, ignited, and the air blower circulates air into the container. Figure 7-3 shows the flows
included in the LCI for air curtain incineration of LCD.
The emissions from burning LCD at its site of generation include those associated with material
and energy input, as well as non-energy emissions released during the preparation of the LCD for
combustion and the combustion of LCD. LCIs were developed for the on-site burning of LCD
through open burning and the use of a firebox ACI. In open burning and ACI scenarios, equipment
is needed to either arrange the LCD into piles in preparation for burning or to load the LCD into
an ACI firebox.
Springsteen et al. (2011) documented the average fuel consumption for loading woody biomass
into a grinder as 0.79 L of diesel fuel per MT of green material. Operating the blower system of
an ACI firebox will also require energy usage. Air Burners' model S-327 ACI specifications were
used to approximate the average fuel consumption for a firebox ACI (Air Burners 2012).
7-3
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Springsteen et al. (2011) also provides diesel consumption for excavators used to load the LCD
into an open burn pile or air curtain incinerator. Fuel use and lifetimes in nonroad equipment were
estimated using the US EPA (2014) Motor Vehicle Emissions Simulator (MOVES). The MOVES
model runs are for fleet average uses of nonroad equipment in the construction and industrial
sectors, and provide total fuel usage for all equipment estimated to be in operation in the US by
type and horsepower class. Types are defined using EPA Source Classification Codes in MOVES.
Total fuel use is divided by reported population to estimate fuel use per unit equipment per year.
Median lifetimes in years are provided in MOVES by equipment classification and engine
horsepower. The amount of a vehicle used is based on the proportion of fuel used and MOVES
estimated fuel use over the equipment's lifetime.
Springsteen et al. (2011) presented emission factors for nitrogen oxides, particulate matter, non-
methane organic compounds (NMOC), carbon monoxide, and methane based on numerous
references, including US EPA's AP-42 sections on open burning and wildfires and prescribed
burning, laboratory studies, pilot, and full-scale studies on conifer (cone-bearing trees) biomass.
In the absence of emissions of carbon dioxide and methane for burning LCD in an ACI, the open
burning emissions were used as a proxy. Emissions of semi-volatile organic compounds (SVOCs),
volatile organic compounds (VOCs), polycyclic aromatic hydrocarbons (PAHs), and criteria
pollutants were collected by Lutes and Kariher (1997) in pilot-scale tests on LCD from two states
in the US (Florida and Tennessee). The results from the burning tests (for which greater than half
of the data were above the level of detection) were averaged and incorporated into the LCI tables
for open burning and ACI air emissions. For data readings below the level of detection, the
detection limit was used for the average estimation. US EPA (1996) AP-42 provided estimates of
burning wood in a trench ACL Therefore, sulfur dioxide and nitrogen oxides in the ACI scenario
from the US EPA (1996) were used. Because LCD incineration frequently occurs at the site of
generation, and in the absence of additional information, no LCD transportation is included in the
LCI datasets presented.
These processes do not include the effects of ash leaching following the burning process, and do
not include the effects of transporting heavy equipment to the site. This process does not include
land-use related changes associated with LCD debris removal and combustion on-site.
7-4
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Fuel combustion
byproducts
t
NOx
t !
CO-,
PM
i
Water
vapor
Heavy equipment
Land clearing debris;
at land clearing
Combustion of land clearing
debris; at open burning
*
Diesel
i
i
Oxygen
Land clearing debris;
combusted; at open burning
Legend
Elementary Flow ^
Technosphere Flow
| Not Included
| Included
System Boundary
Figure 7-2. Unit process flow diagram for land clearing debris at the site of open-burning
combustion.
Fuel combustion
byproducts
J
t
NOx
t
co2
PM
Heavy equipment--
Land clearing debris;
at land clearing
Combustion of land
clearing debris; at air
curtain incineration
Water
vapor
Land clearing debris;
- combusted; at air —~
curtain incinerator
Diesel
t
Oxygen
Legend
Elementary Flow
Technosphere Flow
j Not Included
| Included
System Boundary
Figure 7-3. Unit process flow diagram for land clearing debris at the air curtain incinerator.
7.5 LCI Related to Landfill Disposal
Emissions associated with LCD disposal in a landfill include air emissions from equipment used
for placing LCD in the landfill, emissions associated with landfill construction and operation, and
liquids and gaseous emissions from material decomposition in the landfill. The potential fa"
7-5
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leachate and landfill gas release to the environment depends on the biogeochemical environment
of the landfill, as well as the environmental controls, as discussed in Chapter 2. There are various
sources that provide landfilling emission factors related to equipment usage and landfill
construction and operation; however, none are specific to LCD landfilling. Generalized landfill
construction and operations LCI data are also presented in Chapter 2.
The primary constituent in LCD is wood, and as discussed in Chapter 6, due to its organic nature,
the decay of wood wastes in an anaerobic (i.e., oxygen-poor) environment produces methane; this
methane may be collected by a GCCS, if the landfill has one, and converted to biogenic carbon
dioxide via flaring or energy conversion technology. LCD exposure to precipitation or other
liquids (e.g., landfill leachate) is also expected to result in chemical leaching, and the emissions
are expected to depend on biogeochemical environment (e.g., MSW landfill, CDD landfill).
Details on the gaseous emissions produced from decomposition and leachable emissions of
untreated wood are presented in Chapter 6.
The emissions from wood decomposition and leaching, energy consumption data from landfill
operations, estimated cover soil demand, and an assumed transport distance were used to develop
an LCI process dataset for the disposal of LCD at an unlined CDD or MSW landfill. The gaseous
and liquid emissions were developed based on the untreated wood waste LCI developed in Chapter
6, but were adjusted for the greater moisture content of LCD. Moisture content of wood used by
Jang (2000), Townsend et al. (1999) was assumed to be 10% and a moisture content of 50% was
used for LCD; the methane and carbon dioxide emission for wood products disposal in landfills
(presented in Chapter 6) are based on a moisture content of 10%. The energy use and the associated
emissions from landfill operation (e.g., waste placement, compaction) include diesel use in heavy
equipment and electricity use in landfill buildings (e.g., administrative buildings, workshop);
calculations detailing these emissions and the method of estimating the quantity of cover soil use
are also included in Chapter 2 of this report. Diesel consumption from landfill operations and
electricity consumption from landfill administrative offices and workshop areas were estimated
from Ecobalance (1999) and IWCS (2014), respectively. In the absence of nationwide average
transport data, it was assumed that LCD would be transported 20 km for landfill disposal. While
actual distances may vary, this assumption was used to represent a typical distance between the
LCD site and the nearest landfill. For the purpose of estimating cover soil requirements for the
disposal of LCD at an unlined CDD landfill, the density of LCD was estimated from the bulk
density of unprocessed forest product fuel wood, as provided Angus-Hankin et al. (1995).
7.6 LCI Related to Recycling
7.6.1 LCD Used as Mulch
LCD recycling generally involves processing (i.e., chipping/grinding) prior to end-of-lifeuse, and
can occur at the site of LCD generation, mobile equipment, or LCD can be transported to a larger
processing facility. The LCI information provided in this section incorporates the transportation
of LCD to a large processing facility to produce mulch. Processing of LCD in preparation for
mulch production typically involves loading and operating a grinder. Horizontal grinders or tub
grinders can be used for grinding vegetative debris. Horizontal grinders are better equipped to
handle debris such as tall trees that may be pre-organized prior to being fed into the grinder. Tub
grinders, although they can process materials wider in diameter such as tree stumps, root balls, and
7-6
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brushy debris, require long trees to be cut to fit into the tub of the grinder (ESEI 2014). Land
clearing applications often involve removing and processing large trees; a horizontal grinder was
the grinding equipment used in the LCD processing LCI. Two LCIs were constructed to model the
use of LCD in mulch. One process includes the grinding of LCD into mulch, which is illustrated
by the flow diagram in Figure 7-4.
7.6.1.1 LCD Processing into Mulch
The emissions from processing LCD include those associated with materials and energy (e.g.,
transportation and equipment fuel) input as well as process non-energy emission released during
grinding/chipping and storing the processed materials. Springsteen et al. (2011) documented the
average fuel consumption for grinding woody biomass. The biomass was generated from a
prescribed tree thinning in California and included only non-merchantable forest debris. The
consumption estimates provided by Springsteen et al. (2011) for loading debris into the grinder
with an excavator and grinding the material with a horizontal grinder were 0.79 and 2.92 L of
diesel fuel perMT of green material, respectively; other studies have reported similar grinder fuel
usage (Jones et al. 2010 and Pan et al. 2008). The horsepower rating of each piece of equipment
listed in Springsteen et al. (2011) was used in conjunction with the US EPA (2014) MOVES
database to quantify equipment usage according to fuel consumption.
Land clearing debris;
at collection
Transport, single unit truck;
short-haul; diesel powered
Waste transport;
at local road; unpaved
Stockpiled material;
at drop site
Heavy equipment
Office trailer
t
PM
Fuel combustion
products
*
Surface
water
t
t
Water vapor Methanol
Land use
Processing of land
clearing debris;
mulched; at
processing facility
Water
i
Diesel
Electricity
Land clearing debris;
mulched;
at processing facility
Legend
Elementary Flow
Technosphere Flow ^
¦ Not Included
Partially Included
Included
System Boundary
Figure 7-4. Unit process flow diagram for mulched land clearing debris atthe processing facility.
The process non-energy-related emissions from LCD grinding include particulate matter emission
and liquid emission from wood/wood chip stockpiles. AP-42 presents air emission factors for a
log chipping operation as part of a medium-density fiberboard (MDF) manufacturing operation.
These data were used as a proxy for an LCD wood grinding operation until measurements from
operating facilities become available. However, it appears that unlike LCD/CDD processing
facilities, engineering controls such as a cyclone and/or fabric filter collection are implemented to
control particulate matter emission from chipping operations at MDF manufacturing facilities. The
use of log chipping air emissions as a proxy would, therefore, underestimate particulate matter
emission from an LCD wood processing facility. LCD is assumed to remain at processing facilities
for a short enough duration of time that both gas emissions from biodegradation and leachate
7-7
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emissions from material stockpiling are considered negligible; these emissions are not included in
this dataset.
This dataset utilizes flows from additional data sets to simulate the release of PM as a result of
LCD transport over unpaved roads, and as a result of LCD and mulch stockpiling operations;
"transport; unpaved road" and "stockpiled material; at drop site," respectively. The dataset user
should modify these upstream processes to account for a 69.7 m, one-way length of unpaved road,
and atotal of 4 material drops. Additional process-specific parameters should be updated forthese
upstream processes as information is available.
An evaluation of 11 LCD or LCD/wood processing facility sites across Florida and California was
conducted by reviewing Google Earth aerial imagery and publicly available permit documentation
(from state environmental agency websites) to estimate the average site area and throughput of a
representative nationwide-average LCD processing operation. Land occupation was estimated
assuming a 20-year facility operational life.
Water consumption for grinder PM emissions control was estimated using the midpoint of a range
of water consumption values provided for an industry-popular dust emissions control device.
While a default value of zero is included in this dataset, additional water consumption associated
with unpaved road dust emissions control was parameterized with a range including a maximum
value for near 100% dust emissions control efficiency, based on US EPA (1985).
Electricity consumption was based on power usage information provided by a confidential source
which responded to the CDRA (2015) survey. The confidential source is an LCD and CDD wood
processing facility; year 2013 electricity consumption was divided by the total LCD and CDD
wood processed by the site in 2013.
Some information was unavailable and required assumptions based on industry experience to
estimate. In the absence of additional information, all incoming LCD was assumed to be
transported 20 km to the processing facility. Industry experience suggests that 12' x 56' standaid
office trailers are commonly used for administrative purposes at LCD and other construction and
demolition debris processing facilities. Office trailer utilization was accounted for in this dataset.
A water balance was conducted assuming that all water consumption for the purposes of dust
control is returned to the atmosphere as water vapor, whereas all water resulting from rainfall onto
paved surfaces is collected and managed as stormwater. Average nationwide precipitation was
estimated from NO AA (2016). Due to an absence of information, specific emissions associated
with stormwater discharge to surface water from these facilities was not included in this dataset.
Therefore, the surface water emissions flow is only considered partially accounted for.
7.6.1.2 Land Application of Mulched LCD
This process represents the emissions and environmental burdens associated with the land
application of mulched LCD, at an application site. This process includes the emissions associated
with the transportation and spreading of LCD mulch. This process is from gate to application site,
and does not include the production of mulch. This process accounts for the byproducts from
aerobic mulch degradation. Figure 7-5 shows the flows included in the LCI dataset.
7-8
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As the wood decomposition in this scenario would occur under aerobic conditions, gas emissions
from the land application of mulch were estimated using Beck Friis et al. (2000) which estimated
total carbon emissions from aerobic degradation of green waste. The carbon content of wood
reported by Barlaz (1998) was used for calculation. The moisture content of green lumber is taken
fromBriggs (1994).
The liquids emissions from land application of mulch are expected to be the same as those from
LCD disposal in CDD materials landfills as leaching is primarily influenced by natural
precipitation - leachate emissions are estimated from Synthetic Precipitation Leaching Procedure
(SPLP) measurements by Jang (2000) and Townsend et al. (1999). These results were scaled by
the moisture content of oven-dried wood used in these leaching studies with the moisture content
of green lumber to estimate a more accurate solid material leaching rate. These moisture contents
were taken from Briggs (1994).
Equipment used for mulch placement is estimated based on equipment required to spread mulch
found in RSMeans (2011). Fuel use is estimated based on taking the average ratio of engine
horsepower and fuel consumption (L/hr) for "light" duty for 6 models of skid-steer found in
Caterpillar (2006), and multiplying by 30 HP from RSMeans (2011). Gasoline equipment lifetimes
in nonroad equipment were estimated using the US EPA (2014) MOVES. The MOVES model
runs are for fleet average uses of nonroad equipment in the construction and industrial sectors, and
provide total fuel usage for all equipment estimated to be in operation in the US by type and
horsepower class. Types are defined by MOVES specific implementation of EPA Source
Classification Codes. Total fuel use is divided by reported population to estimate fuel use per unit
equipment per year. Using primary data for nonroad equipment used on site, each is matched to a
MOVES type and horsepower class. Median lifetimes in years are provided in MOVES by
equipment classification and engine horsepower. The amount of a vehicle used is based on the
proportion of fuel used over the equipment's lifetime. It was assumed that the LCD mulch would
be transported 20 km to the mulch end user.
7-9
-------
Transport, single-unit truck;
short-haul; diesel powered
Heavy equipment
Land clearing debris;
mulched;
at processing facility
Stockpiled material;
at drop site
*
Ground
water
Leachate
t
C02
Land application of
land clearing debris
mulch; at land
application
Gasoline
Water
T
Oxygen
Legend
Elementary Flow ^
Technosphere Flow
| Not Included
| Included
System Boundary
Figure 7-5. Unit process flow diagram for the land application of mulched land clearing debris.
7.7 Data Gap Analysis and Opportunities for Additional LCI Data
Table 7-3 summarizes the type of US-based LCI identified from compilations of LCD EOL
management sources. Most of the data identified are recognized as being partial, as they provide
only energy consumption data or emissions from one aspect (e.g., air, water, materials). Overall,
limited EOL-specific LCI are available for LCD, which is likely a result of limited material
management tracking of LCD. Based on a review of the available information, the following data
gaps were identified for compilation of a more comprehensive LCI dataset for LCD EOL
management:
1. Long-term leachable emissions from LCD products placed in a landfill. The liquid
emissions presented in this study are based on SPLPand TCLP tests on untreated wood, which
simulates leaching from disposal in an unlined CDD landfill (or land-application), and MSW
landfill, respectively. The leaching test data of untreated dimensional lumber was used as a
proxy to estimate liquid emissions for LCD due to lack of data. Although woody material is
the primary constituent of LCD, other LCD constituents such as leaves, roots, stems, bark may
impact liquid emissions. The batch leaching data used for estimating liquid emissions
correspond to an L:S ratio of 20 and are, therefore, not representative of complete liquid
emission. Furthermore, the standardized leaching tests simulate leaching associated with
physical and chemical mechanisms and do not simulate leaching associated with biological
decomposition of wood due to the short duration of these tests (18 hours). Future research
should assess leaching over a greater L:S ratio. Research efforts involving biological
decomposition leaching should also consider the potential presence of pesticides or herbicides
that may have historically been used to control insects and vegetation where LCD is generated.
2. Long-term gaseous emission from LCD biodegradation in landfill. The data reported for
branches were used as a proxy for estimating gaseous emissions from anaerobic biodegradation
of LCD disposed of in landfills, due to lack of LCD-specific data. Moreover, the emissions of
7-10
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only two compounds (methane and carbon dioxide) are included in the LCI for landfill disposal
of LCD products. Future research should quantify a larger suite of gaseous emissions from
LCD to assess the impact of large amounts of LCD disposed of in landfills and/or landfills that
collect primarily LCD materials.
3. Water consumption and emissions from LCD processing. Water consumption data for
operating LCD processing facilities were not available. Emissions from a log chipping
operation for the production of medium density fiberboard (MDF) were used as a proxy for
emissions from LCD grinding. Particulate matter emissions from log grinding for MDF
manufacturing were not detected, likely due to air pollution control practices for MDF
manufacturing operations. The LCI for LCD processing does not include particulate matter
emissions or liquid emissions from short-term LCD stockpiling due to the lack of these data
Future research should consider collecting and compiling these emissions data.
4. LCD composting. There is a large body of literature available for composting organics (e.g.,
yard waste, food waste). The data pertaining to LCD composting, however, are lacking. Future
research should quantify air and liquid emissions from LCD composting operations.
5. ACI emissions data for onsite burning of LCD. Research has been conducted on open
burning of LCD and related materials; there have been some studies conducted with ACIs
burning woody materials, and one instance of burning LCD (Lutes and Kariher 1997). While
ACI burning is recommended as a method to reduce emissions compared to emissions from
open burning, there is little data confirming the effectiveness of ACIs burning LCD. Future
research should consider collection and compilation of these data, since burning LCD at the
site of LCD generation is still presumed to be the most commonly used EOL management
strategy.
Table 7-3. Overview of life cycle inventory data sources associated with land clearing debris. "P"
represents partial coverage of the required data, while "X" represents full coverage, and
indicates no data.
Process
AIM 2
NREL
Springsteen et
al. (2011)
Cochran
(2006)
Lutes and
Kariher
(1997)
Onsite Disposal by Burning
P
-
-
-
P
Transport
-
X
-
-
-
Mulch Processing and Use
P
-
P
P
-
Combustion of LCD for
Energy
P
X
P
-
-
7-11
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7.8 References
Air Burners (2012). Fire Box Specification S-327. Palm City, Florida, USA.
Angus-Hankin, C., Stokes, B., and Twaddle, A. (1995). The Transportation of Fuelwood from
Forest to Facility. Biomass and Bioenergy, 9(1-5): 191-203.
Barlaz, M.A. (1998). Carbon Storage during Biodegradation of Municipal Solid Waste
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380.
Boldrin, A., Andersen, J.K., Moller, J., Favoino, E., and Christensen, T.H. (2009). Composting
and Compost Utilization: Accounting of Greenhouse Gases and Global Warming
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CAR (2010). Climate Action Reserve, http://www.climateactionreserve.org/resources/. Accessed
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Cochran, K.M. (2006). Construction and Demolition Debris Recycling: Methods, Markets, and
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Ecobalance (1999). Life Cycle Inventory of a Modern Municipal Solid Waste Landfill. A Report
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ERGI (2001). Open Burning Volume III: Chapter 16. Prepared by Eastern Research Group, Inc.
for the Area Sources Committee Emission Inventory Improvement Program, January
2001.
ESEI(2014). Tub Grinder vs. Horizontal Grinder. http://earthsaverequiprnent.com/Equipment-
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Haug, R.T. (1993). The Practical Handbook of Compost Engineering. Ann Arbor, MI: Lewis
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IWCS (2014). Personal communication with confidential client and Pradeep Jain, Innovative
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Jang, Y. (2000). A Study of Construction and Demolition Waste LeachateFrom Laboratory
Landfill Simulators. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Jenkins, B.M., Baxter, L.L., and Miles, T.R. (1998). Combustion Properties of Biomass. Fuel
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Jones, G., Loeffler, D., Calkin, D., and Chung, W. (2010). Forest Treatment Residues for
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Pan, F., Han, H.S., Johnson, L.R., and Elliot, W.J. (2008). Net Energy Output from Harvesting
Small-Diameter Trees Using a Mechanized System. Forest Products Journal, 58: 25-30.
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Burning, Journal of the Air and Waste Management Association, 61(1): 63-68.
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Waste: A Case Study Using Ash from the Combustion of Wood and Tires. Ph.D.
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Demolition Waste Leachate. Journal of Environmental Engineering, ASCE, 125 (11),
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Review on Biomass Classification and Composition, Co-Firing Issues and Pretreatment
Methods. 2011 ASABE Annual International Meeting.
US EPA (1985). Dust Control at Hazardous Waste Sites. Hazardous Waste Engineering
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US EPA (1991). Nonroad Engine and Vehicle Emission Study - Report. Office of Air and
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USEPA (1996). AP-42, Fifth Edition, Volume I, Chapter 2, Section 2.1: Refuse Combustion,
http://www. epa.gov/ttn/chief/ap42/ch02/ind ex. html
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Identification of Nonhazardous Secondary Materials that are Solid Waste Construction
and Demolition Materials - Land Clearing Debris. February 2011.
http://www.epa.gov/waste/nonhaz/define/rulemaking.htm
US EPA (2014). Motor Vehicle Emission Simulator (MOVES): User Guide for MOVES 2014.
EPA-420-B-14-055. July 2014.
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February 2014.
Wiltsee, G. (1998). Urban Wood Waste Resource Assessment. A Report Prepared by National
Renewable Energy Laboratory Managed by Midwest Research Institute for U.S.
Department of Energy. November 1998.
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8 Portland Cement Concrete
8.1 Introduction
Portland cement concrete (PCC) is a composite material formed from fine aggregates (i.e., sand),
coarse aggregates (e.g., gravel, crushed stone), binder (Portland cement), water, and stabilizers.
Aggregates, cement, and water represent by volume 60-70%, 10-15%, and 15-20% of the concrete
mix, respectively (PCA 2014a). The remaining volume of concrete is entrained air and any
additional stabilizers or other amendments, i.e. admixtures, to enhance desired properties of the
concrete. Concrete is widely used in the construction industry due to its versatility, strength, and
cost. Assuming a density of 150 pounds per cubic foot, nearly 480 MMT of ready-mix concrete is
used annually in the US (PCA 2014b). Ready-mix concrete is the most commonly used type of
concrete and it accounts for about 75% of all concrete used annually (PCA 2014b).
Pavements, bridges, and various components of airports and buildings that have been constructed
from concrete may be rehabilitated or demolished and reconstructed due to wearing or damage
that has occurred over time. Concrete may be removed by different techniques (e.g., blasting,
crushing, cutting, impacting, milling, and splitting), which are determined based on factors such
as cost, project duration, the quality of the concrete, the potential for recycling, transport distances,
and accessibility (Lee etal. 2002, Lechemi etal. 2007, and Woodson 2009).
Once removed, reclaimed PCC may be recycled or disposed of in a landfill. The concrete is
typically processed (e.g. crushing, sorting, metal removal) prior to use in a recycling application.
Figure 8-1 identifies the flow of materials and processes that are included in this report's associated
datasets. Most commonly, recovered concrete is recycled as aggregate (i.e., "recycled concrete
aggregate" (RCA)) in road base, for new concrete mix, or for asphalt pavement mix production
(NAPA 2020). Closed-loop recycling of concrete, where RCA replaces both primary aggregate
and cement, is currently not a common practice.
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End-of-life
Product
Removal
Aggregate as
Site Fill
CDD MRF
Debris as
Site Fill
CDD
Landfill
MSW
Landfill
Figure 8-1. Process flow diagram for the end-of-life management of Portland cement concrete.
8.2 EOL Management
As presented in Figure 8-1, there are three primary EOL management pathways for concrete: use
as RCA, use as general fill, and landfill disposal. CDRA(2014) reports that approximately 127
MMT of concrete are recycled annually; however, the basis of this estimate of recycled concrete
is not well documented. Concrete (with and without rebar, painted and unpainted concrete)
represents approximately 10.8 to 15.2% by mass of CDD materials received at CDD landfills
(CCG 2006, CDM 2009, and RWB et al. 2010). Based on US EPA's estimate of total CDD
landfilled (as presented in Chapter2 of this report), approximately 9.5 to 13 MMT of concrete was
disposed of in landfills in 2011. Based on information from CDRA (2015), Turley (2002), and
Wilburn and Goonan (1998), the total amount of concrete recovered for EOL management in the
US annually ranges from 212 to 254 MMT.
Figure 8-2 presents the distribution of concrete debris used in different applications in the US
based on US EPA's estimate of concrete landfilled, CDRA's estimate of concrete recycled, and
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the distribution of concrete uses reported in Deal (1997). The use of aggregate for road base is the
most common management option for RCA (FHWA 2004, CTCA 2012). Use of RCA in asphalt
mixes is another desirable option as it can improve the stability and surface friction of thepavement
apart from offsetting the production of primary aggregate (Snyder and Rodden n.d.). However, the
use of RCA in asphalt mixes can increase the need for greater asphalt content in the paving mix
due to RC A's absorptive properties (Snyder and Rodden n.d.).
Figure 8-2. Distribution of recycled concrete applications in the US (Deal 1997).
Recycling RCA into new concrete mixes poses challenges, such as decreased workability due to
angular structure, reduced durability due to potential alkali-silica reaction, and reduced
compressive strength when substituted for fine aggregates or substituted for more than 30% of
coarse aggregates (Hansen 1986, Li and Gress 2006, Mclntyre et al. 2009, and Hiller et al. 2011).
In a survey of state concrete recycling practices, 10 out of 30 responding states allowed the use of
crushed concrete for road surface course (CTCA 2012). However, only two states, Alabama and
Texas, reported this use as a common practice.
Recycled concrete can also be used as riprap. FHWA (2004) reported that most states allow
processed recycled concrete to be used as riprap for erosion control, as long as steel reinforcement
has been removed prior to use. Demolished concrete can also be used in fill applications (e.g.,
embankments) as a substitute for natural soil. The "drainage aggregates" includes uses such as
drainage fields and pipe bedding. The "Other" category includes use as railroad ballast and
landscaping rock.
Table 8-1 lists and describes the LCA processes included in this report's associated datasets for
EOL management of concrete. The emissions associated with energy and materials requirements
and process non-energy emissions (e.g., fugitive dust, liquid emissions associated with disposal of
concrete in a landfill) were taken into account in compiling the different LCI datasets.
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Table 8-1. End-of-life management
Process
LCA considerations
Unlined CDD
landfilling of concrete;
at unlined CDD landfill
This process includes material (e.g., equipment, soil, water) and
energy (e.g., fuel, electricity) inputs for placing and compacting
discarded concrete in a CDD landfill along with non-energy
emissions (e.g., dust emissions from equipment operation and liquid
emissions associated with physicochemical degradation of concrete
in a landfill).
MSW landfilling of
demolished concrete; at
MSW landfill
This process includes material and energy inputs from disposal of
concrete in an MSW landfill. It also includes landfill construction,
operation and closure/post-closure care.
Use of recycled
concrete aggregate; as
general fill; at fill site
RCA may be used as primary aggregate substitute in a variety of
applications, including road base construction, asphalt or concrete
mix production, or as riprap. This use of RCA precludes the
production of an equivalent amount of primary aggregate.
Use of concrete debris;
as general fill; at fill
site
Concrete may be used as a soil substitute in fill applications. When
used as a fill material, the concrete will likely not need the
processing and sorting requirements necessary for using concrete as
aggregate. This use of demolished concrete precludes the production
of natural soil.
8.3 LCI Sources
Peer-reviewed literature, government and private industry publications, and various LCA
modeling tools were reviewed to identify LCI datasets pertaining to concrete EOL management
processes. Table 8-2 lists sources reviewed to develop the LCIs presented in this chapter. If LCI
data were not available, process metadata and documentation (e.g., included emission categories,
background dataused to compile the dataset, geographic location and time period of the data) were
reviewed to evaluate the completeness of the dataset. Primary information sources were used
wherever possible.
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Table 8-2. Overview of life cycle inventory data sources associated with Portland cement concrete.
Source
Description
Wilburn and
Goonan (1998)
Wilburn and Goonan (1998) provide energy requirements associated
with crushing/sorting stone, sand and gravel, and concrete. These data
originate from the Portland Cement Association and from an energy
audit of a recycling facility in Denver, Colorado.
CDRA (2015)
The Construction and Demolition Recycling Association (CDRA)
conducted a survey of CDD processing facilities across the US.
US EPA (2019)
EPA's Waste Reduction Model (WARM) presents data on GHG
emissions associated with transporting, recycling, and landfilling
concrete.
Cochran (2006)
Cochran (2006) presents diesel energy requirements for concrete
processing equipment based on a survey of equipment manufacturers.
Jang (2000)
Jang (2000) presents batch and column leaching test data for various
CDD materials, including concrete.
NIST (2018)
BEES
The National Institute of Standards and Technology (NIST) Building for
Environmental and Economic Sustainability (BEES) model enables an
economic and environmental impact comparison across various building
materials, including concrete.
RSMeans (2011)
RSMeans provides data on the cost of construction, which includes
estimates for types and productivity of heavy equipment. This also
includes data on placement and spreading of end products for PCC.
8.4 LCI Related to Removal/Demolition
Concrete demolition generally includes breaking the concrete into manageable chunks for ease of
handling and transportation. In-place concrete characteristics may be analyzed to assess properties
and suitability foruse in targeted applications before demolition (Hiller et al. 2011). Contaminants
such as joint sealant and large portions of asphalt overlay or patch are recommended to be removed
before concrete demolition, but, depending on the RCA application, small amounts of asphalt
contamination are not detrimental (CDRA 2014). The location and nature of the project dictate the
concrete removal method and equipment required (Lechemi et al. 2007 and Woodson 2009).
Dykins and Epps (1987) and NHI (1998) [as cited in Hiller et al. (2011)] describe two general
types of equipment that can be used for breaking up concrete in highway applications to render
concrete into sizes acceptable for crushing: impact breakers and resonant breakers. Impact breakers
use individually weighted drops to break the concrete, and this equipment has greater production
rates than resonant breakers (900 to 1,100 m2/hr compared to 670 m2/hr), which use a high-
frequency, low-amplitude pulse to fracture the concrete. Resonant breakers have the advantage of
producing more uniform slabs and causing fewer disturbances to underlying infrastructure, such
as sewers and utilities.
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After the concrete has been fractured and separated from reinforcement, larger pieces of concrete
(< 2 ft or 0.6 m) can be removed using a backhoe, while a front-end loader can be used to remove
the remaining pieces. The demolished concrete may be processed on site or transported to an off site
recycling/processing facility to produce RCA; alternatively, it may be transported to a landfill for
disposal.
8.5 LCI Related to Disposal
As discussed earlier, based on published literature and US EPA's compilation of state data, it is
estimated that approximately 9% of the concrete generated in the US is disposed of in landfills.
Emissions associated with concrete disposal in a landfill include air emissions from equipment
used for placing concrete in the landfill, emissions associated with landfill construction and
operation, and liquid emissions from material decomposition in the landfill. There are various
sources which provide landfilling emission factors related to equipment use and landfill
construction and operation; however, none are specific to concrete landfilling. LCI data for
generalized landfill construction and operations are presented in Chapter 2. Since concrete is a
non-biodegradable material, the flows included in CDD and MSW landfilling are represented by
Figure 2-5 and Figure 2-8, respectively.
Concrete exposure to precipitation or other liquids (e.g., landfill leachate) is expected to result in
contaminants leaching and the emissions are expected to depend on biogeochemical environment
(e.g., MSW landfill, CDD landfill). Leachable emissions were estimated using the synthetic
precipitation leaching procedure (SPLP) and leaching column data reported by Jang (2000). Jang
(2000) conducted leaching tests on several individual CDD materials, including concrete collected
from a concrete recycling facility in Florida, to assess leaching of conventional water parameters
(e.g., pH, conductance, total oxygen demand (TOD), chemical oxygen demand (COD), ions, and
heavy metals). The liquid-to solid (L:S) ratio of the batch tests in Jang (2000) were much greater
than the L:S ratio in the column tests (20:1 versus 1.3:1); batch data were used for parameters
(calcium, chloride, potassium, and sodium) that were measured above the detection limits (because
of the greater propensity to leach pollutants). Batch test concentrations were multiplied by the total
solution volume and divided by the sample mass to estimate leachability on a per-kilogram-
concrete basis.
For parameters that were either below the detection limit in the batch testing experiment or were
not measured during the batch test, column test data were used to develop leaching LCI.
Leachability of COD and magnesium were calculated from column test data by summing the total
mass of pollutant leached and dividing this mass by the mass of the concrete material in the
column. Nitrate and sulfate emission data reported by the study were not used for developing the
LCI dataset since SPLP extraction fluid contains these anions. TDS data reported by the study
were also excluded to avoid double-counting emissions, as some of the contaminants are included
in TDS measurement.
Concrete leaching data, energy consumption data from landfill operations, estimated cover soil
demand, and an assumed transport distance were used to develop an LCI process dataset for
disposing of concrete at an unlined CDD landfill. The energy use and the associated emissions
from landfill operations (e.g., waste placement, compaction) include diesel use in heavy equipment
and electricity use in landfill buildings (e.g., administrative buildings, workshop); calculations
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detailing these input flows are included in Chapter 2 of this report. The bulk density of loose
concrete used to estimate cover soil requirements was provided by CCG (2006). In the absence of
nationwide average transport data, it was assumed that concrete would be transported 20 km for
landfill disposal.
8.6 LCI Related to Recycling
8.6.1 Concrete Processing
The recycled end use and the project's location dictate the degree of processing necessary before
concrete debris can be beneficially used. For example, no processing may be needed if concrete is
used to replace natural soil fill in a non-load-bearing fill application. RCA production, on the other
hand, requires extensive processing. Concrete may be processed on-site using mobile equipment
if the processed concrete is intended to be used at the site. It may be transported for off-site
processing for future end uses. Concrete commingled with other CDD materials would need to be
segregated prior to processing.
Land occupation was estimated by taking the average of the developed site areas (or if the site
elected to remain confidential, the total reported site areas) and dividing this value by the average
reported material diversion rate, assuming a facility lifespan of 20 years. The average site
developed area was also used to estimate the average length of onsite unpaved roads. This was
calculated as half the square root of the average site developed area, which would represent the
distance from the edge to the center of the developed area of the site (assuming a square site).
This dataset calls in flows from additional datasets to simulate the release of PM as a result of
materials transport over unpaved roads, and as a result of material stockpiling operations. These
PM emissions are accounted for by the "transport; unpaved road" and "stockpiled material; at drop
site" inflows, respectively. The dataset user should modify these upstream processes to account
for a 105-meter, one-way length of unpaved road, and a total of 2 material drops (i.e., discharge
from conveyor, loading into haul trucks). It was assumed that incoming loads of demolished
concrete chunks, due to the relatively small fraction of fines, provide a negligible contribution to
PM emissions during stockpiling operations. Indeed, OSHA also mandates that workers in the
construction industry be exposed to concentrations of airborne crystalline silica no greater than 50
|ig/m3 per 8-hour workday (OSHA 2016).
Further process-specific parameters should be updated as more reliable data for upstream
processes become available.
Water consumption for crusher/screener/conveyance PM emissions control is estimated using the
midpoint of a range of water consumption values provided for an industry-popular dust emissions
control device. While a zero-default value is included in this dataset, additional water consumption
associated with unpaved road dust emissions control was parameterized with a range including a
maximum value for near 100% dust emissions control efficiency (US EPA 1985).
Some assumptions were necessary in areas where information is lacking. In the absence of
additional information, all incoming material was assumed to be transported 20 km to the
processing facility. Industry experience suggests that 12' x 56' standard office trailers are
commonly used for administrative purposes at concrete and other construction and demolition
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debris processing facilities. Office trailer utilization was accounted for in this dataset. Due to an
absence of information, specific emissions associated with stormwater discharge to surface water
from these facilities were not included in this dataset. Therefore, the surface water emissions flow
is only considered partially accounted for in this dataset.
A water balance was conducted assuming that all water consumption for the purposes of dust
control is returned to the atmosphere as water vapor, whereas all water resulting from rainfall onto
paved surfaces is collected and managed as stormwater. Average nationwide precipitation was
estimated from NOAA (2016).
8.6.2 RCA Use as Aggregate
Recycled concrete is most commonly used to replace primary aggregate in the following
applications: road base, subbase, asphalt and cement concrete, and riprap. RCA produced from
demolition activities will likely need to be crushed, processed to remove contaminants such as
steel, and sorted before use in these applications to meet gradation specifications. The primary
emissions resulting from the use of RCA as an aggregate material (after the concrete has been
processed) include leaching to groundwater and surface water. Granular base (unbound)
applications of RCA have been shown to leach calcium carbonate precipitate that can clog drainage
pipes, particularly if there is a large amount of fine material in the RCA, and may restrict use in
various drainage applications (Gupta 1993, Snyder 1995, Steffes 1999). Although there have been
no reported problems with this precipitate for embankment applications, there is the possibility for
highly alkaline precipitates to occur (FHWA 2012). The liquid emissions from unbound RCA are
expected to be similar to those from an inert landfill. While it would take longer for bound RCA
(e.g., use in asphalt pavement or new concrete mixes) to leach contaminants, it is expected that
over an infinite time horizon, these emissions would ultimately be the same.
Leachable emissions were estimated using batch leaching tests (SPLP) and leaching column data
reported by Jang (2000). Jang (2000) conducted leaching tests on several individual CDD
materials, including concrete (size reduced to 1") collected from a concrete recycling facility in
Florida to assess leaching of conventional water parameters (e.g., pH, conductance, TDS, and
COD), ions and heavy metals. The L:S ratio of the batch tests in Jang (2000) was much greater
than the L:S ratio in the column tests; therefore, batch data were used for parameters (calcium,
chloride, potassium, and sodium) that were measured above the detection limits (because of the
greater propensity to leach pollutants). Batch test concentrations were multiplied by the total
solution volume and divided by the sample mass to estimate leachability on a per-kilogram-
concrete basis.
For parameters that were either below the detection limit or not measured in the batch testing
experiment, the column experiment data was used (magnesium and COD). An additional column
leaching experiment is included from Townsend et al. (1999) which included data on carbonate, a
parameter not measured in Jang (2000). Nitrate and sulfate results from Jang (2000) were not
included because the SPLP solution uses sulfuric and nitric acid; therefore, the sulfate and nitrate
results may not represent only leached values.
This dataset calls in a flow from an additional dataset to simulate the release of PM as a result of
material stockpiling. These PM emissions are accounted for by the "stockpiled material; at drop
site" inflow. The dataset user should modify this upstream process to account for one material drop
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(i.e., discharge). Additional process-specific parameters (i.e., wind speed, moisture content) should
be updated for this process if information is available.
Information on equipment and fuel used to place the concrete aggregate was taken from RSMeans
(2011) item 31 23 23.17 0020: placement of gravel as general fill. These data include bulldozer
use and productivity (area per time). Fuel use for the bulldozer from RSMeans (2011) was
estimated using fuel use datafrom five bulldozers given in Caterpillar (2006). Since all the engines
in Caterpillar (2006) were larger than the engine size given in RSMeans (2011), the datafrom
Caterpillar (2006) were fit to a power curve with engine size as the independent variable and fuel
use per engine size as the dependent variable. The fuel use estimate for the equipment, combined
with the equipment productivity from RSMeans (2011), were used to estimate the fuel use per
functional unit. The lifetime of the nonroad equipment (i.e., for equipment utilization per
functional unit) was estimated using the equipment-specific lifetime fuel consumption values
presented intheUS EPA(2014) MOVES database. The MOVES database presents average annual
fuel consumption of specific equipment categories. Individual fuel consumption by each
equipment type was estimated based on the proportion of average fuel use for that equipment type
to the remainder of the fuel usage estimated for the other equipment types.
The use of RCA in an aggregate fill application avoids the need for the production of primary
aggregates. Since the transport distance from the processing facility to the aggregate fill site is
unknown, an average distance of 20 kilometers was assumed. Figure 8-3 presents the flows
included in the LCI dataset for RCA used to replace primary aggregate.
Fuel combustion
byproducts
Ground
water
t
Leachate
Transport, single unit truck;
short-haul; diesel powered
Heavy equipment
Recycled concrete
-aggregate; at concrete
processing facility
Stockpiled material;
at drop site
Use of recycled concrete
aggregate; as general fill;
at fill site
Diesel
i
Water
Legend
Elementary Flow
Technosphere Flow
: 1
| Not Included
| Included
System Boundary
Oxygen
Figure 8-3. Unit process flow diagram of recycled concrete aggregate used as general site fill.
8.6.3 Demolished Concrete Use as Soil Fill Replacement
The use of demolished concrete to replace natural soil fill (e.g. lake fill, embankment fill) would
avoid the emissions resulting from the production and transport of natural soils. Figure 8-4 presents
the flows included in the LCI dataset for demolished concrete used to replace natural soil. While
leachable emissions are assumed to ultimately be the same whether demolished concrete is used
as a substitute for bound aggregate, unbound aggregate, or natural soil, this dataset assumes that
the concrete debris used as a fill material does not need to be size reduced or screened. It also uses
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theRSMeans (2011) labor category of "riprap placement" as opposed to "general fill" since the
size of the placed concrete pieces is assumed to be larger without processing. All other assumptions
and data sources are the same as the previous section on RCA use as an aggregate.
Transport, single unit truck;
short-haul; diesel powered
Heavy equipment
Concrete debris;
demolished;
at collection
Stockpiled material;
at drop site
+
Ground
water
Leachate
Use of concrete debris; as
general fill; at fill site
Diesel
T
Water
Oxygen
Legend
Elementary Flow ,
Technosphere Flow
| Not Included
| Included
System Boundary
Figure 8-4. Unit process flow diagram for concrete debris used as general site fill.
8.7 Data Gap Analysis and Opportunities for Additional LCI Data
Table 8-3 summarizes the type of US-based LCI identified from reviewed concrete EOL
management sources, including Wilburn and Goonan (1998), Jang (2000), and Cochran (2006).
Wilburn and Goonan (1998) and Cochran (2006) only provide partial data on present energy
consumption information for concrete processing. Data sources that are not the primary source of
LCI data have not been included in Table 8-3. Overall, limited EOL-specific LCI are available for
concrete. Based on a review of available information, the following data gaps were identified for
compiling a more comprehensive LCI dataset for concrete EOL management:
1. Leachable emissions from the landfill disposal or use of RCA as aggregate or in the
replacement of natural soils are lacking. Although an estimate of liquid emissions from
concrete disposal inunlined CDD landfills and for use as aggregate and natural soil is included
in the LCI datasets and was developed based on batch and column concrete leaching test data
reported by one study (Jang 2000), these liquid emissions are associated with a maximum L:S
of 20 and do not represent the complete liquid emissions over an infinite time horizon. Future
research should consider estimating the long-term leaching of RCA at an L:S ratio that would
represent complete contaminant leaching.
2. Limited data available for quantifying concrete carbonation. Studies have shown that the
cement portion of concrete can, over time, absorb carbon dioxide in a process called
carbonation. There have been several recent studies (e.g., Padeand Guimaraes 2007, Dodoo
et al. 2009, Collins 2010, and Garcia-Segura et al. 2014) describing and comparing
observations of carbon dioxide uptake in different stages of the life cycle of concrete (e.g.,
calcination of cement, concrete usage phase, demolition of concrete, use of crushed concrete
as an aggregate). Most of these studies estimate carbon dioxide uptake in concrete by using
predictive modeling based on Fick's law of diffusion and a carbon uptake equation developed
by Lagerblad (2005). This modeling approach incorporates the following parameters into the
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estimation of carbon dioxide uptake: the amount of carbonation that has already occurred in
the concrete, the amount of Portland cement in the concrete, the amount of calcium oxide in
cement, the molar weight of oxide (carbon dioxide/calcium oxide), the service life of the
concrete, the exposed concrete surface area, and a carbonation rate coefficient based on the
strength and environmental exposure conditions of the concrete. The EOL management of
concrete has been identified as an important component in calculating carbon dioxide
emissions in the life cycle of concrete (Collins 2010). Concrete EOL management may
enhance carbonation, particularly in recycling applications because concrete is commonly size
reduced when it is recycled; this creates a greater surface area that exposes fresh uncarbonated
carbon for carbonation. Due to lack of data, a carbonation factor was not included in the LCI.
3. Differences in transportation emissions between mobile concrete processing operations
and stationary concrete processing facilities. Concrete recycling can occur at the site of
demolition (e.g., a concrete pavement being demolished and then crushed for use as subbase
onsite) and a mobile processing unit, usually smaller than a concrete processing facility, can
be brought to the construction site. The differences in using materials at the site of demolition
will likely be realized in emissions savings in transportation when comparing the emissions
from the transportation necessary to mobilize the processing equipment to the site and back
and loading up and trucking large amounts of material to a stationary facility. Details on the
distances mobile operations are transported, how they are transported, and average distances
and the methods of transporting (e.g., barge, trucks, train) concrete to a recycling facility would
be valuable in improving transportation emissions estimates to compare each processing
option.
4. Water consumption and LCI for mobile versus stationary concrete processing facilities.
Mobile processing units are usually smaller and likely less efficient than a stationary concrete
processing facility. Data for water usage for controlling particulate matter emissions are
lacking.
5. No data were found to assess the difference in the energy requirement for processing
reinforced and non-reinforced concrete. Amounts of waste produced from processing
concrete were not identified. Although a majority of non-reinforced concrete may result in
very little waste to be disposed of, the recovery rate of concrete from reinforced concrete is
likely to be less because concrete can remain stuck to steel mesh or rebar.
Table 8-3. Overview of life cycle inventory data sources associated with Portland cement concrete.
"P" represents partial coverage of the required data, while "X" represents full coverage, and
indicates no data.
Wilburn and
Goonan
Cochran
Process
(1998)
Jang (2000)
(2006)
Transport
P
-
-
Landfill Leachate
-
X
-
Demolished Concrete Processing
P
-
P
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8.8 References
Caterpillar (2006). Caterpillar Performance Handbook. Edition 36. Cat publication. Caterpillar
inc., Peoria IL.
CCG (2006). Targeted Statewide Waste Characterization Study: Waste Disposal and Diversion
Findings for Selected Industry Groups. A Report Prepared by Cascadia Consulting Group
for the California Integrated Waste Management Board, June 2006.
CDM (2009). Illinois Commodity/Waste Generation and Characterization Study. A Report
Prepared by CDM Smith Commissioned by Illinois Department of Commerce & Economic
Opportunity and Contracted by the Illinois Recycling Association, 22 May 2009.
CDRA(2014). Concrete: Good economic sense, https://cdrecycling.org/materials/concrete.
CDRA (2015). The Benefits of Construction and Demolition Materials Recycling in the United
States. Report prepared for the Construction and Demolition Recycling Association by the
Department of Environmental Engineering Sciences, Engineering School of Sustainable
Infrastructure and Environment, University of Florida. Version 1.1 updated 14 January
2015.
Chen, J., Brown, B., Edil, T.B., Tinjum, J. (2012). Leaching Characteristics ofRecycled Aggregate
used as Road Base. University of Wisconsin System Solid Waste Research Program:
Student Project Report, May 2012.
Cochran, K. M. (2006). Construction and Demolition Debris Recycling: Methods, Markets and
Policy. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Cole, R.J., Kernan, P.C. (1996). Life-Cycle Energy Use in Office Buildings. Building and
Environment, 31 (4), 307-317.
Collins, F. (2010). Inclusion ofCarbonationDuring theLifeCycle ofBuilt andRecycled Concrete:
Influence on Their Carbon Footprint. International Journal of Life Cycle Assessment 15,
549-556.
CTCA (2012). Concrete Recycling: Reuse of Returned Plastic Concrete and Crushed Concrete as
Aggregate. A Report Prepared by CTC & Associates LLC, Preliminary Investigation
Caltrans Division of Research and Innovation for Rock Products Committee: Materials and
QA Sub Task Group of the Concrete Products Task Group. Revised 7 September 2012.
Dodoo, A., Gustavsson, L., Sathre, R. (2009). Carbon Implications of End-of-Life Management
of Building Materials. Resources, Conservation and Recycling 53, 276-286.
Ecoinvent (2014). Swiss Center for Life Cycle Inventories: Ecoinvent Centre. Dataset Information
(UPR): Treatment of Waste Asphalt, Sanitary Landfill, CH.
FHWA (2004). Transportation Applications ofRecycled Concrete Aggregate. Federal Highway
Administration State of the Practice National Review, September 2004.
Garcia-Segura, T., Yepes, V., Alcala, J. (2014). Life Cycle Greenhouse Gas Emissions of Blended
Cement Concrete Including Carbonation and Durability. International Journal of Life
Cycle Assessment, 19,3-12.
Gupta, J. D., Kneller, W. A. (1993). Precipitate Potential of Highway Subbase Aggregates, Report
No. FHWA/OH-94/004 Prepared for the Ohio Department of Transportation, November
1993. http://trid .trb.org/view.aspx?id=405751
Hansen, T. C. (1986). Recycled Aggregates and Recycled Aggregate Concrete: Second State-of
the-Art Report: Developments 1945-1985. Material Structures, 19 (111), 201-246.
Hiller, J. E., Deshpande, Y. S.,Qin, Y., Shorkey, C. J. (2011). Efficient Use ofRecycled Concrete
in Transportation Infrastructure. A Report Prepared by Michigan Technological
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University, Houghton, Michigan for Michigan Department of Transportation, January
2011.
Jang, Y.C. (2000). A Study of Construction and Demolition Waste Leachate from Laboratory
Landfill Simulators. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Kelly, T. (1998). Crushed Cement Concrete Substitution for Construction Aggregates - A
Materials Flow Analysis. U.S. Geological Survey Circular 1177.
Lagerblad, B. (2005). Carbon Dioxide Uptake During Concrete Life Cycle - State of the Art.
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Centre Project, NI-project 03018.
Lechemi, M., Hossain, K.M.A., Ramcharitar, M., Shehata, M. (2007). Bridge Deck Rehabilitation
Practices in North America. J. Infrastruct. Syst. 13, 225-234.
Lee, E. B., Roesler, J., Harvey, J. T., Ibbs, C. W. (2002). Case Study of Urban Concrete Pavement
Reconstruction on Interstate 10. J. Constr. Eng. Manage. 128, 49-56.
Li, X., Gress, D.L. (2006). Mitigating Alkali-Silica Reaction in Concrete Containing Recycled
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Mclntyre, J., Spatari, S., MacLean, H.L. (2009). Energy and Greenhouse Gas Emissions Trade-
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MGE (1997). Demolition Energy Analysis of Office Building Structural Systems. A Report
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NIST (2018). Building for Environmental and Economic Sustainability (BEES) Online 2.0
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NOAA (2016). National Centers for Environmental Information, State of the Climate: National
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Pade, C., Guimaraes, M. (2007). The CO2 Uptake of Concrete in a 100 Year Perspective. Cement
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9 Polyvinyl Chloride
9.1 Introduction
Polyvinyl chloride (PVC) is a thermoplastic material (i.e., a plastic which can be repeatedly
softened and reshaped with the application of heat) commonly used in a variety of building
products such as pipe, vinyl siding, wire insulation, vinyl flooring and window frames. Apart from
being used in construction, PVC is also used in electronics, automotive parts, packaging, and other
consumer products. Without the addition of plasticizers, PVC is a rigid, light-weight plastic. While
some PVC building components may be manufactured from PVC resin without the addition of
plasticizers (e.g., unplasticized PVC (uPVC) pipe, vinyl siding), plasticizers and/or copolymers
are commonly added to PVC to make it more flexible.
PVC can be combined with a variety of different plasticizers and stabilizers at different loading
levels (Daniels 2009). Of the various plasticizers that can be used with PVC, it appears that
phthalates are one of the most common (US EPA 2012a, Stark et al. 2005). However, studies that
have shown adverse health effects in male laboratory animal reproductive system development
(US EPA 2012a) and other concerns such as cancer, and developmental and chronic organ toxicity
(CPSC 2010) have prompted research for alternatives. Stabilizers such as metals and metal
compounds are added to PVC to help prevent thermal and photodegradation. Organotin
compounds are currently one of the most common types of PVC stabilizers in use (CIWMB 2006,
ATSDR2005, The Vinyl Institute 2008).
Due to the wide variety of PVC building materials used in the US and the varying degree of
plasticizers that are incorporated into plastic product, this chapter focuses on PVC resin. Figure
9-1 identifies the flow of materials and processes included in this report's associated datasetsfor
EOL management of PVC.
Figure 9-1. Process flow diagram for EOL management of PVC products without plasticizers.
9.2 EOL Management
According to a national survey of US CDD MRFs (CDRA 2015), it appears that the targeted
recovery of PVC building materials for recycling is not common as of the time of this study. The
majority of PVC products which reach the EOL management stage are landfilled, though some
9-1
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may be incinerated in MSW incinerators since CDD materials are present in MSW (US EPA
2015a). Even though PVC has a relatively high heating value of 19.2 MJ/kg (Shemwell and
Levendis, 2000), it is not targeted for incineration due to its high chlorine content. When
combusted, PVC releases highly caustic hydrogen chloride gas (NREL 1993, Jafari and Donaldson
2009) and dioxins and furans (Kim et al. 2004, Christman et al. 1989), which are known or
reasonably anticipated to be human carcinogens (NTP 2016, US EPA 2020).
Based on the combined average weight fraction of two common PVC building materials (i.e.,
plastic piping, plastic siding/decking) as reported in several CDD composition studies (R.W. Beck
et al. 2010, CCG 2008, CCG 2009, MSW Consultants 2008), PVC materials represent
approximately 0.68% of the CDD waste stream. Using the methodology presented in US EPA
(2015a), approximately 68 million tons of mixed CDD was placed in CDD landfills across the US
in 2011. US EPA (2015a) also estimates that CDD represents approximately 10.5% of the MSW
waste stream. Survey results reported by van Haaren et al. (2010) for a 2008 survey estimated that
nearly 270 million and 26 million tons of MSW were landfilled in MSW landfills and combusted
at waste-to-energy (WTE) plants, respectively. Therefore, assuming that PVC materials represent
approximately 0.68% of mixed CDD, it can be calculated that nearly 68% of PVC materials were
placed in CDD landfills, 29% were placed in MSW landfills, and 3% were incinerated in MSW
WTE facilities. Figure 9-2 shows a summary of EOL management pathways for PVC materials
estimated in 2011.
Figure 9-2. Approximation of 2011 CDD PVC material EOL management pathways. Calculated
from US EPA 2015a and van Haaren et al. 2010.
Table 9-3 lists and describes the LCA processes included in this report's associated datasets for
EOL management of PVC products.
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Table 9-1. End-of-life management processes for PVC.
Process
Description
Unlined CDD
landfilling of polyvinyl
chloride, PVC; at
unlined CDD landfill
The landfill disposal of PVC products includes transport and
compaction at the landfill and consumption of materials for
construction, operation, closure, and post-closure care. While it is not
expected that PVC products will biodegrade to produce landfill gas, it
is possible that leaching of PVC plasticizers and stabilizers could
occur.
MSW landfilling of
polyvinyl chloride,
PVC; at MSW landfill
This process includes material and energy inputs as well as leachate
volumes expected from the disposal of PVC materials in an MSW
landfill. It also includes landfill construction, operation, and
closure/post-closure care.
Incineration of
polyvinyl chloride,
PVC; at MSW waste-to-
energy facility
While not targeted for incineration or combustion, PVC products may
inadvertently be incinerated along with MSW at waste-to-energy
facilities.
9.3 LCI Sources
Peer-reviewed literature, as well as government and private industry publications, were reviewed
to identify LCI datasets pertaining to EOL management of PVC products. Most sources focus on
analyzing air emissions from PVC resin combustion. Table 9-2 lists sources reviewed to develop
the LCIs presented in this chapter.
Table 9-2. Overview of life cycle inventory data sources associated with PVC.
Source
Description
Wang et al. (2001)
One (1) gram samples of PVC powder were combusted in a tube-
type furnace between 600-900° C. This Chinese study focused on the
measurement of chlorinated polycyclic aromatic hydrocarbons
(PAHs). Emissions were collected via a glass wool and glass fiber
filter with a pore diameter of 0.2 [j,m, which was followed by an
adsorption tube packed with XAD-2 resin.
Wang et al. (2003a)
This US study documented the emissions from combustion of 0.5-
gram granulated PVC samples at different temperatures ranging from
500° C - 1000° C. Emissions of PAHs, particulates, and aromatic and
light hydrocarbons were measured. The emissions were not
tabulated, but were presented in a series of graphs - these graphs
were used to estimate individual emission quantities. The
combustion chamber used for this experiment was a horizontal
furnace with two sections. The sample was combusted in the primary
chamber (i.e., a laminar flow muffle furnace) and the resulting
emissions were further combusted in a secondary chamber (i.e.,
9-3
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Source
Description
laminar flow afterburner). Emissions were collected using paper
filters with a 0.45 [j,m pore size followed by XAD-4 resin.
Wang et al. (2003b)
This study is a continuation of previous work by Wang et al. (2001).
It was conducted using the same methodology and experimental
design and analyzed additional chlorinated PAHs.
Boettneret al. (1973)
This US study documented the emissions from combusting 0.5 - 2.0
gram samples of several PVC polymers, identified as A, B, C and
collecting the resulting gases emitted in Saran® bags. An assumption
was made that polymers A, B and C represent PVC resin. For
polymer A, emissions were provided for various temperature ranges
and summed together to get the total emissions from combustion of
polymer A. The emissions resulting from the combustion of
polymers B and C were measured from a combustion temperature of
600° C. The total emissions from all three polymers were averaged.
Although the temperature at incineration plants is generally higher
than 600° C, part of this data was incorporated into the incineration
dataset due to the availability of a large amount of emission data on
semi-volatile compounds.
Conesa et al. (2008)
The combustion chamber used for this Spanish study was a single-
chamber furnace operated at a combustion temperature of 850° C.
PVC resin samples (0.1-0.2 g) free of any additiveswere used for the
study. Combustion emissions were collected using adsorption traps
filled with XAD-2 resin, or in Tedlar bags.
Shemwell and
Levendis (2000)
In this US-based study, 3g samples of PVC powder were combusted
in an electrically-heated drop-tube furnace over a range of bulk
equivalence ratios from 0.5-1.5. The resulting particulate emissions
were passed through a series of particle impact plates and were then
routed through a fiberglass filterto remove particles smaller than 0.4
[j,m. The combustion emissions of PVC at a bulk equivalence ratio of
0.5 (i.e., lean conditions) were assumed to be the closest to MSW
incineration facilities and were used to compile this dataset. It should
be noted that the emissions data were measured at a combustion
temperature of approximately 1,230° C.
Kim et al. (2004)
This South Korean study examines one-stage combustion of 0.5-
gram samples of PVC powder (free of additives) in a vertical electric
muffle furnace, where emissions were collected using a silica glass
microfiber filter followed by XAD-2 resin and a backup Toluene
solvent impinger. Dioxins and furans emissions data measured at
combustion temperatures of600° C and 900° C were averaged for this
dataset.
9-4
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Source
Description
US EPA (1996)
AP-42, 5th Edition, Volume 1, Section 2.1 documents the efficiency
of several MSW combustor air pollution control device (APCD)
technologies in controlling particulate, dioxin and furan, and
hydrochloric acid emissions. The combined efficiency of various sets
of in-line air pollution control devices is provided.
9.4 LCI Related to Disposal
Emissions associated withPVC disposal in a landfill include PVC-specific leachate emissions and
non-PVC-specific emissions related to landfill construction, operation, closure and post-closure
care. PVC is not expected to biologically degrade and is therefore not expected to produce gaseous
emissions (US EPA 2012b). Non PVC-specific emissions associated with landfilling that are not
PVC-specific are discussed in detail in Chapter 2, and are applied based on the loose density of
PVC (King County 2007).
The only leaching data found for PVC materials was from a leaching study by Ejlertsson et al.
(2000) that was conducted on organotin-stabilized PVC. The leached concentration of 7 organotin
compounds at tri-monthly intervals was measured from lysimeters containing inoculum (500 mL),
a prepared model waste (600 g), and PVC sheet or PVC foil (without plasticizers, 60 g) over a 9-
month period. Of the 7 organotin compounds, all but one (i.e., dioctyltin) were observed above the
detection limit in the study controls (i.e., lysimeters that only contained the inoculum and model
waste) over the range of study temperatures. Based on the study results and description of the
experimental methodology, it was not possible to develop an estimate of the total mass of dioctyltin
expected to leach from PVC in a landfill environment.
Ejlertsson etal. (2000) also utilized gas chromatography to identify the presence of vinyl chloride
gas from organotin-stabilized PVC. Vinyl chloride was not detected in any of the gas samples
collected from any of the lysimeters. The detection limit of this analysis was 1 mg vinyl
chloride/m3 of biogas. Based on this analysis, it was assumed that vinyl chloride gas is not
generated in significant quantities from the disposal of PVC in landfills. Additional studies which
estimate landfill gas emissions resulting from the placement of PVC in an unlined CDD landfill or
an MSW landfill were not found.
9.5 LCI Related to Incineration
While PVC is not targeted for energy recovery at solid waste incineration facilities, the combustion
of this material does occur at MSW waste-to-energy (WTE) plants due to the fact that CDD
materials are found in MSW, and may represent approximately 10.5 percent of this waste stream
(US EPA 2015a). Emissions from the combustion of additive-free PVC are documented in peer-
reviewed literature across a range of temperatures, with studies commonly utilizing combustion
temperatures between 500 and 1000° C. Similar to US EPA (2015b), it was assumed that a typical
MSW WTE facility operates at a combustion temperature of approximately 750° C. Therefore,
emissions data from peer-reviewed literature measured from the combustion of PVC at this
temperature was calculated.
9-5
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For those studies which provided emissions information for temperatures that evenly bounded the
target temperature of 750° C (e.g., 700° C and 800° C in Wang et al. (2001)), the average of the
emissions was calculated. For parameter emissions only provided by a single study (e.g., PM2 in
Shemwell and Levendis 2000), this information was incorporated into the dataset "as-is",
irrespective of combustion temperature. In the event the emissions of a certain parameter were
included in more than one source, and if the combustion temperature used in one study was much
closer to 750° C than the combustion temperature used in the other study, the emissions from the
study using a combustion temperature closer to 750° C was preferentially chosen over the
emissions from the other study. For example, carbon dioxide and carbon monoxide from Boettner
et al. (1973) were not utilized in this dataset, since these emissions were available from other
studies conducted closer to 750 °C.
The methodology used in these PVC combustion studies generally shared a number of common
characteristics:
• Combusted PVC samples were typically 0.1 to 2 grams and free of additives.
• Combustion took place in an externally-heated (i.e., electric) furnace.
• Emissions were collected by means of a particulate filter followed by the adsorption of
organic gases on a resin bed.
• Eluents derived from the dissolution of adsorbed gases and/or particulates were typically
analyzed by means of gas chromatography and mass spectrometry.
• Emissions were typically reported per mass of PVC combusted.
Based on Boettner et al. (1973), Conesa et al. (2008), and Shemwell and Levendis (2000), it
appears that there is no residue left behind from the combustion of PVC at or near 750° C. The
quantity of fly ash estimated in this dataset is calculated to incorporate the product of the
uncontrolled particulate emissions (estimated by Shemwell and Levendis (2000) and Wang et al.
(2003a)) and the air pollution control efficiencies as presented for a number of devices in US EPA
(1996). In the absence of additional information, it was assumed the particulate control efficiencies
as presented in US EPA (1996) were the same for all particulates regardless of size. US EPA
(1996) control efficiencies were also used in the dataset to estimate the control of emissions of
hydrochloric acid, benzoic acid (i.e., assumed to be controlled with the same efficiency as
hydrochloric acid) and dioxins and furans.
Information approximating the national average distance that PVC travels from EOL production
to MSW WTE facilities was not found, and instead, an average transport distance of 20 km was
assumed. Figure 9-3 presents the flows included, partially-included, and excluded from the LCI
dataset for PVC incinerated at an MSW WTE facility.
9-6
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Dioxins
and
furans
VOCs
Benzoic Carbon
acid dioxide
Heat
Carbon Hydrogen
monoxide chloride
Transport,
single unit truck;
short-haul;
diesel powered
Incineration of polyvinyl chloride, PVC; at
MSW waste-to-energy facility
Fly ash;
at MSW landfill
i
i
i
i
Diesel
i
?
i
i
i
Water
i
Electricity
Legend
Elementary Flow ^
Technosphere Flow
¦ Not Included
| Included
System Boundary
Figure 9-3. Unit process flow diagram for incineration of PVC products without plasticizers at
waste-to-energy (WTE) facility.
9.6 Data Gap Analysis and Opportunities for Additional LCI Data
Table 9-3 summarizes the type of US-based LCI identified from reviewed PVC EOL management
sources. Overall, only limited EOL-specific LCI are available for PVC. Based on a review of
available information, the following datagaps were identified for compiling a more comprehensive
LCI dataset for EOL management of PVC products:
1. Leachable emissions for the degradation of PVC products in a landfill. While some
limited data were found regarding the leachability of some PVC products (i.e., PVC products
with organotin stabilizers), the leaching data presented could not be converted into leachate
emissions per mass of landfilled PVC. Data also were not found on the leachability of PVC
products with other stabilizers. The limited leachate data found for PVC products was specific
to MSW landfill conditions and was not found for the behavior of PVC products in CDD
landfills.
2. Emissions from the incineration of PVC products at a waste-to-energy plant. Data
regarding emissions from the combustion of PVC products were available, but the operating
conditions were inconsistent across the studies. Experimental conditions such as temperature
and air-to-fuel ratio were similar to the conditions at a waste-to-energy plant, but were never
the same. Specifically, an estimate of hydrogen chloride emissions and a number of semi-
volatile compounds were utilized from an older study (Boettner et al. 1973) which was
9-7
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incrementally conducted across a relatively low range of combustion temperatures (i.e., from
200° C to 600° C for each sample). It is possible this gradual heating process released a very
different set of emissions than what would occur at a typical WTE facility, where MSW is
dropped directly into the combustion chamber. The control efficiencies of air pollution control
devices were available but were limited to particulate matter, acid gases, dioxins and furans. It
was assumed that the rest of the emissions were unaffected by the air pollution control devices.
A detailed set of pre- and post-treatment emissions for all parameters reported from the
combustion of PVC for common arrangements of air pollution control devices at MSW WTE
plants would greatly improve the quality of this dataset.
Table 9-3. Overview of US-based LCI data available. "P" represents partial coverage of the ideal
data, and indicates no data..
Process
Wang et al.
(2003a)
Boettner
et al.
(1973)
Shemwell and
Levendis
(2000)
Landfilling
-
-
-
Incineration
P
P
P
9.7 References
ATSDR (2005). Toxicological Profile for Tin and Tin Compounds. Prepared by the US
Department of Health and Human Services, Public Health Service, Agency for Toxic
Substances and Disease Registry. August 2005.
Boettner, E.A., Ball, G.A., and Weiss, B. (1973). Combustion Products from the Incineration of
Plastics. Prepared forthe Office of Research and Monitoring, US EPA. 1 February 1973.
CCG(2008). Construction & Demolition Waste Composition Study FINAL Report. Seattle
Public Utilities. July 2008.
CCG (2009). King County Waste Monitoring Program. 2007/2008 Construction and Demolition
Materials Characterization Study. February 2009.
CDRA (2015). The Benefits of Construction and Demolition Materials Recycling in the United
States. Report prepared for the Construction and Demolition Recycling Association by the
Department of Environmental Engineering Sciences, Engineering School of Sustainable
Infrastructure and Environment, University of Florida. Version 1.1 updated 14 January
2015.
California Integrated Waste Management Board (CIWMB) (2006). Health Concerns and
Environmental Issues with PVC-Containing Building Materials in Green Buildings.
Produced by the California Environmental Protection Agency Office of Environmental
Health Hazard Assessment. October 2006.
Christmann, W., Kasiske, D., Kloppel, K.D., Partscht, H., and Rotard, W. (1989). Combustion of
Polyvinylchloride - An Important Source forthe Formation of PCDD/PCDF. Chmosphere,
19(1-6): 387-392.
Conesa, J.A., Font, R., Fullana, A., Martin-Gullon, I., Galvez, A., Molto, J., and Gomez-Rico,
M.F. (2008). Comparison Between Emissions from the Pyrolysis and Combustion of
Different Wastes. Journal of Analytical and Applied Pyrolysis.
CSPC (2010). Review of Exposure and Toxicity Data for Phthalate Substitutes. Prepared for the
US Consumer Product Safety Commission by Versar, Inc. 15 January 2010.
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Daniels, P.H. (2009). A Brief Overview of Theories of PVC Plasticization and Methods Used to
Evaluate PVC-Plasticizer Interaction. Journal of Vinyl & Additive Technology, 219-223.
Ejlertsson, J., Horsing, M., and Mersiowsky, I. (2000). Behaviour of PVC Products in Landfilled
Municipal Solid Waste at Different Temperatures. Updated 21 September 2000. Available:
http://www.seepvcforum.eom/system/assets/000/000/139/original_Landfill_Report.pdf.
Accessed 1 May 2017.
Jafari, A.J. and Donaldson, J.D. (2009). Determination of HCL and VOC Emission from Thermal
Degradation of PVC in the Absence and Presence of Copper, Copper(II) Oxide and
Copper(II) Chloride. E-Journal of Chemistry, 6(3): 685-692.
Kim, K., Hong, K., Ko, Y., and Kim, M. (2004). Emission Characteristics of PCDD/Fs, PCBs,
Chlorobenzenes, Chlorophenols, and PAHs from Polyvinylchloride Combustion at
Various Temperatures. Journal of Air and Waste Management Association, 54:5, 555-562.
King County (2007). Product Volume to Weight Conversion Table. http://Lusa.gov/lLj8Fpp
accessed 18 March 2015.
MSW Consultants (2008). Mecklenburg County, North Carolina Construction and Demolition
Debris Composition Study. September 2008.
NREL (1993). Polyvinyl Chloride Plastics in Municipal Solid Waste Combustion: Impact Upon
Dioxin Emissions - A Synthesis of Views, NREL/TP-430-5518. April 1993.
NTP (2016). 14th Report on Carcinogens. Released by the US Department of Health and Human
Services. 3 November 2016.
https://ntp.niehs.nih.gOv/whatwestudy/assessments/cancer/roc/index.html#tocl. Accessed
1 May 2017.
R.W. Beck, CCG, IWCS (2010). Statewide Construction and Demolition Debris Characterization
Study. Georgia Department of Natural Resources, Sustainability Division. June 2010.
Shemwell, B.E. and Levendis, Y.A. (2000). Particulates Generated from Combustion of Polymers
(Plastics). Journal of the Air and Waste Management Association, 50:1, 94-102.
Stark, T.D., Choi, H., andDiebel, P.W. (2005). The Influence of Molecular Weight on Plasticizer
Retention. GFR Magazine, Volume 23, Number 2.
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Accessed 1 May 2017.
USEPA (1996). AP-42, Fifth Edition, Volume I, Chapter 2, Section 2.1 Refuse Combustion.
US EPA (2012a). Phthalates Action Plan. Revised 14 March 2012.
https://www.epa.gov/sites/production/files/2015-
09/documents/phthalates_actionplan_revised_2012-03-14.pdf. Accessed 1 May 2017.
US EPA (2015a). Methodology to Estimate the Quantity, Composition, and Management of
Construction and Demolition Debris in the United States. EPA-600-R-15-111. October
2015.
US EPA (2019). Waste Reduction Model (WARM) Tool: User's Guide. Version 15, May 2019.
Available: https://www.epa.gov/warm/documentation-chapters-greenhouse-gas-emission-
energy-and-economic-factors-used-waste-reduction.
USEPA (2020). Learn about Dioxin. Available: https://www.epa.gov/dioxin/learn-about-dioxin.
van Haaren, R., Themelis, N., and Goldstein, N. (2010). The State of Garbage in America: 17th
Nationwide Survey of MSW Management in the U.S. Biocycle. October 2010.
Wang, D., Piao, M., Chu, S., and Xu, X. (2001). Chlorinated Polycyclic Aromatic Hydrocarbons
from Polyvinylchloride Combustion. Bulletin of Environmental Contamination and
Toxicology, 66:326-333.
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Wang, Z., Wang, J., Richter, H., Howard, J.B., Carlson, J., and Levendis, Y.A. (2003a).
Comparative Study on Polycyclic Aromatic Hydrocarbons, Light Hydrocarbons, Carbon
Monoxide, and Particulate Emissions from the Combustion of Polyethylene, Polystyrene,
and Poly(vinyl chloride). Energy and Fuels, 17:999-1013.
Wang, D., Xu, X., Chu, S., and Zhang, D. (2003b). Analysis and Structure Prediction of
Chlorinated Polycyclic Aromatic Hydrocarbons Released from Combustion of
Polyvinylchloride. Chemosphere, 53:495-503.
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10 Fiberglass Insulation
10.1 Introduction
Fiberglass insulation is installed to minimize the exchange of heat through a building's walls,
roof/ceiling, and floors. Fiberglass insulation is available in a range of "R-" ratings, where a higher
R-value represents a greater thermal resistance and thus greater insulation. Fiberglass insulation
commonly used in residential buildings comes in two primary forms: batting and loose insulation.
Batting is generally available in sizes which are compatible with common stud-spacing intervals,
whereas loose insulation is usually blown into the target location using an application-specific
blower (US DOE 2012). Fiberglass is also used in the manufacture of thermal insulating boards
and acoustic insulation (i.e., sound-proofing) batts (USCB 2005, Owens Corning 2017, Knauf
2017). Fiberglass batting may or may not include a kraft paper facing combined with an adhesive
(commonly asphalt), which acts to reduce the transport of moisture through wall spaces.
Besides a possible paper facing, fiberglass is made up of several inert materials. The predominant
materials are recycled glass cullet and sand. Together these materials represent the majority of the
total mass composition of fiberglass batting (US EPA 2015, CertainTeed 2013, Owens Corning
2012). Other materials used to amend the glass fibers to lower the glass batch melting temperature
and enhance fiber formation are borax, limestone, and soda ash. Table 10-1 presents the
approximate composition of non-faced fiberglass insulation.
Table 10-1. Approximate mass composition of
Component
Mass Composition
Glass Cullet
(recycled)
40%
Sand
28%
Soda Ash
11%
Limestone
8%
Borax
8%
Binder Coatings
5%
i-faced fiberglass insulation (US EPA 2015).
Binder coating is sprayed over the glass fibers immediately afterthey are formed and cooled. The
binder is made up of a number of compounds including urea, ammonia, lignin, phenol-
formaldehyde resin, silane and water (US EPA 1995). The binder provides the matrix which holds
the fibers together until the batt is shaped and cured at temperature so that the binder resin sets.
As described in detail in the next section, the EOL phase of fiberglass insulation is managed
predominantly through landfilling, as depicted in Figure 10-1.
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Figure 10-1. Process flow diagram for EOL management of fiberglass insulation products.
10.2 EOL Management
Fiberglass insulation has been estimated to have a service life in excess of 50 years (NIST 2011).
While it is possible to remove and reuse fiberglass insulation, this is not common practice (US
EPA 2015). Due to its inert nature, fiberglass insulation also is not suitable for composting. An
optional paper facing is the only potentially biologically degradable component in fiberglass
insulation. As described in several publications (CertainTeed 2013, NIST 2011), only 5-30% of
the total mass of insulation is potentially combustible material (i.e., binder coating - a
thermosetting resin that holds the fibers together; or kraft paper facing with adhesive). Therefore,
fiberglass insulation cannot effectively be used for incineration with energy recovery. In addition,
while recycling fiberglass insulation is technically possible, the low density of the material would
mean that the potential energy savings through recycling would be lost during recovered material
transport. Therefore, it does not appear that post-consumer fiberglass is currently recycled; it was
assumed that landfilling is currently the only means by which discarded fiberglass insulation is
managed in the EOL phase.
Table 10-2 lists and describes the LCA processes included in this report's associated datasetsfor
EOL management of fiberglass.
Table 10-2. End-of-life management processes for fiberglass.
Process
Description
Unlined CDD
landfilling of fiberglass
insulation; at unlined
CDD landfill
This process includes gas emissions, leachate emissions, material and
energy inputs, and equipment emissions from the placement of
fiberglass insulation in an unlined CDD landfill. It also includes cover
soil placement and landfill operation, closure, and post-closure care.
MSW landfilling of
fiberglass insulation; at
MSW landfill
This process includes gas emissions, leachate volume emissions,
material and energy inputs, and equipment emissions from the
placement of fiberglass insulation in an MSW landfill. It also includes
cover soil placement and landfill operation, closure, and post-closure
care.
10.3 LCI Sources
Peer-reviewed literature, as well as government and private industry publications, were reviewed
to identify LCI datasets pertaining to EOL management of fiberglass insulation. Because literature
10-2
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describing expected leachate emissions for fiberglass insulation was not found, leachate emissions
from kraft paper asphalt coating were used from Azah (2011), as presented for asphalt shingles in
Chapter 4 and adjusted for the mass of asphalt in both the shingles and fiberglass insulation
products. Table 10-3 lists sources reviewed to develop the LCIs presented in this chapter.
Table 10-3. Overview of life cycle inventory data sources associated with fiberglass insulation.
Source
Description
Azah (2011)
Azah (2011) conducted batch and column leaching tests (SPLP) on
a shingle sample collected from a recycling facility in Florida to
assess polycyclic aromatic hydrocarbons (PAHs) leaching from
shingles. These data were used to estimate PAH emissions from
disposal of fiberglass insulation in unlined CDD landfills.
NIST (2018) BEES-
Generic Fiberglass
Insulation
NIST (2018) presents the density and thickness for fiberglass
insulation products including R-13, R-15, R-19, and R-38 insulation
batting and R-38 loose fiberglass insulation. These densities and
thicknesses were used in conjunction with USCB (2005) and
manufacturer product specifications and data sheets to estimate an
average density of building-related fiberglass products that undergo
EOL management.
USCB (2005)
The values of 2002 product shipments for 5 major groups of
building-related fiberglass insulation products underNAICS 327993
were used as a surrogate to estimate the mass fraction of each of these
products entering the EOL management phase and to calculate a
representative density of the average US building-related fiberglass
insulation product.
10.4 LCI Related to Disposal
The USCB (2005) 2002 Economic Census was the last census to provide a detailed breakdown of
the value of NAICS 327993 mineral wool (i.e., fiberglass) product shipments. USCB (2005)
provided the value of product shipments for 5 major groups of building fiberglass insulation
products: loose fiber; building batts, blankets, and rolls rated R-19 or more; building batts,
blankets, and rolls rated below R-19; acoustical; and board. In the absence of product shipment
tonnage data, the fractional value of product shipments was used as a surrogate to approximate the
mass fraction of product shipments assuming that each group of building fiberglass insulation
products has the same shipment value on a per-ton basis. Building batts, blankets, and rolls R-19
or more were assumed to, by area, equally represent R-19 and R-38 batting. These insulation
batting products are commonly used to insulate floors and ceilings, respectively. Building batts,
blankets, and rolls below R-19 were assumed to represent R-13 insulation batting, which is a
batting product commonly used to insulate exterior walls. The density of the average US building
fiberglass insulation product was then calculated by multiplying the individual density of each
product by the fractional value of its product shipments. The uncompacted density of each of these
products was used from NIST (2018) (i.e., loose, R-13, R-19, R-38), Knauf (2017) (i.e., acoustical
batting, selected as the midpoint density of the five acoustical batts listed), and Owens Corning
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(2017) (i.e., fiberglass board, taken as the average of the two board densities listed by the
manufacturer).
Unlike more rigid CDD materials, it is expected that fiberglass insulation products would likely
compress substantially due to the overburden pressure of additional waste and/or cover soils.
Because the landfill LCI datasets developed as part of this study estimate emissions and
environmental burdens on a compacted density basis, an estimate of the compacted density of
fiberglass insulation was developed. The maximum compaction of fiberglass insulation was
approximated by IWCS (2017), where a section of R-13 kraft-faced insulation batting was
compressed under various pressures up to approximately 0.75 psi. The total compaction (i.e., the
quotient of the uncompacted thickness and the compacted thickness) was plotted as a function of
the quotient of the change in compaction and the change in pressure. IWCS (2017) estimated a
maximum compaction of approximately 7.6 at a pressure below 1 psi. Considering typical landfill
heights in excess of 50 feet, this pressure is significantly less than the expected overburden pressure
that the majority of landfilled waste would experience. Because other fiberglass batting densities
were found to be in the same general range as the R-13 insulation batt, it was assumed these
products would have approximately the same maximum compaction. However, because the
density of fiberglass insulation board was found to be significantly higher than fiberglass batting,
it was assumed that fiberglass insulation board would not significantly compact once placed and
covered in a landfill.
Landfill gas emissions were estimated for fiberglass insulation kraft paper. IWCS (2017) measured
the mass of asphalt-coated kraft paper backing for a roll of R-13 fiberglass insulation as 0.16 kg/m2
of fiberglass insulation batt. The maximum mass fractional ratios provided by CertainTeed (2013)
were used to estimate the fraction of asphalt and paper. The fractional area of each of the 5
categories of building fiberglass insulation products was estimated using the fractional value of
product shipments, and product density and thickness. The average area of kraft paper facing per
kg of US building fiberglass insulation product was then calculated assuming that 50% of all
fiberglass insulation batting products are kraft-faced. In the absence of kraft paper specific gas
emissions data, emissions data provided by Barlaz (1998) and US EPA (2015b) for corrugated
containers was used as a proxy.
Due to an absence of SPLP, TCLP, or comparable data specific to fiberglass insulation, studies
quantifying the leachate emissions from the individual layers of fiberglass insulation (i.e.,
fiberglass batt and binder, kraft paper, and asphalt binder) were sought. Leached emissions of
PAHs from asphalt shingles were utilized from Azah (2011) and adjusted for the asphalt content
of asphalt shingles (i.e., 22% from US EPA (2015c)) and the quantity of asphalt binder in kraft-
faced fiberglass insulation (calculated as approximately 3% from Certainteed (2013) and IWCS
(2017)), assuming that the asphalt in shingles is responsible for the leaching of PAHs. While
several studies were found which estimate leachate emissions from asphalt pavement or reclaimed
asphalt pavement (e.g., Birgisdottiretal. (2007), Townsend and Brantley (1998), and Legretet al.
(2005)), these were not included in this dataset due to the possibility of influences associated with
vehicular traffic (e.g., oil leaks, coolant leaks, tire granules, lead wheel weights).
Additional leaching studies of asphalt were identified in peer-reviewed literature. Bowen et al.
(2000) and Brandt and De Groot (2001) performed a static Dutch standardized test for the leaching
of building materials on a series of 9 different asphalt pavements. However, the list of detected
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PAHs and reported leached concentrations was generally less exhaustive than those included by
Azah (2011). This may have been a result of the relatively low liquid-to-solid ratios used in these
studies (i.e., 4.0:1 and 4.5:1).
10.5 Data Gap Analysis and Opportunities for Additional LCI Data
The following data gaps were identified for the development of LCIs representing the complete
set of EOL management options for fiberglass insulation:
1. A better understanding of the maximum compressibility of discarded fiberglass
insulation. While maximum compaction of R-13 fiberglass was estimated by IWCS (2017),
additional studies which measure the compressibility of common fiberglass insulation products
is critical to developing accurate landfill LCI datasets.
2. A more accurate estimate of the average density of discarded building-related fiberglass
insulation products. Based on the categories of fiberglass insulation products summarized in
USCB (2005), it appears that there are a number of major classes of fiberglass insulation
products, each of which has a different density. A study analyzing the fraction of these different
insulation products entering the EOL management phase would allow the calculation of a more
accurate estimate of a representative discarded fiberglass insulation density.
3. Landfilled fiberglass insulation leachate emissions. The chemical degradation rate and
nature of contaminants released from fiberglass insulation placed in a landfill is unknown.
These could be estimated using SPLP and TCLP tests on representative samples of post-
consumer fiberglass insulation.
4. Landfilled fiberglass insulation landfill gas emissions. The total quantity and types of
landfill gas emissions, including trace constituents, which would be produced as a result of the
biodegradation of fiberglass insulation kraft paper facing is unknown. As discussed
previously, these were estimated assuming gas generation parameters for corrugated containers
as listed inBarlaz (1998) and summarized in US EPA (2015b). However, biological methane
potential (BMP) tests conducted on samples of fiberglass insulation kraft paper would provide
a more accurate estimate of fiberglass insulation landfill gas emissions. In addition, a better
understanding of the average nationwide mass fraction of paper facing discarded with
fiberglass insulation would be necessary for a more reliable estimate of gas emissions.
10.6 References
Azah, E.M. (2011). The Impact of Polycyclic Aromatic Hydrocarbons (PAHs) on Beneficial Use
of Waste Materials. Ph.D. Dissertation. University of Florida, Gainesville, FL, USA.
Birgisdottir, H., Gamst, J., and Christensen, T.H. (2007). Leaching of PAHs from Hot Mix Asphalt
Pavements. Environmental Engineering Science, 24(10):1409-1421.
Brandt, H.C.A and De Groot, P.C. (2001). Aqueous Leaching of Polycyclic Aromatic
Hydrocarbons from Bitumen and Asphalt. Water Resources, 35(17): 4200-4207.
Certainteed (2013). Environmental Product Declaration - Sustainable Insulation: Unfaced and
Kraft Faced Batts. Issued 25 June 2013.
IWCS (2017). Internal Study of R-13 Kraft-faced Fiberglass Insulation Compaction as a Function
of Pressure. May 2017.
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Knauf (2017). Earthwool Internal Wall (Acoustic) Batts.
https://www.knaufinsulation.com.aU/product/earthwool-wall-batt#downloads. Accessed 3
May 2017.
Legret, A., Odie, L., Demare, D., and Jullien, A. (2005). Leaching of Heavy Metals and Polycyclic
Aromatic Hydrocarbons from Reclaimed Asphalt Pavement. Water Research, 39, 3675-
3685.
NIST (2018). Building for Environmental and Economic Sustainability (BEES) Online 2.0
Technical Manual. NIST Technical Note 2032, National Institute of Standards and
Technology, December 2018.
https://nvlpubs.nist.gov/nistpubs/TechnicalNotes/NIST.TN.2032.pdf.
Owens Corning (2012). Environmental Product Declaration - EcoTouch Kraft-Faced Insulation.
Issued 29 October2012.
Owens Corning (2017). Type 706 and Type 707 Series Fiberglas Insulation Boards. Product Data
Sheet. March 2017.
Townsend, T.G. and Brantley, A.S. (1998). Leaching Characteristics of Asphalt Road Waste.
Report #98-2. State University System of Florida, Florida Center for Solid and Hazardous
Waste Management.
USCB (2005). Mineral Wool Manufacturing: 2002, 2002 Economic Census, Manufacturing,
Industry Series. US Department of Commerce, Economic and Statistics Administration.
January 2005.
US EPA (1995). AP-42, Fifth Edition, Volume I, Chapter 11, Section 11.13, Glass Fiber
Manufacturing. Reformatted January 1995.
US EPA (2019). Waste Reduction Model (WARM) Tool: User's Guide. Version 15, May 2019.
Available: https://www.epa.gov/warm/documentation-chapters-greenhouse-gas-emission-
energy-and-economic-factors-used-waste-reduction.
US DOE (2012). Energy Renovations: Insulation. PNNL-20972. Volume 17. Building America
Best Practices Series. Prepared by the Pacific Northwest National Laboratory & Oak Ridge
National Laboratory. May 2012.
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11 Carpet
11.1 Introduction
According to the Carpet America Recovery Effort (CARE), an organization comprised of carpet
manufacturers and entities such as carpet installers and recyclers, approximately 1.58 million MT
of post-consumer carpet were discarded in 2016 (CARE 2017). CARE (2017) reported that
approximately 0.142 million MT (or 9.0%) was diverted from land filling (i.e., recycling, carpet as
an alternative fuel or cement kiln, waste-to-energy). Of the fraction diverted from landfilling,
approximately 54% was recycled, 28% was combusted as alternative fuel or in cement kilns, and
17% was combusted in waste-to-energy (WTE) incineration facilities. Figure 11-1 presents the
fraction of discarded carpet managed through EOL pathways as presented in CARE (2017). Data
on the quantity of discarded carpet padding managed through different EOL pathways was not
found in literature.
Carpetas
Alternative Fuel
forCement Kiln,
Figure 11-1. Fraction of carpet managed through different EOL pathways (CARE 2017).
Carpet is a composite material typically made of four primary components including fiber piling,
primary backing, adhesive, and secondary backing. Carpet backings are primarily made from
polypropylene (PP), while the adhesive—usually containing a limestone filler to provide bulk and
weight—is often styrene butadiene (US EPA 2019, Ucar and Wang 2011, Realff2010, Potting and
Blok 1996). The approximate mass composition of the different components of waste carpet is
provided in Figure 11-2. The fiber is commonly made of nylon (e.g., 6, 6-6), polyester (e.g.,
polyethylene terephthalate (PET)) or olefin (e.g., polyethylene, PP) and is sometimes made of
wool or a wool blend. The national average mass representation of each of the common fiber types
found in US residential carpet is presented in Figure 11-3.
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Adhesive
10%
Limestone
30%
Piling
53%
Backing
7%
Figure 11-2. Mass composition of carpet waste (Ucar and Wang 2011).
Figure 11-3. Residential fiber mix (Realff2011 as cited in US EPA 2019).
There are numerous types of carpet, but the two most common are tufted and woven, with tufted
carpet representing over 90% of carpet construction (CRI2010). Tufted carpet is created through
the insertion of pile yarns through the bottom of the primary backing where the yarn is then looped
and pushed back through the backing - the process is repeated with the same yarn strand at a regular
interval. Piling may be loop, cut, or a combination of the two, where looped pile is a continuous
fiber sewn through the primary backing to form loops on the top and cut piling involves the
severing of each of the individual loops. Shag and plush carpets are examples of cut pile carpets
while Berber is an example of a loop pile. Woven carpet production occurs when backing and pile
yarns are combined into a single fabric. The increased density of the fibers and mechanical unity
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of the fibers with the facing give woven carpet a longer service life than tufted carpet, and also
mean that a secondary backing is not necessary.
A separate though integral part of carpet installations is carpet padding. Carpet padding is primarily
comprised of polyurethane foam with small amounts of adhesive (e.g., a mixture of polyol and
toluene diisocyanate), and has can be recycled. The most common production processes for carpet
padding is rebonding (Zia et al. 2007). During the rebond process, used pieces of carpet padding
may be added to manufacturer scrap foam (e.g., from furniture production) and are ground up into
pieces, mixed with a binder, compressed, steamed, cured, and finally shaped into new carpet
padding.
The discarded carpet and padding EOL management processes discussed in this section are
presented in Figure 11-4, which provides a summary of the EOL management pathways for these
materials.
Figure 11-4. Process flow diagram for end-of-life management of carpet and carpet padding. MRF =
material recovery facility; PU = polyurethane.
11.2 EOL Management
Once carpet has reached the end of its serviceable life, it may be landfilled, incinerated at a waste-
to-energy facility, combusted in acement kiln, orrecycled. While carpet produced with bio-based
fibers (e.g., wool) would be expected to produce landfill gas during the degradation process, the
vast bulk of carpet produced in the US is made of synthetic material. Therefore, only leachate
emissions for landfilled carpet are considered here. Due to the petroleum-based components in
carpet, the material has a mass-based heat content about 20% below bituminous coal (Realff 2010),
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and therefore may be combusted for energy-recovery purposes in a waste-to-energy facility or in
a cement kiln. Carpet is recycled through open-loop pathways, where the specific recycling options
are dependent on the carpet composition, particularly the composition of the fiber piling. As
discussed in US EPA (2016), the fiber facing and PP backing can be recycled into secondary fiber
and pellet precursor materials, some of which may be used in the production of new carpet
components, or which may be used in the manufacture of other secondary plastic materials.
Table 11-1 lists and describes the LCA processes included in this report's associated datasetsfor
EOL management of discarded carpet.
Table 11-1. End-of-life management processes for carpet
Process
Description
Unlined CDD
landfilling of
carpet; at unlined
CDD landfill
This process includes material (e.g., equipment, soil, water) and energy
(fuel, electricity) inputs for placement and compaction of carpet in a CDD
landfill, along with the associated energy and non-energy emissions (e.g.,
dust emissions from equipment operation and liquid emissions from
leaching of carpet).
MSW landfilling of
carpet; at MSW
landfill
This process includes gate-to-gate incineration of carpet at an MSW waste-
to-energy facility.
Unlined CDD
landfilling of carpet
padding; at unlined
CDD landfill
This process includes material and energy inputs and emissions from the
placement of carpet padding in a CDD landfill. It also includes cover soil
placement and landfill operations, closure, and post-closure care.
MSW landfilling of
carpet padding; at
MSW landfill
This process includes material and energy inputs, as well as leachate
volumes expected from the disposal of carpet padding in an MSW landfill,
using the density of carpet padding products reported by CCG (2006). It
also includes landfill construction, operation, and closure/post-closure care.
Incineration of
carpet; at MSW
waste-to-energy
facility
This process describes the step in which carpet is shredded to provide a
more uniform burn and further used as fuel in a cement kiln.
Incineration of
carpet padding; at
MSW waste-to-
energy facility
This process includes gate-to-gate modeling of the incineration of carpet
padding at an MSW waste-to-energy facility.
Processing and
cement kiln
combustion of
carpet; at cement
kiln; 21% carpet,
79% bituminous
coal (mass)
Due to a relatively high heating value resulting from the petroleum-based
fiber and backing constituents, and high limestone content of carpet
backing, carpet may be beneficially used as kiln fuel for cement
production. This process includes the emissions resulting from combustion
of carpet as fuel along with coal in a cement manufacturing facility.
Recycling of carpet
padding; at rebond
polyurethane foam
This process includes post-consumer carpet padding as part of the scrap
foam mix used in a closed-loop process to make new carpet padding.
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Process
Description
production facility
Recycling of
carpet; residential;
at recycling
This process includes the gate-to-gate modeling of recycling residential
carpet at a U.S. carpet recycling facility.
11.3 LCI Sources
Peer-reviewed literature, carpet industry organization publications and technical reports, and
government reports were reviewed to compile emissions inventory information for EOL
management of carpet. Table 11-2 presents a list of key publications that provided this information.
Table 11-2. Overview of life cycle inventory data sources associated with carpet.
Source
Description
Konopa et
al. (2008)
Konopa et al. (2008) presented emissions resulting from the combustion of
carpet at a pilot-scale incinerator at a US EPA research facility.
Lemieux et
al. (2004)
Lemieux et al. (2004) provided a comparative analysis of the emissions
resulting from a cement kiln operating with 100% coal fuel and with an
85% coal and 15% carpet fuel mass mix. .
HLA/CTA1
(1994)
This application for permit to construct a flexible polyurethane foam
manufacturing facility provides a detailed breakdown of the natural gas,
steam, material inputs and outputs, and emissions associated with the
rebond polyurethane foam production process, which can incorporate post-
consumer carpet padding in the new foam production mix.
USCB
(2004)
The 2002 Economic Census provides electricity consumption data for the
entire Urethane and Other Foam Product Manufacturing Industry (NAICS
326150). These data were used as a proxy to quantify electricity use during
the carpeting padding recycling process.
US EPA
(2019) -
WARM
EPA's Waste Reduction Model (WARM) presents data on GHG emissions
associated with source reduction, recycling, combustion, and landfilling
(i.e., collection and placement) of carpet.
NIST (2018)
-BEES
There are a total of 21 different carpet products listed in the National
Institute of Standards and Technology (NIST) Building for Environmental
and Economic Sustainability (BEES) LCAtool for evaluation the impacts
of different building materials. The product manufacturing energy
quantities and types (e.g. electricity, natural gas) as well as the composition
by mass fraction of the different carpet products and product precursors are
included in the tool's documentation.
Bertrand and
Wagner
(1997)
Bertrand and Wagner (1997) combusted individual carpet cushion samples
under controlled conditions to collect and evaluate the total quantity of
particulates emitted. The researchers provided the fraction of particulates
collected that was below 10 microns in diameter.
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Source
Description
Garrido et al.
(2016)
Garrido et al. (2016) analyzed an array of gaseous emissions associated with
the combustion of polyurethane foam samples under pyrolytic and excess
air conditions at different temperatures.
NIST (2008)
NIST (2008) analyzed the combustion of polyurethane foam seat cushion
with several different types of coverings. This study describes the total heat
release forthe combustion of polyurethane foam.
11.4 LCI Related to Disposal
Emissions associated with carpet disposal in a landfill include emissions from landfill construction,
operation, closure and post-closure care, and leachate emissions. Carpet is not expected to
biologically degrade and produce gaseous emissions. The landfilling emissions presented in
Chapter 2 are applied based on the loose density of carpet to estimate carpet-specific disposal
LCIs. The loose density of carpet and carpet padding was referenced from CCG(2006).
The leaching emissions from carpet and padding depend on material composition, service life, and
the use location (residential or commercial) to estimate foot traffic and exposure. It has been
reported, for example, that carpet in a room adjacent to an arsenic-treated deck can become
contaminated with arsenic (Patch et al., 2009). Limited data on leaching were found based on an
extensive review of the publicly-available literature.
Pohland and Rachdawong (1996) conducted TCLP tests on five carpet samples: two composed of
nylon 6 facing, two of nylon 6-6 facing, and one refabricated carpet made of unknown facing with
no backing. The unknown facing data were not included in the dataset due to the absence of carpet
backing. Together, nylon 6 and nylon 6-6 represent 65% of the carpet facing used in theUS (Realff,
2011). These data represent the majority of facing in the US, but as shown in Figure 11-2,it should
be noted that PET and PP facing represent 15% and 20% of carpet facing in the US, respectively,
and this facing is not represented in this developed dataset. An additional limitation of the data
from Pohland and Rachdawong (1996) is that the method detection limits are not provided, so non-
detects could not be included in an average value calculation. The leaching data from the study are
assumed to be representative of carpet leaching in MSW landfills since these were obtained using
the TCLP test. The concentrations of leached parameters reported in this study were multiplied by
the TCLP liquid-to-solid ratio of 20 L/kg to obtain a mass per kilogram leaching value.
Shoaeioskouei (2012) conducted leaching studies in Vancouver on eight carpet samples of
unknown facing using leachate from MSW landfills and distilled water. Forthe lined MSW landfill
dataset, data were taken from the leaching tests conducted at a temperature of 35°C, which, of the
temperatures used (5-35° C) was assumed to be most representative of an MSW landfill. For the
unlined CDD landfill dataset, leaching data from distilled water was used. Parameters studied
included a suite of perfluorinated compounds (PFCs) used in stain-resistant coatings applied to
carpets. Five carpet samples were used to develop a set of composite samples: three from carpets
stored for multiple years in a warehouse, and two from used carpets. Due to the relatively small
sample size, leaching results from all five samples were used even though the samples of
warehouse-stored carpet may not be representative of carpet removed at the end of its serviceable
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life. The total mass of the PFCs present in the leachate/distilled water before and after contact with
the carpet were calculated. The mass leached was obtained by subtracting the mass of the
parameter in the leachate/distilled water before contact with carpet from the mass of the parameter
in the leachate/distilled water after leaching. Some of the PFCs were present in the initial
leachate/distilled water in greater quantities before the carpet leaching study was conducted, and
therefore a value of zero was assigned for these PFCs.
11.5 LCI Related to Recycling
The majority of carpet diverted from landfills are managed by recycling. CARE (2017) estimates
that 76,000 MT of carpet was recycled in 2016 - approximately 53% of all carpet diverted from
landfilling in 2016. Trying to quantify the prevalence of the various recycling processes is difficult
because the only reported carpet recycling statistics are categorized by recovered product end -
market use. According to CARE (2017), 2016 post-consumer carpet was categorized for three
primary markets: engineered resins (72%), carpet fiber (3%), and carpet backing (8%) - other
markets make up the remainder (18%).
Carpet recycling is a challenging process because the material is designed to be durable and not
pull apart, is non-homogenous, has varying amounts of contamination (e.g., dirt, cleaners) and
because there are still a limited number of end markets available for some recovered materials
such as PET fiber (CARE 2017, Mihut et al. 2001). Several different carpet processing
technologies exist including depolymerization, mechanical separation, melt blending and carpet
reconditioning. The composition of commercial carpet is different than residential carpets. US
EPA (2016) categorizes residential carpets into four types based on their face fiber; nylon 6, nylon
6-6, PP and PET.
According to US EPA (2016), 49.25%) (by mass) of the carpet recovered for recycling has end
uses. Styrene butadiene latex (used as an adhesive) and limestone (used as a filler) are assumed to
be unrecyclable. Together, styrene butadiene latex and limestone make up 40% of the weight of
carpet (US EPA 2019). The unrecycled part of the carpet is assumed to be landfilled in the carpet
recycling LCI dataset developed as part of this study. The remaining 10.75% of the carpet is
assumed to be residue from recycling and is also assumed to be landfilled. According to US EPA
(2016), non-energy emissions are not generated during residential carpet recycling.
It was assumed that discarded carpet is transported 20 km from the point of generation to the
recycling facility. It was also assumed that the face fibers and polypropylene backing components
recovered by recycling carpet are used to produce the same type of virgin resins used in the
manufacture of components of new carpet or secondary materials (e.g., molded or extruded
plastics). Figure 11-5 presents the flows included, partially-included, and excluded from the LCI
dataset for carpet recycling.
11-7
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Carpet;
from disposal
Transport,
single unit truck;
short-haul;
diesel powered
Recycling of carpet; residential; at
recycling
Diesel
Natural Gas
Electricity
Polyethylene
terephthalate ~
fiber
Nylon 6,6 fiber—~
- Polypropylene fiber-*-
Nylon 6 fiber—>
-Nylon 6,6 pellets->
- -Residue to landfill- -~
Legend
Elementary Flow
Technosphere Flow
¦ Not Included
| Included
System Boundary
Figure 11-5. Unit process flow diagram for carpet recycling.
An additional LCI dataset was developed to simulate the recycling of carpet padding. HLA/CTA1
(1994) presents mass balance data and heat requirements to operate a rebonded polyurethane foam
(PUF) manufacturing operation for a Florida facility. The percentage of post-consumer carpet
padding incorporated into the scrap foam mix used as a feedstock in the rebond process depends
on the density of the PUF end product, where higher-density products can accept larger
percentages of post-consumer padding. Because the electricity requirements of rebond PUF
production were not provided in HLA/CTA (1994), the total electricity demand of the urethane
and other foam product manufacturing industry (NAICS 326150), as presented in USCB (2004),
was utilized as a proxy.
Figure 11-6 presents the flows associated with the LCI dataset for carpet recycling.
11-8
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Toluene
diisocyanate
(gas emission)
Water
Polyol; at plant--*
Carpet padding;
from disposal
Transport,
single unit truck;
short-haul
diesel-powered
Toluene
diisocyanate
Recycling of carpet padding; at rebond
polyurethane foam production facility
Carpet padding;
recycled; at
-rebond polyurethane*
foam production
facility
Electricity
Water
Natural gas
Legend
Elementary Flow
Technosphere Flow
H Not Included
| Included
System Boundary
Figure 11-6. Unit process flow diagram for carpet padding recycling.
11.6 LCI Related to Incineration
Based on estimates of EOL carpet management in CARE (2017), approximately 24,900 MT of
waste carpet was incinerated in waste-to-energy facilities in 2016. This amount represents
approximately 18% of the total quantity of carpet diverted from landfilling. In a study described
in Konopa et al. (2008), individual charges of 0.46 kg nylon carpet squares were combusted in a
rotary kiln incineration simulator. The carpet used in this study is composed of nylon 6 fiber, which
represents 40% of the residential face fiber mix. While the incinerator used in the study includes
both an afterburner and secondary combustion chamber, these devices were not used during the
study-only the emissions from the primary combustion chamber were quantified. Therefore, the
emission factors presented in the WTE carpet incineration dataset that was developed as part of
this study represent untreated emissions.
Five air emissions were reported by Konopa et al. (2008) using continuous monitoring equipment
including carbon dioxide, carbon monoxide, nitric oxide, sulfur dioxide and total hydrocarbons.
These emissions were reported on a mass-per-energy (g/MJ) basis for 3 different materials
including carpet, coal, and particle board. Five separate carpet charges were fired, where the
dataset-reported emissions represent the average value from all firing events. The emissions from
carpet combustion were multiplied by the heat of combustion of carpet to estimate emissions on a
per kilogram carpet-input basis. The quantity of ash resulting from the incineration of carpet was
estimated as the average ash content of the carpet types through ultimate and proximate analyses
by Lemieux et al. (2004). As transport information was not available, it was assumed that discarded
carpet is transported 20 km from the point of removal to the incineration facility.
The carpet combusted in the incinerator used by Konopa et al. (2008) was co-fired with natural
gas. The study discusses that the specific emissions resulting from the combustion of the carpet
11-9
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charges were found by subtracting the natural gas emission baseline curve (assumed to have been
found from previous testing where natural gas was fired alone) from the total emissions resulting
from firing the kiln with both natural gas and carpet charges. The amount of power which would
be generated from the combustion of carpet was estimated from the waste-to-energy combustion
system efficiency (i.e. 17.8%), as used in USEPA (2016) and assuming anylon 6 fiber, 20% soiled
carpet with a heat of combustion of 17.7 MJ/kg (Realff 2010). This theoretical heat of combustion
compares favorably with the heat of combustion measured experimentally by Konopa et al. (2008).
Figure 11-7 presents the flows included, partially-included, and excluded from the LCI dataset for
carpet incineration at a waste-to-energy plant.
PM HCl C02 THC NOx CO Heat
Heavy
metals
t
Dioxiris
and Chlorine
furans
Carpet;
from disposal
Transport,
single unit truck;
short-haul;
diesel powered
Incineration of carpet; at MSW waste-to-
energy facility
i
i
i
Diesel
i
i
i
i
i
i
Water
i
i
Electricity
Ash
Legend
Elementary Flow ^
Technosphere Flow
¦ Not Included
| Included
System Boundary
Figure 11-7. Unit process flow diagram for carpet incineration at an MSW waste-to-energy facility.
Emissions resulting from the incineration of carpet padding in an MSW waste-to-energy (WTE)
facility are assumed to be unimpacted by the presence of other materials in the waste stream. In
actuality, the specific mixture of emissions resulting from incineration of carpet padding will be a
product of both the combustion of padding and nearby materials. Per US EPA (2015), the average
USMSW waste-to-energy facility is assumed to have an operating temperature of 750° C. For the
purposes of this dataset, fly ash is assumed to represent the collected portion of particulate matter.
This dataset assumes that the collection efficiency for particulates, as provided in US EPA (1996),
is independent of particle size.
With the exception of particulate emissions, all emissions and heat release from carpet padding
was calculated using polyurethane foam as a proxy for carpet padding. However, it should be noted
that carpet paddingis comprised of both scrap polyurethane foam and cured adhesives (HLA/CTA
11-10
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1994). The impact of the presence of these adhesives on emissions and heat content from the
combustion of carpet padding is unknown.
Garrido et al. (2016) conducted a detailed analysis of the gaseous emission released from the
pyrolysis and combustion of polyurethane foam. The researchers combusted 50 mg samples of the
foam at 850° C to measure the mass release of gases including carbon dioxide, carbon monoxide,
nitric oxide, ammonia, methane, a number of additional light hydrocarbons, dioxins and furans,
chlorobenzenes, and PCBs. As described by Bertrand and Wagner (1997), carpet padding samples
were fired over a micro Bunsen burner, however, the flame temperature was not reported. It is
assumed that combustion conditions from this test are reflective of WTE incinerator conditions.
The estimate of bottom ash produced as a result of carpet padding combustion was developed from
seat cushion polyurethane pad burn experiments conducted by NIST (2008). The residual mass
data after the burn test runs were from cushions which had a polypropylene cover. It is assumed
that these covers had a negligible impact on the overall remaining charred residual. Similar to
Bertrand and Wagner (1997), the flame temperature used in NIST (2008) was not reported and it
was assumed that combustion conditions from this test are reflective of WTE incinerator
conditions.
Further assumptions include that the carpet padding is transported 20 km from the point of
generation to the incineration facility. It was assumed that the air pollution control devices
described in US EPA (1996) would only impact emissions of PM, and dioxins and furans. The
dataset allows selection of one of the following air pollution control options: electrostatic
precipitator (ESP); dry sorbent injection (DSI) and ESP; spray dryer (SD) and ESP; DSI and fabric
filter (FF); SD and FF. Operational energy inputs (e.g., electricity for claw operation, diesel for
heavy equipment operation) are not included in this dataset. Other operation and maintenance
consumables, such as scrubber feed material (e.g., lime, bicarbonate) and lubrication for equipment
are not quantified for this dataset. Figure 11-8 presents the flows included, partially-included, and
excluded from the LCI dataset for carpet padding incineration at a waste-to-energy plant.
11-11
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PM C02 THC CO NOx
Heavy
metals
Dioxins
and
furans
S02
Carpet padding;
from disposal
Transport,
single unit truck;
short-haul;
diesel powered
Incineration of carpet padding; at MSW
waste-to-energy facility
Electricity
i
I
I
I
Water
i
i
i
i
i
Diesel
i
i
Bottom ash; at
MSW
->
waste-to-energy
facility
Fly ash; at MSW
-waste-to-energy-*-
facility
Legend
Elementary Flow
Technosphere Flow
| Partially Included
System Boundary
Figure 11-8. Unit process flow diagram for carpet padding incineration at an MSW waste-to-energy
facility.
11.7 LCI Related to Cement Kiln Combustion
Lemieux et al. (2004) presents emissions resulting from the co-combustion of carpet with coal at
a full-scale Portland cement production facility. Two trial runs were performed for the purpose of
a comparative evaluation of emissions resulting from the combustion of different kiln fuel
feedstock - the first trial run used 100% coal while the other used a coal/carpet mix. In the
coal/carpet mix, carpet was used to substitute 15% of the coal fuel feedstock (by energy) or
approximately 21% of the incoming coal by mass. Depending on the sampling method, emissions
were either measured continuously or at discrete times at four equal-elevation points along the
exhaust stack, post air pollution control devices (e.g., baghouse). To promote combustion
uniformity, carpet was shredded prior to combustion, where the estimated amount of electricity
necessary for shredding is estimated as the midpoint of the range presented by Realff (2010) as
0.15 MJ/kg carpet.
While the majority of the emissions from Lemieux et al. (2004) were graphically presented on a
mass-per-volume concentration basis, some (i.e., carbon monoxide, sulfur dioxide, nitrogen
oxides) were presented on a volumetric basis. The exhaust gas flow rate, exhaust gas temperature,
feedstock feed rate, and an assumed exhaust stack pressure of 1 atmosphere were used to estimate
emissions on a per-mass fuel-input basis. Nitrogen oxides were unable to be converted using this
method because the individual contribution of specific nitrogen oxides (e.g., nitric oxide, nitrogen
dioxide) was not available.
While separate emission estimates were provided for a 100% coal fuel feed from the first trial run,
the emission contribution of carpet in the 21% carpet / 79% coal fuel feed mix in the second trial
11-12
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run could not be isolated from the emission contribution of the coal, due to a change in the
combustion profile. The emission codependence is particularly evident in the additional release of
sulfur dioxide from the combustion of the carpet/coal mix; carpet does not contain sulfur, so it
would be expected that sulfur dioxide emissions would decrease if less coal is combusted.
However, the average concentration of sulfur dioxide actually increased by an approximate factor
of six with the combustion of the carpet/coal mix. The studies suggest this is the result of the
change in the thermal profile in the burn zone which favors the formation of sulfur dioxide over
sulfur trioxide (Lemieux et al. 2004, Realff 2010). This emission codependency appears evident
in a number of other compounds, where the decrease in emissions is not explainable by the
reduction in coal combustion alone (e.g., hydrochloric acid, chromium, selenium). While carpet
has a noticeably higher nitrogen content than coal, combustion temperatures and the potential
creation of thermal nitrogen oxides (i.e., resulting from the thermal combination of atmospheric
nitrogen and oxygen) are described as the most important factor in preventing nitrogen oxides
release. The emissions reported in the dataset are controlled and were measured after air pollution
control devices including a cyclone, spray tower, and fabric-filter baghouse (Lemieux et al. 2004).
Figure 11-9 presents the flows included, partially included, and excluded from the LCI dataset for
carpet incineration at a waste-to-energy plant. In the absence of additional information, it was
assumed that discarded carpet is transported 20 km from the point of generation to the cement kiln,
where it is first shredded before it is fired in the kiln.
t
C02 CO Chloride HeaV PM S02 HCI
metals
Dioxins
and
furans
THC NOx
Carpet;
from disposal
Bituminous coal;
— -~
at mine
Transport, barge;
average fuel mix
Transport, train;
diesel powered
Transport, single unit
- truck; short-haul; ~
diesel powered
Processing and cement kiln combustion
of carpet; at cement kiln; 21% carpet,
79% bituminous coal (mass)
Electricity
Legend
Elementary Flow
Technosphere Flow ^
| Not Included
H Partially Included
| Included
System Boundary
Figure 11-9. Unit process flow diagram for carpet (21% by mass) and coal (79% by mass)
combustion in a cement kiln.
11.8 Data Gap Analysis and Opportunities for Additional LCI Data
Figure 11-4 summarizes thetypeof datapresented by various sources reviewed for the compilation
of carpet EOL management LCIs. There are currently no sources that present complete LCIs for
11-13
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EOL carpet management options. Sources that present data on carpet incineration only include air
emissions measured during pilot-scale or trial operations (i.e., Konopa et al. (2008)). In addition,
solid waste emissions (e.g., fly ash) and water (e.g., stormwater runoff) are not provided. Lemieux
et al. (2004) provides partial air emissions which were derived from a pilot-scale incineration unit.
WARM documentation (US EPA 2019) focuses on greenhouse gas emission factors. Emissions
from landfilling include tailpipe emissions resulting from waste collection; landfill operations
equipment and leachate emissions are not considered. As included in US EPA (2016), emissions
from manufacturing and waste-to-energy incineration of carpet only account for GHG air
emissions. While HLA/CTA(1994) includes energy (i.e., natural gas) and amass balance of inputs
and outputs associated with polyurethane foam production, it does not include an estimate of
process electricity consumption. The leachable emissions from landfilled carpet, as provided in
Shoaeioskouei (2012), were limited to perfluorinated compounds.
Therefore, the following data gaps were identified for the development of LCIs representative of
the complete set of EOL management options for carpet:
1. Landfilled carpet and carpet padding leachate emissions. Limited data were found on the
leached emissions expected from carpet placed in a landfill. Specifically, CDD landfill leachate
emissions were limited to perfluorinated compounds based on a study that used distilled water
as the leaching solution over a test period of 6 hours and at a temperature of 15° C
(Shoaeioskouei 2012). These emissions may not be completely representative of those that
would be expected from the placement of carpet in a landfill, nor do they correspond to
emissions expected from carpet padding.
2. Emissions resulting from the incineration of carpet and carpet padding at a waste-to-
energy recovery facility. As described previously, carpet has a relatively high heat of
combustion and therefore can provide electricity when incinerated at an energy recovery
facility. Konopa et al. (2008) lists the uncontrolled emissions from the combustion of carpet
in a pilot-scale incineration facility. These emissions are not representative of those expected
from a full-scale WTE operation since a full-scale operation would be expected to have air
pollution control technology which would substantially reduce its air emissions.
In addition, air emissions were estimated based on individually combusted charges of carpet.
It is unlikely that carpet at WTE facilities would be combusted in isolation from other wastes.
Therefore, it may be unrealistic to try to extract the emissions due to the combustion of carpet
from the emissions due to the combustion of a complete waste stream of which carpet is a
component.
Finally, with the exception of particulates, emissions datafrom the combustion of polyurethane
foam was used as a proxy for carpet padding incineration emissions. While emissions of sulfur
dioxide (and potentially heavy metals) are detectable in the emissions produced from carpet
combustion, these data are currently missing from the literature.
3. Emissions resulting from the combustion of carpet at a cement kiln. It was not possible to
separate cement kiln emissions due to the combustion of carpet from emissions due to the
combustion of coal. Therefore, the dataset included in this study reports the emissions
associated with firing a specific coal/carpet mix. To fully analyze the emissions associated with
11-14
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4.
using discarded carpet as a cement kiln fuel, multiple combined-fuel scenarios should be
investigated.
Energy, material inputs, and emissions resulting from carpet recycling. No LCI data were
found for carpet recycling processes. Information is needed for each type of carpet recycling,
including mechanical separation, depolymerization, melt blending, and carpet reconditioning.
5. Electricity consumption for carpet padding recycling. No data were found that would allow
an estimate of process electricity consumption for the rebond polyurethane foam production.
6. Transport. Some of the literature suggested that carpet may travel significant distances from
the point of discard to various stages in a given recycling process (Mihut et al. 2001). Since a
20-km transport distance was assumed herein, more detailed information is needed to bolster
the analysis.
Table 11-3. Overview of US LCI data utilized for dataset development. "P" represents partial
Process
Konopa
et al.
(2008)
Lemieux
et al.
(2004)
US
EPA
(2016)
Pohland
and Rach-
dawong
(1996)
Shoaeio-
skouei
(2012)
HLA/
CTA
(1994)
USCB
(2004)
Bertrand
and
Wagner
(1997)
Garrido
et al.
(2016)
NIST
(2008)
Carpet
Landfilling
-
-
-
P
P
-
-
-
-
-
Cement Kiln
Fuel
-
P
-
-
-
-
-
-
-
-
Carpet Waste-to-
Energy
P
-
P
-
-
-
-
-
-
-
Carpet Padding
Waste-to-Energy
-
-
-
-
-
-
-
P
P
P
Carpet Recycling
-
-
P
-
-
-
-
-
-
-
Carpet Padding
Recycling
-
-
-
-
-
P
P
-
-
-
11.9 References
Bertrand, C.M. and Wagner, J.P. (1997). Evaluation of toxic emissions and residues from the
controlled combustion of selected foamed plastics. Polymer-Plastics Technology and
Engineering, 36(1): 67-88.
CARE (2019). Carpet America Recovery Effort, https://carpetrecovery.org.
CCG (2006). Targeted Statewide Waste Characterization Study: Waste Disposal and Diversion
Findings for Selected Industry Groups. A Report Prepared by Cascadia Consulting Group
for the California Integrated Waste Management Board, June 2006.
CRI(2017). The Carpet Primer. The Carpet and Rug Institute, Inc. Dalton, Georgia. https://carpet-
rug.org.
Garrido, M.A., Font, R., and Conesa, J.A. (2016). Pollutant emissions from the pyrolysis and
combustion of viscoelastic memory foam. Science of the Total Environment, 577, 183-
194.
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Harding Lawson Associates - Cross/Tessitore & Associates (HLA/CTA) (1994). Application for
permit to construct air pollution source. Application No. AC48-21 4902, Orlando, Florida,
30 June 1994.
Konopa, S.L., Mulholland, J.A., Realff, M.J., and Lemieux, P.M. (2008). Emissions from carpet
combustion in pilot-scale rotary kiln: Comparison with coal and particle-board
combustion. Journal of the Air & Waste Management Association, 58(8): 1070-1076.
Lemieux, P., Stewart, E., Realff, M., and Mulholland, J.A. (2004). Emissions study of co-firing
waste carpet in a rotary kiln. Journal of Environmental Management, 70, 27-33.
Mihut, C., Captain, D.K., Gadala-Maria, F., and Amiridis, M.D. (2001). Review: Recycling of
nylon from carpet waste. Polymer Engineering and Science, 41(9): 1457-1470.
Ohlemiller, T.J. and Shields, J.R. (2008). Aspects of theFire Behavior of Thermoplastic Materials.
NIST Technical Note 1493: National Institute of Standard sand Technology, January 2008.
NIST (2018). Building for Environmental and Economic Sustainability (BEES) Online 2.0
Technical Manual. NIST Technical Note 2032, National Institute of Standards and
Technology, December 2018.
https://nvlpubs.nist.gov/nistpubs/TechnicalNotes/NIST.TN.2032.pdf.
Patch, S.C.; Ullman, M.C.; Maas, R.P.; Jetter, J.J. (2009). A pilot simulation study of arsenic
tracked from CCA-treated decks onto carpets. Science of the Total Environment, 407(22):
5818-5824.
Pohland, F.G.; Rachdawong, P. (1996). Use of post-consumer carpet products during landfill
management of solid wastes. Water Science and Technology, 34(7-8): 429-436.
Potting, J.; Blok, K. (1996). Life-cycle Assessment of Four Types of Floor Covering. Journal of
Cleaner Production, 3(4):201-213.
Shoaeioskouei, S. (2012). Perfluorinated Compounds in Landfill Leachate from Discarded
Carpets, Masters Thesis, University of British Columbia: Vancouver, Canada.
Ucar, M.; Wang, Y. (2011). Utilization of recycled post consumer carpet waste fibers as
reinforcement in lightweight cementitious composites. International Journal of Clothing
Science and Technology, 23(4): 242-248.
US EPA (2019). Waste Reduction Model (WARM) Tool: User's Guide. Version 15, May 2019.
Available: https://www.epa.gov/warm/documentation-chapters-greenhouse-gas-emission-
energy-and-economic-factors-used-waste-reduction.
Zia, K.M.; Bhatti, H.N.; Bhatti, I.A. (2007). Methods for polyurethane and polyurethane
composites, recycling and recovery: A review. Reactive and Functional Polymers, 67(8):
675-692.
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12 Clay Bricks
12.1 Introduction
Clay bricks represent a relatively small fraction of the total CDD material stream and are
commonly generated from the demolition of buildings, structures, and pavements. The quantity
of clay bricks produced from demolition will vary depending on the building composition. Most
clay bricks are produced from common clay and shale, a material obtained from mining and
excavation which must go through an extensive drying and extrusion process prior to kiln firing.
Clay bricks are primarily manufactured for structural construction purposes, which require the use
of common face brick. With over 3.3 billion clay bricks being produced in 2008 in the US, this
quantity accounts for 60% of the nationwide total annual production of all brick types (USCB
2011). According toUSGS, close to 15.9 MMT of common clay was mined and nearly 57% of
this was used for clay brick production (USGS 2008). Figure 12-1 shows the EOL management
options for clay brick.
Figure 12-1. Process flow diagram for end-of-life management of clay bricks.
12.2 EOL Management
Less than 2% of CDD is comprised of bricks and approximately 11 MMT is discarded (US EPA
2020). Disposal via landfill is the dominant EOL management option for clay bricks. Only limited
amounts of discarded clay bricks are recycled. Due to concerns about structural strength, reuse of
salvaged bricks in load-bearing applications is not recommended (Webster 2002). Salvaged clay
bricks are sometimes reused in non-structural applications such as brick fireplaces, hearths, and
patios. Reza (2013) and Cavelline (2012) reported that typical brick recycling practices include
reuse as a replacement for aggregate in structural fills or pavements. Recovered clay bricks should
be processed prior to use as aggregate.
Table 12-1 lists and describes the LCA processes included in this report's associated datasetsfor
EOL management of clay bricks. Primary aggregate production and general material transport
datasets are relevant for multiple materials in this report and are discussed in detail in Chapter 2.
12-17
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Table 12-1. End-of-life management processes for clay brick.
Process
Description
Unlined CDD
landfilling of clay
brick; at unlined CDD
landfill
This process includes the placement and compaction of bricks and
their long-term physiochemical decomposition in a CDD landfill
environment.
MSW landfilling of
clay brick; at MSW
landfill
This process includes material and energy inputs as well as leachate
volumes expected from the disposal of clay bricks in an MSW
landfill, using the average, minimum and maximum density values
reported in six studies (Mohajerani et al., 2016; D. Eliche-Quesada,
2015; Phonphuak et al., 2011; Demir, 2006; Demir et al., 2005;
Sutcu and Akkurt, 2009). It also includes landfill construction,
operation, and closure/post-closure care.
Crushing of clay brick;
at crushing facility
This process includes the emissions and other environmental
impacts associated with the operation of a clay brick crushing
processing facility per kilogram of clay brick material. It accounts
for diesel and heavy equipment utilization; particulate matter (PM)
emissions associated with crushing, screening, stockpiling and
unpaved road transport; land occupation; electricity consumption;
and water consumption (for PM control from crushing, screening,
and conveyor transfer points).
Reuse of clay brick; at
construction
This process includes the reuse of clay bricks for new construction.
This process does not account for any other emissions that would
result from cleaning or preparing waste clay bricks for reuse.
Use of clay brick; as
soil fill
This process includes the placement of clay brick as a soil fill.
Emissions from mortar, including leaching emissions, are not
included in this dataset.
Crushing of clay brick;
as soil fill
This process includes the placement of crushed clay brick as soil fill.
Some fraction of the material likely represents mortar affixed to
bricks after demolition, although emissions from mortar, including
leaching emissions, are not included in this dataset.
12.3 LCI Sources
Peer-reviewed literature and government and private industry publications were reviewed to
identify available LCI datasets pertaining to clay brick EOL management. WARM documentation
was the only source of US-based datafound for EOL management datafor clay bricks. Table 12-2
lists data sources reviewed to compile LCI presented in this chapter. IfLCI data were not available,
process metadata and documentation were reviewed to evaluate the completeness of applicable
datasets (e.g., which emissions categories were included, background data used to compile the
dataset, geographic location, and time period of the data). Primary data sources were used
wherever possible.
12-18
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Table 12-2. Overview of life cycle inventory data sources associated with clay bricks.
Source
Description
US EPA (2019) -
WARM
WARM presents data on GHG emissions associated with the source
reduction, transport, and landfilling (i.e., collection and placement) of
clay bricks. While clay brick recycling was mentioned, no LCI data were
presented.
12.4 LCI Related to Disposal
The primary EOL management method for bricks in the US is landfill disposal.). Similar to all the
other materials modeled in WARM (2019) estimates GHG emissions released from fossil fuel
combustion resulting from the transport to and placement of demolished clay bricks at a landfill.
Since clay bricks are an inert material, they do not undergo biological decomposition.
Review of the literature only yielded international leaching test results for clay bricks. Karius and
Hamer (2001) conducted a series of leaching tests (pH static tests with a L:S-ratio of 10) on bricks
made from clay and dredging sediments from local water bodies to assess the impact of the
sediments on leaching of 20 contaminants, including metals and sulfur. The tests were conducted
on crushed bricks (grain size of 50-30,000 |im) made with and without sediment. In general, the
heavy metals from the sediment bricks leached at the upper end of the concentration ranges for the
bricks made without sediments.
12.5 LCI Related to Recycling
12.5.1 Clay Brick Demolition
Doka(2009) presented building materials demolition-, recycling-, and disposal-related LCI as part
of the Ecoinvent database based on management practices in Switzerland. Energy consumption
and air emission estimates from the study were derived from other studies that were done in the
European Union. Equipment demolition efficiencies (i.e., the time spent per volume of waste
demolished) and fuel consumption rates were compiled from literature to estimate the energy
required todemolish brick wall, gypsum board, and cement-fiber slab as 0.0359 MJ/kg. Particulate
matter is a major non-fuel air emission associated with demolition activities. Doka (2009) included
an air emissions factor of 80 mg PMio/kg of demolition waste for all building construction,
demolition, and renovations activities. AP-42 provides air emission factor calculation methods for
various heavy construction operations, which include dust generation activities from the
demolition of buildings and removal of debris as a function of various factors such as site-specific
conditions and equipment used (US EPA 1995).
12.5.2 Clay Brick Sorting
Prior to recycling, discarded clay bricks in the mixed CDD waste stream would undergo sorting
operations in which bricks were separated from other materials. While Doka (2009) reported
energy requirements of CDD materials sorting and size reduction specific to European practices,
US-specific data regarding brick processing are lacking. Please see Chapter 2 for more details
regarding the development of an LCI dataset for modeling and allocating the environmental
burdens of a mixed CDD processing facility.
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12.5.3 Clay Brick Use as Aggregate
While clay brick recycling is not common in the US, limited beneficial applications of recovered
clay brick as aggregate have been documented (Cavalline 2012, Reza 2013). The Minnesota
Department of Transportation conducted a study in 2013 to test the feasibility of reusing bricks in
aggregate for road base and showed that the material met departmental specifications (Reza 2013).
Cavalline (2012) explored the potential use of clay brick rubble as a possible replacement for
aggregate in building and pavement concrete mixtures with a focus on the mechanical and
engineering properties of brick aggregate. Cavalline (2012) reported that recycled brick may
provide acceptable performance when used in pavement and shows promise for use in structural
applications. The recovered bricks would need to be size-reduced for use as aggregate. US-specific
energy requirement and emission data specific to clay brick processing are lacking.
Similar to other CDD materials that may be processed and beneficially reused as aggregate, the
use of demolished clay brick aggregate would offset the production and transport of primary
aggregate materials. LCI datasets for primary aggregate production are presented and detailed in
Chapter 2 of this report. Due to the lack of gaseous and liquid emission and energy requirement
data, LCI for processing and use of clay bricks as recycled aggregates were not developed.
12.6 Data Gap Analysis and Opportunities for Additional LCI Data
Most LCI data on the EOL management of clay bricks are not specific to US practices. US EPA
(2012a) was the only source of information that provided US-specific data, but these data only
included GHG emissions factors for source reduction and landfilling of clay bricks. Based on a
review of government publications, peer-reviewed literature, and industry data, the following US-
specific LCI data gaps were identified with respect to EOL management of clay bricks:
1. Energy requirements for sorting and processing clay bricks at a CDD processing facility.
Clay bricks that are part of a mixed CDD stream are recovered at a CDD processing facility.
Recovery may include separating, grinding, and fractioning operations, depending on the end-
use market. Although diesel consumption data for sorting common CDD materials are
available (see Chapter 2), energy requirements for size-reduction and associated emissions
specific to brick are lacking.
2. Long-term leachable emissions from bricks placed as aggregate (e.g., in a fill) or in a
landfill. No US-specific leaching data for clay bricks were available. Development of these
data are critical to improve EOL modeling of clay bricks.
12.7 References
Cavalline, T.L. (2012) Recycled Brick Masonry Aggregate Concrete: Use ofRecycled Aggregates
from demolished Brick Masonry Construction in Structural and Pavement Grade Portland
Cement Concrete. Ph.D. Dissertation. University of North Carolina, Charlotte, NC, USA.
Doka, G. (2009). Life Cycle Inventories of Waste Treatment Services, Part V: Inventory of
Building Material Disposal. Ecoinvent Report No. 13, Swiss Centre for Life Cycle
Inventories, Dubendorf, April 2009.
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Ecoinvent (2014). Swiss Center for Life Cycle Inventories: Ecoinvent Centre. Dataset Information
(UPR): Treatment of Waste Asphalt, Sanitary Landfill, CH.
Karius, V. and Hamer, K. (2001). pH and Grain-Size Variation in Leaching Tests with Bricks
Made of Harbour Sediments Compared to Commercial Bricks. The Science of the Total
Environment, 278(1-3): 73-85.
Reza, F. (2013). Use of Recycled Bricks in Aggregates. A Report Prepared by Farhad Reza,
Published by the Minnesota Department of Transportation. Mankato, MN, USA. August
2013.
US EPA (1995). AP-42, Fifth Edition, Volume I, Chapter 13: Miscellaneous Sources, Section
13.2.3: Heavy Construction Operations.
US EPA (2019). Waste Reduction Model (WARM) Tool: User's Guide. Version 15, May 2019.
Available: https://www.epa.gov/warm/documentation-chapters-greenhouse-gas-emission-
energy-and-economic-factors-used-waste-reduction.
US EPA (2020). Advancing Sustainable Materials Management: 2018 Fact Sheet. Office of
Resource Conservation and Recovery. EPA 530-F-20-009. Washington, DC.
USCB (2011). Clay Construction Products - Summary 2010. MQ327D(10)-5. U.S. Census
Bureau, Washington, D.C., USA. May 2011. http://Lusa.gov/WmLqW4. Accessed 20
March 2014.
USGS (2008). Clays Statistics and Information. Mineral Commodity Studies.
http://on.doi.gov/lpj6Pri. Accessed 16 July 2014.
Webster, M. (2002). The Use of Salvaged Structural Materials in New Construction. Presentation
Posted on the U.S. Green Building Council Website, November. As cited in
http://l .usa.gov/lAuuNax.
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13 CopperWire
13.1 Introduction
Copper wire is found in building electrical equipment, telecommunications, and other electronic
systems. The quality and quantity of wire determine whether it goes to a landfill, is recycled, or is
processed for further use. Some percentage of copperwire is commonly removed from CDD sites
before it is able to be processed for end-of-life disposition. Since this activity falls outside most
state tracking systems, this percentage is typically not accounted for in material tracking studies.
Figure 13-1 depicts the flow of materials and processes included in this report's associated datasets
for EOL management of copper wire.
Figure 13-1. Process flow diagram for end-of-life management of copper wire.
13.2 EOL Management
Table 13-1 lists and describes the LCA processes included in this report's associated datasets for
EOL management of copper wire.
Table 13-1. End-of-life management processes for copper wire.
Process
Description
Unlined CDD
landfilling of copper
wire; 12-2; with
ground; at unlined
CDD landfill
This process includes material and energy inputs and emissions from
the placement of 12-2 (with ground) copper building wire in a CDD
landfill using the density of copper wire as calculated from
Cerrowire (2013). It also includes cover soil placement and landfill
operation, closure, and post-closure care.
MSW landfilling of
copper wire; 12-2; with
ground; at MSW
landfill
This process includes material and energy inputs as well as leachate
volumes expected from the disposal of 12-2 (with ground) copper
building wire in an MSW landfill using the density of copper wire,
as calculated from Cerrowire (2013). It also includes landfill
construction, operation, and closure/post-closure care.
Incineration of copper
wire; 12-2; with
ground; at MSW waste-
to-energy facility
This process includes incineration of copperwire (12-2, with
ground, non-metallic sheathing) at a US MSW waste-to-energy
facility.
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Process
Description
Processing of copper
wire scrap; chopped,
insulation removed; at
processing
This process simulates the energy requirements to chop and remove
insulation from copper wire scrap in preparation for secondary cast
copper production.
13.3 LCI Related to Disposal
As presented by CD A (2016), building wire has the greatest end-market share of all wire types,
representing approximately 36% of all wire mill products sold according to a preliminary estimate
of 2015 wire end-market consumption. The type 12-2 wire is a common copper building wire used
in residential applications and can handle up to a 20-amp load, which is enough power to run a
variety of household appliances (Cerro Wire, 2014). Type 12-2 wire is comprised of two conductor
wires and an additional ground wire; all three wires are size 12 American Wire Gauge (AWG).
The conductor wires may be insulated with PVC, while the ground wire is wrapped in paper. All
three wires are wrapped in a paper sheath which is surrounded by a PVC jacket. Based on 12 AWG
wire weights for bare copper building wire and 12-2 building wire (Southwire 2013, Cerro Wire
2013), it is assumed that copper and insulation (e.g., PVC and paper) make up approximately 72%
and 28% of the weight of 12-2 wire, respectively. However, the data regarding amount of paper
sheathing in 12-2 wire are lacking.
Because the specific copper wire modeled by this dataset incorporates a paper sheath, it is possible
that the landfill disposal of this material may contribute to landfill gas emissions. However,
because the paper sheath is located in the outer PVC jacket and because details were not found on
the mass fraction of paper per length of 12-2 copper wire, landfill gas production is not calculated
for this dataset. Due to an absence of SPLP data (or other leaching studies that estimate expected
leachate emissions in an CDD landfill environment), leachate emissions from the CDD landfilling
of this material are also not included in this dataset. However, it should be noted that copper wire
PVC insulation and jacketing may contain phthalate-based plasticizers and lead (used as a heat
stabilizer) and these components could potentially leach out of the insulation and jacketing in a
landfill environment (US EPA 2008).
For water balance purposes, precipitation infiltration is assumed to occur over 20 years using the
average precipitation from EREF (1999). In the absence of additional data, it is assumed that all
water applied to road surfaces for the purpose of dust control evaporates. The bulk material
densities used in this dataset represent uncompacted densities; compacted waste densities for
individual CDD materials were not found in literature. Engine classification of the equipment types
reported by EREF (1999) was estimated based on equipment make and model reported in operation
plans for several CDD landfills.
13.4 LCI Related to Incineration
Emissions resulting from the incineration of copper wire in an MSW waste-to-energy (WTE)
facility are assumed to be unaffected by the presence of other materials in the waste stream. In
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actuality, the specific mixture of emissions resulting from incineration of copper wire will be a
product of the combustion of both copper wire and other co-located materials.
Emissions were estimated for copper wire incineration by aggregating emissions from the
combustion of copper wire insulation and jacketing components individually. Specifically,
emissions were estimated for the incineration of PVC, ditridecyl phthalate (DTDP), trioctyl
trimellitate (TOTM), wax, bisphenol A, and lead sulfophthalate based on the mass fraction of each
of these components as found in copper wire, summing to a total 100%.
Per US EPA (2015), the average US MSW waste-to-energy facility is assumed to have an operating
temperature of 750 0 C. For the purposes of this dataset, fly ash is assumed to represent the collected
portion of particulate matter generated from the combustion of PVC. This dataset assumes the
collection efficiency for a number of air pollution control devices as listed in US EPA (1996). It
also assumes that the collection efficiency for benzoic acid vapor is the same as the collection
efficiency for hydrochloric acid vapor, as described in US EPA (1996).
13.5 Data Gap Analysis and Opportunities for Additional LCI Data
1. Copper wire collection. The modeling of LCI data has an inherent mass flow issue in that
copper wire is often removed from CDD sites before it can be collected and processed. As this
removal is not tracked in most statewide systems, more resourceful or statistical analyses could
help fill this data gap.
2. Landfilling of copper wire sheath. There is very little data available on leachate or other
emissions associated with landfilling of copper wire components, such as the surrounding
sheath. Development of such data would improve landfill modeling for copper wire
components.
3. Co-location of incinerated materials. Copper wire is rarely incinerated on its own in practice;
however, that is the current modeling standard. Development of co-incineration scenarios
would help build more robust LCI datasets.
13.6 References
CDA (2016). Annual Data 2016: Copper Supply & Consumption - 1995-2015. Report by the
Copper Development Association, Inc.
Cerrowire (2014). Application Charts. What Type and Gauge of Wire Should I Use?
https://www.cerrowire.com/products/resources/tables-calculators/applications-charts.
Accessed 21 April 2015.
Cerrowire (2013). Building Wire, NM-B Wire, Citrex® Nonmetallic-Sheathed Cable. Revised
January 2013. https://www.cerrowire.com. Accessed 21 April 2015.
EREF (1999). Life Cycle Inventory of a Modern Municipal Solid Waste Landfill. Report prepared
by Ecobalance, Inc. for the Environmental Research and Education Foundation, June 1999.
MA TURI (2002). Environmental, Health and Safety Issues in the Coated Wire and Cable Industry,
Technical Report No. 51. The Massachusetts Toxics Use Reduction Institute, University
of Massachusetts Lowell. April 2002.
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Southwire (2013). Bare Copper Wire and Cable. Bare Copper Conductor. Solid and Stranded.
https://www.southwire.com/wire-cable/copper-bare-covered/bare-copper-wire-and-
cable/p/CUBARE7. Accessed 21 April 2015.
USEPA (1996). AP-42, Fifth Edition, Volume I, Chapter 2, Section 2.1 Refuse Combustion
US EPA (2008). Wire and Cable Insulation and Jacketing: Life-Cycle Assessments for Selected
Applications, EPA 744-R-08-001. June 2008.
US EPA (2019). Waste Reduction Model (WARM) Tool: User's Guide. Version 15, May 2019.
Available: https://www.epa.gov/warm/documentation-chapters-greenhouse-gas-emission-
energy-and-economic-factors-used-waste-reduction.
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14 Vinyl Composition Tile
14.1 Introduction
Vinyl composition tile (VCT) is a floor material known for its durability and long lifespan,
especially in high-traffic commercial spaces. Most VCT products are made of polyvinyl chloride
(PVC) and calcium carbonate (limestone). Some VCT varieties include recycled materials. At the
end-of-life phase, VCT is transported to either a CDD or MSW landfill, or is incinerated in a waste-
to-energy (WTE) facility. Figure 14-1 illustrates the end-of-life disposition processes for VCT.
Figure 14-1. Process flow diagram for end-of-life management of vinyl composition tile (VCT).
14.2 EOL Management
Table 14-1 lists and describes the LCA processes included in this report's associated datasetsfor
EOL management of vinyl composition tile.
Table 14-1. End-of-life management processes for vinyl composition tile (VCT).
Process
Description
Unlined CDD
landfilling of vinyl
composition tile, VCT;
at unlined CDD landfill
This process includes material and energy inputs and emissions from
the placement of VCT in a CDD landfill, using the density of VCT
as calculated from NIST (2018). It also includes cover soil
placement and landfill operation, closure, and post-closure care.
MSW landfilling of
vinyl composition tile,
VCT; at MSW landfill
This process includes material and energy inputs as well as leachate
volumes expected from the disposal of VCT in an MSW landfill,
using VCT density calculated from NIST (2018). It also includes
landfill construction, operation, and closure/post-closure care.
Incineration of vinyl
composition tile, VCT;
at MSW waste-to-
energy facility
This gate-to-gate process models incineration of VCT at a US MSW
waste-to-energy facility.
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14.3 LCI Related to Disposal
This dataset estimates the volume of leachate emissions expected from the MSW landfilling of
VCT and assumes environmental burdens resulting from landfill operation over a 10-year
operational life and a 30-year post-closure care period. In the absence of additional data, it is
assumed that all water applied to road surfaces for the purpose of dust control (equivalent to 0.116
L/kg waste) evaporates. The density of individual VCT is assumed to be representative of the
compacted density of this material; data on the landfill-compacted density of VCT were not found
in literature. EREF (1999) reported fuel (diesel and gasoline) consumption rates for construction,
operations, and closure/post-closure phases of MSW, which were distributed among the various
equipment types. Engine classification of the equipment types reported by EREF (1999) was
estimated based on equipment make and model reported in operation plans for several MSW
landfills.
Leachate treatment and leachate mass emissions are not included in this dataset. Leaching studies
of VCT which estimate total leachable emissions associated with VCT placed in an MSW landfill
were not found in literature. However, both fugitive (volumetric) leachate emissions and the
quantity of leachate collected for treatment are included. In the absence of leaching studies which
would allow an estimate of leachate quality, fugitive leachate and collected leachate flows are only
considered partially accounted for. It should be noted that this dataset does not account for
electricity use during the post-closure period (e.g., leachate collection sumps).
14.4 LCI Related to Incineration
Emissions resulting from the incineration of VCT in an MSW waste-to-energy (WTE) facility are
assumed to be unaffected by the presence of other materials in the waste stream. In actuality, the
specific mixture of emissions resulting from incineration of VCT will be a product of the
combustion of both VCT and other co-located materials. It should be noted that the byproducts of
the combustion of individual components of VCT (i.e., lime, hydrogen chloride gas) may react
and produce net emissions that cannot be accurately estimated assuming aggregation of the
emissions from the combustion of individual VCT components.
In the absence of literature which describes the emissions resulting from the combustion of actual
VCTs, emissions were estimated for VCT incineration by aggregating emissions from the
combustion of individual VCT components. Specifically, emissions were estimated for the
incineration of limestone, PVC (used as a surrogate for vinyl chloride), vinyl acetate, benzyl butyl
phthalate (BBP), and diisononyl phthalate (DINP) based on the mass fraction of each of these
components, summing to 100% coverage.
Per US EPA (2015), the average US MSW waste-to-energy facility is assumed to have an operating
temperature of 750° C. For thepurposes of this dataset, fly ash is assumed to represent the collected
portion of particulate matter. Finally, this dataset assumes that the collection efficiency for a
number of air pollution control devices provided for hydrochloric acid vapor according to US EPA
(1996) is the same collection efficiency for benzoic acid vapor.
In terms of the transportation phase, it was assumed that VCT waste is transported 20 km from the
point of generation to the incineration facility.
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Emissions resulting from the combustion of non-PVC components of VCT were not available,
therefore, complete combustion (or calcination of limestone) was assumed and the resulting
emissions calculated stoichiometrically for these components. However, it is likely that the
combustion of these non-PVC components would contribute to some of the same emissions that
have been measured for the combustion of PVC. Specifically, combustion of non-PVC carbon-
based components (e.g., vinyl acetate, plasticizers) could potentially contribute to emissions of
furans, VOCs, benzoic acid, heat and carbon monoxide. Therefore, for purposes of LCI modeling,
these emission flows are considered partially included in this dataset.
It should be noted that some limited emissions information was identified for the combustion of
vinyl floor tiles. However, these data was either found for floor tiles with an unknown composition
(i.e., Boettneret al., 1973) or with a composition different from VCT (i.e., Potting and Blok, 1996).
For emissions resulting from the combustion of PVC, with the exception of benzoic acid, it was
assumed that the air pollution control devices described in US EPA (1996) would only impact
emissions of PM, hydrochloric acid, and dioxins and furans. The modeled dataset allows selection
of one of the following air pollution control options: electrostatic precipitator (ESP); dry sorbent
injection (DSI) and ESP; spray dryer (SD) and ESP; DSI and fabric filter (FF); SD and FF. Based
on US EPA (1996), the default selection is SD with FF.
Operational energy inputs (e.g., electricity for claw operation, diesel for heavy equipment
operation) are not included in this dataset. Other operation and maintenance consumables, such as
scrubber feed material (e.g., lime, bicarbonate) and lubrication for equipment are not quantified
for this dataset.
The following is a summary of data treatment for emissions from the combustion of PVC.
Emission modeling sources primarily used gas chromatography in conjunction with mass
spectrometry (GC/MS) to identify the compounds present in the combustion products from
incineration of PVC resins. For studies that reported emission results at different combustion
temperatures, emissions were selected or averaged to best approximate emissions expected at 750°
C. For example, if a set of emissions is given for combustion at 700° C and 800° C, the average of
the two sets was used.
During a study by Boettnar et al. (1973), samples ranging from 0.5g to 2g of different PVC
polymers were heated from 200° C to 600° C in a thermogravimetric analysis apparatus, where
samples were prevented from contacting the flame. The temperature was increased incrementally
over this range and emissions were collected over the entire heating process. Total emissions
resulting from this heating process may vary significantly from emissions resulting from the direct
exposure of PVC to 750° C (i.e., the expected temperature in an MSW WTE incinerator).
Availability of specific emissions data determined which studies were used in modeling the LCI
datasets.
Medium molecular weight PVC resin samples of O.lg to 0.2g were incinerated at temperatures
from 450-1050° C by Conesa et al. (2008) in a single-chamber furnace. Combustion gases were
passed through a XAD-2 resin adsorption trap and were also separately analyzed after being
collected in Tedlar bags to avoid resin adsorption of benzene and toluene. The gases were analyzed
by means of gas chromatography in conjunction with thermal conductivity detection or flame
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ionization detection. Eluents extracted from the XAD-2 resin were analyzed using a number of
different GC/MS techniques. Emission data from combustion at 850° C were used in this study.
Three (3)-gram samples of PVC powder were combusted in an electrically-heated drop tube
furnace by Shemwell and Levendis (2000) to analyze the sizes of emitted particulate matter.
Emissions given for lean fuel conditions were selected since it was assumed that a WTE facility
will operate with excess air. Emissions were passed through a series of particle impact plates and
a fiberglass filter to remove particles smaller than 0.4 [j,m. Emissions data taken from combustion
near 1230° C were selected from this study.
A split-cell electric muffle furnace was used by Wang et al. (2003a) to incinerate 0.5g samples in
the primary furnace chamber with a furnace wall temperature set from 500-1000° C. A portion of
the emissions were then directed to a secondary furnace chamber with a wall temperature set to
1000° C. The experiment had two exhaust points: one immediately after the primary furnace and
the other after the secondary furnace (i.e., the afterburner). Particulate emissions were collected
by means of paper filters with a nominal pore size of 0.45 [j,m. Gas-phase PAHs were adsorbed on
XAD-4 resin. Gas emissions were analyzed by use of GC coupled with flame ionization detection.
Eluents extracted from the resin were analyzed by means of GC/MS. Particulates were quantified
by weighing the filter papers.
One-stage combustion of 0.5-gram samples of PVC powder was conducted by Kim et al. (2004)
in a vertical electric muffle furnace, where emissions were collected using a glass microfiber filter
followed by XAD-2 resin and a backup toluene solvent impinger. Emissions datafrom combustion
at 600 and 900° C were averaged from this study.
One (1) gram samples of PVC powder were combusted in a tube-type furnace by Wang et al.
(2001) and (2003b) between 600-900° C. Emissions were collected by via a glass wool and glass
fiber filter with a pore diameter of 0.2 [j,m which was followed by an adsorption tube packed with
XAD-2 resin. Emissions data from combustion at 700° C and 800° C were averaged from this
study for inclusion in thedataset.
Based onBoettner etal. (1973), Conesa et al. (2008) and Shemwell and Levendis (2000), it appears
that there is no residue left behind from the combustion of PVC at or near 750° C. The quantity of
fly ash estimated in this dataset is a product of the uncontrolled particulate emissions estimated by
Shemwell and Levendis (2000) and Wang et al. (2003a) and the air pollution control efficiencies
as presented for a number of devices in US EPA (1996).
Emissions resulting from the calcination of limestone were estimated based on the stoichiometry
presented in Worrell et al. (2001). Zhong and Bjerle (1993) show that limestone is nearly 80%
calcined at a temperature of 900° C and a pressure of 1 atmosphere within 30 seconds of reaction
time. This dataset assumes 100% calcination under typical MSW WTE operating conditions. The
products of calcination are carbon dioxide and lime (i.e., CaO).
Emissions resulting from the combustion of vinyl acetate, BBP, and DINP were not found in
literature. Emissions were estimated assuming complete combustion of VCT components since
the flash point of each material (as presented in NIH (2017)) is well below the assumed operating
temperature of an MSW WTE facility. The combustion products (i.e., carbon dioxide and water
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vapor) of these 3 materials were calculated stoichiometrically. The molecular weights and
formulas for each of these VCT components were found inNIH (2017).
14.5 Data Gap Analysis and Opportunities for Additional LCI Data
1. Leachate treatment and emissions. Leachate emissions due to placement of VCT in a landfill
are missing from the literature, as are electricity inputs during the landfill post-closure period.
2. Emissions from incineration of non-PVC components of VCT. Currently, it is assumed that
non-PVC components contribute some of the same emissions as PVC components and thus
the LCIs model a stoichiometric split. More specific combustion and emissions release data on
non-PVC components should be investigated in future research.
14.6 References
Boettner, E.A., Ball, G.A., and Weiss, B. (1973). Combustion Products from the Incineration of
Plastics. Prepared forthe Office of Research and Monitoring, US EPA. 1 February 1973.
EREF (1999). Life Cycle Inventory of a Modern Municipal Solid Waste Landfill. Report
prepared by Ecobalance, Inc. for the Environmental Research and Education Foundation,
June 1999.
National Institute of Health (2017). PubChem Open Chemistry Database. Developed by the
National Center for Biotechnology Information under the U.S. National Library of
Medicine, https://pubchem.ncbi.nlm.nih.gov. Accessed 25 July 2017.
NIST (2018). Building for Environmental and Economic Sustainability (BEES) Online 2.0
Technical Manual. NIST Technical Note 2032, National Institute of Standards and
Technology, December 2018.
https://nvlpubs.nist.gov/nistpubs/TechnicalNotes/NIST.TN.2032.pdf.
Potting, J. and Blok, K. (1996). Life-cycle Assessment of Four Types of Floor Covering. Journal
of Cleaner Production, 3(4):201-213.
Shemwell, B.E. and Levendis, Y.A. (2000). Particulates Generated from Combustion of Polymers
(Plastics). Journal of the Air and Waste Management Association, 50:1, 94-102.
USEPA (1996). AP-42, Fifth Edition, Volume I, Chapter 2, Section 2.1 Refuse Combustion
Wang, Z., Wang, J., Richter, H., Howard, J.B., Carlson, J., and Levendis, Y.A. (2003a).
Comparative Study on Polycyclic Aromatic Hydrocarbons, Light Hydrocarbons, Carbon
Monoxide, and Particulate Emissions from the Combustion of Polyethylene, Polystyrene,
and Poly(vinyl chloride). Energy and Fuels, 17:999-1013.
Worrell, E., Price, L., Martin, N., Hendriks, C., and Media, L.O. (2001). Carbon Dioxide
Emissions from the Global Cement Industry. Annual Review of Energy and the
Environment, 26:303-329.
Zhong, Q. and Bjerle, I. (1993). Calcination Kinetics of Limestone and the Microstructure of
Nascent CaO. Thermochimica Acta, 223:109-120.
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15 Summary and Future Research Needs
15.1 Summary
The objective of the work presented in this report was to assess the body of knowledge regarding
CDD life-cycle data and to compile US-specific LCI models for distinct CDD material categories
from publicly available sources. In the previous chapters, available LCIdataforprocesses common
to all CDD (Chapter 2), as well as LCI datafor six CDD materials (Chapters 3-14) were presented.
While the 12 materials examined (asphalt pavement, asphalt shingles, gypsum dry wall, wood
products, land clearing debris, Portland cement concrete, PVC, fiberglass insulation, carpet and
carpet padding, clay bricks, copper wire, and vinyl composition tile) do not represent every
component of CDD, they do comprise the vast majority of CDD materials generated annually in
the US.
As described in the introduction to this report, CDD has not received the degree of attention with
respect to environmental emissions or other life-cycle considerations that other waste streams
have. Thus, for many LCI categories, US-specific data were not available from publicly available
sources. In addition, some of the available data that were used to develop an LCI category were
incomplete or needed to be estimated. Therefore, each chapter concludes with a description of LCI
data gaps.
This final chapter summarizes the data gaps highlighted in each of the chapters. Table 15-1
presents a summary of processes, associated energy and material inputs, and emissions. An "X" in
this table denotes that data were included in the LCI category, though it does not indicate that all
flows were able to be quantified. Product manufacturing LCIs were evaluated only for those CDD
materials that are commonly recycled in a closed loop.
Table 15-1. Summary of CDD life cycle inventory processes included in this report Refer to
Appendix B for the creation of bridge processes for additional data and emissions.
Material
Process
Energy
Input
Material
Input
Emissions1
Air
Water
Soil
Asphalt Pavement
Processing at facility
X
-
-
-
-
Use of reclaimed asphalt
pavement as general fill
X
X
-
-
-
Bridge to hot mix asphalt
-
-
-
-
-
Asphalt Shingles
Processing at facility
X
-
X
X
-
Use as general fill
-
-
-
X
-
Bridge to asphalt shingles
-
-
-
-
-
Gypsum Dry wall
Processing at dry wall facility
X
X
X
X
-
Disposal in CDD landfill
X
X
X
X
-
Disposal in an MSW landfill
X
X
X
X
-
Bridge to gypsum drywall
-
-
-
-
-
Wood Products
Application of mulch to land
X
-
X
X
-
Disposal of wood (treated,
untreated) in CDD landfill
X
X
X
X
X
Disposal of wood (treated,
untreated) in MSW landfill
X
X
X
X
-
15-1
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Material
Process
Energy
Input
Material
Input
Emissions1
Air
Water
Soil
Disposal of ash in CDD
landfill
X
X
X
X
-
Disposal of ash in MSW
landfill
X
X
X
X
-
Bridge to lumber
-
-
-
-
-
Land Clearing Debris
(LCD)
Processing for size reduction,
mulching
X
X
X
X
Combustion via open burning
X
-
X
-
-
Combustion via air curtain
incinerator
X
-
X
-
-
Disposal in CDD landfill
X
X
X
X
X
Disposal in MSW landfill
X
X
X
X
-
Portland Cement
Concrete
Use of debris as general fill
X
X
-
X
-
Use of aggregate as general
fill
X
X
-
X
-
Disposal in CDD landfill
X
X
X
X
-
Disposal in MSW landfill
X
X
X
X
-
Bridge to concrete mixes
-
-
-
-
-
Polyvinyl Chloride
(PVC)
Disposal in CDD landfill
X
X
X
X
-
Disposal in MSW landfill
X
X
X
X
-
Incineration with energy
recovery
X
-
X
-
-
Bridge to PVC pipe
-
-
-
-
-
Fiberglass Insulation
Disposal in CDD landfill
X
X
X
X
-
Disposal in MSW landfill
X
X
X
X
-
Carpet and Padding
Recycling of carpet at carpet
recycling facility
X
-
X
-
-
Recycling of padding at
rebond facility
X
X
X
-
-
Disposal of carpet in CDD
landfill
X
X
X
X
-
Disposal of padding in CDD
landfill
X
X
X
X
-
Disposal of carpet in MSW
landfill
X
X
X
X
-
Disposal of padding in MSW
landfill
X
X
X
X
-
Incineration of carpet with
energy recovery
X
-
X
-
-
Incineration of padding with
energy recovery
X
-
X
-
-
Incineration of carpet in kiln
with bituminous coal
X
X
X
-
-
Clay Bricks
Crushing at crushing facility
X
X
X
X
-
Crushing for soil fill
X
X
-
X
-
Use as soil fill
X
X
-
X
-
Reuse in new construction
X
X
-
-
-
15-2
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Material
Process
Energy
Input
Material
Input
Emissions1
Air
Water
Soil
Disposal in CDD landfill
X
X
X
X
-
Disposal in MSW landfill
X
X
X
X
-
Copper Wire
Processing and insulation
removal
X
X
-
-
-
Incineration with energy
recovery
X
-
X
-
-
Disposal in CDD landfill
X
X
X
X
-
Disposal in MSW landfill
X
X
X
X
-
Vinyl Composition Tile
(VCT)
Incineration with energy
recovery
X
-
X
-
-
Disposal in CDD landfill
X
X
X
X
-
Disposal in MSW landfill
X
X
X
X
-
Bridge to vinyl flooring
-
-
-
-
-
Mixed CDD
Recovery of CDD at mixed
CDD processing facility
X
X
X
X
-
'if no emissions are listed for any of the environmental media (i.e., air, water, or soil), refer to Appendix B and the appropriate
bridge processes in the dataset for additional data linking to process emissions.
15.2 Data Gaps and Future Research Opportunity
Based on the data gaps highlighted earlier in the report and that can be inferred from the table
above, future data-gathering and research opportunities have been identified. The following
sections highlight major LCI data categories pertaining to CDD materials and summarize their
associated data gaps and identified research needs.
EOL management practices of CDD materials
Of all the CDD materials reviewed during this study, only the EOL management practices of
asphalt pavement were found to be substantially well-documented and quantified; NAPA has been
conducting a US-wide annual survey of paving mix producers since 2009 to track uses of asphalt
pavement reclaimed from road construction and maintenance projects. Although USGS also
compiles and reports the amounts of reclaimed asphalt pavement (RAP)- and PCC-derived
recycled aggregates in the US based on a survey of aggregate producers and CDD contractors, the
data are incomplete because of survey limitations. For other materials, data on the amount of
materials managed via different EOL management options (e.g., recycled versus disposed) are
very limited.
Many state environmental agencies track the statewide amount of CDD materials landfilled
annually, but only four states (Florida, Massachusetts, Nevada, and Washington) appear to closely
track the amount of materials recycled. Due to this lack of EOL management recycling data and
because of the interest of multiple government agencies (e.g., USGS, US DOE, US EPA, FHWA)
and industry organizations (e.g., NAPA and CDRA) in analyzing this information, there is an
opportunity for collaborative research on the quantification of CDD materials EOL management
practices.
15-3
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CDD Materials Transport
For many of the LCI presented in this report, a uniform transport distance of 20 km was assumed
between the point of generation and the next step of management (e.g., processing facility,
landfill). Different transport distances would impact the results of an LCA, but such detailed data
are not currently available. This data gap could be addressed in several ways. First, the US Census
Bureau's commodity flow survey, which provides estimates of the distances various commodities
are transported via different modes, represents an opportunity to include waste haulers, recyclers,
and material processors in the future surveys to estimate the transport distances of CDD materials.
The current commodity flow survey includes "waste and scrap," but details regarding the materials
included and the universe of entities surveyed for the analysis are not available. A second option
would be to conduct direct research of facilities by compiling average or typical transport distances
for a variety of CDD management facility sizes (i.e., material quantity accepted) in different
geographic areas (e.g., each of the 10 US EPA regions).
Long-Term Liquid Emissions from Materials Deposited in Landfills
None of the US-specific LCA models (WARM, MSW-DST) include liquid emissions from the
disposal of materials at a CDD landfill. As discussed in the report, some MSW waste categories
can be used as proxies for estimating liquid emissions from a few of CDD materials, and some
European databases (e.g., Ecoinvent) include liquid emissions from CDD material disposal in a
sanitary landfill. Although liquid emissions based on laboratory leaching data(SPLP and TCLP)
for specific CDD components were presented in this analysis, this approach has limitations and
additional research must be conducted to provide a more comprehensive and realistic view of
liquids emissions at operating facilities. Example research areas include:
• identifying realistic L:S ratios for establishing a leaching test framework to assess liquids
emissions from an LCA perspective
• a larger listing of leached chemicals
• biological processes that result in leaching in addition to physico-chemical processes
• aggregated CDD materials rather than specific components, as the leaching behavior of a
specific CDD component may differ in the presence of another CDD component (or, in
like fashion, MSW components)
• the impact of the management of sludge from wastewatertreatment plant used for treating
leachate on the overall release of metals into the environment.
Long-Term Gaseous Emissions from Materials Deposited in Landfills
Gaseous emission estimates presented in this report only included methane, carbon dioxide, and
(to a lesser extent) hydrogen sulfide, and these estimates have several notable limitations. Methane
generation potentials (based on research conducted at North Carolina State University) for
branches and old corrugated cardboard (OCC) were used as a proxy to estimate the methane and
biogenic carbon dioxide emissions for wood/LCD and gypsum drywall (paper fronting and
backing), respectively. The researchers at North Carolina State University estimated methane
generation potential of various MSW constituents based on bioassays conducted in 2-L reactors.
These data have been widely used by various LCA models (e.g., WARM, MSW-DST). Although
multiple studies investigated hydrogen sulfide emissions from drywall disposal in landfills, the
15-4
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hydrogen sulfide generation potential, specifically resulting from the disposal of drywall and
recovered screen material (RSM) in CDD landfills, has not been reported. Hydrogen sulfide
generation estimates have been obtained as the result of either laboratory testing or from the bulk
disposal of MSW (which may include other sulfur-containing materials). Measurements specific
to CDD materials from larger-scale studies should be considered for future research to provide a
better estimate of the emissions of major (i.e., carbon dioxide, methane) and minor (hydrogen
sulfide, non-methane organic compounds) landfill gas constituents.
Several studies have documented how the cement in concrete can absorb atmospheric carbon
dioxide over time in a process called carbonation. This mechanism of carbon dioxide uptake was
not included in the LCI presented for concrete. Future studies should consider measurements of
carbon dioxide uptake by concrete disposed of in landfills or used in other recycling applications.
Long-Term Performance of Recycled Material-Derived Products and Services
Although the use of recycled materials to replace primary resource extraction would generally
reduce the overall impact on the environment, additional factors may reduce the anticipated
benefits of recycling. For example, pavement made from recycled concrete aggregate and/or RAP
may have a shorter service life compared to pavements manufactured entirely from primary
materials. Additional research should attempt to quantify the serviceable life of materials
manufactured from recycled materials on a per-mass-recycled basis, and account for this lifespan
difference in developing and updating LCI process datasets.
Decommissioning and Disposal Burdens
Similar to capital equipment burdens, a majority of LCI do not quantify the impacts of
facility/equipment decommissioning/disposal for the same reason described above. However, the
manner in which a process-dedicated piece of equipment is managed at the EOL may have a
significant impact on the overall emissions associated with that process. For example, if all the
steel recovered from landfill operations equipment (e.g. compactors, excavators, dozers) was
recycled for the production of new landfill operations equipment, the capital equipment burdens
associated with virgin iron ore extraction and smelting would be avoided.
Operation and Maintenance Consumable Burdens
While it is likely that the bulk of emissions resulting from the operation of a particular process
would occur as a result of energy use, almost all equipment requires the replacement of various
fluids, filters, and worn mechanical components over the course of its service life. The
environmental burdens resulting from the production of these consumable materials should be
accounted for during the future development of LCI. Until these emissions are quantified, it is not
possible to estimate their impact on the overall emissions associated with that particular process.
15.3 References
Wilburn, D.R. and Goonan, T.G. (1998). Aggregates from Natural and Recycled Sources:
Economic Assessments for Construction Applications - A Materials Flow Analysis. US
Geological Survey Circular 1176, US Geological Survey and US Department of the
Interior.
15-5
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Cochran, K.M. (2006). Construction and Demolition Debris Recycling: Methods, Markets and
Policy. Ph.D. Dissertation, University of Florida, Gainesville, FL, USA.
Stripple, H. (2001). Life Cycle Assessment of Road - A Pilot Study for Inventory Analysis, 2nd
Revised Edition. A Report Prepared by the IVL Swedish Environmental Research Institute
for the Swedish National Road Administration, March 2001.
https://trid.trb.org/view/689935. Accessed 20 February 2014.
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16 Acknowledgements
EPA would like to acknowledge the reviewers of several LCI process datasets, including Richard
Bergman of the U.S Forest Service Forest, Anders Damgaard of the Technical University of
Denmark (DTU), and Trine Henriksen, also of DTU. In addition, datasets were reviewed internally
by EPA LCA experts.
16-1
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17 AppendixA: LCI Review Template
Table A-l illustrates the review guidance sheet provided to reviewers of the CDD LCIs.
Conference calls were convened with external reviewers to demonstrate the use of the review
sheets and to answer any questions. The overarching theme of the review questions can be
summarized as, "Does the LCI provide the end user with enough information to make an informed
decision regarding use of the modeled process?" Each statement in the review guidance can be
rated as "acceptable as is", "minor revisions needed", or "major revisions needed". These ratings
and additional reviewer comments were considered in the revisions of the LCIs and modeled
processes.
Table A-l. Guidance template provided to LCI reviewers.
LCI Review: Process name
Statement to evaluate
Spreadsheet tab
Rating (drop-down menu)
Comments
Metadata fields sufficient for end user, i.e. LCA modeler, to
make informed modeling decisions.
General information
acceptable as is
minor revisions needed
major revisions needed
Metadata fields sufficient for end user, i.e. LCA modeler, to
make informed modeling decisions.
Modeling and validation
acceptable as is
minor revisions needed
major revisions needed
System diagram clearly describes the process as seen in
industry.
Calculations
acceptable as is
Typical material inputs are included.
InputsOutputs
minor revisions needed
Typical energy inputs are included.
InputsOutputs
major revisions needed
Typical emissions to air are included.
InputsOutputs
Typical emissions to water are included.
InputsOutputs
Typical emissions to soil are included.
InputsOutputs
Typical wastes requiring further treatment are included.
InputsOutputs
Parameters are used when appropriate, i.e. when the data
values could vary depending on circumstances.
Parameters
Based on your expertise, are there any other issues with
the data inventory that need to be addressed before the
data are published to the LCA Commons?
All tabs
Review completed by: Reviewer Name, Affiliation Ratings (drop-down)
Review form designed by: Briana Niblick, LCA Center of Excellence, U.S. EPA. acceptable as is
Please return to: niblick.briana@epa.gov. minor revisions needed
Federal LCA Commons contact: lca@epa.gov. major revisions needed
17-1
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18 AppendixB: Bridge Process Creation
This section contains a description of how CDD output flows were matched to USLCI and USEEIO input
flows in preparation for creating bridge processes to these background databases.1 Bridge processes
provide a path for connecting flows from any database to a flow in CDD.
Electricity
Output flow (kWh) matched to corresponding electricity LCI input flow (MWh).
Fuel
Output flows (kg) matched to corresponding USLCI flows (m3). Density values are taken from ACC,2011
documentation for fuel models used in USLCI. The density of each fuel is listed below:
Natural gas, processed, at plant: 0.7373 kg/m3
Diesel, at refinery: 867.3 kg/m3
Gasoline, at refinery: 739 kg/m3
Liquefied petroleum gas, at refinery: 542 kg/m3
Transport
Output flow (tkm) matched to corresponding USLCI flow (tkm).
Construction Equipment
Output flows (items) matched to corresponding USEEIO flow ($).
Using the equipment type and horsepower range in the output flow name, a corresponding machine was
identified on Caterpillar's (CAT) website.2 Price data for these machines were gathered through
communication with a representative from a CAT dealership located in Massachusetts. The prices
represent the most basic form of the machine, without any add-ons. Additionally, no national average
value exists, so these costs are region-specific (in this case, New England). Attempts were made to create
a national average by collecting cost data from various regions across the U.S., but CAT dealers were
unwilling to share prices to an entity located outside their geographic region.
Four output flows did not have a corresponding machine on CAT's website:
Terminal tractor; 175
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Terminal tractor; 300
-------
vvEPA
United States
Environmental Protection
Agency
PRESORTED STANDARD
POSTAGE & FEES PAID
EPA
PERMIT NO. G-35
Office of Research and Development (8101R)
Washington, DC 20460
Official Business
Penalty for Private Use
$300
Recycled/Recyclable Printed on paper that contains a minimum of
50% postoonsumer fiber content processed chlorine free
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