EPA/600/R-09/138F | March 2011 | www.epa.gov
United States
Environmental Protection
Agency
The Effects of
Mountaintop Mines and Valley Fills
on Aquatic Ecosystems of the
Central Appalachian Coalfields
United States Environmental Protection Agency
Office of Research and Development, National Center for Environmental Assessment
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EPA/600/R-09/138F
March 2011
The Effects of Mountaintop Mines and Valley
Fills on Aquatic Ecosystems of the Central
Appalachian Coalfields
NOTICE
This document has been reviewed in accordance with U.S. Environmental Protection Agency
policy and approved for publication. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC 20460
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DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental Protection
Agency policy and approved for publication. Mention of trade names or commercial products
does not constitute endorsement or recommendation for use.
ABSTRACT
This report assesses the state of the science on the environmental impacts of mountaintop
mines and valley fills (MTM-VF) on streams in the Central Appalachian Coalfields. These
coalfields cover about 48,000 square kilometers (12 million acres) in West Virginia, Kentucky,
Virginia, and Tennessee, USA. Our review focused on the impacts of mountaintop removal coal
mining, which, as its name suggests, involves removing all—or some portion—of the top of a
mountain or ridge to expose and mine one or more coal seams. The excess overburden is
disposed of in constructed fills in small valleys or hollows adjacent to the mining site.
Our conclusions, based on evidence from the peer-reviewed literature, and from the U.S.
Environmental Protection Agency's Programmatic Environmental Impact Statement released in
2005, are that MTM-VF lead directly to five principal alterations of stream ecosystems:
(1) springs, and ephemeral, intermittent, and small perennial streams are permanently lost with
the removal of the mountain and from burial under fill, (2) concentrations of major chemical ions
are persistently elevated downstream, (3) degraded water quality reaches levels that are acutely
lethal to standard laboratory test organisms, (4) selenium concentrations are elevated, reaching
concentrations that have caused toxic effects in fish and birds and (5) macroinvertebrate and fish
communities are consistently degraded.
Preferred citation: U.S. EPA (Environmental Protection Agency). 2011. The Effects of Mountaintop Mines and
Valley Fills on Aquatic Ecosystems of the Central Appalachian Coalfields. Office of Research and Development,
National Center for Environmental Assessment, Washington, DC. EPA/600/R-09/138F.
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CONTENTS
LIST OF TABLES vi
LIST OF FIGURES vii
LIST OF ABBREVIATIONS AND ACRONYMS ix
FOREWORD x
PREFACE xi
AUTHORS, CONTRIBUTORS AND REVIEWERS xii
ACKNOWLEDGMENTS xv
1. EXECUTIVE SUMMARY 1
2. INTRODUCTION 4
2.1. OPERATIONS USED IN MTM-VF 7
2.2. REGULATORY CONTEXT 10
3. LOSS OF HEADWATER RESOURCES 15
3.1. BACKGROUND 15
3.2. ESTIMATING THE EXTENT OF HEADWATER ECOSYSTEM LOSS 16
3.3. LOSS OF HEADWATER ECOSYSTEM BIOTA 20
3.4. LOSS OR ALTERATION OF HEADWATER ECOSYSTEM FUNCTIONS 24
3.4.1. Transformation and Removal of Nutrients and Contaminants 24
3.4.2. Storage and Export of Woody Debris 25
3.4.3. Organic Matter Processing 25
3.4.4. Habitat 26
4. IMPACTS ON WATER QUALITY 27
4.1. ALTERATION OF STREAMFLOW 27
4.2. CHANGES IN CHEMICAL TRANSPORT 30
4.2.1. pH, Matrix Ions and Metals 30
4.3. OTHER WATER QUALITY VARIABLES 39
4.3.1. Water Temperature 39
4.3.2. Nutrients 40
4.3.3. Dissolved Oxygen 40
4.4. CHANGES IN SEDIMENTATION 40
4.5. CHANGES IN SEDIMENT CHEMISTRY 42
4.6. CUMULATIVE IMPACTS 43
5. AQUATIC TOXICITY TESTS 45
5.1. TOXICITY TESTS USING WATER OR SEDIMENTS DOWNSTREAM OF
MTM-VF 45
5.2. TOXICITY TESTS ON WATER FROM OTHER ALKALINE COAL
MINING EFFLUENTS 46
in
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CONTENTS (continued)
5.3. TOXICITY OF MAJOR IONS: K+, HCO3 , Mg2+, Cl , SO42 , Na+, Ca2+ 47
5.3.1. Mount etal. (1997) 48
5.3.2. van Dam etal. (2010) 48
5.3.3. LasierandHardin(2010) 48
5.3.4. Soucek (2007a, b); Soucek and Kennedy (2005) 49
5.3.5. Meyer etal. (1985) 49
5.3.6. Skaar et al. (2006) 49
5.4. COMPARING TOXICITY TESTS ON MAJOR IONS TO OBSERVATIONS
DOWNSTREAM OF MTM-VF 50
5.5. TOXICITY OF TRACE METALS IN WATER 52
5.5.1. Selenium 52
5.5.1.1. Selenium Dynamics in Aquatic Ecosystems 53
5.5.1.2. Invertebrates 54
5.5.1.3. Fish 54
5.5.1.4. Birds 55
5.5.2. Manganese and Iron 55
5.6. TOXICITY OF TRACE METALS IN SEDIMENT 56
6. IMPACTS ON AQUATIC ECOSYSTEMS 59
6.1. EFFECTS ON BIOLOGICAL COMPOSITION 59
6.1.1. Benthic Macroinvertebrates 59
6.1.1.1. Benthic Macroinvertebrate Indices 59
6.1.1.2. Benthic Macroinvertebrate Diversity 64
6.1.1.3. Benthic Macroinvertebrate Density 65
6.1.1.4. Benthic Macroinvertebrate Functional Groups 65
6.1.1.5. Benthic Macroinvertebrate Taxa 65
6.1.2. Fish 69
6.1.3. Amphibians, Particularly Salamanders 70
6.2. EFFECTS ON ECOLOGICAL FUNCTION 71
6.3. BIOLOGICAL CONDITION 71
6.4. RELATIONSHIP BETWEEN BIOLOGICAL METRICS AND
ENVIRONMENTAL FACTORS 73
6.4.1. Ion Concentration 73
6.4.2. Specific Metals and Selenium 76
6.4.3. Organic and Nutrient Enrichment 77
6.4.4. Instream Habitat 77
6.4.5. Disturbance and Loss of Upland Habitat 79
6.5. CUMULATIVE EFFECTS 79
7. RECLAMATION, MITIGATION, AND RECOVERY 80
7.1. RECLAMATION OF MTM-VF SITES 80
7.1.1. Overview 80
7.1.2. Reclamation with Grasses and Pasture 82
7.1.3. Forestry Reclamation Approach (FRA) 83
iv
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CONTENTS (continued)
7.2. MTM-VF MITIGATION EFFORTS 84
7.2.1. Overview 84
7.2.2. On-Fill Mitigation Efforts 85
7.2.2.1. Constructed Channels 85
7.2.2.2. Natural Channel Design 87
7.2.2.3. Erosion Control Structures and Constructed Wetlands 88
7.2.3. Below-Fill Mitigation Efforts 88
7.2.3.1. Riparian Restoration and Stream Channel Enhancement 88
8. SUMMARY, INFORMATION GAPS AND RESEARCH OPPORTUNITIES 90
8.1. A CONCEPTUAL MODEL OF THE IMPACTS OF MTM-VF 90
8.2. CONCLUSIONS 93
8.2.1. Loss of Headwater Resources 94
8.2.2. Impacts on Water Quality 94
8.2.3. Toxicity Impacts on Aquatic Organisms 95
8.2.4. Impacts on Aquatic Ecosystems 96
8.2.5. Cumulative Impacts of Multiple Mining Operations 96
8.2.6. Effectiveness of Mining Reclamation and Mitigation Efforts 96
8.3. INFORMATION GAPS, ASSESSMENT NEEDS AND RESEARCH
OPPORTUNITIES 97
8.3.1. Update the MTM-VF Inventory and Surveys of Impact Extent 97
8.3.2. Quantify the Contributions of Headwater Streams 98
8.3.3. Improve Understanding of Causal Linkages 98
8.3.4. Develop Tests Using Sensitive Taxa 99
8.3.5. Conduct Mesocosm and Microcosm Experiments with Indigenous Taxa 101
8.3.6. Further Investigate Selenium and Sediments 101
8.3.7. Quantify Cumulative Effects 102
8.3.8. Quantify Longitudinal Effects 103
8.3.9. Quantify Effects on Stream Hyporheic Zones 103
8.3.10. Quantify Functionality of Constructed Streams and Mitigation Efforts 103
8.3.11. Expand the Scope of Review to Include Evidence from Non-Peer-
Reviewed Sources and Terrestrial Impacts 104
LITERATURE CITED 105
APPENDIX A: LITERATURE SEARCHES 124
APPENDIX B: REGULATORY ISSUES RELATED TO MTM-VF OPERATIONS 128
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LIST OF TABLES
1. Watershed areas above the toe of valley fills in permits approved between 1985 to
2001 18
2. Water quality variables in unmined streams versus streams below valley fills 31
3. Water quality parameters for unmined or reference streams or streams
downstream from mined, filled, or filled and residential watersheds in West
Virginia 33
4. Alkalinity, pH, and metals in control streams and streams downstream from filled
watersheds in West Virginia 34
5. Range of dissolved oxygen, pH, and conductivity values for sites in eastern
Kentucky 34
6. Seasonal mean (standard deviation) of conductivity (uS/cm) for the four classes of
streams 37
7. Substrate measures in streams located in different land use classes 41
8. Proportion of sediments that were sand and fines (mean [standard error]) in paired
sites 41
9. Range of sediment concentrations of metals and arsenic (mg/kg) in streams
downstream from the sedimentation ponds below valley fills in 2002 and 2004
and from a reference site in 2002 42
10. Polycyclic aromatic hydrocarbons, arsenic, and metals detected in sediments of
larger streams in the Kanawha Basin 44
12. Comparison of measured sediment concentrations with probable effects levels 57
13. Summary of research examining the relationship between mountaintop mining
and ecological characteristics in downstream habitats 61
14. Average ion concentration (reported as specific conductance) in MTM-VF and
reference streams reported in conjunction with biological data 75
VI
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LIST OF FIGURES
1. A summary diagram of the principal observed and expected effects of MTM-VF
on aquatic ecosystems 2
2. The central Appalachian coalfields 5
3. An overview of activities and sources associated with MTM-VF 6
4. A watershed view of a mountaintop mine and valley fill 8
5. Small stream watershed before and after mountaintop mining and creation of a
valley fill 9
6. Satellite images of the 40-km2 Hobet 21 mine (Boone County, WV), and the
Washington DC area, at the same scale 11
7. Permit boundaries for surface and underground mines in southwestern West
Virginia 12
8. Earth movement by humans and streams 13
9. Observed and expected effects of stream loss and burial and riparian forest
clearing on aquatic ecosystems 17
10. Map showing loss of headwater streams to MTM-VF 19
11. Hot spots of rarity-weighted species richness in the United States 22
12. Observed and expected effects of MTM-VF on streamflow characteristics 28
13. Observed and expected effects of MTM-VF on total dissolved solids, metals, and
pH 35
14. Observed and expected effects of MTM-VF on sediments, nutrients, and
temperature 39
15. Ions expected to contribute to effects in toxicity tests of water sampled
downstream of MTM-VF 50
16. Selenium transformation, transfer and effects expected in aquatic ecosystems
downstream of MTM-VF 53
17. Macroinvertebrate and fish responses associated with MTM-VF 60
18. Conceptual model depicting stages of change in biological conditions in response
to an increasing stressor gradient 72
vn
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LIST OF FIGURES (continued)
19. Macroinvertebrate and fish responses associated with stressors and treatment
ponds in studies of MTM-VF 74
20. Mn and Fe deposits on a caddisfly collected downstream of a mountaintop mine
and valley fill 78
21. Observed and expected effects of on-site reclamation and stream mitigation
efforts 81
22. Observed and expected effects of MTM-VF on aquatic ecosystems 91
Vlll
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LIST OF ABBREVIATIONS AND ACRONYMS
AWQC
BCG
CaCO3
CaMg(CO3)
CWA
ECX
EPA
EPT
FECX
FRA
GIS
GLIMPSS
HBI
HCO3
mi
KSO4
LCX
LOEC
MBI
MgSO4
MHRW
MTM-VF
NaHCO3
NPDES
OSMRE
PAH
PEIS
SMCRA
IDS
USAGE
WVDEP
WVSCI
ambient water quality criterion
biological condition gradient
calcite
dolomite
Clean Water Act
effect concentration for x% of the tested organisms
U.S. Environmental Protection Agency
Ephemeroptera, Plecoptera, and Trichoptera
field-based effect concentration for x% of the tested organisms
Forestry Reclamation Approach
geographic information system
genus-level index of most probable stream status
Hilsenhoff Biotic Index
bicarbonate
Index of Biotic Integrity
potassium persulphate
lethal concentration for x% of the tested organisms
lowest-observed-effect concentration
macroinvertebrate bioassessment index
magnesium sulfate
moderately hard reconstituted water
mountaintop mines and valley fills
sodium bicarbonate
National Pollutant Discharge Elimination System
Office of Surface Mining Reclamation and Enforcement
polycyclic aromatic hydrocarbons
Programmatic Environmental Impact Statement
Surface Mining Control and Reclamation Act
total dissolved solids
United States Army Corps of Engineers
West Virginia Department of Environmental Protection
West Virginia Stream Condition Index
IX
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FOREWORD
Headwater streams and watersheds in Appalachia are keystone components of the
region's ecology. They are sources of clean, abundant water for larger streams and rivers, are
active sites of the biogeochemical processes that support both aquatic and terrestrial ecosystems,
and are characterized by exceptional levels of plant and animal biodiversity. The benefits of
healthy headwaters are cumulative as the critical ecological functions of many small streams
flowing into the same river system are necessary for maintaining ecological integrity.
The practice of mountaintop mining and valley fills, which has become increasingly
common in Appalachian states, can have major environmental consequences for the mountain
ecosystem, the nearby valleys, and the downstream water quality. There is a growing body of
evidence in the scientific literature that valley fills from mountaintop mining are having
deleterious ecological effects. Recent published reports (reviewed herein) show that as water
quality deteriorates downstream of a valley fill, the biota within the stream are likewise affected.
The mining of coal in the United States is highly regulated. Mountaintop mining, in
particular, involves multiple statutes and agencies at both the federal and state levels. The two
key federal laws are the Surface Mining Control and Reclamation Act (SMCRA, 25 U.S.C.
§ 1201) and the Clean Water Act (CWA, 33 U.S.C. § 1252). The key entities at the federal level
are the Office of Surface Mining Reclamation and Enforcement (OSMRE), the
U.S. Environmental Protection Agency (EPA), and the U.S. Army Corps of Engineers (USAGE).
On June 11, 2009, in a Memorandum of Understanding, these agencies committed to a series of
activities to improve the regulation of mining practices under existing statutory authorities.
This assessment report is one of several actions EPA has initiated to better understand the
ecological impacts of mountaintop mining. For this report, the EPA Office of Research and
Development has reviewed and assessed the published peer-reviewed literature on the aquatic
impacts associated with mountaintop mining. This report was externally peer reviewed by
EPA's Science Advisory Board (SAB) and reflects the SAB's comments and suggestions. In
addition, comments received from the public, the mining companies, and environmental groups
were evaluated in preparing this final report. This final peer-reviewed assessment will inform
the EPA as the Agency continues to implement its regulatory responsibilities under the Clean
Water Act.
Michael W. Slimak, PhD, Associate Director
National Center for Environmental Assessment, Office of Research and Development
U.S. Environmental Protection Agency, Washington, DC
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PREFACE
This assessment of the effects of mountaintop mines and valley fills on aquatic
ecosystems was requested by the U.S. Environmental Protection Agency's (EPA's) Office of
Water and regional offices. It will be used to inform the EPA's reexamination of its reviews of
Appalachian surface coal mining operations under the Clean Water Act, the National
Environmental Policy Act, and the Environmental Justice Executive Order (E.O. 12898). The
report was prepared by the National Center for Environmental Assessment in EPA's Office of
Research and Development.
The assessment reviews and evaluates evidence from peer-reviewed sources published up
through December 2010 and the Programmatic Environmental Impact Statement and its
associated appendices published in 2003 and 2005. The external review draft released April
2010 (EPA/600/R-09/138A) was reviewed by EPA staff and panel of the EPA's Science
Advisory Board (SAB) that convened July 20 to 22, 2010 (EPA-SAB-11-005, available online at
www.epa.gov/sab). In addition, hundreds of comments from the mining companies, other
government agencies, nonprofit environmental and scientific organizations, and private citizens
were received through the docket or at the SAB panel meeting. Comments from all of these
sources were considered and used to improve the clarity and scientific rigor of the document.
XI
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AUTHORS, CONTRIBUTORS AND REVIEWERS
AUTHORS
Susan B.Norton, PhD1
Michael Griffith, PhD2
Laurie Alexander, PhD1
Amina Pollard, PhD3
Glenn W. Suter II, PhD2
Stephen D. LeDuc, PhD1
^.S. Environmental Protection Agency, National Center for Environmental Assessment,
Washington, DC
9 _
U.S. Environmental Protection Agency, National Center for Environmental Assessment,
Cincinnati, OH
3U.S. Environmental Protection Agency, Office of Water, Washington, DC
CONTRIBUTORS
Kate Schofield, PhD
U.S. Environmental Protection Agency, National Center for Environmental Assessment,
Washington, DC
Stefania Shamet, JD
U.S. Environmental Protection Agency, Region 3, Philadelphia, PA
REVIEWERS
R. Hunter Anderson, PhD
U.S. Environmental Protection Agency, Office of Research and Development, National Center
for Environmental Assessment, Cincinnati, OH
Theodore R. Angradi, PhD
U.S. Environmental Protection Agency, Office of Research and Development, National Health
and Environmental Effects Research Laboratory, Mid-Continent Ecology Division, Duluth, MN
Paolo D'Odorico, PhD
Department of Environmental Sciences, University of Virginia, Charlottesville, VA
xn
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AUTHORS, CONTRIBUTORS AND REVIEWERS (continued)
Ken M. Fritz, PhD
U.S. Environmental Protection Agency, Office of Research and Development, National
Exposure Research Laboratory, Cincinnati, OH
Kyle J. Hartman, PhD
Division of Forestry, West Virginia University, Morgantown, WV
Brent R. Johnson, PhD
U.S. Environmental Protection Agency, Office of Research and Development, National
Exposure Research Laboratory, Cincinnati, OH
Teresa Norberg-King, PhD
U.S. Environmental Protection Agency, Office of Research and Development, National Health
and Environmental Effects Research Laboratory, Mid-Continent Ecology Division, Duluth, MN
Caroline Ridley, PhD
U.S. Environmental Protection Agency, Office of Research and Development, National Center
for Environmental Assessment, Washington, DC
Science Advisory Board Panel on Ecological Impacts of Mountaintop Mining and Valley
Fills
Duncan Patten, Chairman, PhD
Montana State University, Bozeman, MT
Elizabeth Boyer, PhD
Pennsylvania State University, University Park, PA
William Clements, PhD
Colorado State University, Fort Collins, CO
James Dinger, PhD
University of Kentucky, Lexington, KY
Gwendelyn Geidel, PhD
University of South Carolina, Columbia, SC
Kyle Hartman, PhD
West Virginia University, Morgantown, WV
Alexander Huryn, PhD
University of Alabama, Tuscaloosa, AL
Xlll
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AUTHORS, CONTRIBUTORS AND REVIEWERS (continued)
Lucinda Johnson, PhD
University of Minnesota Duluth, Duluth, MN
Robert Hilderbrand, PhD
Appalachian Laboratory, University of Maryland Center for Environmental Science, Frostburg,
MD
Thomas W. La Point, PhD
University of North Texas, Denton, TX
Samuel N. Luoma, PhD
University of California-Davis, Sonoma, CA
Douglas McLaughlin, PhD
National Council for Air and Stream Improvement, Kalamazoo, MI
Michael C. Newman, PhD
College of William & Mary, Gloucester Point, VA
Todd Petty, PhD
West Virginia University, Morgantown, WV
Edward Rankin, MS
Ohio University, Athens, OH
David Soucek, PhD
University of Illinois at Urbana-Champaign, Champaign, IL
Bernard Sweeney, PhD
Stroud Water Research Center, Avondale, PA
Philip Townsend, PhD
University of Wisconsin-Madison, Madison, WI
Richard Warner, PhD
University of Kentucky, Lexington, KY
xiv
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ACKNOWLEDGMENTS
We gratefully acknowledge our colleagues in Region 3 and the National Center for
Environmental Assessment, especially Greg Pond, Maggie Passmore, Bette Zwayer, Maureen
Johnson, C. Richard Ziegler, Jeff Frithsen, Michael Troyer, David Bussard, Michael Slimak, and
Peter Preuss. Chris Russom, Chuck Stephan, and Dave Mount of the National Health and
Environmental Effects Research Laboratory, Mid-Continent Ecology Division, provided
valuable assistance on literature searches and interpretation of the toxicity of ions. We thank the
many reviewers who provided comments through the public docket and Science Advisory Board
review process. Contract support was provided by Tetra Tech under contract EP-C-07-068.
Cover photographs by Maggie Passmore.
xv
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1. EXECUTIVE SUMMARY
Mountaintop mines and valley fills (MTM-VF) lead directly to five principal alterations
to stream ecosystems (see Figure !):(!) springs, and ephemeral, intermittent streams, and small
perennial streams are permanently lost with the removal of the mountain and from burial under
fill, (2) concentrations of major chemical ions are persistently elevated downstream, (3) degraded
water quality reaches levels that are acutely lethal to standard laboratory test organisms,
(4) selenium (Se) concentrations are elevated, reaching concentrations that have caused toxic
effects in fish and birds, and (5) macroinvertebrate and fish communities are consistently and
significantly degraded. These conclusions are based on evidence, described in this report, from
the peer-reviewed literature and from the U.S. Environmental Protection Agency (EPA)
Programmatic Environmental Impact Statement (PEIS) released in 2005. Our review focused on
the impacts on mountaintop removal coal mining, which, as its name suggests, involves removal
of all or some portion of the top of a mountain or ridge to expose and mine one or more coal
seams. The excess overburden is disposed of in constructed fills in small valleys or hollows
adjacent to the mining site.
Evidence shows that concentrations of chemical ions are, on average, about 10 times
r\
higher downstream of MTM-VF than in streams in unmined watersheds. Sulfate (SO 4 ),
bicarbonate (HCOs ), calcium (Ca2+), and magnesium (Mg2+) are the dominant ions in the
mixture, but potassium (K+), sodium (Na+), and chloride (Cl~) are also elevated. These ions all
contribute to the elevated levels of total dissolved solids (TDS) typically measured as specific
conductivity and observed in the effluent waters below valley fills. Downstream ion
concentrations were accurately predicted using a simple dilution model, indicating that
concentrations decrease primarily when diluted by a cleaner source of water—for example, an
unmined tributary. Water from sites having high chemical ion concentrations downstream of
MTM-VF is acutely lethal to invertebrates in standard aquatic laboratory tests, and models of ion
toxicity based on laboratory results predict that acute toxicity would be expected from the ions
alone. Benthic macroinvertebrate assessments of condition frequently score "poor quality" or
"biologically impaired" at sites downstream of MTM-VF. Declines in macroinvertebrate indices
were observed at ion concentrations well below those associated with effects in tests of mine
effluent using standard laboratory organisms.
Selenium concentrations are also elevated downstream of MTM-VF. Selenium can
bioaccumulate through aquatic food webs—especially in ponds and reservoirs where retention is
high and food webs are long. Elevated levels have been found in fish in this mining region.
More than half of the sites surveyed downstream of MTM-VF exceeded the chronic-duration
Ambient Water Quality Criterion (AWQC) for selenium. Selenium has been associated with
1
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mountamtop mines
& valley fills
overburden&
coal removal
overburden
handling
T" watercontactwith
overburden
stream loss & burial
stream habitat
'r total dissolved solids
/r deformities of
survival, growth&
reproduction of standard
toxicity test organisms
aquaticfauna
s|/ quality & quantity of aquatic communities
Figure 1. A summary diagram of the principal observed and expected effects
of MTM-VF on aquatic ecosystems.
increased death and deformities in fish and reduced hatching in birds in studies of coal
overburden effluents in other regions.
Permits already approved from 1992 through 2002 are projected, when fully
implemented, to result in the loss of 1,944 km of headwater streams. This represents a loss of
almost 2% of the stream miles in the focal area (KY, TN, WV, and VA), a length that is more
than triple the length of the Potomac River, just during this 10-year-period. We found no studies
that updated the MTM-VF inventory conducted as part of the PEIS in 2002, but both mine
footprint and stream losses were projected to double over 2002 levels by 2012. An updated
inventory that would support statistically sound estimates of cumulative stream loss is a critical
information need.
2
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Reclamation practices (e.g., contouring and revegetation) were common in all of the
reviewed studies. The data indicate that reclamation partially controls the amount of soil erosion
and fine sediments transported and deposited downstream. The acidic drainage that is often
associated with coal mining is largely neutralized through reactions with carbonate minerals
within the valley fills or treatment in the sediment retention ponds. Yet, because ions, metals,
and selenium below MTM-VF were elevated in the reviewed studies, we conclude that past and
current management efforts do not improve all aspects of water quality. Additionally, there is no
substantive evidence in the literature or PEIS that onsite mitigation by constructed channels or
wetlands has replaced or will replace the lost ecosystem functions and biodiversity.
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2. INTRODUCTION
The purpose of this report is to assess the state of the science on the environmental
impacts of MTM-VF on streams in the Central Appalachian Coalfields.l As defined in the PEIS,
the coalfields cover about 48,000 square kilometers (12 million acres) in West Virginia,
Kentucky, Virginia and Tennessee, USA (see Figure 2) (U.S. EPA, 2003, 2005).
The Central Appalachian Coalfields have a long history of mining. Current mining
practices, including MTM-VF, employ methods to control the acid mine drainages that have
been a historic and continuing source of water quality degradation. The purpose of this report is
to evaluate evidence of the impacts of MTM-VF on headwater and downstream systems despite
improvements in acidic discharges. It is prompted by EPA's reexamination of how best to
implement environmental laws—especially the Clean Water Act (CWA), that are relevant to
surface mining (see Section 2.2).
We evaluated six potential consequences of MTM-VF (see Figure 3):
• Loss of headwater resources (see Section 3)
• Impacts on water quality (see Section 4)
• Impacts from aquatic toxicity (see Section 5)
• Impacts on aquatic ecosystems (see Section 6)
• Cumulative impacts of multiple mining operations
(see subsections of Sections 3, 4, and 6)
• Effectiveness of on-site reclamation and mitigation activities (see Section 7)
We reviewed the impacts on terrestrial ecosystems from the narrow perspective of their effects
on aquatic ecosystems. Our review of reclamation and mitigation practices was limited to their
effectiveness in improving on-site aquatic ecosystems. We did not evaluate the impacts of
MTM-VF on cultural or aesthetic resources, or human health.
lrThe derivation of the study boundary is described further in Chapter 4 of the PEIS (U.S. EPA, 2003, 2005).
4
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Central Appalachian Coalfields
State Boundaries
/ County Boundaries
Figure 2. The central Appalachian coalfields.
Source: EPA (2003, 2005).
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mountaintop mines
& valley fills
upland& riparian road construction heavy equipment overburden &
forest clearing & maintenance operations coal removal
overburden
handling
V
V y
f bare compacted soils J
regrading&
replanting
4- riparian cover
\/
\|/ y
f streamburial j^ f valleyfill J fbackstacksj
stream loss
V
surface runoff
( stream N ..
I reconstruction J ^
\/
^watershed
erosion
sediment retention
&treatment ponds
compaction of
valleyfill surface
I.
See Section 4:
IMPACTS ON WATER QUALITY
I
I
I
watercontact
with overburden
See Section 3:
LOSS OF HEADWATER RESOURCES
LE
GEND
human activity
f source J
additional step in
causal pathway
treatment or
reclamation
See Section 5:
AQUATICTOXICITY TESTS
See Section 6:
IMPACTS ON AQUATIC ECOSYSTEMS
CD See Section 7:
RECLAMATION & MITIGATION
Figure 3. An overview of activities and sources associated with MTM-VF.
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We used two sources of information for our evaluation: (1) the peer-reviewed, published
literature and (2) the PEIS and its associated appendices (U.S. EPA, 2003, 2005). Only a few
peer-reviewed papers have studied water quality or stream ecosystems in headwaters directly
affected by or downstream of MTM-VF in the Central Appalachian Coalfields (see Appendix A).
This report draws from these papers and from the relevant research findings of laboratory studies
and observational studies from other locations and mining activities. We also discuss the
findings published in the PEIS, which was published as two separate documents: the Draft,
published in 2003, and the Final, published in 2005. The final PEIS included responses to
comments on the draft and newer research results but did not include a revision of the original
material. When citing results from the many appendices of the PEIS, we specified the source to
make it easier for readers to find the original material. Finally, authoritative textbooks were used
as a source of background information and general scientific knowledge.
2.1. OPERATIONS USED IN MTM-VF
Mountaintop removal mining, like other surface mining practices, removes the soil and
rock over a coal seam (i.e., the overburden) to expose the coal. This overview of the processes
used in MTM-VF summarizes the description in the PEIS (U.S. EPA, 2003, 2005; see Figure 3).
The mountain or ridge top is prepared for mining by building access roads, clearing all trees and
stockpiling topsoil for future use in reclamation. Then, explosives are used to blast the entire top
of the mountain or ridge to expose and mine one or more coal seams (see Figure 4). As much as
300 vertical meters (1,000 ft) of overburden are removed. With it, any springs, ephemeral,
intermittent, and small perennial streams on the mountain's surface are also removed. The
overburden removed during mountaintop mining cannot all safely be put back into place because
of the overall volume of the material and because the volume increases when the rock is broken
up. Some of the overburden is stored on the mined surface in backstacks and used to recontour
the surface. The excess overburden is disposed of in constructed fills in valleys or hollows
adjacent to the mined site. These fills bury additional springs and ephemeral, intermittent and
small perennial streams.
Both water flow and sediment discharges are altered by MTM-VF (see Figure 5). The
heavy equipment used to mine and move the overburden compacts the bare soils, forming a
large, relatively impervious surface on the mined site that increases surface runoff. Surface
runoff is diverted into ditches and sediment ponds, replacing natural subsurface flow paths.
Water flows out of the ditches through notches, or is directed toward the valley fill. Depending
on the construction and degree of compaction of the valley fill, the water then either percolates
through porous fill material or flows through ditches and coarser rock drains within, under, or
beside the fill. The effluent that emerges downstream of the ditches and below the downgradient
7
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Former
Mountain Contour
Figure 4. A watershed view of a mountaintop mine and valley fill (no
consistent scale).
8
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i i i
Sensitive taxa
Tolerant taxa
- Water Table
Soil Horizon (organic/mineral)
Coal Layer
Overburden
Figure 5. Small stream watershed before and after mountaintop mining and
creation of a valley fill (simplified view). Monuments added for scale. Scales
differ between upper and lower halves of diagram.
A. Before mining. The figure on the left side of the diagram illustrates the natural topography, geologic
strata, and soil layers associated with small mountain streams in eastern coalfields. Stream valleys (natural
depressions in the landscape that conduct channelized streamflow) are the most obvious topographic feature
of the watershed. However, most of the water in small watersheds flows underground though a complex
system of local aquifers (a) soil layer interflows (b) and minu -. stress fractures in geologic strata of the
parent mountain (c). Overland flow and subsurface flows (indicated by arrows) form seeps or springs
(d) and channelized flows (e) that integrate features of the entire landscape, including riparian vegetation and
diverse, in-stream biological communities.
B. After mining. On the right side, the same watershed is shown after the mountain rock layers have been
removed, crushed, and deposited in the stream valley. Flat surfaces of remaining rock layers are less
permeable, producing higher surface runoff into a flood control channel (f) and valley fill (g) height is
approximate). Infiltration though valley fills of water exposed to larger total surface area of porous
unweathered rock (h) produces higher channelized flows and higher concentrations of dissolved ions and
trace metals downstream, where biological communities shift towards tolerant taxa (i). Subsurface
flowpaths in the intact geologic strata vary, depending on the types of rock in them, but water tables can
'back up' against the valley fill, elevating the water level in the fill, as shown here (]), increasing baseflows
and exposure to valley fill materials.
Photographs of macroinvertebrates by Greg Pond.
-------
edge (i.e., the toe) of the valley fill is discharged into constructed channels and then to ponds that
are also used as treatment basins, for example, to settle solid particles, precipitate metals, or
regulate pH.
After the coal is removed, the extraction area is graded and planted to control sediment
runoff. The sediment retention pond can be eventually removed, and the stream channel is
recreated under the footprint of the pond.
The coal is transported from the mine using trucks, conveyers, or rail to a processing site,
where it is washed prior to transport to market. The impacts of coal processing, slurry ponds,
and transport are not discussed in this report.
Mines can be as large as some cities (see Figure 6) and can use several different types of
mining, including underground methods such as room and pillar or long-wall mining and surface
methods such as contour, area and high-wall mining, in addition to mountaintop removal.
Though these other forms of mining can also produce fills, valley fills from mountaintop mining
operations are expected to be larger because of the volume of material involved. The active life
of a mine increases with size; larger mines can be active between 10 and 15 years.
The density of all coal mining activity (surface and underground) can be quite high in
some parts of the region (see Figure 7). Current statistics on the spatial extent of MTM-VF are
r\
unavailable. As of 2002, the footprint of surface mine permits was estimated at 1,634 km
(U.S. EPA, 2002) or about 3.4% of the land cover in the central Appalachian coalfields. As of
2001, permits for 6,697 valley fills were approved (U.S. EPA, 2002).
Surface mining and reclamation have been identified as the dominant driver of land
cover/land-use change in the central Appalachian coalfields and have produced significant
changes in the region's topography, hydrology, vegetation, groundwater, and wildlife
(Townsend et al., 2009; Loveland et al., 2003; U.S.EPA, 2003, 2005). Coal mining in this region
was identified as the greatest contributor to earth-moving activity in the United States (Hooke,
1999; see Figure 8).
2.2. REGULATORY CONTEXT
MTM-VF are permitted by state and federal surface mining and environmental protection
authorities. Individual mines are regulated under the Surface Mining Control and Reclamation
Act (SMCRA) by the Office of Surface Mining Reclamation and Enforcement (OSMRE) and by
delegated states under OSMRE oversight. In addition, several specific sections of the CWA
apply. These are implemented by the EPA, the U.S. Army Corps of Engineers (USAGE), and
individual states authorized to implement portions of the CWA. Although a complete listing and
interpretation of the regulations that affect MTM-VF operations are beyond the scope of this
paper, Appendix B provides a brief discussion of how water quality standards are implemented
through the CWA in the context of MTM-VF.
10
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A.
B
Figure 6. Satellite images of the 40-km Hobet 21 mine (Boone County, WV)
(Panel A), and the Washington DC area (Panel B), at the same scale.
Source: Google Maps (2009).
11
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Wesl Virginia Environmental I >rotee<
County Boundaries
Mining Permit Boundaries
Hoi
Imtctiv*
U«t,r,ow,/flth«i
Figure 7. Permit boundaries for surface and underground mines in
southwestern West Virginia. The Hobet 21 mine is shown in middle left near
Point a.
Source: WVDEP (2009a). Colors modified to improve legibility.
12
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Figure 8. Earth movement by humans and streams. Maps of the United
States showing, by variations in peak height, the rates at which earth is moved in
gigatonnes per annum in a grid cell measuring 1° (latitude and longitude) on a
side, by (A) humans and (B) rivers.
Source: Hooke (1999), used with permission from the publisher.
Two CWA permits are relevant to MTM-VF. The USAGE issues a permit pursuant to
Section 404 of the CWA (33 U.S.C. § 1344) for the discharge of dredged and/or fill material.
This permit includes the valley fill itself and the fill necessary to create a sediment pond below
the valley fill. The second permit is issued by either the EPA or an authorized state pursuant to
Section 402 of the CWA (33 U.S.C. § 1342). The Section 402 program is also known as the
National Pollutant Discharge Elimination System (NPDES). The NPDES permit includes the
discharge from the sediment pond and any stormwater associated with the mining activity.
13
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Both permitting programs prohibit activities or discharges that cause or contribute to
violations of numeric or narrative state water quality criteria. While numeric criteria protect a
water body from the effects of specific chemicals, narrative criteria protect a water body from the
effects of pollutants that are not easily measured, or for pollutants that do not yet have numeric
criteria, such as chemical mixtures, or suspended and bedded sediments. Examples of narrative
standards that are particularly relevant to evaluating MTM-VF impacts include
• From West Virginia: No significant adverse impact to the chemical, physical, hydrologic,
or biological components of aquatic ecosystems shall be allowed (WV § 47-2-3).
• From Kentucky: Total dissolved solids or conductivity shall not be changed to the extent
that the indigenous aquatic community is adversely affected (401 KAR 10:031,
Section 4(f)).
• From Kentucky: "Adversely affect" or "adversely change" means to alter or
change the community structure or function, to reduce the number or proportion
of sensitive species, or to increase the number or proportion of pollution tolerant
aquatic species so that aquatic life use support or aquatic habitat is impaired
(401 KAR 10:001, Section 1(5)).
14
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3. LOSS OF HEADWATER RESOURCES
Headwater streams dominate surface flows in the United States and comprise 70-80% of
the total stream miles in the eastern coal mining states (Leopold, 1964; U.S. EPA, 2003, 2005).
Headwater stream ecosystems occur on all mountains in the eastern coalfields and in all valleys
that receive the excess overburden from mountaintop mining. Impacts include the loss of
headwater streams on the removed mountaintops; burial of streams in the actual footprint of the
valley fills; and potential fragmentation of remaining stream and riparian habitats.
3.1. BACKGROUND
The term "headwaters" refers to the numerous small channels that form the origins of a
stream or river network. Headwaters are characterized by small drainage area, shallow channels,
and variable flow. They include hillside springs and seeps, creeks permanently or seasonally
connected to local or regional groundwater sources, and transitional channels that flow only
during periods of rainfall or snowmelt. Variation in the timing and duration of flow, and the
relative contributions of groundwater and stormwater inputs are used to classify headwater
streams as perennial, intermittent, or ephemeral (Hewlett, 1982). Perennial headwaters are
predominantly groundwater-fed and have continuous surface or subsurface flow except in
exceptionally dry periods; intermittent streams flow seasonally (e.g., winter, spring) when
groundwater levels are elevated; and ephemeral streams receive no groundwater input and flow
only in response to precipitation events (e.g., rainfall, snowmelt) (Johnson et al., 2009). A single
stream channel can have reaches in all three flow duration classes.
An alternative way to classify headwater streams is by stream order (Strahler, 1957).
Streams without tributaries are first-order streams, second-order streams are formed when two
first-order streams join, third-order streams are formed when two second-order streams join, and
so forth. First- and second-order streams are typically classified as headwaters
(Meyer and Wallace, 2001; Gomi et al., 2002; Benda et al., 2005). Because stream order is
usually determined by using maps, determinations will vary with the scale and accuracy of the
map (Leopold, 1994).
The hydrology and setting of headwaters influence their function, especially the
transformation and transport of water, organic matter, sediment, and other materials downstream
(Paybins, 2003; Freeman et al., 2007; Nadeau and Rains, 2007). Flow properties are influenced
by drainage area, climate, topography, channel morphology, underlying geology, and other local
and regional factors. Field classification of headwater streams by stream order or flow duration
15
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class (ephemeral, intermittent, or perennial) requires an understanding of the geographic setting
(Paybins, 2003; Fritz et al., 2006).
In the central Appalachian region, natural headwaters are forested, high-gradient streams
that occur on hilltops and in intervening valleys. They are typically dendritic (branched pattern
similar to tree roots) in structure with channels confined by underlying rock layers. In unmined
areas, hilltop streams receive inputs from precipitation, overland flow, and local hilltop aquifers.
These local aquifers are formed by shallow groundwater perched on low-permeability rock
layers above coal seams. Lateral discharge into hillside streams and springs, often at coal
outcrops (see Figure 5), is the primary direction of flow from such aquifers, which can have
residence times as brief as a week (Hawkins et al., 1996). Vertical discharge through rock
fractures (secondary porosity) and intact layers (primary porosity) also occurs but typically at
much slower flow rates. Vertical discharge occurs through deep zones of unweathered rock and
has long residence times, while lateral discharge circulates on near-surface zones of weathered
rock with shorter residence times. Both types of flow connect headwaters and local aquifers to
regional rivers and groundwater. Hilltop stream channels and aquifers slow runoff into valleys,
reducing erosion and contributing to flood control (Callaghan et al., 2000).
Here we summarize the available data on headwater ecosystem loss and burial (see
Section 3.2); and potential impacts to headwater biota (see Section 3.3) and headwater ecosystem
function (see Section 3.4 and Figure 9).
3.2. ESTIMATING THE EXTENT OF HEADWATER ECOSYSTEM LOSS
The OSMRE inventoried valley fills in the central Appalachian coalfields to estimate the
number of headwater stream miles lost to mountaintop mining and valley fills, based on permit
data and a 0.12-km2 (30-acre) minimum watershed size. This study found that in the 17-year
period from 1985 to 2001, approximately 1,165 km (724 mi) of headwater streams were
permanently buried under valley fills in West Virginia, Kentucky, Virginia, and Tennessee
(U.S. EPA, 2003, 2005). In a cumulative impact study, the EPA (U.S. EPA, 2002) reassessed
the number of stream miles lost by including those that were lost to other mining activities
(blasting, backfilling, etc.) in addition to valley fill footprints. In the revised estimate, 1,944 km
(1,208 mi) of streams were approved to be lost due to mountaintop removal, valley fills, and
associated activities from 1992 to 2002 (U.S. EPA, 2003, 2005). This means that more than 2%
of the total stream miles and 4% of first- and second-order stream miles in the PEIS study area
were approved for permanent loss or burial during this 10-year period. More current statistics
were unavailable at the time this report was written, but both mine footprints and stream losses
were projected to double over 2002 levels by 2012 (U.S. EPA, 2002).
16
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mountain top removal
\/
V
backstacking
& valley fill
removal of soil & surface
channel material
removal of geologic layers
& perched aquifers
V
r stream loss J
ephemeralstream
lessor burial
intermittent stream
lessor burial
V
V
•I- intermittent &
smallperennial
stream habitat
•I- hillslope/stream
exchanges
\
/
A sediment
(\i- amphibians^ C^ macroinvertebrates^)
perennialstream
lessor burial
A groundwater
surface water
connectivity
•I- organic matter
processing
amphibians^
Figure 9. Observed and expected effects of stream loss and burial and riparian forest clearing on aquatic
ecosystems. See Section 3 for more discussion and evidence.
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Headwater streamflows are tightly coupled with catchment and hillslope processes
(Gomi et al., 2002). Headwater catchment burial estimated as the watershed area above the toe
of valley fills on permits approved from 1985 through 2001 is another metric for assessing
headwater ecosystem loss (see Table 1).
Table 1. Watershed areas above the toe of valley fills in permits approved
between 1985 to 2001
State
KY
TN
VA
WV
Total
Average watershed area
above valley fill toe
km2 (acres)
0.23
0.22
0.33
0.39
(56.30)
(54.85)
(81.05)
(97.28)
Largest watershed
area above valley fill
toe
km2 (acres)
15.28
1.17
5.01
6.59
(3,777)
(288)
(1,238)
(1,628)
Total watershed area
approved for valley fill
km2 (acres)
1,138.57
12.21
172.5
451.14
1,774.42
(281,347)
(3,017)
(42,629)
(111,479)
(438,472)
Source: Chapter III of EPA (2003, 2005).
At the time of this report, data to quantify the area impacted by valley fill permits
approved prior to 1985 or since 2001, or predict cumulative losses from planned MTM-VF
activities, are not available in the MTM-VF PEIS or peer-reviewed literature. Data to quantify
headwater stream loss by flow duration class (e.g., ephemeral, intermittent, and perennial) are
not available; however, in a study of 36 headwater streams in southern West Virginia for which
valley fill permits were pending or approved, Paybins (2003) estimated that the median
r\
watershed area for intermittent flows was 0.06 km (14.5 acres) and the median watershed size
for perennial flows was 0.17 km2 (40.8 acres). The same study cites digital geographic
information system (GIS) data of valley fills estimated from permit maps by the West Virginia
Department of Environmental Protection (WVDEP) showing that the median size of permitted
r\
fills in southern West Virginia is 0.05 km (12.0 acres, comparable to the median intermittent
stream drainage), and the maximum size of permitted fills is 1.94 km2 (480 acres). Statewide
averages of valley fill area reported in the MTM-VF PEIS (see Table 1) range from
0.22-0.39 km2 (54.36-97.28 acres). Estimates of valley fill area from both reports suggest that
headwaters in all three duration classes are being permitted for burial by valley fills. Valley fill
footprints estimated from permit maps may be smaller or larger than the actual fill areas. Some
18
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areas approved for fill may not be used; in other cases, areas from permit maps underestimate the
actual area of fill (see Figure 10).
Key
Streams
Valley fill area estimated from permit maps
Valley fill area estimated from elevation data
Permit boundaries
Figure 10. Map showing loss of headwater streams to MTM-VF. This
diagram depicts the loss of stream miles and channel complexity that can result
from extensive mountaintop mining and valley filling. Solid blue lines inside
valley fill areas represent buried streams. Note that some headwaters above filled
areas are disconnected from the rest of the stream network.
Source: Modified from Figure 12 in Shank (2004).
Estimating headwater stream loss in terms of area or miles of stream impacted in
watersheds above a size threshold is a useful beginning, but it does not address the full extent of
affected headwaters, or loss of other aquatic ecosystems. For example, the current estimate does
not include unmapped streams, springs, seeps, and wet areas that may occur in watersheds less
r\
than 0.12 km (30 acres) in size, or headwaters disconnected from the stream network by valley
fills (see Figure 10). Small stream channels often are not designated on United States Geological
Survey topographic maps (Hansen, 2001) and difficult to detect on aerial photographs; thus,
accurate inventories of them are difficult and surveys frequently underestimate their true extent
(Fritz et al., 2006). Similarly, estimates based on stream miles or catchment area do not include
impacts due to the loss of headwater wetlands and forested vernal pools, which provide refuge
and habitat for breeding, hunting, foraging by amphibians, reptiles, and aquatic or semiaquatic
invertebrates. In the West Virginia portion of the study area, the projected loss of riparian habitat
r\
from MTM-VF is 30.72 km , 3.2% of the riparian habitat in the study area. Approximately 42%
19
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of these projected losses occur in headwater (first- and second-order) streams (U.S. EPA, 2002).
Data for forested wetlands, and for riparian habitats lost in other MTM-VF states, were not
available at the time of this report.
Further, current estimates of impacts do not account for potential long term effects of
landscape-scale changes in land cover associated with mountaintop removal mining. Effects of
spatially extensive changes in land-cover on regional biodiversity can persist for decades. For
example, Harding et al., (1998) found that past land-use activity (primarily deforestation and
agriculture) resulted in long term modifications to and reductions in stream fish and invertebrate
diversity, despite preservation of forest patches and reforestation of stream riparian zones.
3.3. LOSS OF HEADWATER ECOSYSTEM BIOTA
The biodiversity of the central Appalachians is of national and even global significance.
The southern Appalachian and most of the central Appalachian Mountains were a refuge for
organisms during the last glacial period, which ended 10,000 years ago (McKeown et al., 1984;
Soltis et al., 2006; Zeisset and Beebee, 2008; Potter et al., 2010). The New and Kanawha Rivers
in West Virginia were the headwaters of the Teays paleodrainage, a major preglacial route offish
dispersal from the Appalachian Mountains to the Mississippi River, and they are home to many
endemic (regionally unique) species today (Hocutt et al., 1978; Stauffer and Ferrerri, 2002).
New stream species or potential species are still being discovered in central Appalachia, which
includes areas of notable biodiversity identified by NatureServe (see Figure 11). For example,
Berendzen et al. (2008) found that the roseyface shiner (Notropis rubellus), thought to be a
single, widespread species, includes at least four species in the Central Highlands and Lowlands
with an endemic genetic lineage above Kanawha Falls. Kozak et al. (2006) found high levels of
genetic diversity in the common two-lined salamander (Eurycea bislineata complex), including
an endemic lineage in southern Virginia and West Virginia. Nearly 10% of global salamander
species diversity is found within streams of the southern Appalachian Mountains
(Green and Pauley, 1987), which are near the southern extent of the Teays River paleodrainage.
A phylogeographic study of the North American giant salamander or hellbender
(Cryptobranchus alleganiensis) found evidence of close genetic relationships among populations
living in the central Appalachian, southern Appalachian, and Ozark Mountains, which were
connected via preglacial river systems (Routman et al., 1994). The biodiversity of this region is
a valuable natural resource in economic, cultural, aesthetic, scientific, and educational terms
(Hughes and Noss, 1992; Cairns and Lackey, 1992; Dudgeon et al., 2006; Meyer et al., 2007).
We assume that most of the organisms inhabiting a headwater stream and riparian area
are eliminated when that headwater basin is buried or blasted during the mining process. It is
possible that some microorganisms persist in or colonize buried stream channels, but we found
20
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no studies of these systems. Surveys conducted as part of the PEIS, studies of stream biota
found in unmined Appalachian streams, and relevant studies of headwaters in other temperate
regions provide background on potential biodiversity impacts due to MTM-VF. This
information is discussed below.
Headwater habitats are spatially and temporally dynamic and support diverse biological
communities (Gomi et al., 2002; Meyer et al., 2007; Clarke et al., 2008). Small but biologically
significant differences in light, hydrology, water chemistry, substrate, sediments, food resources,
gradient, and precipitation across small streams within the same river network offer a wide
variety of habitats and niches for aquatic and semiaquatic plants, animals and microorganisms
(Southerland, 1986; Vinson and Hawkins, 1998; Meyer et al., 2007). Communities in small,
permanent or intermittent streams differ from those found in larger streams and rivers (Vannote
et al., 1980; Morse et al., 1993, 1997; Hakala and Hartman, 2004; Stauffer and Fererri, 2002).
Flow permanence and duration are likely to influence aquatic community structure, because the
relative abundance (number of individuals) of species with lower resistance or resilience to
drying is expected to decrease as surface water flows become more intermittent (Fritz and
Dodds, 2004; Arscott et al., 2010). Many headwater stream taxa are adapted to variable flows,
and even ephemeral and intermittent streams can support diverse and abundant invertebrate
assemblages (Feminella, 1996; Williams, 1996; Kirchner et al., 2003). Kirchner et al. (2003)
sampled 36 intermittent and perennial headwater streams in West Virginia and Kentucky that
were scheduled for burial by MTM-VF and collected approximately 73 genera and 41 families of
aquatic invertebrates. Many of the genera were found in both intermittent and perennial stream
types. Similarly, Collins et al. (2007) found that subsurface invertebrate community composition
was comparable in intermittent and perennial stream reaches of a stream having surface flows
over only 30% of its length in summer.
In addition to the differences in taxonomic structure of invertebrate assemblages in
headwater streams, the functional role of aquatic invertebrates also differs from larger streams
and rivers (Vannote et al., 1980) because of the closer contact between stream and forested
terrestrial ecosystems. Particularly in densely forested Appalachian catchments, large amounts
of leaf litter fall into small streams in autumn. This leaf litter supplies much of the energy for the
detrital food webs in these streams (Wallace et al., 1997; Meyer and Wallace, 2001; Cross et al.
2006; Wipfli et al., 2007). Many invertebrates in small streams consume whole, decomposing
leaves (i.e., shredders) or organic particles created when these leaves are broken apart (i.e.,
collectors). In turn, predators, including other invertebrates, salamanders, and fish, feed on these
invertebrates.
21
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Low
High
Source: NatureServe and its Natural Heritage
member programs July 2008
Produced by National Geographic Maps and NatureServe
December 2008
Figure 11. Hot spots of rarity-weighted species richness in the United States. The central Appalachian
Mountains—including the central Appalachian coalfields—have been identified as one of the most significant hot spots
for biological diversity in the United States.
Source: NatureServe and its Natural Heritage member programs, July 2008 (National Geographic Maps and NatureServe, 2008). Used with permission.
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Headwater streams support diverse algal and fungal communities. In studies of two
Appalachian headwater streams, more than 30 species of diatoms and more than 40 species of
fungi were recorded (Gulis and Suberkropp, 2004; Greenwood and Rosemond, 2005). Many
algal and fungal species' ranges extend from the Mid-Atlantic Highlands to the southern extent
of the Appalachian Mountains (e.g., Ponader and Potapova, 2007), and similarly high levels of
diversity are expected in central Appalachia (Pan et al., 2000). Diatoms (silica
skeleton-producing algae) and fungi are important food sources for fish and aquatic insects.
Fungi produce enzymes that are essential to the rapid decomposition of organic matter (e.g.,
wood and leaf litter). The breakdown of plant matter by fungi and other microbes makes energy
and nutrients in difficult-to-digest vegetation accessible to fish and invertebrates (Gulis et al.,
2006).
Appalachian headwater streams also support diverse and abundant assemblages of
amphibians. Salamanders are the most common vertebrates in headwaters and may often be the
major predator of the aquatic invertebrates (Davic and Welsh, 2004). Many stream salamanders
require ephemeral and intermittent streams in forested habitats to maintain viable populations
(Petranka, 1998; Davic and Welsh, 2004). Among the Appalachian plethodontids, species vary
in their preferences for ephemeral, intermittent, or perennial headwaters to the extent that life
stage and taxonomic information could be used to estimate hydroperiod at the collection sites
(Johnson et al., 2009). Many amphibian species are most abundant in very small permanent or
intermittent streams because these reaches are too small to support predatory fish (Petranka,
1983; Davic and Welsh, 2004). For example, in a radio-telemetry study of black-bellied
salamanders (Desmognathus quadramaculatus) in one spring-fed stream, Peterman et al. (2008)
estimated the population density to be 11,294 salamanders per hectare (2.47 acres), or 99.30 kg
per hectare of biomass. Some species of salamanders split their lives between forests and
headwaters and depend on a close connection in order to move between the two (Petranka,
1998). Cool, moist soils and large, woody debris in the forested riparian zones of small streams
provide suitable habitat for salamanders (Petranka, 1998).
Although some of these species occur in larger streams downstream from MTM-VF,
Appendix F of the PEIS lists 22 fish, 1 salamander, 1 bird, 38 mussels, 7 snails, and 6 aquatic
invertebrates that are considered threatened, endangered, or species of special concern in the
central Appalachians that are associated with streams (FWS, 2003). Some 31 plants and
21 terrestrial and 50 cave-dwelling invertebrates are also listed.
Loss or burial of headwater streams and associated riparian and subterranean ecosystems
can result in fragmentation of remaining habitats by increasing geographical distance among
populations. Subdivided populations are smaller in size, and thus more susceptible to loss of
genetic diversity and to adverse effects of environmental change, placing them at higher risk of
23
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local extinction (Gilpin and Soule, 1986; Frankham, 2005). The effects of fragmentation may be
increased by the dendritic structure of stream networks (Pagan, 2002). Loss of aquatic
organisms from MTM-VF-impacted streams has been documented (e.g., Pond et al., 2008; Pond,
2010), but, to our knowledge, effects of habitat fragmentation on surviving populations have not
been studied in the central Appalachians. Relevant studies from other areas indicate potential for
detrimental effects on species that disperse frequently among headwater streams—including
salamanders (Grant et al., 2010) and insects (reviewed by Hughes et al., 2009); and species that
disperse longitudinally through stream networks, especially fish (Pagan, 2002; Letcher et al.,
2007).
For species capable of overland dispersal, network configuration and terrestrial land use
in headwater catchments can increase or decrease connectivity or isolation of stream populations.
Alexander et al. (2011) found that tree cover in first-order watersheds was the best predictor of
regional genetic diversity in the common mayfly Ephemerella invaria, which is closely related to
ephemerellid species in central Appalachia (Alexander et al., 2009). Forest clearing increases
the dispersal distance between the two ecosystems and is expected to decrease the abundance of
salamanders in small streams that remain at a site (Petranka et al., 1993; Maggard and Kirk,
1998). Grant et al. (2009) found higher occupancy by salamanders in less-developed catchments
of headwaters connected to other headwaters (i.e., more highly branched headwater networks).
3.4. LOSS OR ALTERATION OF HEADWATER ECOSYSTEM FUNCTIONS
As with the loss of biota, we assume that most ecosystem functions performed by a
high-gradient, forested Appalachian headwater stream are lost when it is buried or removed.
Some functions, such as water conveyance and export of dissolved solids, might continue under
fills in a quantitatively or qualitatively altered state. At the time of this report, no ecosystem
studies of buried streams were found in the published literature. Evidence of stream function in
channels constructed on valley fills is reviewed in Section 7.
Data to quantify regional loss of stream function due to mountaintop removal and valley
fill mining are not available at the time of this report. Here, we briefly review ecologically
important functions that were likely to have been operating in central Appalachian headwater
streams lost to blasting or burial.
3.4.1. Transformation and Removal of Nutrients and Contaminants
Due to the small size of headwater streams, their contributions to ecosystem function at
the watershed scale are often overlooked. However, the individual and cumulative effects of
headwater streams can be substantial (Meyer and Wallace, 2001; Benda et al., 2004; Freeman et
al., 2007; Meyer et al., 2007). Nutrients are taken up and transformed more rapidly in
24
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headwaters, where waters slowed by woody debris have longer contact times with biologically
and chemically reactive benthic substrates and hyporheic zones2 of small, shallow channels
(Alexander et al., 2000; Bernhardt et al., 2005). Peterson et al. (2001) estimated that 50-60% of
the inorganic nitrogen entering a stream is retained or transformed in the headwaters, reducing
downstream nutrient loads by half. This estimate is likely conservative because denitrification, a
process that microbes perform in the substrates and hyporheic zones of natural stream channels
and riparian areas (Payne, 1981), removes N from the stream in the form of N gases and is not
included in the estimate by Peterson et al. (2001). Riparian buffers have a central role in
nitrogen removal, which is affected not only by buffer width and riparian vegetation, but also by
soil type, subsurface hydrology, chemistry, and interstitial microbial communities in the riparian-
hyporheic zone (Pusch et al., 1998; Mayer et al., 2007).
In addition to reducing excess nutrients, natural headwaters can remove metal
contaminants including copper (Cu), zinc (Zn), manganese (Mn), and iron (Fe)
(Schorer and Symader, 1998). In contrast, outflows from filled headwaters typically are net
exporters of toxicants to downstream segments (see Section 4). The loss of natural ecosystem
function and the export of toxicants act in combination to increase risks to water quality below
MTM-VF.
3.4.2. Storage and Export of Woody Debris
In their natural state, forested headwaters typically transport little sediment or woody
debris by fluvial processes and act as sediment reservoirs for periods spanning decades to
centuries (Benda et al., 2005). Substrate and organic debris dams provide habitat and slow the
flow of water through headwaters, creating more contact time for processing organic matter,
nutrients, and toxicants and regulating runoff in normal rain events. Woody debris could be of
particular importance in this, because both phosphorus and ammonium can travel further
downstream before being taken up by benthic organisms when wood is removed from a
headwater stream (Webster et al., 2000).
3.4.3. Organic Matter Processing
Forested headwaters also receive and process large volumes of organic matter from
upland and riparian vegetation (Wipfli et al., 2007). Organic material enters headwater streams
through litter fall from riparian vegetation, surface runoff of particulate and dissolved material
and subsurface movement (Cummins et al., 1989; Wallace et al., 1999). Once introduced,
2Hyporheic zone: the subsurface ecotone below and adjacent to the stream channel, where surface water and ground
water mix and exchange solutes. Much of the streamflow and biogeochemical processing in streams occur
underground. The hyporheic zone also supports a rich variety of aquatic flora and fauna (Boulton et al., 1998).
25
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organic material can be retained in the headwater stream, transformed through feeding of
organisms in the headwater stream, or transported downstream (Webster et al., 1999; Hall et al.,
2000; Wipfli et al., 2007). Macroinvertebrates and detritus from headwater streams supports the
biomass of animals, plants, and fungi found in downstream segments (Wipfli and Gregovich,
2002; Brittain and Eikeland, 1988). The organic matter transported from headwaters to
downstream segments is largely in the form of fine particulate organic matter (FPOM, 0.63 jim
to 1.0 mm diameter) or dissolved (DOM) organic matter (Wipfli et al., 2007). Kaplan et al.
(2008) found that only 8.6% of dissolved organic carbon (DOC) was labile and was taken up
within a stream reach, while the remaining bioavailable DOC was semi-labile and is usually
transported out of the reach to larger streams. The loss of trophic subsidies from headwater
streams may lead to lower secondary productivity in downstream habitats.
3.4.4. Habitat
Headwaters and associated interstitial habitats provide refugia for macroinvertebrates
during floods or spates and speed the recovery of aquatic communities when flow conditions
improve (Angradi, 1997; Angradi et al., 2001). These areas could facilitate a 'rescue effect'
where there is the potential for recolonization from undisturbed sites, and the presence of this
source of colonists can be a strong determinant of population resilience (Brown and
Kodric-Brown, 1977). Headwaters also serve as nurseries and spawning grounds for amphibians
and fish, including the brook trout (Salveliniisfontinalis), creek chub (Semotilus atromaculatus),
blackside dace (Phoxinus cumberlandensis), southern redbelly dace (Phoxinus erythrogaster\
arrow darter (Etheostoma sagitta), and orangethroat darter (Etheostoma spectabile [Meyer et al.,
2007]). Today, most brook trout populations in West Virginia occur outside of the coal mining
region, but there is overlap in their historic range. In a study of one West Virginia watershed,
Petty et al. (2005) estimated that >80% of all brook trout spawning occurred in small streams
(watersheds <3 km2), including headwaters draining areas less than 0.25 km2.
26
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4. IMPACTS ON WATER QUALITY
In this section, we report the results of a number of studies that have assessed the changes
in the physicochemical attributes of streams downstream of MTM-VF. Although much of this
information might also apply to constructed channels and other water-containing structures on
the valley fills and the mined site, there are few data in the PEIS or the peer-reviewed literature
on these constructed systems. The physicochemical attributes we review below include
alteration of streamflow, water chemistry, sedimentation of stream substrates, and sediment
chemistry. Alterations of these attributes are the potential causes of the effects observed
downstream of MTM-VF, which are described in Sections 5 and 6 of the report.
4.1. ALTERATION OF STREAMFLOW
Four factors can affect streamflow below valley fills. First, trees and other vegetation are
removed from both the mined area and the area of the valley fill, the reclaimed surface might be
planted with grasses, and, if planted, trees are generally slow to regrow on the reclaimed mined
area and valley fill (see also Section 7). This reduces evapotranspiration rates from the
watershed because transpiration is a function of the type and abundance of active vegetation
(Dickens et al., 1989; Messinger, 2003). Second, the valley fill forms an unconsolidated aquifer
in the watershed that stores a portion of any water that infiltrates into it (Dickens et al., 1989;
Wunsch et al., 1999). This water comes from recharge along the periphery of the spoil body
where surface-water drainage might be caught, from groundwater intercepted from adjacent
bedrock aquifers, or from precipitation falling on the fill. Third, compaction of the fill surface
by heavy equipment can reduce infiltration of precipitation and increase overland runoff
(Negley and Eshleman, 2006). Fourth, when a headwater stream is lost (see Section 3),
attributes that influence surface flow (e.g., woody debris, surface water/ground water
connections) are also lost (see Figure 12). Valley fills can act like a headwater aquifer and
provide a more constant source of flow during the dry parts of the year. Comparing adjacent
mined and unmined watersheds, monthly mean flow in the mined watershed was greater than
that in the unmined watersheds during summer, autumn and early winter (July to January), when
soil and aquifer moisture levels were reduced (see Figure 5, Messinger and Paybins, 2003).
Wiley et al. (2001) found the 90% duration flows3 at sites below valley fills were 6 to 7 times
greater than the 90% duration flows found at unmined sites. Moreover, daily streamflows from
sites below valley fills were generally greater than those in unmined watersheds during periods
3The 90% duration flow is the streamflow (nrVsec) equaled or exceeded at a site 90% of the time, a measure of the
baseflow.
27
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to
oo
> heavy equif
operatic
)ment road construction
ns & maintenance
J \^ ) .<
ove
r~r\
\
/
^
/ \/
X /_ X
[access roads ] f bare compacted 1 /-
x, / 1 inik J I
\
N / x
/ V
1s surface runoff
V
1s peak stormf low
\ / \
f
| overburden
1 handling
rburden& L
al removal
j \
\
bias
/
seeSEDIMENT& NUTRIENTS
diagram
\
/ f valle
V x
° 1 / \
/ compaction of
U/alleyfill surface J ^
\/
^ N
A subsurface .
^ hydrological structure J ^^ Airmit
\
t
/
yfill
/
ration
/
r ^i
upland forest
clearing
V J
)
\
/
•^ evapotranspiration
\
basef low from toe
ofvalleyfill
/
^ A water
' temperature
Figure 12. Observed and expected effects of MTM-VF on streamflow characteristics. See Section 4.1 for
additional details and evidence.
-------
of low streamflow (Wiley et al., 2001). Both Green et al. (2000) and Armstead et al. (2004)
observed that the streams below valley fills continued to have surface flows during the summer
and fall of a year when a drought occurred, but several of their unmined sites did not have
surface flows.
Storm intensity changes the relative effect of the valley fill on downstream flows. Intense
storms can produce greater stormflows in watersheds with MTM-VF compared to unmined
watersheds, but stormflows associated with precipitation from lower intensity storms might be
ameliorated by valley fills. Messinger and Paybins (2003) found that a mined watershed had
greater peak flows during severe storms than an unmined watershed. Unit peak flow4 was greater
in the mined watershed following summer thunderstorms when rainfall intensity exceeded
2.5 cm/hour (Messinger, 2003). In contrast, unit peak flow was lower in the mined watershed
following low-intensity, long-duration rainfall events—particularly in the winter.
Wiley and Brogan (2003) found that peak discharges after an intense storm were greater
downstream of valley fills than in unmined watersheds. Peak discharges were estimated by
applying the slope-area method5 (Benson and Dalrymple, 1967) to measurements of high water
marks observed after flooding associated with a 7.6 to 15 cm rainfall in southeastern West
Virginia over a 5- to 6-hour period. Six sites were studied: three below valley fills and three in
unmined watersheds. At two of the three sites downstream of valley fills, the estimated peak
discharges were equivalent to floods that would naturally occur only once every 50 to
>100 years. Peak discharges at the sites in unmined watersheds had less severe estimated flood
recurrence intervals of 10 to 25 years (Wiley and Brogan, 2003). Peak discharges at the third
site downstream of a single, large valley fill without active mining had an estimated flood
recurrence interval of <2 years, much less severe than the other two mined sites. The differences
might be due to unaccounted for differences in rainfall among the watersheds or differences in
mine and valley fill attributes. Thunderstorms can cause locally variable rainfall, particularly in
mountainous terrains (Barros and Lettenmaier, 1994; Roe, 2005).
Some current state regulations generally require that MTM-VFs be designed to reduce
such increases in downstream flooding (WVDEP, 2009b). Recent research by Taylor et al.
(2009a, b) suggests that alternate spoil placement methods that reduce compaction of the mining
spoil might increase infiltration of rainfall and ameliorate the increases in peak discharges.
4Unit peak flow is discharge per unit area of watershed, m3/sec/km2.
5With the slope-area method, the maximum flood height is estimated from the physical evidence left by the flooding,
the high water marks. Then the cross-sectional area and wetted perimeter (i.e., the length of the part of the perimeter
of the channel cross-section [stream bed and banks] below the water surface) of the stream channel are measured at
that flood height. The slope of the stream bed is also measured, and Manning's n, an index of the roughness of the
stream bed, is estimated. The peak discharge is then calculated using these variables.
29
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4.2. CHANGES IN CHEMICAL TRANSPORT
4.2.1. pH, Matrix Ions and Metals
Almost invariably, coal mining exposes pyrite, a ferric sulfide mineral formed in
association with coal (Caruccio et al., 1977; Altschuler et al., 1983; Casagrande, 1987; Younger,
2004). In the presence of water and oxygen (62), pyrite is oxidized in a reaction catalyzed by
autotrophic bacteria to form the strong acids characteristic of acid mine drainage
(Stumm and Morgan, 1996):
(2+) - (3+) 2
+
Fe+S2 + 3.75 O2 + 3.5 H20 -> Fe+(OH)3 + 2 SO4 + 4 H+ (Eq. 1)
However, in the presence of sufficient carbonate minerals, such as calcite (CaCO3) and
dolomite [CaMg(CO3)], the acidity can be neutralized (Rose and Cravotta, 1998):
2 CaCO3 + 2 H+ -> 2 Ca2+ + 2 HCO3 (Eq. 2)
2 CaMgCO3 + 2 H+ -> Ca2+ + Mg2+ + 2 HCO3 (Eq. 3)
The effluent waters from valley fills are generally not acidic and can be somewhat
alkaline (Bryant et al., 2002; Merricks et al., 2007). The pH is generally 7.0 or greater
(Bryant et al., 2002; see Tables 2-5). The alkaline pH has been attributed to exposure of the
water to carbonate minerals within the valley fill that originate from fragmentation of the
noncoal formations that form the overburden or are added during construction of the valley fill
(Sobek et al., 1978; Banks et al., 1997; Skousen et al., 1997). Other methods that can moderate
pH include physically isolating the pyritic materials within the mine or valley fill (Skousen et al.,
2000; Hawkins, 2004) and treatment within the sediment retention pond (Hartman et al., 2005;
see Table 4 and Figure 13).
Iron forms relatively insoluble compounds, such as Fe(OH)3, under more alkaline
conditions and might not be found in elevated concentrations in the effluent waters below valley
fills (Bryant et al., 2002; see Table 2). However in some conditions, such as during higher flows,
Fe can remain elevated (Hartman et al., 2005; see Table 4).
Most other metals, such as cadmium (Cd), chromium (Cr), Cu, lead (Pb), and Zn,
coprecipitate with or sorb to the iron compounds (Kimball et al., 1995; Lee et al., 2002;
Larsen and Mann, 2005) and were not found (in one study) at elevated concentrations in the
effluent waters (Bryant et al., 2002; see Table 2). Exceptions to this are Mn and nickel (Ni),
which can be elevated in the effluent waters below valley fills (Bryant et al., 2002; Hartman et
al., 2005; see Tables 2 and 4). Mn can occur in association with siderite (FeCO3) in shales
30
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Table 2. Water quality variables in unmined streams versus streams below valley fills. Variables are ordered by
ratio of their median concentration in unmined streams to those with valley fills (Filled). Units are mg/L, unless
indicated otherwise.
Variable
SO42~*
Ca, total*
Mg, total*
Hardness*
TDS*
Mn, total
Conductivity (uS/cm)*
HCO3 *
Se, total*
Alkalinity
K, total*
Na, total*
Mn, dissolved
Cl*
Acidity
Ni, total
NO37NO2~*
pH (standard)*
Acidity, hot
Al, dissolved
Sb, total
As, total
Be, total
Unmined
Median
12.6
4.88
4.10
29.1
50.5
0.005
66.4
—
0.0015
20.0
1.58
1.43
O.005
<2.5
2.5
—
0.81
6.8
<2.5
O.050
0.0025
0.001
0.0005
Mean
16.0
7.50
4.30
42.0
—
0.034
62.0
25.5
0.0015
—
1.60
2.40
0.021
2.8
—
O.010
0.40
7.1
—
0.093
—
—
—
Range
11.0-21.6
2.70-12.0
2.30-7.00
17.0-72.0
—
0.005-0.083
34.0-133
7.44^2.7
0.0015
—
1.30-2.00
0.70-5.50
O.005-0.055
<2.5-4.0
—
O.010
0. 10-0.90
6.1-8.3
—
O.050-0.19
—
—
—
Filled
Median
523
104
86.7
617
847
0.044
585
—
0.012
150
8.07
4.46
0.044
4.5
4.2
—
0.95
7.8
<2.5
O.050
0.0025
0.001
0.0005
Mean
696
138
122
801
—
0.14
1,020
223
0.011
—
9.90
12.6
0.11
4.6
—
0.014
3.4
7.9
—
0.096
—
—
—
Range
155-1,520
38.0-269
28.0-248
225-1,620
—
0.009-9.0
159-2,540
13.1-612
0.0015-0.037
—
3.00-19.0
2.60-39.0
0.0065-0.85
<2.5-ll
—
O.010-0.059
0.80-17
6.3-8.9
—
O.050-0.27
—
—
—
Detection
limit
5.0
0.10
0.50
3.3
5.0
0.010
—
NA
0.003
5.0
0.75
0.50
0.01
5.0
2.0
0.02
0.10
—
5.0
0.10
0.005
0.002
0.001
-------
Table 2. Water quality variables in unmined streams versus streams below valley fills (continued)
Variable
Cd, total
Cr, total
Co, total
Cu, total
Pb, total
Hg, total
Total organic carbon
P, total
Au, total
Th, total
V, total
Ba, total
Dissolved oxygen
Dissolved organic carbon
Total suspended solids
Fe, total
Fe, dissolved
Zn, total
Al, total
Unmined
Median
0.0005
0.0025
O.0025
0.0025
O.001
0.0001
1.4
0.10
O.005
0.001
O.005
0.029
13.6
2.45
5.75
0.42
0.22
0.0060
0.15
Mean
—
—
—
0.0029
0.0012
—
—
0.10
—
—
—
0.040
—
—
—
0.18
0.074
0.010
Range
—
—
—
0.0025-0.005
O.OO 10-0.0021
—
—
0.10
—
—
—
0.015-0.072
—
—
—
0.065-0.47
O.050-0.19
0.0033-0.023
Filled
Median
0.0005
0.0025
O.0025
0.0025
O.001
0.0001
1.4
0.10
O.005
0.001
O.005
0.025
11.0
1.95
4.25
0.19
0.10
0.0025
O.10
Mean
—
—
—
0.0026
0.0012
—
—
0.10
—
—
—
0.041
—
—
—
0.28
0.092
0.0091
Range
—
—
—
0.0025-0.0034
O.OO 10-0.0040
—
—
0.10
—
—
—
0.022-0.068
—
—
—
0.066-0.65
O.050-0.28
0.0025-0.027
Detection
limit
0.001
0.005
0.005
0.005
0.002
0.0002
1.0
0.10
0.01
0.002
0.01
0.020
—
1.00
5.00
0.10
0.10
0.005
0.10
to
Sources: Bryant et al. (2002) and Pond et al. (2008). An asterisk (*) next to the variable name indicates that the mean concentration in streams below valley fills
was statistically significantly greater than that in unmined streams atp = 0.05. Median concentrations (Bryant et al., 2002) are from 9 unmined sites and 21 filled
sites, each sampled about six times from August 2000 to February 2001. Means and ranges (Pond et al. 2008) are from sites having biological data; 7 unmined
sites and 13 filled sites, except for pH and conductivity, which were measured at 10 unmined sites and 27 filled sites. In Pond et al. (2008), HCO3
concentrations were reported as CaCO3 (Personal Communication from M. A. Passmore, U.S. EPA Region III, Wheeling, WV, 2009), and were converted to
HCO3 by multiplying by 1.22. Concentrations below detection are shown as
-------
Table 3. Water quality parameters for unmined or reference streams or streams downstream from mined,
filled, or filled and residential watersheds in West Virginia
Variable
Conductivity
(uS/cm)
pH (standard)
Dissolved O2
(mg/L)
Hardness
(mg/L)
Green et al. (2000)
Unmined
58-140
59 (38-178)
7.1-7.5
7.5 (5.7-9.4)
6.5-13.3
10.9 (5.6-15.2)
—
Filled
643-1,232
850 (159-2,500)
7.1-7.9
7.7 (5.9-8.5)
7.5-13.0
10.0 (5.8-14.5)
—
Filled/
Residential
538-1,124
843 (155-1,532)
7.1-8.3
8.0 (6.4-8.7)
8.5-14.0
9.4(7.3-16.1)
—
Mined
172-385
187 (90-618)
6.7-8.4
7.4 (6.0-8.7)
8.7-12.7
10.2 (7.4-14.5)
—
Merricks et al. (2007)
Reference
247 ± 87
7.2 ±0.36
—
86 ±20
Filled
923 ± 380-
2,720 ± 929
7.93 ±0.18-
8.37 ±0.47
—
544 ± 226-
1,904 ± 596
Hartman et al. (2005)
Reference
47.6 ± 2.4-
259.7 ±30.6
6.5 ±0.6-
7.0 ±0.4
8.5 ±0.8-
13.4 ±0.4
—
Filled
502.0 ± 98.4-
1,479.0 ±110.6
7.2 ± 0.6-
7.5 ±1.0
9.1 ±1.0-
13.0 ±0.6
—
Sources: Green et al. (2000) (range of means among seasons, overall mean, overall range), Merricks et al. (2007) (range of means and standard deviations) and
Hartman et al. (2005) (range of means and standard deviations).
-------
Table 4. Alkalinity, pH, and metals in control streams and streams
downstream from filled watersheds in West Virginia
Parameter
Alkalinity*
pH (standard)
Na*
K*
Mg*
Ca*
Cu*
Ni*
Mn*
Fe*
Zn
Al
Reference
Mean
12.8
7.2
2.9
3.3
23
37
0.00080
0.0076
0.019
0.016
0.0027
0.012
Range
0.400-46.8
6.7-7.7
0.80-3.1
1.5-5.1
2.2-52
2.6-67
0.00020-0.0019
<0.00030-0.018
0.0016-0.046
0.0014-0.030
0.0014-0.0047
0.0090-0.019
Filled
Mean
163
7.7
10
10
86
130
0.0012
0.025
0.062
0.047
0.0028
0.019
Range
16.2-319
6.9-8.2
3.9-22
1.8-14
4.9-130
5.9-200
0.00050-0.0018
<0.00030-0.051
0.0020-0.17
<0.00050-0.082
0.00090-0.0086
0.00090-0.064
Units are mg/L unless indicated otherwise. If the concentration was less than the detection limit, the value is shown
as < the detection limit. An asterisk (*) marks those measures where the fill streams were statistically significantly
greater (p < 0.05) than the reference streams.
Source: Hartman et al. (2005).
Table 5. Range of dissolved oxygen, pH, and conductivity values for sites in
eastern Kentucky
Parameter
Dissolved oxygen (mg/L)
pH (standard)
Conductivity ((^S/cm)
Reference (« = 4)
9.1-9.6
7.1-7.4
30-66
Filled (« = 8)
8.4-9.7
7.2-8.2
420-1,690
Source: Howard et al. (2001).
34
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overburden handling
V V
[ valley fill J ( backstacks J
f compaction of ]
U/alley rill surface J x>
V
^ watercontact with
overburden
<
V V V
'T- 1-ntal rlk<;nl\/prl snlirk -T" nH T* metals
(water)
•fMg2+ -fCa2+ -fK+ -tCI- -fHCCV -f Ni T"
•f hardness T" Na+ -fSO42-
LEGEND
human activity modifyingfactor
[ source J f treatment or |
> / ^ reclamation )
additional stepin i 1
causal pathway proximate
*> ' 1 stressor
z [sediment retention &|
L treatmentponds J
Y
<" "> ^ metals
(sediment)
Se -t Fe -fMn
Figure 13. Observed and expected effects of MTM-VF on total dissolved
solids, metals, and pH. Water quality variables shown were significantly
different from reference sites in at least one of the reviewed studies. See
Section 4.2 for additional details and evidence.
within the overburden and is more soluble in the more alkaline waters (Larsen and Mann, 2005).
Aluminum (Al) is found primarily associated with clay minerals in soils and is not soluble unless
the pH is less than 4.9 (Nordstrom and Ball, 1986).
Sulfate (SC>42 ), calcium (Ca2+ from calcite-type minerals), magnesium (Mg2+ from
dolomite-type minerals), and HCOs (from both calcite and dolomite), which are formed in the
above reactions (see Eq. 1-3), are commonly present at elevated concentrations in the effluent
waters and dominate the mixture of ions in these waters in relative concentrations (Bryant et al.,
2002; Hartman et al., 2005; see Tables 2 and 4). In addition, other water-soluble compounds
within coal or overburden can be solubilized by the above reactions or just by the increased
exposure to water in the fragmented overburden (Yudovich and Ketris, 2006a; Vesper et al.,
2008). These ions, including K+, Na+, and Cl~, and Se, occur at elevated concentrations in the
effluent waters (see Tables 2 and 4), but at concentrations at least one order of magnitude less
than those observed for SC>42 , Ca2+ , Mg2+, and HCOs .
35
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All these ions are components of the elevated specific conductivity, a measure of the
stream's ability to conduct an electrical current, which reflects the concentration of dissolved
ions in the water (measured in units of microSiemens per cm, |iS/cm), and TDS observed in
these waters (Green et al., 2000; Howard et al., 2001; Bryant et al., 2002; Bodkin et al., 2007;
Merricks et al., 2007; Pond et al., 2008; see Tables 2-5). In these studies, measured conductivity
at sites downstream from valley fills ranged from 159-2,720 |iS/cm and at unmined reference
sites from 30-260 |iS/cm. Hardness is another measure of these dissolved ions—particularly the
divalent ones like Ca2+ and Mg2+. These aggregate measures of total ions are coarser than the
individual ion concentrations but are relatively simple to measure. TDS theoretically would be
simply the sum of the dissolved ion concentrations (mg/L), but, in practice, it also includes some
particulates that pass through the filter used to separate dissolved solids from suspended solids
(APHAetal., 1998).
In terms of individual dissolved ion concentrations, conductivity is a product of the molar
concentration of each ion (mmol/L), the absolute value of its charge (meq/mmol), and an
ion-specific, equivalent conductance (X°) that is 80.0, 59.5, 53.1 and 44.5 |iS/cm2/meq for the
dominant ions, SO42 , Ca2+, Mg2+, and HCO3 , respectively, at 25.0°C (APHA et al., 1998;
Talbot et al., 1990; Pawlowicz, 2008). The equivalent conductances for the less elevated ions,
K+, Na+, and CF, are 73.5, 50.5,76.4 |iS/cm2/meq, respectively, at 25.0°C (APHA et al., 1998).
For example
Conductivity [SO42~] = Molar Cone. (mmol/L) x 2 meq/mmol x 80 uS/cm2/meq. (Eq. 4)
Considering the greater concentrations of the four dominant ions, SC>42 , Ca2+, Mg2+, and
HCOs , and the greater charge of three of these ions, it follows that these ions also dominate the
ion mixture in their contribution to the conductivity.
Most studies have not assessed the seasonal variability of water chemistry at these sites,
but Green et al. (2000) present data for five consecutive seasons from 1999 to 2000. There
appears to be little seasonal pattern to pH, but mean conductivities were greatest in all four
watershed types during the summer sampling period, possibly because of seasonally reduced
discharges (see Table 6). In particular, mean conductivity exceeded 1,000 uS/cm in streams in
filled and filled/residential watersheds during the summer sampling period. In all seasons,
conductivities at sites in filled and filled/residential watersheds were an order of magnitude
(10 times) greater than at reference sites in unmined watersheds (see Table 6). Pond et al. (2008)
observed conductivities up to 2,540 uS/cm in streams from mined watersheds.
36
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Table 6. Seasonal mean (standard deviation) of conductivity (uS/cm) for the
four classes of streams
Season
Spring 1999
Summer 1999
Autumn 1999
Winter 2000
Spring 2000
Unmined
64(19)w = 9
140 (54) n = 2
91 (59)» = 2
73 (29) n = 9
58(28)« = 10
Filled
946 (614) n= 15
1,232 (643) n= 15
958 (430) n= 14
836 (425) n= 14
643 (382) n= 15
Filled/residential
652 (237) n = 6
1,124 (282) n = 6
984 (221) n = 6
844 (173) n = 6
438 (249) n = 6
Mined
172 (90) n = 4
3 85 (202) n = 3
260 n = 1
254 (171) n = 3
192 (155) n = 5
The number of sites («) analyzed is also given.
Source: Green et al. (2000).
Moreover, only one study has assessed the downstream extent of elevated ion
concentrations related to these sites but only within a single stream drainage. Johnson et al.
(2010) measured conductivity at intervals of 100 to 500 m along the main stem of Buckthorn
Creek (Breathitt, Perry, and Knott Counties, Kentucky) downstream from a valley fill and along
several mined and unmined tributaries. Maximum conductivities along Buckthorn Creek were
3,190 |iS/cm in spring (May 2006) and 11,810 |iS/cm in summer (September 2005), and
measured conductivities gradually decreased in a downstream direction—except downstream
from the mined tributaries. Downstream of mined tributaries, conductivities increased, while
downstream of unmined tributaries, conductivities decreased. The increase or decrease was
related to the watershed size of the tributary, which was used as a surrogate measure of the
relative discharge of the tributary. This suggests that most of the decrease in conductivities was
the result of dilution by low conductivity, unimpacted waters. Downstream from the first valley
fill along the mainstem of Buckthorn Creek (i.e., a distance of 20 km), conductivities never
decreased below 400 |iS/cm in spring and 2,000 |iS/cm in summer.
Kirk and Maggard (2004) sampled a third-order stream, Trough Creek, at two stations:
one was upstream of the stream's confluence with two smaller tributaries where mountaintop
mines and valley fills were developed, and the second was downstream of both confluences.
Sampling was conducted in spring (i.e., April) and autumn (i.e., October) from October 1995
(before mining began in February 1996) until April 2003. The downstream sites exhibited
variation—particularly for SC>42~, TDS, and conductivity—between the two seasons that appears
to be related to dilution associated with seasonal variation in discharge of Trough Creek
(Kirk and Maggard, 2004). Conductivity in spring increased from 64 |iS/cm in 1996 to over
37
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400 |iS/cm in 2002 and 2003, while in autumn, conductivity increased from 242 |iS/cm in 1996
to over 1,600 |iS/cm in 2002 and 2003 (Kirk and Maggard, 2004).
The relative ion composition of the effluent is expected to be fairly consistent among
valley fills in the region. Although the names of coals vary by state, mountaintop mines in the
central Appalachians primarily mine a similar series of coals that occur above the level of the
permanent draining streams. Specific coals include the No. 6 Block (called Richardson and
Skyline in Kentucky), Stockton (called Broas in Kentucky), Coalburg (called Peach Orchard in
Kentucky), and in some cases the Winifrede (called Hazard and Haddix in Kentucky), Chilton
(called Cropland and Taylor in Kentucky), and Fire Clay (Neuzil, 2001; Tewalt et al., 2001).
These coals occur in the lower Allegheny (called Charleston in West Virginia and upper
Breathitt in Kentucky) and Pottsville (called middle to upper Kanawha in West Virginia and
middle to lower Breathitt in Kentucky) formations, both of which formed during the middle
Pennsylvanian period. The intervening noncoal formations, which form the overburden removed
by mountaintop mining and placed into the valley fills, are siltstones, shales, and sandstones with
a few limestones, mostly of marine origin. Because of the mixing of overburden formations
within the valley fill and the geochemical reactions that create the ion mixture, the relative ion
composition of the effluent is unlikely to differ significantly among valley fills in the region.
However, the maximum conductivities might differ substantially (Green et al., 2000; Howard et
al., 2001; Bryant et al., 2002; Bodkin et al., 2007; Merricks et al., 2007; Pond et al., 2008), in
part as a function of the amount of dilution by unaffected, lower conductivity waters (Johnson et
al., 2010).
Se is enriched in coal relative to other rocks (Coleman et al., 1993). In coals mined by
MTM, average Se concentrations range from 3.9 to 7.1 |ig/kg in West Virginia (Neuzil et al.,
2005) and 3.8 to 6.6 |ig/kg in Kentucky (Eble and Hower, 1997). It appears to be mainly
associated with the organic fraction of the coal, where it substitutes for organic sulfur. It can
also be associated with pyrite or with other accessory minerals, clausthalite (lead selenide)
(Coleman et al., 1993; Finkelman, 1994; Hower and Robertson, 2003, Yudovich and Ketris,
2006b). Reflecting these different modes of occurrence, the correlations between Se and the
organic content of the coal as measured by loss on ignition or total sulfur (e.g., reflecting the
pyrite content of the coal) are variable among regions (Coleman et al., 1993), but Neuzel et al.
(2007) found significant correlations between Se and the organic content of coal-bearing strata
(i.e., noncoal formations adjacent to coal measures that form the overburden) from the central
Appalachians, but not between Se and total sulfur. Neuzel et al. (2007) did not find a correlation
between the Se concentration in these rocks and the Se concentration in leachates.
38
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4.3. OTHER WATER QUALITY VARIABLES
Other water quality variables investigated include water temperature, nutrients,
particularly nitrogen and phosphorus, and dissolved oxygen (see Figure 14).
overburdens
coal removal
f blasting
overburden
handling
\
/
fvallevfill
"I
road construction
& maintenance
upland forest
clearing
riparian forest
clearing
1 1 1
\
/ \
1
see FLOW ALTERATION
diagram
fsediment retention &
I treatment ponds
see FLOW ALTERATION
diagram
LEGEND
human activity
Figure 14. Observed and expected effects of MTM-VF on sediments,
nutrients, and temperature. See Sections 4.3 and 4.4 for additional details and
evidence.
4.3.1. Water Temperature
Valley fills reduce the annual variation in water temperature. Comparing mean daily
water temperatures between an unnamed tributary of Ballard Fork near Mud, West Virginia; a
stream downstream from a valley fill; and a reference site, Spring Branch near Mud, West
Virginia; Wiley et al. (2001) found that mean stream temperatures were warmer downstream of
the valley fill during the autumn, winter, and spring, with the greatest difference being in
February. In the summer, the mean stream temperatures downstream from the valley fill were
cooler than those in the reference site. Moreover, the range of variation both annually and within
different seasons was less downstream from the valley fill. The minimum and maximum
temperatures downstream of the valley fill were 3.3°C and 16.5°C, respectively, while those in
the reference stream were below 0°C and 20.0°C.
39
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4.3.2. Nutrients
Bryant et al. (2002) found generally low median concentrations of nitrate (NO3") plus
nitrogen dioxide (NO2 ) in streams from unmined watersheds and below valley fills, with some
samples having concentrations less than the detection limit of 0.10 mg/L. However, the mean
concentration of NOs plus NO2 was slightly greater in the streams below valley fills (Pond et al.,
2008), and a maximum concentration of 17 mg/L was observed. Bryant et al. (2002) speculated
that this could be caused by use of nitrogen-containing explosives at these sites or by spreading
nitrogen-containing fertilizers during reclamation. Phosphorus (P) was not detected in any
samples with a detection limit of 0.10 mg/L (Pond et al., 2008).
4.3.3. Dissolved Oxygen
In the studies that have measured dissolved oxygen, concentrations in unmined streams
and streams in either mined and valley fill streams have been reasonably high and similar among
the different types of watersheds (see Tables 3 and 5; Green et al., 2000; Howard et al., 2001;
Bryant et al., 2002; Hartman et al., 2005). Published concentrations range from 6.5 to
13.0 mg/L. However, no studies have looked at diurnal variation of dissolved oxygen in these
streams.
4.4. CHANGES IN SEDIMENTATION
Sediment retention ponds built downstream of valley fills are intended to capture sand
and finer-sized particles that are produced by the fragmentation of the overburden and washed
downstream from the toe of the valley fill (U.S. EPA, 1979; see Figure 14). Despite this,
Wiley et al. (2001), using a modified Wolman (1954) pebble count for the bankfull channel,6
found that the percentage of particles less than 2 mm (i.e., sand and fines) was elevated in stream
reaches downstream from valley fills and any sediment retention ponds (i.e., median = 60%,
interquartile range = 56-65%) when compared to unmined streams (i.e., median = 24%,
interquartile range = 15-34%).
Similarly, Green et al. (2000), using methods from EPA's Environmental Monitoring and
Assessment Program for the wetted channel7 (Kaufmann and Robison, 1998), found that mean
substrate sizes were smaller in filled or filled/residential streams downstream from sediment
retention ponds compared to unmined streams, and the mean percentage of sand and fines was
6The bankfull channel is the entire channel, which is submerged at bankfull discharge—the point just before the
streamflow begins to spread out onto the stream's flood plain at high flows. As a result, this approach measures
some substrate that is dry during baseflow, which is when these channel characteristics are usually measured.
7The wetted channel is the portion of the channel that was submerged at the time these channel characteristics were
measured.
40
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greater. However, mean substrate sizes were largest at sites described as being downstream of
other types of mining without valley fills (i.e., generally older contour mines) (see Table 7).
Table 7. Substrate measures in streams located in different land use classes
Substrate measure: mean
(standard deviation)
Mean substrate size class (unitless)
Calculated mean substrate size
(diameter in mm)
% <2 mm diameter (sand and fines)
Unmined
(» = 9)
3.7(0.3)
53
16.9(9.9)
Filled
(n = 15)
3.5(0.5)
38
20.7(12.9)
Filled/
residential
(» = 6)
3.6(0.8)
42
29.7(24.1)
Other
mined
(» = 4)
4.0(0.3)
109
8.0 (9.2)
Source: Green et al. (2000).
Hartman et al. (2005) did not find any clear pattern of fine sediments in a study that
compared pairs of mined and unmined sites using samples taken in December with a scoop
sample separated with modified Wentworth sieves (McMahon et al., 1996; see Table 8). In two
cases, the proportions of sand and fines were similar; in the third case, it was greater in the filled
site; and in the fourth case, it was greater in the reference site. However, there appears to have
been a significant nonmining disturbance in this last control site, Big Buck Fork. In this study,
the filled sites were upstream from the sediment retention ponds (Hartman et al., 2004).
Table 8. Proportion of sediments that were sand and fines (mean [standard
error]) in paired sites
Site names (reference/impaired)
W. Br. Atkins Creek/E. Br. Atkins Creek
Big Buck Fork/Hill Fork
Bend Branch/Rockhouse Creek
N. Br. Sugar Tree Creek/S. Br. Sugar Tree Creek
Reference
0.35 (0.00)
0.78 (0.03)
0.25 (0.07)
0.27 (0.02)
Filled
0.46(0.10)
0.50 (0.06)
0.23 (0.02)
0.50 (0.04)
Source: Hartman et al. (2005).
Much of the fine sediment, though, might come from the streambanks rather than the
mined area or the valley fill. Using stable isotopic signatures of carbon (i.e., 513C) and nitrogen
41
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(i.e., 515N), Fox (2009) attempted to identify the sources of fine sediments in streams in four
forested catchments. Of the catchments, one had an active mountaintop mine, one had an
inactive mountaintop mine, one had some pre-SMCRA surface mining, and one had no surface
mining. In the mined watersheds, there were two potential sources of fine sediments:
streambank erosion and surface erosion from the mined area and valley fill. Fox (2009)
concluded that about 50% of the fine sediments in the inactive mined watershed and 40% in the
actively mined watershed were from streambank erosion. The streambank erosion could occur
because of the alteration in stream baseflows and peak flows caused by mining and creation of
the valley fills (see Section 4.1).
4.5. CHANGES IN SEDIMENT CHEMISTRY
Data on sediment chemistry in larger streams downstream of valley fills are limited to a
study by Merricks et al. (2007), who measured metals and arsenic. They sampled three to six
stations at 100- to 150-m intervals in each of three streams downstream from sedimentation
ponds below valley fills in West Virginia and a single reference site (see Table 9).
Table 9. Range of sediment concentrations of metals and arsenic (mg/kg) in
streams downstream from the sedimentation ponds below valley fills in 2002
and 2004 and from a reference site in 2002
Metal or arsenic
Al
As
Cd
Cu
Fe
Hg
Mn
Se
Zn
Reference— 2002
(« = 1)
11
—
—
0.018
51
—
1.4
—
—
Downstream from
valley fill— 2002
(« = 11)
9-20
—
—
0.012-0.122
49-158
—
1.6-17
—
—
Downstream from
valley fill— 2004
(» = 18)
2-28
0.015-0.070
0.005-0.015
—
10-151
0.006-0.015
1.0-41
0.001-0.011
0.1-2.5
The reference site was only sampled in 2002, and the analytes measured differed between the 2 years. The
unmeasured analytes are indicated by —.
Hg = mercury.
Source: Merricks et al. (2007).
42
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Sediment concentrations of metals and arsenic were generally greater at one stream,
Lavender Fork, which was downstream from a reclaimed, 6-year old valley fill and that also had
the greatest measured stream water conductivities. Sediment concentrations also generally
decreased with increasing distance below the sedimentation ponds.
4.6. CUMULATIVE IMPACTS
In terms of downstream water chemistry, the primary cumulative impact of MTM-VF
and other mining methods for coal in the region affected by MTM-VF has been elevated
concentrations of SC>42 and conductivity. In larger streams of the Kanawha basin, Paybins et al.
(2000) found that one-fourth of all water samples exceeded a SC>42 concentration of 250 mg/L
and 70% of the water samples collected downstream of coal mines exceeded a regional
background concentration of 21 mg/L that was calculated from data for basins with no history of
r\
coal mining. Moreover, the median concentration of SCM had increased by 1.6 times in these
streams between 1980 and 1998, and conductivity had increased by 1.2 times (Paybins et al.,
2000). SC>42 and some of the other ions contributing to conductivity are conservative ions in
water, meaning that there are no chemical or biological processes that alter these ion
concentrations in the waters. Any changes in SC>42 concentrations are the outcome of mixing of
waters with differing SC>42 concentrations (Cooper et al., 2000). Therefore, the increased SC>42
r\
and conductivity are associated with increased sources of water with elevated SC>4 and
conductivity within the Kanawha basin. MTM-VF appears to be these sources because other
r\
land disturbances, such as residential development, are not origins of elevated SO4 and
conductivity.
Conversely, while total Fe, total Mn, and total Al in many larger streams within mined
basins exceeded regional background concentrations of 129, 81, and 23 ug/L, respectively, the
median concentrations of total Fe and total Mn had decreased between 1980 and 1998 by
approximately one-third and one-half, respectively, and pH had increased (Paybins et al., 2000).
As discussed previously, these metals are not as soluble under more alkaline conditions, and their
decrease might reflect the increase in pH associated with the increased number of valley fills,
which are the sources of alkaline waters within the Kanawha basin.
In the absence of other direct evidence on the cumulative effects of the changes in water
chemistry associated with MTM-VF on downstream water quality, it should be noted that
headwater streams, such as those affected by MTM-VF, have a large influence on downstream
water quality. Alexander et al. (2007) found that first-order, headwater streams contributed
70% of the mean annual water volume in second-order streams and 55% of the volume in higher-
order rivers. For nitrogen, a nutrient that is not as conservative as the ions associated with
43
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MTM-VF, these first-order streams contributed 65% of the flux in second-order streams and
40% of the flux in higher-order rivers (Alexander et al., 2007).
Johnson et al. (2010), working in a single stream drainage with some tributaries directly
affected by valley fills and others that were not, measured conductivities upstream and
downstream of the tributary confluences. They found that the cumulative level of conductivity
values could be predicted using a simple function that combined tributary concentrations with
watershed area. Watershed area was a surrogate for volume of discharge from the tributary.
In terms of sediment contaminants, Paybins et al. (2000) found significant concentrations
of polycyclic aromatic hydrocarbons (PAH) at several stations within the Kanawha River basin
(see Table 10). Most of these PAHs appear to be constituents of particles of coal that occur in
sediments because of the extensive coal mining and transport of coal in the region (Paybins et al.,
2000). Downing-Kunz et al. (2005) found sediment concentrations of coal ranging from 1 to
53 g/kg in streams draining more southern parts of the central Appalachian coalfields in
Kentucky. PAHs are a natural component of coal (Chapman et al., 1996; Paybins et al., 2000),
but these PAHs are unlikely to be bioavailable to benthic invertebrates or fish (Carlson et al.,
1979; Ahrens and Morrissey, 2005; Yang et al., 2008). Arsenic (As) and metals were also
detected in sediments (see Table 10) of the Kanawha River. However, the source of these
sediment contaminants is less clear.
Table 10. Polycyclic aromatic hydrocarbons, arsenic, and metals detected in
sediments of larger streams in the Kanawha Basin
Chemical and units of concentration
benz [a] anthracene (fig/kg)
dibenz[a,h] anthracene (ug/kg)
2,6-dimethylnaphthalene (fig/kg)
fluoranthene (ug/kg)
fluorene (fig/kg)
naphthalene (ug/kg)
phenanthrene (fig/kg)
As (mg/kg)
Cr (mg/kg)
Pb (mg/kg)
Ni (mg/kg)
Zn (mg/kg)
Number of detects/number of
samples
12/13
4/13
10/13
13/13
7/13
9/13
13/13
13/13
13/13
13/13
13/13
13/13
Range of detections
5-800
40-200
50-500
30-1,100
60-300
3-700
9-900
4-20
60-110
20-50
50-100
200-600
Source: Paybins et al. (2000).
44
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5. AQUATIC TOXICITY TESTS
In this section, we report on results of toxicity tests relevant to evaluating water quality
downstream of MTM-VF. Toxicity tests expose organisms under laboratory conditions to
ambient media (i.e., water or sediment samples), whole effluents, reconstituted effluents, or
specific effluent constituents. Toxicity tests are valued because they can measure the effect of
the mixture as a whole, including antagonistic and synergistic effects. They also help distinguish
the effects of water quality from other stressors (e.g., habitat quality, flow regime changes,
temperature). Toxicity tests have been used as the basis for deriving water quality criteria and
permitting industrial and waste water effluents.
The most common standard toxicity tests used to evaluate the effects of effluents measure
the survival of the crustacean Ceriodaphnia dubia after 48 hours of exposure and the survival of
fathead minnows (Pimephales promelets) after 96 hours of exposure. Both of these tests have
significant limitations for evaluating MTM-VF effects: neither C. dubia nor P. promelas are
native to the streams of the study area, and the standard test durations are much shorter than the
exposures experienced by organisms downstream of MTM-VF operations. There are likely more
sensitive responses than death. In particular, because ions are so influential in regulating
membrane permeability during fertilization and egg development, effects on reproduction would
be expected (Zotin, 1958; Ketola et al., 1988). On the other hand, toxicity tests might
overestimate effects if organisms in the field are able to acclimate to exposures that slowly rise
over time. Still, the standard survival tests provide a useful benchmark for understanding toxic
potential. Other tests, which are more difficult and time consuming to run, can be used to
extrapolate short-term tests on survival to longer-term exposures, sublethal responses, and other
species.
5.1. TOXICITY TESTS USING WATER OR SEDIMENTS DOWNSTREAM OF
MTM-VF
One study (i.e., Merricks et al., 2007) tested media from three streams downstream of
MTM-VF in the central Appalachian coalfields. Water and sediment collected from some, but
not all, sites downstream of valley fills produced significant toxicity in laboratory organisms.
Water was tested using C. dubia. Results were reported as the percent dilution that killed
one-half of the test organisms over 48 hours (48-hour LCso). Three streams were tested. The
frequency of toxicity was highest in Lavender Fork; undiluted water from three of eight sites
sampled killed 50% or more of the test organisms. Lavender Fork also had the highest specific
conductivity levels; the undiluted water at the three toxic sites averaged 3,050, 2,497, and
2,657 jiS/cm. Specific conductivity measurements were available for two of the five sites from
45
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Lavender Fork that did not result in 50% or greater mortality; specific conductivity
measurements (2,720 and 2,667 |iS/cm) were comparable to the toxic sites. Only 1 of 20 sites
from the other two streams was sufficiently toxic to kill 50% or more of the test organisms.
Specific conductivity measurements in these streams ranged from 923 to 1,643 |iS/cm. There
was no obvious relationship between toxicity and water column measurements of trace metals
(e.g., Al, Fe, Mn, Zn, and Se).
Merricks et al. (2007) also conducted toxicity tests on sediments with another crustacean
Daphnia magna. The organisms were exposed to sediments for 10 days; results were reported as
percent survival and reproduction. Sediments from two of eight sites on Lavender Fork
significantly reduced survival or reproduction ofD. magna. Sediments from 3 of 19 sites on the
other two tested streams produced reduced survival or reproduction. Of the three streams,
Lavender Fork generally had the highest concentrations of trace metals in sediments (i.e., Al, Fe,
Cu, Cd, mercury [Hg], Se, As, Mn, and Zn). Concentrations of major ions or other chemicals
were not measured. Because of the way the sediment chemistry results were grouped for
summary, it is difficult to quantitatively relate them to the toxicity test results.
Asian clams (Corbiculaflumined) were deployed at monitoring stations (Merricks et al.,
2007). Growth was significantly greater below the treatment ponds and decreased downstream,
indicating that the ponds increased the food available to the clams. Significant mortality was
observed at 1 of 16 test sites. The authors attributed the mortality to Al and Cu, which had been
detected in a previous, unpublished study at water concentrations of 223 and 7.6 |ig/L,
respectively.
5.2. TOXICITY TESTS ON WATER FROM OTHER ALKALINE COAL MINING
EFFLUENTS
In a series of studies, Kennedy et al. tested the toxicity of mining effluents from Ohio
using C. dubia and the mayfly Isonychia bicolor (Kennedy et al., 2003, 2004, 2005). The
effluents originated from a surface mine, an underground coal mine, and a preparation facility.
Discharges from the underground mine and preparation facility were treated in a settling pond to
r\ ,
neutralize pH and reduce Mn, resulting in an effluent with high SO4 , Na , and Cl
concentrations and a mean hardness of 770 mg/L as CaCOs. Toxicity tests using C. dubia were
conducted following EPA protocols and used moderately hard reconstituted water (MHRW)8 to
dilute the effluent. Survival of C. dubia in 48-hour tests significantly decreased relative to
controls at a mean specific conductivity of 6,040 jiS/cm (Kennedy et al., 2003). Decreased
8MHRW was used as diluent in this study and many of the other studies discussed in section. MHRW has low
chloride concentrations (mean of 1.9) and a Ca:Mg molar ratio of 0.88; hardness ranges from 80-100 mg/L as
CaCO3 (Smith etal., 1997).
46
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survival in 7-day tests was observed at a mean specific conductivity of 4,730 jiS/cm. Decreased
reproduction in 7-day tests was observed at a mean conductivity of 3,254 jiS/cm, about 1.9 times
lower than the 48-hour results for survival (Kennedy et al., 2005). Tests on simulated effluent
made using only the major ions (i.e., no heavy metals) agreed well with the whole effluent,
providing evidence that the toxicity was caused by the ions, rather than an unmeasured toxicant
(Kennedy et al., 2005).
The same field-collected effluent was tested with a nonstandard test species, the mayfly,
/. bicolor, (Kennedy et al., 2004) in 7-day tests. In these tests, water from an unpolluted
reference stream was filtered and used as dilution water for the tests. Toxicity was greater at the
warmer temperature tested (20°C vs. 15°C); those results are reported here. Survival of
Isonychia significantly decreased relative to controls at specific conductivities of 1,562, 966, and
987 jiS/cm for three tests. These conductivities are about 3 times lower than those that reduced
Ceriodaphnia reproduction in 7-day tests using the same dilution water but a higher temperature
of25°C.
A sodium-dominated effluent from a coal-mine processing impoundment was tested
using C. dubia and the mayfly /. bicolor in 7-day tests using MHRW to dilute the effluent
(Echols et al., 2010). Conductivities corresponding to the lowest-observed-effect concentration
(LOEC) ranged from 1,508 to 4,101 for Isonychia survivorship compared with 2,132-4,250 for
Ceriodaphnia reproduction. Effects on Isonychia reproduction were not studied. Chapman et al.
(2000) tested a high sulfate alkaline coal mine effluent from Alaska in 10-day tests using the
insect Chironomus tentans. No effects on chironomid survival were found, but dry weight was
reduced approximately 45% in synthetic effluent (2,089 TDS/L). The researchers also tested the
effects of synthetic effluent on rainbow trout using two exposures: eggs were exposed for 4 days
starting immediately after fertilization, and swim-up fry were exposed for 7 days. No adverse
effects were seen in embryo viability or fry survival in the highest synthetic effluent
concentrations tested (2,080 TDS/L).
5.3. TOXICITY OF MAJOR IONS: K1", HCO3", MG2+, CL", SO42", NA+, CA2+
Laboratory studies that vary ion mixtures provide additional insight into which ions
might be driving toxicity and how interactions might be producing observed effects. We report
on the results of six of these groups of studies. Then, we compare the experimental results to ion
concentrations reported downstream of MTM-VF operations to gauge whether ion
concentrations would be expected to cause toxicity.
47
-------
5.3.1. Mount et al. (1997)
Mount et al. (1997) tested the acute-duration toxicity of over 2,900 ion solutions using
two crustacean species (i.e., C. dubia and/), magna) and the fathead minnow (P. promelas).
C. dubia was the most sensitive of the three. The toxicity of ion mixtures varied greatly with
composition; total ion concentrations corresponding to acute LCso for C. dubia ranged from
390 mg/L to over 5,610 mg/L. For/1, promelas, LCso values ranged from 680 to 7,960 mg/L.
I r\, r\
The authors reported relative toxicity as K > HCOs ~ Mg > Cl > SC>4 . They also
developed regression models that could be used to predict the 48-hour acute toxicity of
field-collected samples. In the models, the effects of the anions and cations were generally
additive with two notable exceptions: solutions with high concentrations of multiple cations had
lower toxicity than expected based on concentration addition, and Na+ and Ca2+ did not add any
explanatory value after the other ions were included in the model.
The regression models have been used to predict the toxicity of several complex
effluents. Tietge et al. (1997) used them to predict the acute toxicity of the ionic component of
production waters from fossil fuel extraction (Tietge et al., 1997). Toxicity of the Ohio coal
mine effluent (described above) to C. dubia was less than expected based on the equations,
although estimates were within a factor of 2 (Kennedy et al., 2005). Soucek (2007b) found that
the model overestimated the toxicity of high hardness solutions to C. dubia by a factor of about 5
(10% survival predicted vs. 50% survival observed).
5.3.2. van Dam et al. (2010)
In experiments using Australian test organisms and very low calcium water,
fji r\
van Dam et al. (2010) confirmed that Mg was a more toxic ion than SC>4 , and that increasing
concentrations of Ca2+ reduced Mg2+ toxicity. All toxicity tests were conducted using water with
94-
much lower Ca concentrations than observed downstream of MTM-VF (see Table 2).
5.3.3. Lasier and Hardin (2010)
Lasier and Hardin (2010) developed regression models to predict the toxicity of effluents
r\
from anion concentrations and hardness. They tested the toxicity of SC>4 , Cl , and HCOs to
C. dubia using a three brood reproductive endpoint over 9 days. All of the tests were conducted
at lower hardness levels (maximum of 93 mg/L) and higher chloride concentrations (minimum of
85 mg/L) than observed downstream of MTM-VF (see Table 2).
At the highest hardness tested (93 mg/L), LOECs were 1,250 mg SO427L, 650 mg C17L,
and 450 mg HCOs /L. Concentrations corresponding to a 25% inhibition of reproduction were
1,060 mg SO427L, 456 mg Cl~ 456/L, and 379 mg HCO37L. Increasing hardness decreased the
r\
toxicity of SO4 and Cl but had an insignificant effect on the toxicity
48
-------
5.3.4. Soucek (2007a, b); Soucek and Kennedy (2005)
Soucek (2007a, b) and Soucek and Kennedy (2005) conducted a series of 48-hour tests
on SC>42 using MHRW dilution water and varying levels of other ions and hardness. At the
r\
highest hardness tested (600 mg/L), the 48-hour LCso value for C. dubia was 3,288 mg SC>4 /L
(Soucek and Kennedy, 2005). In all tests, the crustacean Hyalella azteca was the most sensitive
test organism, followed by C. dubia, the bivalve Sphaerium simile and the insect C. tentans9
r\
H. azteca was particularly sensitive to SC>4 at low Cl concentrations. At Cl concentrations of
1.9 mg/L, H. azteca was four times more sensitive to SC>42 than C. dubia (Soucek, 2007b)10.
Toxicity decreased as Ca increased relative to Mg concentrations (Soucek and Kennedy, 2005).
Toxicity also decreased with increasing hardness, although the ameliorative effects of hardness
appeared to level off above 500 mg/L as CaCOs.
r\
In three-brood, 7-day tests on C. dubia, sublethal effects of SC>4 occurred at
concentrations 2.5 times lower than those that reduced survival (Soucek, 2007a). The LOEC
r\
was 899 mg SC>4 /L for a reproductive endpoint (mean number of neonates per female)
compared with 2,216 mg/L for percent survival. Other sublethal effects were investigated using
24-hour tests; significant declines in feeding rates and oxygen consumption were observed in
C. dubia exposed to 1,000-mg SO427L.
5.3.5. Meyer et al. (1985)
Meyer et al. (1985) tested four salts using 48-hour tests on D. magna and 96-hour tests on
P. promelas. High hardness dilution water was used (563 mg/L as CaCOs). D. magna was more
sensitive to all of the salts than P. promelas. The relative toxicity of the salts was magnesium
sulfate (MgSO4) > NaCl > NaNO3 > Na2SO4.n The LC50 values calculated for MgSO4 were
4,300 mg/L and 7,900 mg/L for D. magna and P. promelas, respectively. All of these values are
well above concentrations reported downstream of MTM-VF (see Table 2).
5.3.6. Skaar et al. (2006)
Skaar and coauthors (2006) conducted acute-duration tests of sodium bicarbonate
(NaHCOs) on fathead minnow (P. promelas) and pallid sturgeon (Scaphirhynchus albus) using
test water that simulated conditions in the Tongue and Powder Rivers of Montana. (Limited tests
were also conducted using white sucker but are not included in this summary). Hardness values
9C. tentans has since been renamed Chironomus dilutes.
These SO42~ results might be unreliable because the synthetic test media contained Cl" concentrations that,
although similar to those observed downstream of MTM-VF, were likely insufficient to maintain healthy Hyalella
cultures (U.S. EPA 2011).
nMgSO4 = magnesium sulfate; NaCl = sodium chloride; NaNO3 = sodium nitrate; Na2SO4 = sodium sulfate.
49
-------
in the tests were not reported. 96-hour LCso values using Powder River water were 840 mg
HCOs/L and 1,193 mg HCOs/L for 4-day-old S. albus and P. promelas, respectively.12 Earlier
life stages appear to be more sensitive: 96-hour exposures to 580 mg HCOs /L resulted in
55% mortality to fathead minnow embryos exposed shortly after fertilization. Longer exposures
also appear to increase toxicity. In 30-day tests using fathead minnows, survival declined at
concentrations above 290 mg HCOs /L. No decline in growth was noted in surviving fish,
however, indicators of kidney damage increased slightly.
5.4. COMPARING TOXICITY TESTS ON MAJOR IONS TO OBSERVATIONS
DOWNSTREAM OF MTM-VF
r\ o n, ,
Evidence from the laboratory toxicity tests suggests that SO4 , HCO , Mg , and K are
the principal contributors to the toxicity of the mixture (see Figure 15). Ca2+ in the effluent and
receiving water is expected to might reduce the toxicity of the mixture, possibly by mitigating
r\ fj,
the toxicity of SO4 and Mg . Ionic regulation by organisms depends on the relative
proportions of all ions. For this reason laboratory manipulations of one or a few ions at a time
are difficult to extrapolate to exposures encountered by organisms in the field. For example, the
relatively low concentrations of ions such as Na+ and Cl in effluents downstream of MTM-VF
might also be contributing to the overall toxicity of the mixture. Increasing our understanding of
the responses of native freshwater organisms to different mixtures of ions and overall ionic
strength is a high priority research need (see Section 8).
t Mg2+
tK+
Ca2+ineffluent&
receiving water
>. j
t HC03- t S042-
\
X
\
/
4/ survival, growth & reproduction in
Ceriodaphnia dubia toxicity tests
LE
GEND
proximate
stressor
' response ^
modifying factor
Figure 15. Ions expected to contribute to effects in toxicity tests of water
sampled downstream of MTM-VF. See Sections 5.3 and 5.4 for additional
details and evidence.
12 Concentrations reported as NaHCO3 were converted to HCO3" concentrations by multiplying them by 0.726.
50
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Applying the Mount et al. (1997) regression models to ion concentrations reported
downstream of MTM-VF suggests that the ion mixture at some sites might reach acutely lethal
levels to C. dubia. The models predict minimal mortality of C. dubia (1%) at mean
concentrations of each ion summarized in Table 4 (mean specific conductance of
1,023 |iS/cm).13 However, applying the assumption that ion concentrations are strongly
correlated, we also calculated predicted toxicity using the maximum reported concentrations for
each ion (maximum specific conductance of 2,540 |iS/cm). More than 75% mortality is
predicted at these maximum concentrations. The results of the model indicate that SC>42
fj, _i_
concentrations contributed the most to the predicted toxicity, followed by HCOs , Mg , and K .
OF concentrations contributed minimally. The interaction between cations (Mg2+and Ca2+) and
r\
SC>4 reduced predicted toxicity substantially. The models predict minimal mortality (1%) for
P. promelas even at maximum concentrations.
Model predictions of toxicity are generally consistent with the observed C. dubia toxicity
test results reported by Merricks et al. (2007). Five sites tested by Merricks et al. (2007) had
specific conductivity measurements comparable or greater than the maximum specific
conductivity summarized in Table 2 (2,540 |iS/cm). If the relative proportion of ions was the
same in Merricks et al. (2007) as in Pond et al. (2008), we would expect these high conductivity
sites to produce greater than 75% mortality. Three of these five sites exhibited 50% or greater
mortality in 48-hour tests. Of the 11 sites with substantially lower specific conductivity readings
(all less than 1,643 |iS/cm), only 1 exhibited greater than 50% mortality in the toxicity tests.
Using the anion plus hardness model developed by Lasier and Hardin (2010), C. dubia
reproduction would be expected to be unaffected even at the highest anion concentrations
reported in Table 2. However if, as the authors suggest, hardness does not reduce the toxicity of
HCOs , then reproduction would be expected to be 88% of controls at the mean concentration
observed and 47% of controls at the maximum concentration of HCOs . The concentrations of
HCOs shown in Table 2 reach levels at which effects were observed in chronic tests on the
fathead minnow P. promelas (Skaar et al., 2006). However, the other ions in the tested mixture
were quite different than those reported in Table 2; for example, sodium levels were higher.
13For C. dubia, proportion surviving (P) in 48-hour tests was calculated as logit (P) = ln[P/(l - P)] = 8.83 +
(-0.0299 x [K+]) + (-0.00668 x [Mg2+]) + (-0.00813 x [Cl']) + (-0.00439 x [SO42D + (-0.00775 x [HCO3D +
(-0.446 x 2) + (0.00870 x 2 x [K+]) + (0.00248 x 2 x [CT]) + (0.00140 x 2 x [SO42D (Mount et al., 1997).
Concentrations are as reported in Pond et al. (2008) except for HCO3 . HCO3" concentrations were reported as
CaCO3 (Personal Communication from M. A. Passmore, U.S. EPA Region III, Wheeling, WV, 2009) and were
converted to HCO3 concentrations by multiplying by 1.22.
For P.promelas, proportion surviving (P) in 96-hour tests was calculated as logit (P) = ln[P/(l - P)] = 4.70 +
(-0.00987 x [K+]) + (-0.00327 x [Mg2+]) + (-0.00120 x [Cl~]) + (-0.000750 x [SO42D + (-0.00443 x [HCCV])
(Mount et al., 1997). Concentrations are as reported in Pond et al. (2008) except for HCO3 . HCO3 concentrations
were reported as CaCO3 (Personal Communication from M. A. Passmore, U.S. EPA Region III, Wheeling, WV,
2009) and were converted to HCO3 concentrations by multiplying by 1.22.
51
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The toxicity tests on other alkaline mine effluents discussed in Section 5.3.1 suggest that
effects to other organisms should be expected at concentrations below those that affect
Ceriodaphnia. Tests using the mayfly /. bicolor and the amphipod H. azteca found effects on
survival at concentrations 3-4 times lower than those affecting Ceriodaphnia. If effects on
reproduction in these organisms are similarly more sensitive than survival, effects would be
expected at most sites downstream of MTM-VF.
The relatively high sensitivity of mayflies to ions in alkaline mine effluent is consistent
with relative sensitivity of mayflies to other salts. Mayflies were the most sensitive order of
invertebrates tested in 72-hour laboratory studies of NaCl on South African invertebrate species
(Kefford et al., 2004). In studies on metal salts in experimental streams and toxicity tests from
the United States, the most sensitive invertebrates tend to be mayflies (Warnick and Bell, 1969;
Clark and Clements, 2006). In studies on artificial seawater (dominated by NaCl) from
Australia, the most sensitive species also were mayflies (Kefford et al., 2003).
Finally, there is some evidence that younger organisms might be more sensitive. The
concentrations associated with effects from the sodium-dominated mine effluent tested using the
mayfly Isoynchia were lowest in tests using the smallest and presumably youngest organisms
(Echols et al., 2010). In tests with bicarbonate, 7-day-old H. azteca were two times more
sensitive than 14-day old organisms (Lasier et al., 1997). In studies of metal salts (Cu, Cd, and
Zn), in experimental streams, (Kiffney and Clements, 1996) toxicity increased as organism size
decreased. Just-fertilized embryos of P. promelas were about 1.5 times more sensitive to sodium
bicarbonate than 4-day old larvae (Skaar et al., 2006).
5.5. TOXICITY OF TRACE METALS IN WATER
5.5.1. Selenium
Se is a metalloid element that is a micronutrient and, at higher exposures, a toxicant.
Selenium from coal ash and coal mine wastes has resulted in elevated Se concentrations in
surface waters and toxicity to aquatic organisms (Orr et al., 2005). Se is unusual in that its
toxicity results from complex processes of transformation and bioaccumulation, analogous to
mercury toxicity (see Figure 16). Environmental exposures of animals are primarily dietary, and
effects on sensitive early life stages are due primarily to maternal transfer. The current chronic
AWQC for Se is 5.0 |ig/L, and the median, mean, and range of Se concentrations in streams
draining valley fills are 12.5, 10.6, and <1.5-36.8 |ig/L, respectively (Bryant et al., 2002;
Pond et al., 2008). The chronic-duration criterion is relevant because the discharge from mining
operations is a chronic source. This section discusses effects of Se on aquatic invertebrates, fish,
and birds, emphasizing studies of waters receiving coal overburden leachates because the valley
fills are filled with coal overburden.
52
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mineralSe
in overburden
f acidic ]
I leachate J
1s inorganic Se
dissolved in stream
V
V
1s particulate Fe & Mn
(neutralized acid drainage)
V
sorbed inorganic Se
1s organo-Se
in microbes, plants & detritus
1s organo-Se
in herbivores & detritivores
V
organo-Se
in predators
mortality &
deformities
1s organo-Se
in secondary predators
Figure 16. Selenium transformation, transfer and effects expected in aquatic
ecosystems downstream of MTM-VF. See Section 5.5.1 for additional details
and evidence.
5.5.1.1. Selenium Dynamics in Aquatic Ecosystems
The complex dynamics of Se have been recently summarized by Chapman et al. (2010),
Luoma and Presser (2009), and Presser and Luoma (2010). Selenium, leached from coal and
organic overburden, enters streams in valley fill effluents (see Section 4.3.1). Dissolved oxy
anions of selenate (Se+4) and selenite (Se+6) are taken up by microbes, algae, and plants and
converted to organic forms. In the streams like those below MTM-VF, the primary community
that can perform this conversion is the periphyton growing on rocks and woody debris, and the
conversion rates are relatively low. However, uptake, conversion, and retention of Se are more
53
-------
efficient in lentic systems such as the reservoirs that occur downstream in some watersheds.
Alternatively, in streams with iron oxyhydroxides or manganese dioxide due to neutralization of
acidic leachates, significant sorption of Se to those amorphous minerals might occur. Herbivores
and detritivores (largely aquatic macroinvertebrates) accumulate Se by grazing and collecting
organic particles. These primary consumers are in turn consumed by predators including fish,
amphibians, and birds. Secondary predators, such as largemouth bass or herons, can further
accumulate Se, but such large predators occur primarily in larger water bodies. In sum, Se can
bioaccumulate and even biomagnify where retention is high and food chains are long. In the
region of concern, these conditions could occur in reservoirs and potentially in riparian wetlands.
Se bioaccumulation is expected to be lower in the streams below MTM-VF operations.
However, these conclusions are based on general knowledge of Se dynamics and not on specific
studies in the region of concern.
5.5.1.2. Invertebrates
A review of the literature estimated that the range of thresholds for sublethal toxicity in
aquatic invertebrate genera is 1-30 |ig/L (DeBruyn and Chapman, 2007), which are similar to
the concentrations observed downstream of MTM-VF. A recent study showed that dietary
selenium is bioaccumulated by the mayfly Centroptilum triangulifer and suggested that
reproductive effects occur at aqueous exposures of 13.9-jig/L dissolved Se (Conley et al., 2009).
These results are consistent with data from streams draining Canadian coal mines that found a
>50% decline in the abundance of some taxa in the range of 5-100 |ig/L (DeBruyn and
Chapman, 2007). In outdoor artificial streams dosed with Se, isopods (Caecidotea) and
oligochaete worms (Tubifex) were severely reduced in abundance at 30 |ig/L and statistically
significantly reduced at 10 jig/L (Swift, 2002). However, the abundances of baetid mayfly
nymphs (Baetis, Callibaetis), damselfly nymphs (Enallagma), and chironomid larvae were not
statistically significantly reduced—even at 30 |ig/L.
5.5.1.3. Fish
Numerous studies have shown severe effects of Se on fish reproduction in the field as
well as in the laboratory, and effects on fish are the basis for the national criterion (U.S. EPA,
2004). Cutthroat trout embryos from a pond at a coal mine in British Columbia with 93 |ig/L Se
showed effects ranging from deformities of larvae to mortality (Rudolph et al., 2008). The
probability of mortality was correlated with Se concentrations in the embryos. These trout are
much less sensitive than other species such as bluegill sunfish. In the artificial stream study,
bluegill sunfish exhibited mortality and characteristic skeletal deformities at all concentrations,
including 2.5 |ig/L, although the effects were not statistically significant at that lowest level
54
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(Swift, 2002). Creek chubs and blacknose dace from the Coal, Little Coal, Big Coal, and Mud
River watersheds in West Virginia contained Se from <0.48 to 6.89 mg/kg dry weight
(Paybins et al., 2000). Fish from 3 of 22 of these waters had concentrations >5 mg/kg, putting
them at "moderate hazard" for toxic effects based on the scale developed by Lemly (1993).
5.5.1.4. Birds
Se has caused reproductive failure and gross deformities in birds that forage in
Se-contaminated waters, but their sensitivity is highly variable (Ohlendorf et al., 2003). Birds
foraging in streams receiving leachate from coal mine overburden in the Elk River, British
Columbia watershed showed reproductive effects, but they were less severe than expected given
the high Se concentrations (8.1-34.2 |ig/L) (Harding et al., 2005). In particular, spotted
sandpipers experienced a reduction in egg hatchability from 92% in reference streams to 78% in
streams receiving overburden leachate. Spotted sandpipers forage in streams in the Appalachian
Range, but Louisiana waterthrushes occur more commonly in the area of concern and forage on
aquatic invertebrates, so they would be similarly exposed. The authors suggest that the low level
of effects relative to other Se-contaminated waters was due to low bioaccumulation, which was
due to the low rates of biotransformation and uptake in those streams. Piscivorous birds
(primarily Belted Kingfishers and Great Blue Herons) could be at risk from Se-contaminated fish
(see Section 5.5.1.3). The 10th percentile effective concentration for hatchability in dietary
exposures of mallard ducks (a surrogate species for the piscivorous birds) to Se in dry diet was
4.87 mg/kg (Ohlendorf et al., 2003). Five of the 22 fish samples from 13 streams analyzed by
Paybins et al. (2000) for Se from the Coal, Little Coal, Big Coal and Mud River watersheds
exceeded that endpoint.
5.5.2. Manganese and Iron
Maximum concentrations of Mn reported downstream of MTM-VF are substantially
lower than those associated with effects in the few available toxicity tests. Maximum
concentrations of dissolved Mn reported in Pond et al. (2008) were 0.853 mg Mn/L. Tests using
C. dubia in hard water (hardness =184 mg/L) yielded a mean 48-hour LCso of 15.2 mg Mn/L for
C. dubia and a 96-hour LCso value for H. azteca of 13.7 mg Mn/L (Lasier et al., 2000). In 7-day
tests, C. dubia reproduction (number of young per female) was inhibited 50% at mean
concentrations of 11.5 mg Mn/L. In 62-day life-cycle tests using brown trout, concentrations
associated with a 25% inhibition in survival or growth were 5.59 mg/L and 8.68 mg/L at
hardness levels of 150 and 450 mg/L, respectively (Stubblefield et al., 1997).
In a study of biochemical effects, concentrations of chemicals involved with cellular
redox regulation were reduced at concentrations lower than reported by Pond et al. (2008):
55
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glutathione levels were reduced in caddisflies (Hydropsyche bettenf) exposed to 0.05 mg Mn/L,
and cysteine levels were reduced in mayflies (Maccaffertium modestum) exposed to 0.10 mg
Mn/L (Dittman and Buchwalter, 2010). However, no overt toxic effects were reported in this
study, or in companion bioaccumulation tests that exposed a wide variety of Appalachian stream
insects to concentrations up to 0.40 mg Mn/L (Dittman and Buchwalter, 2010).
Maximum concentrations of total and dissolved Fe of 0.65 mg/L and 0.28 mg/L,
respectively, have been observed downstream of MTM-VF (see Table 2). Iron toxicity decreases
as pH increases (see the recent review by Phippin et al. [2008]). The pH of water downstream of
MTM-VF ranged from 6.3-8.9 (see Table 2). Observed iron concentrations are similar to the
96-hour tolerance limit concentration of 0.32 mg FeSCVL at pH 7.5 reported for Ephemerella
sp. survival in a study conducted prior to standardized toxicity test protocols (Warnick and Bell,
1969). Other organisms tested in that study were less sensitive. LCso values were 16.0 mg Fe/L
for both the stonefly Acroneuria lycorias and the caddisfly Hydropsyche betteni in 9- and 7-day
tests, respectively. No differences in survival, feeding, or escape activity were observed in
experiments exposing field-collected larval mayflies (Leptophlebia marginata) to up to
50 mg Fe/L at pH 7 for about 30 days (Gerhardt, 1992). The effect concentration for 50% of the
tested organisms (ECso) for reduced escape activity was calculated as 70 mg Fe/L at
circumneutral pH (between 5.95 and 6.74) (Gerhardt, 1994). In studies using/), magna at pH of
7.5, reproduction declined 16% at 4.38 mg Fe/L (Biesinger and Christensen, 1972).
Observed iron concentrations reach levels that exceed several of the family-level
benchmarks for total Fe derived from field observations of benthic macroinvertebrates from
West Virginia (see Table 11). Benchmark values (called field effect concentrations, FEC2os)
corresponded to a 20% decline in the organism numbers compared with reference sites and were
estimated from the 90th percentile quantile regression relationship between total Fe and numbers
of organisms collected from different families. However, because the benchmark derivation did
not control for stressors that covary with iron, the benchmarks might reflect the effects of other
stressors in addition to iron.
5.6. TOXICITY OF TRACE METALS IN SEDIMENT
Only two studies measured concentrations of trace elements in sediments below
MTM-VF. Most concentrations were below available consensus-based screening levels (see
Table 12). The consensus-based screening levels are based on analysis of paired sediment
chemistry and toxicity test results from field studies and should be interpreted as concentrations
at which effects in toxicity tests are frequently observed. Zinc and Ni concentrations in
Kanawha Valley sediments exceed the probable effects levels and warrant further investigation.
Toxicity of Zn and Ni is a function of particle size, organic carbon content, pH, and acid volatile
56
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Table 11 Field-based 20% effect concentrations (FEC2o) for iron
(Linton et al., 2007)
Macroinvertebrate family
Leptophlebiidae
Emphemerellidae
Philopotamidae
Psephenidae
Heptageniidae
Elmidae
Baetidae
Tipulidae
FEC2o (mg total iron/L)
0.21
0.43
0.44
0.48
0.66
1.13
1.48
7.05
Table 12. Comparison of measured sediment concentrations with probable
effects levels
Chemical
Al
As
Cd
Cr
Cu
Fe
Pb
Mn
Hg
Ni
Se
Zn
Concentration downstream
ofMTM-VF(mg/kg)
(Merricks et al., 2007)a
3-32
0.015-0.070
0.005-0.045
0.019-0.122
<48.5-157.6
1-41
0.006-0.015
0.1-2.5
2.0-2.5
Concentration in Kanawha
Valley sediments (mg/kg)
(Paybins et al., 2000)b
4-20
60-110
20-50
50-100
200-600
Consensus probable effects
level (mg/kg)
(MacDonald et al., 2000)c
33
4.98
111
149
128
1.06
48.6
459
"Data from Table III and Figure 3 combined.
bData from figures in appendix.
°We note that the concentrations reported in Merricks et al. (2007) are substantially lower than ranges of values
reported in Paybins et al. (2000) or used to develop the probable effects levels (e.g., see Smith et al., 1996),
suggesting that any comparisons should be made with caution.
Blank cells indicate that the metal was not measured, or there is no probable effects level available.
57
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sulfides (Di Toro et al., 2001; Doig and Liber, 2006). It is difficult to interpret the observed
concentrations without measurements of the factors that influence toxicity, or, alternatively,
pore-water concentrations.
58
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6. IMPACTS ON AQUATIC ECOSYSTEMS
In this section, we report on changes in stream community composition associated with
MTM-VF. In contrast to the toxicity tests discussed in Section 5, field studies are our primary
resource for this section.They are directly relevant to both the exposures and biota of interest.
Macroinvertebrate and fish assessments consistently indicate degraded biological conditions
downstream of MTM-VF.
6.1. EFFECTS ON BIOLOGICAL COMPOSITION
Mountaintop mining and associated valley fills in a watershed are associated with
degraded community compositions in downstream habitats. Though there are relatively few
studies on the direct ecological effects of mountaintop mining, the findings are unambiguous
(Howard et al., 2001; Stauffer and Ferreri, 2002; Fulk et al., 2003; Armstead et al., 2004;
Kirk and Maggard, 2004; Hartman et al., 2005; Merricks et al., 2007; Pond et al., 2008).
Although a number of different biological responses have been associated with the effects of
MTM-VF (see Figure 17), all the relevant studies reviewed found that mayfly (i.e., the insect
order Ephemeroptera) populations were consistently lower in streams draining watersheds with
MTM-VF than in streams draining watersheds with intact forest. Associated with the extirpation
of mayfly species, biological assessment metrics indicate degraded conditions immediately
downstream of MTM-VF s.
6.1.1. Benthic Macroinvertebrates
6.1.1.1. Benthic Macroinvertebrate Indices
All surveys that used multimetric and aggregate taxonomic indices observed degraded
biological conditions in streams affected by mining and valley fills (see Table 13). Fulk et al.
(2003) used the West Virginia Stream Condition Index (WVSCI) to analyze benthic
macroinvertebrate data from 34 streams in West Virginia. The index is composed of several
metrics that are responsive to environmental and chemical stress, e.g., EPT (Ephemeroptera,
Plecoptera and Trichoptera) taxa, total taxa, and percent EPT were expected to decrease with
increasing stress and percent Chironomidae; Hilsenhoff biotic index (HBI) and percent of the top
two dominant taxa were expected to increase with increasing stress. Four classes of streams
were compared (1) no mountaintop mining upstream (n = 9), (2) upstream valley fills (n = 15),
(3) mountaintop mining in watershed (n = 4), and (4) upstream valley fill and residential
development in the watershed (n = 6). Fulk et al. (2003) found that benthic macroinvertebrate
indices were lower in streams with upstream valley fills. With the exception of the fall of 2000,
59
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mountain top removal mining
& valley fill
\|/ total invertebrate
\|/ invertebrate
relative
dominance
functionalgroups
\|/ other
invertebrate taxa
-J/ clingers ; CTfilterers
•
^-—
\|/ scrapers ) ( \|/ shredders
\|/ benthic
invertivores
\|/ taxa
richness
\|/ macroinvertebrate
index quality
\|/ fish index
quality
Figure 17. Macroinvertebrate and fish responses associated with MTM-VF. Responses were significantly
different from reference sites in at least one of the reviewed studies. Both negative and positive responses of
non-insects and midges were observed and are not shown. See Section 6.1 and Table 13 for additional details and
evidence.
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Table 13. Summary of research examining the relationship between
mountaintop mining and ecological characteristics in downstream habitats
Reference
Fulk et al.,
2003
Hartman et al.,
2005
Experimental design
Fish survey comparing MTM-VF streams
(n -17) to unmined reference streams
(n - \A)
Benthic macroinvertebrate surveys (spring)
comparing MTM-VF streams (n = 9 in
1999 n ~ 10 in 2000) to unmined reference
streams (n= 15)
Benthic macroinvertebrate survey
comparing streams with valley fills (n = 4)
(» = 4)
Ecological response
FishlBI
Benthic invertivores
Native Cyprinidae (minnows)
richness
% gravel spawners
% predators
Intolerant species richness
% normative fish
% macro-omnivores
% tolerant species
% Cottidae (sculpins)
Invertebrate IBI
Total taxa richness
EPT taxa richness
Hilsenhoff Biotic Index
% 2 dominant taxa
% EPT taxa
% Chironomidae
Coleoptera density
Diptera density
Ephemeroptera density
Odonata density
Plecoptera density
Trichoptera density
Total density
EPT density
Chironomidae density
Noninsect density
Collector density
Scraper density
Shredder density
Observed effect"
Lower
Lower
Lower
No difference
No difference
No difference
No difference
No difference
No difference
No difference
Lower
Lower
Lower
No difference
No difference
No difference
No difference
Lower
No difference
Lower
Lower
No difference
No difference
No difference
No difference
No difference
Lower
No difference
Lower
Lower
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Table 13. Summary of research examining the relationship between
mountaintop mining and ecological characteristics in downstream habitats
(continued)
Reference
Armstead et al.,
2004
Howardetal.,
200 lb
Merricks etal.,
2007
Experimental design
Benthic macroinvertebrate survey
comparing MTM-VF streams (n = 14,
watersheds without mining activity (n = 9,
winter; n = 10, spring)
Benthic macroinvertebrate survey
comparing streams in mined watersheds
mining activity (n = 4)
Benthic macroinvertebrate survey
comparing streams with valley fills (n = 4)
to a reference stream without vallev fill
(« = 1)
Ecological response
Total density
Taxa richness
Hilsenhoff's Biotic Index
EPT density
EPT richness
% EPT
% 2 dominant taxa
% Chironomidae
% Ephemeroptera
% Plecoptera
% Trichoptera
Taxa richness
EPT index
Biotic index
% clinger
% Ephemeroptera
% chironomids + oligochaetes
KYMBI
Total richness
EPT richness
Ephemeroptera richness
Plecoptera richness
Trichoptera richness
Hilsenhoff Biotic Index
% Chironomidae
% EPT
% Ephemeroptera
% Plecoptera
% Trichoptera
% collector-filterer
% shredder
Observed effect"
No difference
No difference
Higher
No difference
Lower
Lower
Increase
No difference
Decrease
Decrease
No difference
Lower
Lower
Higher
Lower
Lower
Higher
Lower
No difference
No difference
Lower
No difference
No difference
Higher
No difference
Lower
Lower
Lower
Higher
Higher
Lower
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Table 13. Summary of research examining the relationship between
mountaintop mining and ecological characteristics in downstream habitats
(continued)
Reference
Pond et al.,
2008
Pond, 2010
Stauffer and
Ferreri, 2002
Experimental design
Benthic macroinvertebrate survey
comparing MTM-VF streams (n = 27) to
Benthic macroinvertebrate survey
comparing MTM-VF streams (20) to
unmined reference streams (44)
Fish communities were compared in streams
with valley fills (n - 9) to streams without
mining activity (n = 9)
Ecological response
Total richness
EPT richness
Ephemeroptera richness
Plecoptera richness
WV genus biotic index
WV family biotic index
Shannon H'
% Orthocladiinae
% Chironomidae
% Ephemeroptera
% Plecoptera
% EPT
Ephemeroptera richness
% Ephemeroptera
Fish species richness
Benthic fish richness
Observed effect"
Lower
Lower
Lower
Lower
Lower
Lower
Lower
Lower
Lower
Lower
No difference
Lower
Lower
Lower
Lower
Lower
aComparing the mean values from the reference and downstream and/or mined sites, where "lower" indicates that
the mined/valley fill site has a significantly lower metric value than the reference site (significance as determined
by statistical analyses in original study).
bThe original study did not present statistical analyses on these comparisons.
IBI = Index of Biotic Integrity; MBI = Macroinvertebrate Bioassessment Index.
the macroinvertebrate index was significantly different among stream classes, and the differences
were caused by fewer total taxa and fewer EPT taxa in streams with valley fill. While unmined
sites were typically classified as "very good," streams with upstream valley fills had WVSCI
scores ranging from "fair" to "good," indicating that stream sites with valley fill were degraded
compared to unmined sites. Similar results were observed in an assessment of the biological
condition of streams in Kentucky. Howard et al. (2001) used the Kentucky Macroinvertebrate
Bioassessment Index (MBI) (Pond and MacMurray, 2002), which includes four components of
macroinvertebrate community condition. Streams with mining activity in the watershed in
Kentucky had lower MBI ranks than streams in watersheds without mining (mined streams had a
rank of "poor," and reference streams were "good"). In a time-series study (i.e., April and
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October samples from 1996 to 2003) of a third-order stream affected by two smaller tributaries
with MTM-VF, Kirk and Maggard (2004) generally observed WVSCI scores in the "fair" to
"good" range at the downstream site compared to scores in the "good" to "very good" range at
an upstream reference site, after a period of drought from fall 1998 through spring 1999. During
the early part of the study (i.e., 1996 to 1998), the WVSCI scores were similar between the sites
in the "good" to "very good" range.
Merricks et al. (2007) found a HBI score of 1.91 in a reference stream, indicative of
excellent water quality, and HBI values ranging between 5.64 and 5.7, indicative of fair water
quality, below valley fills and ponds in three streams. A similar pattern of increased HBI was
observed during sampling in two seasons by Armstead et al. (2004).
In a comparison of streams with and without mining in the watershed, Pond et al. (2008)
observed that streams below fills had a significantly lower macroinvertebrate biotic index score
than those without fills using both a genus-level index of most probable steam status (GLIMPSS,
2.4 vs. 4.5) and a family-level biotic index (WVSCI, 3.4 vs. 4.3). Using the genus-level index,
93% (25 of 27) of sites downstream of mining activity exhibited scores indicative of biological
impairment, compared with none (0 of 10) of the sites that were in reference (unmined)
watersheds. Using the family-level index, 63% (17 of 27) of downstream of mining activity
exhibited scores indicative of biological impairment, compared with none (0 of 10) of sites that
were in reference (unmined) watersheds (Pond et al. 2008).
6.1.1.2. Benthic Macroinvertebrate Diversity
In most cases, lower taxonomic diversity was observed at sites downstream of MTM-VF.
A pattern of lower macroinvertebrate richness in streams with mining in the watershed was
found in Kentucky (mean of 31 at mined sites and 43 at reference sites, Howard et al., 2001), in
West Virginia (mean generic richness of 21.7 at mined sites and 31.9 at unmined sites,
Pond et al., 2008), and in a combination of sites from both states (median family richness of
12-13 at sites with fills and 18-21 at unmined sites in spring, Fulk et al., 2003). In a time-series
study (i.e., April and October samples from 1996 to 2003) of a third-order stream affected by
two smaller tributaries with MTM-VF, Kirk and Maggard (2004) generally observed lower
macroinvertebrate richness at the downstream site compared to an upstream reference site,
particularly since 1999. In contrast, Armstead et al. (2004) and Merricks et al. (2007) found no
significant difference in taxonomic richness between streams with valley fills and streams
without valley fills in the watershed.
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6.1.1.3. Benthic Macroinvertebrate Density
No difference was found in the total density of macroinvertebrates between streams with
valley fills and reference streams (Armstead et al., 2004; Hartman et al., 2005). In a time-series
study (i.e., April and October samples from 1996 to 2003) of a third-order stream affected by
two smaller tributaries with MTM-VF, Kirk and Maggard (2004) generally observed greater
macroinvertebrate densities at the downstream site compared to an upstream reference site early
in the study (i.e., 1996-2000) but lower densities in the later part of the study (2000-2003).
6.1.1.4. Benthic Macroinvertebrate Functional Groups
MTM-VF were associated with changes in the functional composition of
macroinvertebrate communities. Typically, macroinvertebrate communities in headwater
streams are dominated by shredders, which feed on leaf detritus (e.g., Vannote et al., 1980). In
the case of mining activities, shredder density metrics (Hartman et al., 2005) and proportion of
the community (3% in streams with mining and 50% in a reference stream, Merricks et al., 2007;
11% in streams with valley fills and settling ponds; 22% in unmined streams, Armstead et al.,
2004) were lower in streams below fills. Other changes include lower percentage of the
community as clingers (i.e., organisms that cling to rocks in riffles) in mined watersheds than in
watersheds without mining (Howard et al., 2001). Also, a scraper (i.e., organisms that feed on
attached algae) density metric was lower in streams with valley fills in the watershed than it was
in streams without valley fills (Hartman et al., 2005). The percentage of the community as
collector-filtering macroinvertebrates (i.e., organisms that feed on suspended particulate organic
matter, including algae) was greater in streams downstream from both the fills and settling ponds
(Merricks et al., 2007; Armstead et al., 2004), but a collector density metric showed no
difference between streams below fills and reference streams in another study
(Hartman et al., 2005).
6.1.1.5. Benthic Macroinvertebrate Taxa
Specific changes in macroinvertebrate taxonomic composition are described below.
6.1.1.5.1. Coleoptera
The only study that included Coleoptera populations in their assessment found that a
density metric of Coleoptera was lower in streams below valley fill than in streams without
valley fills in the watershed (Hartman et al., 2005).
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6.1.1.5.2. Diptera
The effects of MTM-VF on Diptera population characteristics were mixed. In some
cases, there were no observed effects of fills or mining on the watershed. For example, perhaps
owing to moderate degradation in the reference sites, density metrics for Diptera and
Chironomidae, a family within the insect order Diptera, showed no difference between streams
downstream of valley fills and those without (Hartman et al., 2005). Armstead et al. (2004) and
Merricks et al. (2007) had similar findings, where the percentage of the community comprised of
Chironomidae showed no difference between sites downstream of valley fills and a stream
without fills. In another study, the percent Chironomidae was greater in streams with mining in
the watershed than in streams with no mining (27% in mined and 13% in unmined streams,
Pond et al., 2008), but in a time-series study (i.e., April and October samples from 1996 to 2003)
of a third-order stream affected by two smaller tributaries with MTM-VFs, percent
Chironomidae did not consistently increase or decrease at the downstream site compared to the
upstream reference site. A combined measure of the percent Chironomidae and Oligochaeta was
higher in streams in mined watersheds compared to streams in watersheds without mining (63%
in mined and 3% in reference streams, Howard et al., 2001). The family Chironomidae includes
both tolerant and intolerant taxa, which might account for the equivocal results.
6.1.1.5.3. Ephemeroptera
Ephemeroptera population characteristics showed the most definitive changes associated
with mining activities, being consistently lower in streams affected by MTM-VF.
Ephemeroptera density metrics were lower in sites downstream of valley fills than in streams
without fill (Hartman et al., 2005). Pond (2010) found decreases in the abundances of individual
mayfly genera, such as Ameletus, Dmnella, Ephemerella, Cinygmula, Epeorus, and
Mccaffertinum in mine-impacted streams. The proportion of the community as Ephemeroptera
was lower in impacted streams. Howard et al. (2001) found an average of 1% in streams with
mountaintop mining in the watershed and 55% in reference streams. Four additional studies
report similar observations of proportion. Merricks et al. (2007) found 3% Ephemeroptera in
streams with mountaintop mining in the watershed and 17% in reference streams in West
Virginia. Pond et al. (2008) found 7% Ephemeroptera in streams with mountaintop mining in
the watershed and 45% in streams with no mining in West Virginia. Pond (2010) found
2% Ephemeroptera in streams with mountaintop mining in the watershed and 45% in streams
reference streams in Kentucky. Armstead et al. (2004) found 4% (winter) or 16% (spring)
Ephemeroptera in streams with valley fills and 30% (winter) or 42% (spring) Ephemeroptera in
reference streams in West Virginia. In a time-series study (April and October samples from
1996 to 2003) of a third-order stream affected by two smaller tributaries with MTM-VF, Kirk
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and Maggard (2004) found reduced % Ephemeroptera at the downstream site compared to an
upstream reference site, particularly in the years after mining ended. Likewise, Ephemeroptera
richness was significantly lower in mine-impacted streams (Merricks et al., 2007; Pond et al.,
2008; Pond, 2010). Also, using nonmetric scaling ordination and a nonparametric multiresponse
permutation procedure, Pond (2010) found that Ephemeroptera assemblages in reference streams
were significantly dissimilar (i.e., the assemblages were not the same) from those in
mine-impacted streams.
6.1.1.5.4. Odonata
An Odonata density metric was significantly lower at sites downstream of valley fills
than it was in streams without upstream valley fills (Hartman et al., 2005).
6.1.1.5.5. Plecoptera
The evidence for MTM-VF Plecoptera populations is weaker. In studies of taxa richness,
Plecoptera richness was lower in streams with mining in the watershed (2.7 genera) than in
streams without mining (6 genera) (Pond et al., 2008). Another study found no significant
difference in Plecoptera richness between sites downstream of valley fills compared to those
without upstream fill (Merricks et al., 2007).
A similar discrepancy was found in studies of relative abundance (i.e., percent
Plecoptera). Merricks et al. (2007) found lower Plecoptera percentages in sites downstream of
valley fills (4.5% in mined streams and 52% at a reference site). Armstead et al. (2004) found
decreased percentages in streams with valley fills sites in spring (11% in valley fills streams and
21% in reference streams) but found no difference in winter. Pond et al. (2008) did not detect a
difference in percent Plecoptera between streams with mountaintop mining in the watershed and
streams with no mining in the watershed. No difference was observed in a Plecoptera density
metric between streams with and without valley fills in Hartman et al. (2005).
6.1.1.5.6. Trichoptera
MTM-VF had mixed effects on Trichoptera populations in streams. When the stream
reach just downstream of the settling pond was sampled, the proportion of the macroinvertebrate
assemblage that was in the order Trichoptera was greater than the stream reaches upstream of the
pond and downstream of the valley fill or streams without mining (20% in mined streams just
downstream of the settling pond, 4.1% in mined streams upstream of the settling pond, and 3.7%
in reference watersheds, Merricks et al., 2007). As the distance downstream of the settling pond
increased, the proportion of the macroinvertebrate assemblages that was Trichoptera decreased.
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Two other studies found no difference among streams downstream of fills and those without fills
(Armstead et al., 2004; Hartman et al., 2005).
6.1.1.5.7. Ephemeroptera, Plecoptera, and Trichoptera (EPT) indices
Most field studies reported a reduction in commonly used measures of sensitive
macroinvertebrates, the aggregated EPT metrics, at sites downstream of MTM-VF. EPT
taxonomic richness was lower in two studies (Pond et al., 2008: EPT generic richness of 17.9 at
unmined sites and 8.9 at filled sites; Armstead et al., 2004: EPT taxa richness of 9 at unmined
sites and 6 at filled sites; Fulk et al., 2003: EPT family richness of 12-13 at unmined sites and 9
at filled sites in spring) and mixed results (decreased EPT richness in two valley fill streams and
no differences in two other valley fill streams) in another (Merricks et al., 2007). Hartman et al.
(2005) observed no differences in EPT richness between mined and unmined streams. An EPT
index was lower in streams in mined watersheds compared to measures in streams in watersheds
without mining activity (an average of 8.9 in mined sites and 21 in reference sites)
(Howard et al., 2001), and the percentage of the community comprised of EPT taxa was lower at
sites downstream of MTM-VF (Armstead et al., 2004; Merricks et al. 2007; Pond et al. 2008). In
a time-series study (i.e., April and October samples from 1996 to 2003) of a third-order stream
affected by two smaller tributaries with MTM-VF, percent EPT was generally lower at the
downstream site compared to an upstream reference site, particularly in the period since 1999
(Kirk and Maggard, 2004).
The mixed effects of mining on EPT aggregate measures likely reflect legacy land-use
differences, differences in location of sample sites (e.g., sampling close to a pond) and
taxonomic shifts within and among insect orders. Particularly important in these effects are
taxonomic shifts because of differing sensitivity among the three orders: Ephemeroptera,
Plecoptera, and Trichoptera. As described previously Plecoptera and Trichoptera, in general, do
not appear to be as sensitive to the effects of MTM-VF as Ephemeroptera. These differences in
sensitivity have been observed for other stressors, such as metals and low pH (Griffith et al.,
1995;Luomaetal., 2010).
6.1.1.5.8. Noninsect benthic macroinvertebrates
A density metric of noninsect macroinvertebrates was significantly lower in at sites
downstream of valley fills than in streams without fills (Hartman et al., 2005). A combined
measure of the percent Chironomidae and Oligochaeta was higher in streams in mined
watersheds compared to streams in watersheds without mining (63% in mined and 3% in
reference streams, Howard et al., 2001). However, bioassessment surveys, such as
Hartman et al. (2005), Howard et al. (2001), and the other studies discussed previously, do not
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generally sample one group of noninsect benthic macroinvertebrates, river mussels of the family
Unionidae, because of their highly clumped distribution in stream systems (Neves and Widlak,
1987). Many river mussels are already threatened or endangered, because of sedimentation,
construction of dams, and other alterations of rivers of this region and elsewhere (Layzer et al.,
1993; FWS, 2004).
6.1.1.5.9. Macroinvertebrate taxa dominance
Dominance metrics generally measure shifts in relative abundance from more sensitive
species to more tolerant species. Dominance of a community by a few organisms is expected to
increase with stress (e.g., Yuan and Norton, 2003). Armstead et al. (2004) observed that the
percentage of the macroinvertebrate assemblage that was the two most numerically dominant
taxa increased in streams with valley fill (i.e., 65 to 67%) as compared with reference streams
(i.e., 50 to 54%).
6.1.2. Fish
Fish community attributes, such as species richness, are widely used to evaluate stream
condition (Karr, 1981; Angermeier et al., 2000). Species richness and the number of benthic fish
species were consistently associated with site quality in Mid-Atlantic Highland streams
(Angermeier et al., 2000), where both attributes declined with increasing degradation
(Barbour et al., 1999). Mountaintop mining for coal and creation of valley fills has had a
harmful effect on the composition of stream fish communities. Comparison of streams without
mining in the watershed and sites downstream of valley fills in Kentucky (five unmined sites and
seven filled sites) and West Virginia (four unmined sites and two mined sites) indicates that
streams affected by mining had significantly fewer total fish species and fewer benthic fish
species than streams without mining in the same areas (Stauffer and Ferreri, 2002). A similar
pattern of fewer taxa in streams affected by mining was observed with species richness (median
of 6 in sites downstream of valley fills and 12 in unmined streams, Stauffer and Ferreri, 2002).
For example, in Kentucky, Stauffer and Ferreri (2002) observed sites downstream of valley fills
had a median richness of 7 fish species, compared to a median richness of 12 fish species in
streams without mining in the watershed. In these streams, the number of benthic fish species
was also lower in the sites downstream of valley fills (median = 1 benthic species) than in the
streams without mining in the watershed (median = 6 benthic species). Ferreri et al. (2004)
conducted another study that compared stream reaches downstream of valley fills in the Mud
River basin (i.e., 8 sites) with stream reaches without mining in the Big Ugly Creek basin (i.e.,
5 sites), both tributaries to the Guyendotte River of southwestern West Virginia and found the
same pattern of fewer total fish species (i.e., median = 17 in unmined reaches versus median = 8
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in mined reaches) and fewer benthic fish species (i.e., median = 6 in unmined reaches versus
median = 1.5 in mined reaches).
Fulk et al. (2003) used the Mid-Atlantic Highlands Index of Biotic Integrity (IBI) to
analyze fish data from 27 streams in West Virginia. The index is composed of several metrics
that are responsive to environmental and chemical stress, e.g., native intolerant taxa, native
Cyprinidae taxa, native benthic invertivores, percent Cottidae, percent gravel spawners, and
percent piscivore/invertivores were expected to decrease with increasing stress, and percent
macro-omnivore, percent tolerant fish and percent exotic fish were expected to increase with
increasing stress. In their study, Fulk et al. (2003) classified streams (i.e., no mining in the
watershed, sites downstream of valley fills, mountaintop mining in the watershed, sites
downstream of valley fills and with residential development in the watershed) and compared fish
metrics and the composite IBI among stream classes. IBI scores from the sites downstream of
valley fills were significantly lower than scores from sites without mining in the watershed by an
average of 10 points, indicating that fish communities were degraded in sites downstream of
valley fills. In their analysis, Fulk et al. (2003) found the reduced index score was caused by
fewer minnow species (median = 5.0 in unmined versus median = 4.3 in streams with valley
fills) and benthic insectivores (median = 6.0 in unmined versus median = 4.9 in streams with
valley fills) in sites downstream of valley fills compared to streams without mining in the
watershed. Index scores were also lower at sites with mining in the watershed compared to
scores from streams without mining in the watershed. Watershed size was also an important
factor in this analysis. Sites with mining and valley fills in small watersheds (<10 km2) showed
r\
more degradation than sites with mining and valley fills in large watersheds (>10 km )
(Stauffer and Ferreri, 2002; Fulk et al., 2003).14
6.1.3. Amphibians, Particularly Salamanders
It is well-known that salamanders are an important part of the aquatic vertebrate
assemblage in the central Appalachians (Davic and Welsh, 2004), particularly in the small
intermittent and permanent streams impacted by MTM-VF. Despite this and the suggestion that
salamanders be used for biomonitoring elsewhere (Welsh and Ollivier, 1998; Welsh and Droege,
2001; Ohio EPA, 2002), only one field study has been conducted to study the effects on
salamanders, and most field studies have concentrated on the more commonly used fish and
macroinvertebrates (see above). Williams and Wood (2004) quantified salamander diversity and
abundance in four second- and third-order reaches downstream from valley fills and in four
intermittent, first- and second-order reference reaches that were unimpacted by MTM-VF.
14Because larger watersheds typically have greater fish diversity than smaller watersheds, both studies adjusted their
analyses to account for the potential effect of watershed size.
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While overall species richness was similar in the streams (i.e., 7 species in valley fill streams
versus 8 species in reference streams), mean salamander abundance per 35-m stream reach was
greater in the reference streams (i.e., 25.7 ± 14.4 [mean ± SE], 8-66 [range]) than in the valley
fill streams (15.9 ± 9.5, 0-76). Salamander abundance was particularly correlated with the
number of rocks in the stream reach (r = 0.63, p < 0.001). Using leaf litter bags, 20 salamander
larvae were captured in the reference streams versus only 3 salamander larvae in the valley fill
streams.
6.2. EFFECTS ON ECOLOGICAL FUNCTION
To date, few studies have assessed the impact of MTM-VF on ecological function (e.g.,
biogeochemical cycling) in downstream habitats. One study, Fritz et al. (2010), compared
in-channel, standing crop biomass of coarse benthic organic matter (CBOM) and breakdown
rates and invertebrate colonization of oak leaf litter in channels associated with valley fills with
that in natural, forested streams unaffected by MTM-VFs. Three classes of channels or streams
in terms of flow duration were identified: ephemeral or intermittent channels constructed on
valley fills or permanent channels downstream of valley fills. They found that standing crop
biomass of CBOM was similar between channels associated with valley fills and natural streams
in the permanent and intermittent reaches, but CBOM was greatest in the natural ephemeral
streams and least in the constructed ephemeral channels on the valley fills. Leaf litter breakdown
rates, whether per day or per degree day, were similar in both types of ephemeral channels (i.e.,
on valley fills versus natural). In intermittent or perennial channels, leaf litter breakdown rates
were greater in natural, forested stream than in channels on or downstream from valley fills. The
densities of all invertebrates or just invertebrate shredders on the leaf packs were also greater in
the intermittent or perennial natural forested streams than in the intermittent channels on the
valley fills or the permanent channel downstream from the valley fill. Because differences in
water temperature among the sites are removed by the use of cumulative degree-days and the
differences in flow duration are partitioned among the three classes of channels, the differences
in leaf breakdown rates appear to be most related to conductivity. Beyond this study, additional
research is needed to better understand the effects of MTM-VF on ecological function in
downstream sites.
6.3. BIOLOGICAL CONDITION
The biological effects downstream of MTM-VF are consistent with generic narrative
descriptions of moderately to severely degraded condition (Davies and Jackson, 2006). The
biological condition gradient (BCG) is a framework that identifies 10 attributes of stream
ecosystems indicative of biological status ranging from pristine, natural condition (Tier 1) to
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severely degraded condition (Tier 6) (Davies and Jackson, 2006; see Figure 18). Evidence was
available to evaluate 3 of the 10 ecological attributes described in the BCG. Sensitive taxa,
specifically insect Order Ephemeroptera, are markedly diminished downstream of MTM-VF
(Tier 5). The spatial and temporal extent of detrimental effects is between the reach- and
catchment-scale (Tiers 4 to 6). Finally, the burial of the headwaters and the construction of
channels correspond with a loss of ecosystem connectance between 'some' and 'complete'
(Tiers 4 to 6). The attributes identified in the BCG highlight many data gaps—including
documenting the extent of regionally endemic taxa, reporting the relative tolerance of taxa to the
stressors specific to the region, identifying the presence of nonnative organisms, reporting the
condition of organisms and measuring ecosystem functions in both reference and MTM-VF
streams. The BCG provides a general framework and is intended to be locally calibrated by state
and regional scientists and resource managers. Local calibration would provide a useful
framework for describing the effects of MTM-VF and restoration efforts on stream condition.
Natural
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6.4. RELATIONSHIP BETWEEN BIOLOGICAL METRICS AND ENVIRONMENTAL
FACTORS
Five environmental variables associated with mining and valley fills are commonly
considered to potentially affect the ecological condition of downstream habitats: ion
concentration, heavy metal concentration, organic enrichment, changes to instream habitat, and
changes to upstream land use/land cover. This section describes associations between these
variables and biological characteristics (see Figure 19).
6.4.1. Ion Concentration
All studies report elevated ion concentration in MTM-VF (see Table 14), and most show
strong negative relationships between biological metrics and specific conductance and/or the
concentrations of individual ions (Howard et al., 2001; Stauffer and Ferreri, 2002; Fulk et al.,
2003; Hartman et al., 2005; Merricks et al., 2007; Pond et al., 2008; Pond, 2010; Timpano et al.,
2010).
Several studies of other types of mining discharges have reported associations with
conductivity. Conductivities ranging from 500-8,000 |iS/cm had a significant negative
correlation with the number of pollution sensitive taxa in benthic macroinvertebrate assemblages
(Soucek et al., 2000; Kennedy et al., 2003). In a study that followed effects downstream of an
alkaline mine discharge in a tributary of the Monongahela River in southwestern Pennsylvania
(Kimmel and Argent, 2010), fish species richness and total abundance declined with increases in
conductivity, although effects based on comparing species dissimilarity of the assemblage to a
reference site (sensu Courtemanch and Davies, 1987) were most apparent only when
conductivity levels exceeded 2,300 |iS/cm (Kimmel and Argent, 2010). However, the authors
also state that tributaries of the Monongahela River, in general, support a relatively
pollution-tolerant fish assemblage (i.e., only 2% classified as intolerant) because of the historical
impacts of coal mining, sewage discharges, agriculture, and urbanization (Kimmel and Argent,
2006). Ephemeroptera richness was negatively correlated with specific conductivity
(Hartman et al., 2005). Though Merricks et al. (2007) did not assess conductivity-
macroinvertebrate relationships among sites, they noted that sites with the highest levels of
conductivity, ranging between 2,657 to 3,050 |iS/cm, lacked Ephemeroptera. Pond et al. (2008)
performed the most complete analysis of ions and observed strong negative relationships
between specific conductance and their biological assessment measures, GLIMPSS (r = -0.90)
and WVSCI (r = -0.80), total generic richness (r = -0.74), EPT generic richness (r = -0.88),
number of Ephemeroptera genera (r = -0.90), the number of Plecoptera genera (r = -0.75), and
percent Ephemeroptera (r = -0.88). Of the sites having greater than 500 |iS/cm specific
conductance, 100% (20 of 20) had genus-level macroinvertebrate index scores indicative of
73
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; sediment retention
^treatment ponds
\/
chlorophyll a
total dissolved solids
tK+
tCa2+
t HCO3-
^ metals
(water&sediment)
tSe
t Ni
'Mvln
t Fe
precipitation
onto substrate
coating of benthic
organisms
s|/ Plecoptera
richness
macroinvertebrate
index quality
Figure 19. Macroinvertebrate and fish responses associated with stressors and treatment ponds in studies of
MTM-VF. Responses shown were observed in at least one of the reviewed studies. Manganese and iron precipitates
were directly observed. Other responses were either significantly different from sites upstream of ponds or were
strongly correlated with the stressor (i.e., absolute correlation coefficient >0.7). See Section 6.4 for additional details
and evidence.
-------
Table 14. Average ion concentration (reported as specific conductance) in
MTM-VF and reference streams reported in conjunction with biological
data. Range values are included when reported by the source literature. Standard
error values were not reported in the source literature.
Source
Hartman et al., 2005a
Howard etal., 2001
Merricks et al., 2007a'b
Pond et al., 2008
Pond, 2010
Stauffer and Ferreri, 2002°
Timpano et al., 2010
Units of
measure
muhm/s [sic]
|imhos/cm
US/cm
|iS/cm
|iS/cm
|imhos/cm
|iS/cm
Filled
n
4
8
3
27
20
8
17
Mean (range)
1,051
994 (420-1,690)
1,653 (991-2,720)
1,023 (159-2,540)
940 (161-2,390)
1,716(513-2,330)
13d (25-970)
Reference
n
4
4
1
10
44
9
3
Mean (range)
150
47 (30-66)
247
62 (34-133)
51 (16-159)
164(125-210)
NRd
"Averages calculated from reported values.
bValues taken from Site 1, which is the first site below valley fill and pond.
°Values reported were limited to the Mud River watershed.
dThe median value and not the mean was reported for the mined sites. No values were reported for the reference
sites.
biological impairment; 85% (17 of 20) had family-level macroinvertebrate index scores
indicative of biological impairment (Pond et al. 2008).
In an analysis of streams downstream from mining and valley fills in Virginia, Timpano
et al. (2010) observed strong negative relationships between specific conductance and the biotic
metrics: EPT taxa richness (r = -0.76), Ephemeroptera taxa richness (r = -0.71), Plecoptera taxa
richness (r = -0.72), total taxa richness (r = -0.50), and collector taxa richness (-0.58). Also, a
metric that increases with impairment, percent abundance of five most dominant taxa, exhibited
a positive relationship with specific conductance (r = 0.64, Timpano et al., 2010). In an analysis
of Kentucky streams, Pond (2010) found specific conductance to be negatively correlated with
percent abundance of Ephemeroptera (r = -0.72). Using a time-series data set (i.e., April and
October samples from 1996 to 2003) from just two sites in a third-order stream upstream and
downstream of its confluences with two smaller tributaries with MTM-VFs, Kirk and Maggard
(2004) found a weaker negative correlation with specific conductivity and WVSCI scores
(r = -0.34) but a strong negative correlation with percent abundance of Ephemeroptera
75
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Pond et al. (2008) further demonstrated a decline in number of Ephemeroptera taxa and
their proportion of the assemblage when conductivity levels exceeded approximately 500 jiS/cm.
Using a nonparametric changepoint analysis using the deviance reduction method, Pond (2010)
found a threshold-type response for percent abundance of mayflies centered on a specific
conductance of 175 |iS/cm (confidences limits: 124-336 |iS/cm). IDS or specific conductance
also had strong negative correlations with biological metrics (Stauffer and Ferreri, 2002;
Pond et al., 2008; Timpano et al., 2010). Although the strength of correlations of richness
metrics and macroinvertebrate indices with individual ions varied somewhat, strong negative
correlations (i.e., absolute value of r > 0.7) were found in at least one of these studies with
HCOs , Ca2+, SO42 , Mg2+, and K+. While these studies do not provide enough detail to
elucidate the mechanistic relationship of biological degradation to ion concentration, the pattern
clearly suggests a strong association between the two.
Additional insights on possible mechanisms can be found in the physiological literature
on osmoregulation (e.g., Bradley, 2009). Although earlier literature emphasized regulation of
osmotic pressure and cell volume as mechanisms by which salts affect freshwater organisms
(e.g., Kapoor, 1979; Pierce 1982), regulation of internal ionic concentrations has been
emphasized more recently. Freshwater vertebrates, including fish and amphibians, as well as
invertebrates, such as aquatic insects, crayfish, and unionid mollusks, are invariably hyper-
regulators (Kirschner, 1970; Dietz and Branton, 1975; Dietz, 1979; Goss et al., 1992; Harvey,
1992; Henry and Wheatly, 1992; Cooper, 1994; Wheatly and Gannon, 1995; Ehrenfeld and
Klein, 1997; Perry, 1997; Kirschner, 2004). These animals maintain internal concentrations of
ions, such as Na+, K+, Cl , Mg2+, and SC>42 , that are greater than the concentrations of these ions
in unimpaired freshwaters. These ions are moved into the organism against concentration
gradients using several mechanisms. In particular, H+ (or combined with ammonia to form
NH4+) and HCOs are by-products of respiration and are exchanged for Na+ and Cl ,
respectively, to move these ions into the organism (Evans, 1975; Dietz, 1979; Grosell et al.,
2002; Kirschner, 2004). In experiments with goldfish (Carassius auratus), the addition of
HCOs to the external medium inhibited the uptake of Cl (Maetz and Garcia Romeu, 1964).
High concentrations of HCOs downstream of valley fills might similarly be inhibiting uptake of
Cl and export of HCO3 . In addition to reducing internal Cl concentrations, the excess internal
HCOs might also alter the acid-base balance within the organisms (Goss et al., 1992; Henry and
Wheatly, 1992).
6.4.2. Specific Metals and Selenium
Though contributing to overall ion concentration, the concentrations of individual metals
were negatively correlated with many of the biological metrics in streams. Hartman et al. (2005)
76
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found strong negative correlations (r ranged from -0.70 to -0.98) between macroinvertebrate
metrics and metals. For example, Ephemeroptera richness was negatively correlated with Fe,
Mn, and Ni; EPT taxa richness was negatively correlated with Mn and Ni. That study, as well as
Merricks et al. (2007), reported that metal concentrations were higher in mining streams with
biological degradation than in reference streams. Ephemeroptera and Plecoptera richness were
both strongly and negatively correlated with Se concentrations (Pond et al., 2008). These results
suggest that elevated metal concentrations associated with mine-impacted streams might
contribute to differences in stream biota.
6.4.3. Organic and Nutrient Enrichment
Two studies suggest a possible association between differences in biota and organic
enrichment in streams affected by MTM-VF (see Figure 19). Pond et al. (2008) found strong
correlations (i.e., absolute Spearman correlation coefficients greater than 0.7) between
NC>2 + NOs and the number of Ephemeroptera taxa; Plecoptera taxa; the sum of Ephemeroptera,
Plecoptera, and Trichoptera taxa; and the sum of all taxa. In addition, the relative abundance of
Ephemeroptera was strongly and negatively correlated with NC>2 + NOs. However, total
phosphorus levels were below detection limits at all sites. Merricks et al. (2007) evaluated
changes in the composition of the structural and functional composition of the macroinvertebrate
assemblages downstream of the settling ponds in mined streams along with growth of the filter-
feeder, Corbicula (Asiatic clam) to assess potential organic enrichment. Stations just
downstream of the ponds had significantly greater proportions of collector-filterers compared to
stations upstream of the ponds, and these proportions decreased further downstream. Similar
patterns in the proportion of the macroinvertebrate assemblage that were Trichoptera and in the
growth of Corbicula were observed. Moreover, water column chlorophyll a concentrations were
generally high in the settling ponds, indicating the presence of algae in the pond that is food for
filter-feeders like Corbicula and many Trichoptera species (e.g., the Hydropsychidae).
Merricks et al. (2007) also noted that the HBI was elevated at all fill-influenced sites compared
to a reference site. The Fffil was developed to respond to a nutrient and organic pollution
gradient, but it is also responsive to other stressor gradients, including increased fine sediments
and specific conductivity (Paybins et al., 2000).
6.4.4. Instream Habitat
There was little evidence of an association between changes in macroinvertebrate
community metrics and characteristics of instream habitat at sites downstream of MTM-VF. As
discussed in Section 6.1.1.4, decreases in macroinvertebrates that cling to rocks (clingers) were
observed in one study (Howard et al., 2001). Similar decreases have been associated with
77
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increases in fine sediments in a regional study, suggesting that the observed declines might be
associated with changes in sediments. (Pollard and Yuan, 2009). However, the field studies we
reviewed found no systematic relationship between macroinvertebrate metrics and habitat
assessment measures including measures of fine sediments and turbidity (Howard et al., 2001;
Hartman et al., 2005; Merricks et al., 2007), which might suggest that habitat characteristics
were not all that different between reference and mined stream sites. Total rapid bioassessment
procedure habitat scores were correlated with macroinvertebrate indices in the study by
Pond et al. (2008). However, individual subscores show only weaker correlations, and the
investigators did not observe excessive sedimentation in sites downstream of valley fills.
Iron can precipitate out of the water onto organisms and substrates, clogging gills and
degrading habitat by cementing together sediment particles and promoting the growth of
Fe-depositing bacteria (Vouri, 1995). Fritz et al. (2010) reported the presence of FeSC>4
precipitates and Fe-depositing bacteria at all study sites downstream of MTM-VF; neither was
observed at reference sites. Pond (2004) noted the presence of both Mn and Fe precipitates on
organisms downstream of MTM-VF (see Figure 20). We did not find any studies documenting
the types or prevalence of precipitates downstream of MTM-VF or any studies that distinguished
the effects of these precipitates from water quality impacts.
Figure 20. Mn (black) and Fe (orange) deposits on a caddisfly collected
downstream of a mountaintop mine and valley fill.
Source: Pond (2004).
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6.4.5. Disturbance and Loss of Upland Habitat
In addition to the effects described above, changes in upland and headwater areas of the
watershed could influence macroinvertebrate composition in downstream habitats. Headwater
streams are critical to downstream ecological condition and their alteration, as in mountaintop
mining and valley fill activities, could impact the integrity and the sustainability of downstream
habitats. As reviewed in Section 3.4, headwater streams provide downstream habitats with
water, nutrients, food, and woody debris (Gomi et al., 2002; Wipfli et al., 2007). Moreover, the
physical structure of headwater streams in the landscape can affect populations and communities
of stream organisms by influencing the movement of sediment, chemicals, and individuals to
downstream reaches within the network (Lowe et al., 2006).
Watershed characteristics and activities greatly affect the structure and the function of
streams. Houser et al. (2005) showed that intact riparian zones were not sufficient to protect
streams from the effect of upland disturbance. They examined the effects of upland soil and
vegetation disturbance on ecosystem respiration and found lower ecosystem respiration rates in
streams with higher levels of upland disturbance. This is relevant because mountaintop removal
represents a significant disturbance to the vegetation and soil characteristics in upland areas. As
a result, even when downstream reaches and associated riparian areas of a stream appear intact,
as in the case of MTM-VF, they could incur significant impacts from mountaintop removal
occurring upstream.
6.5. CUMULATIVE EFFECTS
There is little evidence in the peer-reviewed literature of cumulative impacts of mining
on downstream ecology. Fulk et al. (2003) found no evidence of additive effects of multiple
mines on the fish IBI. In another MTM-VF study, Pond et al. (2008) reported no evidence of a
significant relationship between the number of upstream valley fills and macroinvertebrate
indices.
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7. RECLAMATION, MITIGATION, AND RECOVERY
In the following section, we address reclamation and mitigation efforts following
MTM-VF (see Figure 21). In particular, we examine (1) the dominant post-SMCRA reclamation
practice of seeding with grasses; (2) the Forestry Reclamation Approach; (3) the creation of
channels on valley fills; (4) the use of erosion control structures and creation of wetlands; and
finally (5) enhancing stream structural heterogeneity and riparian areas below the valley fill. We
discuss these practices as they relate to streamflow, water quality, and aquatic communities.
This discussion is limited to on-site reclamation and mitigation techniques. For a discussion of
off-site mitigation efforts, see Chapter III and Appendix D of the PEIS (U.S. EPA, 2003, 2005).
7.1. RECLAMATION OF MTM-VF SITES
7.1.1. Overview
Prior to SMCRA (30 U.S.C. § 1231, etseq.) passed in 1977, mining practices left large
areas of unstable land and eroding hillslopes that impaired streams and created human health
risks from mudslides and pollution. That history, plus concerns about the much larger volumes
of blasted rock and debris being produced by then-new mountaintop removal technology, led to
early SMCRA enforcement priority on stability and flood control. Reclamation techniques
developed prior to 2000 focused on regrading, soil compaction, fast-growing herbaceous
vegetation and stabilization, rather than reforestation or stream restoration for protection of water
quality. Since 2000, reclamation techniques have been developed that seek to restore some of
the productivity and ecological function of native forests. These techniques, termed the Forestry
Reclamation Approach (FRA), are based in part on research and extension efforts of the
Appalachian Regional Reforestation Initiative, a coalition of groups formed to promote
reforestation of Eastern coal mine sites (http://arri.osmre.gov) (Skousen et al., 2009).
Under SMCRA, reclamation is not considered complete until water quality leaving the
site complies with CWA standards without additional treatment. Among other requirements,
SMCRA and OSMRE regulations stipulate that mine operators:
minimize disturbances and adverse impacts on fish, wildlife, habitat, and hydrologic
balance;
recover the approximate original contour and vegetation in mined areas; and
restore or approximate the original stream channels and riparian vegetation in permanent
constructed stream diversions (30 U.S.C. 1260 and 1265; 30 C.F.R. 816.43, cited in the
PEIS, Chapter II) (U.S. EPA, 2003, 2005).
80
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;on-site reclamation & ^
stream enhancement J
oo
other watershed characteristics,
including: geology, current &
historical land use, mining intensity
on-fi recamation
be ow-fi stream enhancement
channel
enhancements
stream
reconstruction
stormwater retention
riparian
enhancements
flood & erosion control
structures (including
sedimentponds)
Channel Design
Channel Design
grass
reclamation
wetland
creation
forest
reclamation
floodplain
connections
f meanders, pools, riffles
I steps, substrate
creation of
L ephemeralchannels J
Forestry
Reclamation
Approach
(FRA)
m-stream structures
weirs, root wads, etc
creation of
intermittent channels
inputs to stream
Aflow& sediment
regime
T perennialstream
habitat
T intermittent
A decomposition
^temperature
lenticamphibians
& invertebrates
T tolerantinvertebrates"^
Figure 21. Observed and expected effects of on-site reclamation and stream mitigation efforts. See Section 7 for
more discussion and evidence.
-------
To insure the implementation of an approved reclamation plan, activation of a mining permit
requires posting of reclamation bonds. If the coal operator forfeits its reclamation
responsibilities, these bonds provide funds for the government to complete the work (SMCRA,
30 U.S.C. § 1259). Reclamation bonds are released upon inspection in three phases:
Phase 1: released after completion of backfilling, placement of homogenized or
crush-rock topsoils and contour regrading.
Phase 2: released upon completion of revegetation activities.
Phase 3 (final): released after the mine site has been inspected and accepted as being
satisfactorily reclaimed to the approved postmining land use (i.e., meets all performance
standards and the approved permit plan) (U.S. EPA, 2003, 2005).
Historically, release of bonds at a given site typically occurred within 5 years after completion of
reclamation and was based on percentage of land covered by fast-growing grasses or legumes
(Holl and Cairns, 1994; U.S. EPA, 2003, 2005).
7.1.2. Reclamation with Grasses and Pasture
Reclamation techniques developed post-SMCRA traditionally focused on regrading, soil
compaction, fast-growing herbaceous vegetation, and stabilization. This was done primarily for
erosion control. As a result, vegetative cover at most reclaimed mine sites consists of rapidly
growing grasses and legumes, serving as a low-cost, low-risk option for reclamation bond release
(Loveland et al., 2003). Pasture and hay lands planted to meet the legal requirements of
reclamation could be viable while maintained but might collapse when agronomic practices are
neglected after bond release (Barnes et al., 2003).
Soil compaction and competition from herbaceous plants slows the reestablishment of
forests on reclaimed mine sites (Bradshaw 1997; Skousen et al., 2009). Minesoils are mixtures
of soils, debris, and fractured rock overburden that are spread on reclaimed surfaces to support
plant growth (U.S. EPA, 2003, 2005). Compaction of minesoils with the use of heavy equipment
during valley fill and reclamation is identified as one of the chief factors limiting the
establishment, growth, and survival of trees (Skousen et al., 2006, 2009). As a result,
establishment of mid- to late-successional trees could take decades (Skousen et al., 2009).
Evidence suggests that the reclamation approach of heavy soil compaction and planting
with grasses largely fails to ameliorate either the hydrologic or water quality impacts of
MTM-VF. Ferrari et al. (2009) modeled the flood response in the mined watershed of Georges
Creek in Western Maryland and found parallels to what would be expected for urban settings
82
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with large areas of impervious surface. Infiltration rates in reclaimed sites are typically 1 to
2 orders of magnitude smaller than for undisturbed forest (Negley and Eshleman, 2006). The
field studies of downstream conditions reviewed in Section 4 (Impacts on Water Quality) were
conducted 3 to 24 years after permitting, and field studies reviewed in Section 6 (Impacts on
Aquatic Ecosystems) were done 3 to 15 years after reclamation. The results indicate that
reclamation efforts post-SMCRA, while providing stabilization, do not eliminate the deleterious
effects of degraded water quality associated with effluent from MTM-VF.
Reclamation with grasses can also alter aquatic communities by decreasing shade and
organic inputs to the stream. In small forested watersheds, overhanging trees provide organic
matter inputs, while simultaneously reducing photosynthesis by autotrophic organisms (Vannote
et al. 1980). This dual effect makes organic inputs the primary source of energy flow into the
food web of these streams. In a small headwater stream near Louisville, Kentucky, for example,
macroinvertebrate communities relied almost exclusively on leaf inputs (Minshall 1967).
Excluding litter artificially from the riparian zone changed the food web structure of a
mountainous headwater stream in North Carolina (Wallace et al. 1997).
Finally, non-native grasses and legumes are used for reclamation at most sites (U.S. EPA
2003, 2005, Table 3.J-1) and the disturbance of riparian zones has been linked with increased
invasion of non-native plant species (Richardson et al. 2007). Non-native riparian species may
in turn modify inputs to the stream (Richardson et al. 2007). However, leaf litter from Japanese
knotweed (Fallopiajaponica), an invasive species found in central Appalachia, did not alter
macroinvertebrate assemblages and leaf decomposition rates in an Idaho stream (Braatne et al.
2007).
7.1.3. Forestry Reclamation Approach (FRA)
Recognizing the limitations of these reclamation efforts, the FRA is a set of techniques
that has emerged from research conducted over the last several decades to promote the regrowth
of forests on reclaimed mine lands (Skousen et al., 2009). Forests were the dominant cover in
Appalachia prior to mining activity, and the FRA is being increasingly employed. There are
considerable upland benefits to reestablishment of native forests, and there is a large body of
literature on tree establishment and growth on reclaimed mine sites (e.g., Ashby, 1997;
Andrews et al., 1998; Brinks et al., 2011). This report, however, focuses exclusively on the
effects on water quantity, quality, and aquatic communities.
As noted previously, the establishment of trees on reclaimed sites requires
low-compaction soils. FRA techniques call for the loose placement of spoil to at least a depth of
4 ft for the establishment of trees (Skousen et al., 2009). Taylor et al. (2009a, b) measured the
hydrological characteristics of five test cells created on a surface mine on the Cumberland
83
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Plateau (Ecoregion 68, southwestern Appalachians) of Kentucky. Each test cell contained one of
three different types of mining spoil consisting of weathered or unweathered sandstone or a
mixture of sandstones and shale. The spoil was loosely dumped following specifications
outlined in the FRA (Taylor et al., 2009b). They found that loose placement of mine spoils
resulted in greater precipitation infiltration rates and low peak discharge rates. These discharge
rates were lower than those from a small, steeply sloped, forested catchment (Taylor et al.,
2009a, b). Thus, FRA techniques could help mitigate high downstream peak flows and flooding
risk in reclaimed watersheds (Taylor et al., 2009b). To our knowledge, there has been no peer-
reviewed study of the impacts of the FRA on water quality. Water infiltrating through the loose
spoil will potentially still have elevated concentrations of ions and some metals. However, the
effects of forest reclamation on water quality need to be empirically tested.
Besides influencing streamflow, forest reclamation in riparian areas provides shade,
temperature moderation, and critical organic inputs to aquatic food webs. Further, the planting
of trees rather than grasses promotes habitat for stream-side amphibians. For example, a study of
reclaimed mine sites replanted with grasses found a general decline in salamander populations
with a concomitant increase in reptiles, particularly snakes, compared to intact forests (Wood et
al. 2001). This was likely due to the drier conditions of the grasslands (Wood et al. 2001).
Given the positive impacts of replanting native forests, we recommend further study regarding
the effects of the FRA on water quality and biota.
7.2. MTM-VF MITIGATION EFFORTS
7.2.1. Overview
In addition to reclamation of postmining sites, placing overburden into stream valleys, if
deemed allowable, necessitates mitigation plans. The USAGE uses CWA Section 404(b)(l) to
evaluate proposals to convert waters of the United States to dry land (U.S. EPA, 2003, 2005).
The preferred alternative is to avoid placing fill in streams; where avoidance is not possible, fills
must be minimized. In either case, the proposal must not result in significant degradation to
waters of the United States. USAGE requirements on Section 404 CWA permits strive for no net
loss of aquatic functions (U.S. EPA, 2003, 2005).
Mitigation plans are developed prior to the start of mining and involve the use of stream
assessment methods to evaluate stream quality before and after impact. Fills resulting in
minimal impact, as determined by local regulatory agencies and the USAGE, can be authorized
by a Nationwide Permit to expedite the review process. Fills causing more than minimal impact
undergo a more detailed individual review. Further, the cost of compensatory mitigation is
higher, and permits are at greater risk of being delayed or denied when valued aquatic resources
are at stake (U.S. EPA, 2003, 2005).
84
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Mitigation activities can be conducted on- or off-site of the permit area. On-site
mitigation activities are preferred, but off-site projects might be the only option for lost
mountaintop ecosystems and are common in MTM-VF permits. Compensatory mitigation plans
include a variety of actions, as indicated below:
• Creating channels using natural stream design techniques to replace streams that have
been filled.
• Restoring riparian resources (e.g., revegetation, wetland creation, and floodplain
connectivity).
• Enhancing or improving existing stream channels (e.g., riffles/pools, dredging, sinuosity,
and bank stabilization).
• Improving fisheries habitat (e.g., shading, increasing habitat heterogeneity, and aeration
through riffles or other natural means).
• Controlling sediment and pollution (e.g., reclamation of abandoned mine lands and
remediation of other adverse environmental conditions within the watershed, anoxic
limestone drains, drums, flumes, and other passive treatment systems).
• Reforesting areas adjacent to mining sites.
• Removing stream encroachments (e.g., roads, crossings, ponds, or other fills).
Below, we address several of these on-site mitigation activities. We examine the extent to which
created channels on the fill might replace functions of streams lost from MTM-VF and address
the potential for natural stream design and wetland creation to reduce or minimize MTM-VF
impacts. Lastly, we discuss stream and channel enhancement and riparian improvements below
the fill.
7.2.2. On-Fill Mitigation Efforts
7.2.2.1. Constructed Channels
As described in Section 2.1, valley fills generally have rock drains and either a central
channel or a set of perimeter channels on or along the fill, which then discharge into sediment
retention ponds. The combination of ditches, channels, and ponds are designed to convey runoff
for large (e.g., 100-year) storm events. Published studies are generally lacking on whether these
ditches and channels can adequately replace the hydrology of natural streams buried by valley
fill. Despite this lack of study, some conclusions can be drawn from studies of intact, forested
catchments. It should be recognized that stream restoration techniques were developed to restore
one or more features of an existing stream with its basic structure intact, not to create streams
85
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starting from scratch (Palmer, 2009). Creation of intermittent and perennial streams on filled
areas is difficult, in general, due to the inability to capture sufficient ground-water flows to provide a
source of streamflow. Lastly, there is no evidence that these channels improve the water quality
impacts of MTM-VF or provide habitat invertebrate communities in intermittent or perennial
reaches.
The hydrology of a small-order stream is heavily dependent on its connections with its
riparian areas, surrounding uplands, and, in perennial or intermittent streams, groundwater.
When it rains on an intact, forested watershed, water infiltrates into the soil and moves vertically
and laterally depending on subsurface strata (see Figure 5). Further, natural forest soils typically
have extensive networks of macropores (spaces in the soil larger than 0.05 mm in diameter),
created by old root channels, earthworm and animal burrowing, and freeze-thaw events. These
macropores are sites of preferential water flow through the soil, important for the movement of
water to the stream channel (Tsuboyama et al., 1994). Steep hillslopes convey water by these
shallow subsurface flow paths to the stream (McGuire et al., 2005; McGuire and McDonnell,
2010). In areas of porous bedrock, local groundwater inputs to the stream are also important
(Winter, 2007). These natural linkages from the surrounding landscape influence the stream by
providing not only stormflow but baseflow during dry periods. The hydrologic flow paths and
the amount of time in which the water moves through these flow paths (i.e., residence time) also
significantly alter the chemistry of the water entering the stream.
When a mountain is leveled and a valley filled, the hillslopes, subsurface flows, and
groundwater exchanges that supported its small streams are permanently dismantled. The degree
to which a buried stream was an expression of these connections is likely to determine whether
its hydrology can be replaced by a recreated channel. Created channels on valley fills are likely
ephemeral or intermittent because they are elevated above the local water table and do not
receive significant, year-round groundwater inputs. During a storm event, the created channels
might act similarly to an ephemeral stream by conveying water and materials. Where a
headwater stream had shallow subsurface and groundwater connections, the hydrology of that
stream cannot be replaced. For example, the lost hydrologic function of a perennial stream
receiving year-round groundwater inputs cannot be replaced by an ephemeral or intermittent
channel. Replacing a perennial stream with an ephemeral channel would cause a shift away from
continuous streamflow to one punctuated only by stormflows.
Besides their hydrologic effects, constructed channels and diverted surface flows on
valley fills fail to restore water quality and the biological diversity of natural headwaters,
particularly in intermittent and perennial reaches. Fritz et al. (2010) found that intermittent and
ephemeral constructed channels did not reduce the conductivity of water sampled at the base of
the valley fill when compared to forested, reference streams. Further, though peer-reviewed
86
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evidence is limited, the available data show that biological assemblages colonizing reclaimed
surface waters differ from those of natural headwaters in predictable ways: loss of headwater-
specific taxa; increase in tolerant taxa; and presence of taxa adapted to ponds or turbid,
slow-moving water (Kirk, 1999; Green et al., 2000; U.S. EPA, 2003, 2005). Total invertebrate
density observed on leaf packets was significantly higher in forested streams versus constructed
channels or streams at the base of the valley fill (Fritz et al., 2010). Similarly, total taxa richness,
EPT richness, and litter decomposition rates (a measure of stream function), were generally
higher in forested, perennial, and intermittent reaches compared to perennial streams at the base
of the valley fill or intermittent constructed channels on the fill (Fritz et al., 2010). In contrast,
decomposition rates and leaf-packet invertebrate diversity did not differ between ephemeral
forested streams or ephemeral channels (Fritz et al., 2010).
In addition to invertebrates, salamanders can be useful indicators of water quality due to
their small home ranges and relatively stable populations when undisturbed (Welsh and Ollivier,
1998). Many stream salamanders require ephemeral and intermittent streams in forested habitats
to maintain viable populations (see Section 3.3). In one study conducted in southern West
Virginia, streamside salamander abundance was found to be significantly higher in reference
streams compared to those below valley fills (Williams and Wood, 2004). Moreover, using leaf
litter bags, 20 salamander larvae were captured in the reference streams versus only 3
salamander larvae in the valley fill streams (see Section 6.1.3) (Williams and Wood, 2004). In
contrast to the overall findings of the study, one restored reach supported a large salamander
population, possibly due to a strong positive correlation with the numbers of rocks in the stream
(Williams and Wood, 2004).
7.2.2.2. Natural Channel Design
Creating channels using natural channel design techniques is one mitigation activity that
is currently being investigated for the purposes of reestablishing streams on mined lands
(Harman et al., 2004). Natural channel design attempts to use the characteristics found in
undisturbed streams to promote channel stability and habitat for aquatic organisms. For
example, steep headwater streams typically exhibit vertical drop and scour-pool features
(Rosgen, 1994). Constructing similar stream-bed morphologies in created channels might more
closely mimic natural surface flows.
One of the main goals of natural channel design is to balance sediment export and
accumulation, preventing excessive erosion and rapid channel migration (Harman et al., 2004).
Given this goal, it may be reasonable to expect natural channel design to reduce sediment export.
However, there is no evidence to suggest a mechanism by which natural channel design might
significantly reduce elevated conductivity, selenium, or metal concentrations. Absent an
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improvement in water quality, the enhancement of stream habitat through natural stream design
might have little impact on stream biota (see Section 7.2.3.1).
7.2.2.3. Erosion Control Structures and Constructed Wetlands
Both erosion control structures and constructed wetlands can reduce sediment runoff,
stormflows, and improve water quality. Erosion control structures include riprap or rock-lined
channels and sediment ponds. Sediment ponds can improve water quality by removing
sediments, suspended solids, and metals (U.S. EPA, 2003, 2005; for a discussion of the effects of
ponds on sediment loads, see Section 4.4). Despite this potential, water chemistry data indicate
that these ponds fail to prevent downstream water quality degradation (see Section 4), given that
sediment ponds are generally present on MTM-VFs as currently constructed.
Constructed wetlands also have the potential to improve water quality. Wetlands can
store water, reducing peak stormflows—though this can vary by wetland type
(Bullock and Acreman, 2003). Constructed wetlands have also been found to reduce heavy
metals, excess nutrients, and total suspended solids (Scholes et al., 1998; Hench et al., 2003;
Sheoran and Sheoran, 2006; Vymazal, 2007). Heavy metal removal via wetlands has been
shown to be effective in acid mine drainage treatment (Sheoran and Sheoran, 2006). However,
constructed wetlands in a waste-water treatment facility failed to reduce conductivity values, and
wetlands might become less effective in removing materials over time (Hench et al., 2003). In a
study of wetlands on mined lands, almost all the constructed wetlands reduced sediments, while
a smaller number provided additional water quality and wildlife benefits (U.S. EPA, 2003,
2005). Wildlife found in wetlands and sediment ponds are generally lentic organisms, not those
typically found in Appalachian headwater streams (U.S. EPA, 2003, 2005).
7.2.3. Below-Fill Mitigation Efforts
7.2.3.1. Riparian Restoration and Stream Channel Enhancement
Besides mitigation efforts on the valley fill itself, mitigation activities also target the
stream reach issuing from below the valley fill. These activities include riparian restoration and
stream channel enhancement. Riparian restoration can consist of planting of riparian forests or
reconnecting the stream to its floodplain. Stream channel enhancement includes natural channel
design techniques of adding structural heterogeneity (e.g., adding boulders and logs, creating
riffles and pools) or reforming stream meanders. The planting of riparian forests increases bank
stability and reduces erosion, while also adding shade and reducing high temperature extremes.
Unconfined stream banks that are devoid of vegetation are more susceptible to channel widening
and erosion (Naiman and Decamps, 1997), and woody debris or other instream structures can
dissipate water energy and store sediments. Further, riparian zone inputs of leaf litter and wood
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are critical to aquatic food webs, particularly in headwater catchments (see Section 3)
(Tank et al., 2010), and provide critical habitat for amphibians (Wood et al. 2001).
To our knowledge, there is no peer-reviewed, published literature addressing whether
riparian restoration or stream enhancement improve water quality and biological assemblages
after MTM-VF. These approaches are likely to increase organic matter retention and reduce
sedimentation and erosion—but not alleviate elevated ion, selenium or metal concentrations.
Therefore, they are likely to have a limited effect on the recovery of instream biota. Though the
disturbance of urbanization differs from MTM-VF, many studies in urban settings found that
adding structural heterogeneity and restoring riparian areas had limited impact on stream biota
when water quality remained degraded (Northington and Hershey, 2006; Sudduth and Meyer,
2006; Bernhardt and Palmer, 2007). The placement of boulders in urban channels has been
shown to increase residence times of organic matter (Lepori et al. 2005). Harrison et al. (2004),
however, concluded that restoring structural heterogeneity failed to effectively restore
macroinvertebrate communities, likely due to the over-riding influence of poor water quality. In
a similar fashion, in an acid mine drainage study, water chemistry limited the recovery of benthic
organisms beyond any effect of chemical precipitates on the stream substrata (DeNicola and
Stapleton, 2002). Biotic assemblages might be able to recover from abrupt downstream
disturbances relatively quickly (i.e., a few months to a few years, Wallace, 1990) if connectivity
to undisturbed reaches is maintained (Detenbeck et al., 1990; Niemi et al. 1990; Blakely et al.
2006). However, the recovery of stream biota will be limited in cases where poor water quality
persists (Niemi et al. 1990).This literature suggests that successful MTM-VF mitigation
approaches will need to address water quality before other restoration efforts have a substantial
effect.
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8. SUMMARY, INFORMATION GAPS AND RESEARCH OPPORTUNITIES
8.1. A CONCEPTUAL MODEL OF THE IMPACTS OF MTM-VF
Mountaintop mining is a type of surface mining that is currently used, particularly in the
central Appalachian coalfields, to mine relatively low-sulfur coal. This review focuses on the
effects of mountaintop removal operations, which use explosives to remove all or portions of the
entire mountain or ridge top, to expose and mine one or more coal seams.
Aquatic ecosystems downstream of MTM-VF differ in significant ways from streams that
receive little human influence. Observations of which organisms are lost and how the streams
they live in are altered can improve our understanding of how MTM-VF result in these impacts.
Our review of the peer-reviewed, published literature and the PEIS is summarized here in a
conceptual model that traces the impacts of MTM-VF on aquatic ecosystems (see Figure 22).
Impacts begin with the preparation of the mountaintop or ridge top site. Access roads are
built, all trees are cleared, and some topsoil is stockpiled for future use in reclamation. Then,
explosives are used to blast the top of the mountain or ridge to expose and mine one or more coal
seams. The coal that is removed is processed and transported to market; we did not include the
impacts of these latter processes in our review. Instead, we follow the fate and impacts of the
excess overburden and the mined site that remains.
When the mountaintop is removed, so are the springs and ephemeral, intermittent, and
small perennial streams that comprise the headwaters of rivers. When trees are removed from
the valley fill footprint prior to construction of the fill, it also removes habitat for amphibians
that move between forest and stream during their life cycle. The excess overburden is disposed
of in constructed fills in small valleys or hollows adjacent to the mountaintop site. When the
valley fill is constructed, the headwaters beneath the footprint are buried, and most organisms
that lived there are killed. These headwaters support diverse biological communities of aquatic
invertebrates, such as insects, and vertebrates, including fish and salamanders, that are often
distinct from the species found in further downstream in the stream system. Coldwater,
permanent headwater reaches can be spawning, and nursery areas for native brook trout
(Salvelinus fontmalis, and headwater reaches are also primary habitat for other native fish such
as the creek chub (Semotilus atromaculatus), blackside dace (Phoxinus cumberlandensis\
southern redbelly dace (Phoxinus erythrogaster\ arrow darter (Etheostoma sagitta), and
orangethroat darter (Etheostoma spectabile; Meyer et al., 2007). Intermittent and permanent
headwater reaches, particularly those too small to support fish, support numerous amphibian
species. This particularly includes salamanders, of which nearly 10% of the global diversity is
found in streams of the southern Appalachians. These streams also provide habitat for diverse
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Figure 22. Observed and expected effects of MTM-VF on aquatic ecosystems (narrative description in
Section 8.1). (Figure formatted for printing on 11" by 17" paper.)
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assemblages of aquatic insects, some of which are unique to these headwater reaches. When a
headwater stream is buried under a valley fill, most of the organisms that inhabit it are
eliminated. As multiple streams in a mountain range are buried, the distance between the
headwaters that remain becomes greater, hindering the movement of biota that is required to
sustain populations. The hyporheic habitat is also buried, eliminating the interface of
groundwater and surface water that harbors the microbial community responsible for much of the
nutrient processing. Export of nitrogen downstream is expected to increase.
Both the water and sediment discharged into downstream ecosystems are altered by
MTM-VF. Water runoff is increased when the forest is cleared for the mine and valley fills. The
compacted, bare soils, which result from the removal of the overburden and coal, form a large
impervious surface that increases surface runoff. Depending on the degree to which they have
been compacted, the valley fills can ameliorate the effects of moderate precipitation events on
high flows, because they temporarily store water from surface flows and direct precipitation.
However, precipitation from more intense storms might produce greater stormflows, if
compaction of the fill surface and mined area by heavy equipment reduces infiltration of
precipitation and increases overland runoff.
Surface runoff is diverted into ditches and sedimentation ponds, replacing natural
subsurface flow paths. Under most circumstances, the sedimentation ponds appear to be
effective at settling out fine sediments; habitat measures were not strongly related to
macroinvertebrate responses. The ponds themselves change the predominant source of energy in
the downstream systems from tree leaves to algae. Organisms that feed on leaves (shredders)
decline; organisms that feed on algae (filterers) increase.
The overburden in backstacks and valley fills has increased surface area available for
water contact with rock particles, and the water that emerges has higher concentrations of major
ions and some trace metals, including selenium. Native mayflies are consistently among the first
to disappear as concentrations of ions and trace metals increase. Most studies have found strong
negative correlations between the biotic metrics for fish or macroinvertebrates and specific
r\
conductance, total dissolved solids, the concentration of individual ions, like SC>4 , and other
measures of the elevated concentrations of various ions observed in streams below valley fills.
These studies have also generally found negative correlations between the various biotic metrics
and some measures of metals, but trace metals are generally less elevated in streams below
valley fills than dissolved ions.
Discharges with high concentrations of ions reduced survival of the standard toxicity tests
using C. dubia. Concentrations of selenium in some streams have been measured at levels that
have been shown to cause fish and bird deformities in other streams. Fe and Mn deposits have
been observed on invertebrates, suggesting that concentrations are elevated under some
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circumstances. Ni concentrations in sediments downstream of MTM-VF exceed empirical
screening values.
The loss of the headwaters has ramifications for the ecosystems downstream. The loss of
headwater invertebrate taxa removes a source of food. Fish that specialize in feeding on
invertebrates (benthic invertivores) decrease downstream. Headwaters are active sites of organic
matter processing and nutrient uptake. The loss of headwater invertebrates and microbial
communities reduces dissolved carbon exports, an important food resource for downstream
biota, and increases nutrient loads in downstream waterbodies.
After the coal is removed, the extraction area and valley fill are graded and planted to
control sediment runoff. There is evidence that erosion is partially controlled, and this mitigates
but does not completely eliminate the amount of fine sediments deposited downstream.
8.2. CONCLUSIONS
This section summarizes our major conclusions of potential consequences of MTM-VF in
six categories:
(1) Loss of headwater resources
(2) Impacts on water quality
(3) Impacts on aquatic toxicity
(4) Impacts on aquatic ecosystems
(5) Cumulative impacts of multiple mining operations
(6) Effectiveness of on-site mining reclamation and mitigation activities
We formed our conclusions by reviewing evidence from two sources of information:
(1) the peer-reviewed, published literature and (2) the PEIS and its associated appendices
(U.S. EPA, 2003, 2005). Only a few peer-reviewed papers were found that studied water quality
or stream ecosystems in headwaters or downstream of MTM-VF in the central Appalachian
coalfields. Our conclusions are based on evidence from these papers and relevant research
findings from laboratory studies and observational studies from other locations and mining
activities. We also discuss the findings published in the PEIS, which was published as two
separate documents; the Draft, published in 2003, and the Final, published in 2005. Our
conclusions are consistent with those presented in another recent review of the ecological effects
of MTM-VF (i.e., Palmer et al., 2010).
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8.2.1. Loss of Headwater Resources
Based on permits approved from 1992 to 2002, more than 1,900 km of headwater streams
are scheduled to be permanently lost or buried because of MTM-VF. These streams represent
more than 2% of the stream miles in the study area (KY, TN, WV, and VA), and their total
length is more than triple the length of the Potomac River. The literature on headwater stream
hydrology and ecology suggests that, while significant, these numbers are likely to underestimate
the true magnitude and extent of impairment to regional biodiversity and ecosystem function
caused by the loss and burial of headwater catchments.
8.2.2. Impacts on Water Quality
Changes in water quality observed in streams downstream of MTM-VF include alteration
of flow and temperature regimes; increased fine sediments; and increases in ions, some metals,
and nitrogen.
Flows in streams below valley fills were generally more constant in both discharge and
temperature than in unimpacted streams. Valley fills influence downstream water quality by
acting like aquifers that store at least some of the water that enters from groundwater, surface
drainage, or direct precipitation. The removal of vegetation—particularly plants that have deep
roots—from the mined area and the area covered by the fills increases flow by decreasing
evapotranspiration.
Valley fills ameliorated the effects of moderate precipitation events on high flows, likely
because they temporarily store water from surface flows and direct precipitation. However, there
is some evidence that precipitation from more intense storms results in greater stormflows
because of compaction of the fill surface and mined area by heavy equipment that reduces
infiltration of precipitation and increases overland runoff.
Effluent waters below valley fills were often alkaline. Most valley fills contain sufficient
carbonate minerals to neutralize the acid produced by pyrite oxidation that has historically
caused water quality problems from coal mining. In addition, the sediment retention ponds can
be used as treatment basins to neutralize pH. As a result, the metals that are not soluble under
higher pH conditions, such as Fe, Cd, Cr, Cu, Pb, Zn, and Al were generally not elevated in
effluent waters below valley fills. Under some conditions, such as during higher flows,
particulate forms of less soluble metals, such as Fe, may be washed out of the valley fills.
Other ions were consistently observed at greatly elevated concentrations in the discharges
from valley fills. SO42 , HCOs , Ca2+, and Mg2+ are the dominant ions, but others include K+,
Na+, and Cl . These ions all contribute to the elevated levels of total dissolved solids, typically
measured as specific conductivity observed in the effluent waters below valley fills. Selenium
concentrations are also elevated. Selenium can bioaccumulate through aquatic food webs, and
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elevated levels have been found in fish tissues of the mining region. More than half of the sites
surveyed exceeded the chronic AWQC for selenium. Selenium reaches concentrations that have
been associated with effects in fish and birds in studies of mining effluents from other regions.
Despite the construction of sediment retention ponds below valley fills, several but not all
studies found increased fine sediments in stream reaches downstream. Concentrations of NOs
plus NC>2 were also slightly elevated downstream.
8.2.3. Toxicity Impacts on Aquatic Organisms
Results of laboratory toxicity tests using the crustacean Ceriodaphnia dubia predict that
acute lethality will occur at the high end of specific conductivity observed downstream of MTM-
VF operations. This expectation was confirmed in a study that observed reduced survival of C.
dubia in short-term tests using water from sites with high specific conductivity. Laboratory tests
of major ions reported effects on reproduction at concentrations 2-3 times lower than effects on
survival. Evidence from other alkaline mine effluents suggests that effects to organisms should
be expected at concentrations below those that affected C. dubia. Tests using the mayfly
Isonychia bicolor and the amphipod Hyalella azteca reported effects on survival at
concentrations 3-4 times lower than those affecting C. dubia. If effects on reproduction in these
organisms are similarly more sensitive than survival, effects from ions would be expected at
most sites downstream of MTM-VF.
Results from laboratory studies that varied the mixture of ions indicate that the interplay
among ions is complex. The ion mixture reported downstream of MTM-VF sites is dominated
by SC>42 , Ca2+, HCOs , and Mg2+. High concentrations of Ca2+ and overall hardness might be a
mitigating factor. If, as expected, the relative proportions of ions are generally consistent
downstream of MTM-VF in the central Appalachian coalfields, specific conductivity (|iS/cm)
may be the best indicator to use to predict when adverse effects would occur.
Se concentrations reported from waters in the study area were high enough to warrant
concern. In some streams, they exceeded the chronic AWQC and fall in the range of
concentrations that have caused toxic effects on aquatic invertebrates, fish, and birds in the field,
including in waters receiving valley fill and overburden dump leachates at coal mines in Canada.
Although Se bioavailability of selenium is difficult to predict, measurements from fish tissue
indicated that the Se is elevated in a form that is bioavailable. The greatest risks from exposure
and effects would be expected in downstream ponds and reservoirs where Se retention is high
and food webs are long.
Other toxicants were also high enough to warrant further investigation. Fe and Mn
deposits have been observed on macroinvertebrates. Ni and Zn concentrations in sediments are
higher than empirical screening level values.
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8.2.4. Impacts on Aquatic Ecosystems
All surveys that used multimetric and aggregate taxonomic indices reported degraded
biological conditions in streams downstream of MTM-VF. Both fish and macroinvertebrate
communities were affected. Within the communities, changes were observed in organisms
grouped by food source: benthic macroinvertebrates that feed on leaf detritus declined, benthic
macroinvertebrates that feed on algae increased, and fish that eat benthic macroinvertebrates
declined. Changes were also observed in organisms grouped by habit; macroinvertebrates that
cling to rocks in riffles declined. All studies showed a reduction in mayflies. Trends observed
for other taxonomic groups were less striking. Declines of the aggregate indices and
mayflieswere most strongly correlated with increased concentrations of ions and selenium and
were observed at ion concentrations well below those associated with increased mortality of
standard laboratory organisms in short-term tests using mine effluent.
8.2.5. Cumulative Impacts of Multiple Mining Operations
Few studies were found that investigated the cumulative impacts of multiple mining
operations. Specific conductivity and SC>42 levels were elevated in larger streams of the
Kanawha basin, downstream of multiple mining operations. Concentrations increased between
1980 and 1988 as more areas were mined. Johnson et al. (2010) found that the cumulative level
of conductivity values could be predicted using a simple function that combined tributary
concentrations with watershed area. This suggests that conductivity levels will accumulate
unless mixed with a more dilute source of water—for example, and umined tributary. Pond et al.
(2008) showed strong relationships between macroinvertebrate assemblages and conductivity.
However, neither Pond et al. (2008) nor Fulk et al. (2003) found additional decreases in fish or
macroinvertebrate multimetric indices with greater than one upstream mine or valley fill
upstream.
8.2.6. Effectiveness of Mining Reclamation and Mitigation Efforts
The results of the water quality studies indicate that reclamation efforts partially
controlled the amount of soil erosion and fine sediments transported downstream. However,
there is no evidence that reclamation efforts altered or reduced the ions or toxic chemicals
downstream of valley fills. Ion concentrations have either remained constant or increased over
time. Given that the alterations of the stream ecosystems reported in Sections 4 and 6 were
observed after sites were reclaimed for 3 to 24 years, the effects would be expected to persist.
Preliminary results suggest that FRA techniques allow increased infiltration and smaller peak
stormflows, possibly decreasing the risk of downstream flooding. Further study on water quality
impacts of this reclamation approach is needed.
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Channels created on valley fills are no longer associated with the same geologic
structures and hydrologic flow paths that preceded MTM-VF. Stream restoration techniques
were developed to restore one or more features of an existing stream, not to create entirely new
streams. During precipitation events, created channels might act similarly to an ephemeral
stream by conveying water and materials, but the hydrology of a buried perennial or intermittent
reach is unlikely to be replaced. There is no evidence that these channels, regardless of
permanence of flow, improve water quality, and invertebrate diversity and decomposition rates
are generally lower in constructed channels or at the base of the fill compared to forested
perennial or intermittent streams. In contrast, invertebrate diversity might not differ in
constructed channels and ephemeral streams. Increasing stream structural heterogeneity and
restoring riparian areas below the fill are likely to have a limited effect on aquatic biota, given
the failure to address the underlying problem of water quality.
8.3. INFORMATION GAPS, ASSESSMENT NEEDS AND RESEARCH
OPPORTUNITIES
The evidence in our review is consistent enough that we have a high degree of confidence
in our conclusions. Still, our review uncovered a number of information gaps that could be filled
by research. Filling these can improve our quantitative understanding of how MTM-VF impacts
aquatic ecosystems, potentially leading to more effective regulatory and mitigation approaches.
Future assessments should consider the comparative risks of MTM-VF. However, the
comparisons should be to real alternatives that might be implemented by real decisions. For
example, risks from MTM-VF might be compared with those associated with other coal sources,
or risks associated with electricity generated by burning coal from MTM-VF might be compared
with those associated with sources of electricity other than coal combustion. Alternatively, if a
land-use decision is being made, MTM-VF risks might be compared with other uses of
mountaintops and headwater streams such as logging or tourism.
8.3.1. Update the MTM-VF Inventory and Surveys of Impact Extent
The most recent data available in the published literature on the extent and potential
additional development of MTM-VF mines in the central Appalachian coalfields are those
compiled for the PEIS (U.S. EPA, 2003, 2005). These data were only for MTM-VF mines
developed between 1985 and 2002, when at least some mines had been developed as early as
1967 (U.S. EPA, 1979), and permitting and construction of MTM-VF have continued since then.
Therefore, the inventory of filled valleys and of stream miles buried by those valley fills should
be updated. Remote sensing and GIS, combined with field sampling, would make this possible
(Townsend et al., 2009). The updated inventory of MTM-VF can be used to design a statistically
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robust estimate of the extent of impacts within the region, based on a probabilistic sampling
design.
To support quantification of cumulative effects (see Section 8.3.7), it could be
appropriate to extend these surveys beyond the linear stream miles directly altered by
mountaintop mining and buried by valley fills to consider the fractal nature of these stream
networks at larger scales. This would involve quantifying the proportional area of individual
larger watersheds of a certain size affected by mountaintop mines and valley fills. If sufficient
comprehensive remote sensing or aerial photo interpretation is or becomes available, the
landscape alterations by MTM-VF might be measured in terms of the area and volume of earth
movement, the change in vegetation cover and type, and the proximity of these activities to
stream channels.
8.3.2. Quantify the Contributions of Headwater Streams
It would be desirable to more fully understand the role of the headwater streams buried
by valley fills in the retention and cycling of nutrients, such as nitrogen and phosphorus, and the
downstream transport of trophic resources, such as leaf litter, small particulate organic matter
produced from the leaf litter, and dissolved organic carbon, algae, and animal prey. This
understanding would allow assessors to better understand and model the cumulative effects of
burying these headwater streams on stream function (i.e., nutrient transport and cycling,
processing and transport of organic matter) and other ecosystem services supplied by these
stream systems.
It would be desirable to more fully understand the metapopulation and metacommunity
linkages among different headwater streams and between these headwaters and downstream
reaches. This information would increase understanding of the effects of burying these
headwater streams on regional biodiversity—including the cumulative effects of this practice of
burying headwater drainages in this region of the Appalachians.
8.3.3. Improve Understanding of Causal Linkages
Our understanding of the causal linkages between MTM-VF and stream ecosystems
could be improved by bringing together additional data. Sources of data include reports that we
were unable to include in this report because we could not confirm that they had been
peer-reviewed, and additional monitoring data that might also be available from various states,
particularly West Virginia and Kentucky. Questions that might be answered include
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(1) At what concentrations of major ions and trace metals do different taxa disappear?
(2) Which downstream organisms in addition to Ephemeroptera are most affected by valley
fills?
(3) How do these effects differ among different insect orders and between insects and
noninsect aquatic taxa?
(4) How do the species within these large orders change? Some evidence indicates that
headwater species are replaced by more downstream species below valley fills.
(5) Are there observable effects on individuals of sensitive taxa?
When selecting such monitoring data, care should be taken that the sampling was timed
so that the common species can be sampled if they are present. This is not a concern for fish, but
many macroinvertebrates are present for part of the year as eggs or larval instars that are too
small to be sampled by the standard net mesh sizes used to sample benthic macroinvertebrates.
8.3.4. Develop Tests Using Sensitive Taxa
Declines in macroinvertebrate indices were observed at ion concentrations well below
those associated with increased mortality of standard laboratory test organisms in short-term
tests, indicating that some native organisms are more sensitive to the constituents in mining
effluents. Quantitative estimates of the concentrations at which effects occur could be improved
by testing effluents using a life-cycle test, especially with vertebrate and invertebrate species
found in these headwater streams. Increasing understanding of toxic mechanisms could help
interpret the responses observed from biological surveys. For invertebrates, we would
recommend an Ephemeroptera species or a physiologically similar aquatic insect. An example
of a full life cycle with a species of Ephemeroptera is described by Sweeney et al. (1993) and
Conley et al. (2009). Tests using these insects would help verify that the differences in
sensitivity between laboratory tests using C. dubia and field observations of Ephemeroptera
declines are due to differences in sensitivity to the ions, rather than a confounding factor. For
fish and amphibians, it would be desirable to perform reproductive toxicity tests with waters like
those found below valley fills using headwater taxa, such as dace, brook trout, or sculpins.
The vertebrate and invertebrate fauna found in headwater streams of the southern
Appalachian Mountains are adapted to waters characterized by low hardness, total dissolved
solids, ionic strength, and conductivity and by neutral to slightly acidic pH. The streams below
valley fills are altered such that the waters are characterized by high hardness, total dissolved
solids, ionic strength and conductivity, and slightly alkaline pH. These waters also have
r\ fj: fj:
relatively high concentrations of individual ions, such as SC>4 , HCOs , Ca , and Mg . These
multiple changes in the dissolved constituents in these waters are likely to have interactive
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effects on aquatic fauna and are not duplicated well by any laboratory test data found in the
published literature. Moreover, the species and life stages used in the laboratory tests found in
the published literature differ from the native fauna of these streams.
Most of the invertebrates that have been used in laboratory toxicity tests of the effects of
conductivity, total dissolved solids, or the individual effects of SC>42 and other dissolved ions
are Crustacea. This includes the cladocerans, C. dubia and D. magna, and the amphipod,
H. azteca. These Crustacea have very different evolutionary histories compared to the aquatic
insects that dominate the headwater streams. In the case of Crustacea and other wholly aquatic
groups like Mollusca, their evolutionary ancestors moved directly into freshwater environments
from marine or estuarine environments (Thorp and Covich, 1991). Some species, such as D.
magna (Martinez-Jeronimo and Martinez-Jeronimo, 2007), have populations found in brackish
waters. Aladin and Potts (1995) describe Cladocera as strong osmoregulators. Hence, in
addition to not being found in these headwater streams, the standard invertebrate test species do
not appear to be sensitive to the sorts of major ions leaching from valley fills.
The evolutionary ancestors of insects moved from marine or estuarine environments into
terrestrial environments. Then, in turn, the evolutionary ancestors of aquatic insects, such as
Ephemeroptera, Plecoptera, Trichoptera, and Odonata moved from terrestrial environments to
freshwater environments (Merritt and Cummins, 1996). As a result, aquatic insects possess very
different mechanisms for osmotic regulation compared to the wholly aquatic groups. In the
aquatic insects found in these streams, osmotic regulation is accomplished, in part, by tissues
called chloride cells or chloride epithelia, which are involved in ion absorption, an important
adaptation in the low ionic strength, freshwater habitats where aquatic insects are found
(Kominick, 1977). In addition, the insects differ from the test species in that their eggs develop
externally, so they are directly exposed to contaminated waters. This suggests that Crustacea are
not appropriate surrogates for these aquatic insects in laboratory toxicity tests, particularly those
that test the effects of the alterations in water chemistry associated with valley fills.
Even insect species, like C. tentans and /. bicolor, might not be good surrogates. Aquatic
Diptera possess anal papillae, which though different in structure are functionally equivalent to
chloride epithelia. In the case of/, bicolor, its natural distribution is in larger streams with
greater conductivities than those found in the streams affected by MTM-VF (Kondratieff and
Voshell, 1984). Also, the bioassays testing /. bicolor used relatively large (~9 to 14 mm in
length), late instar nymphs in 7-day tests, where molting and survival were the only measurement
endpoints. Other life stages and measurement endpoints appropriate to the survival of these
mayfly populations could be more sensitive to the chronic stresses imposed by the observed
changes in water chemistry.
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8.3.5. Conduct Mesocosm and Microcosm Experiments with Indigenous Taxa
An alternative approach to the single species bioassays would be the use of mesocosm
and microcosm experiments to further investigate the causal relationships between elevated
conductivity and the ions associated with MTM-VF and benthic macroinvertebrate assemblages.
Such an approach has been described by Clements (2004), who investigated metals associated
with mine drainage from hardrock metal mining. As described by Clements (2004), plastic trays
filled with substrates from the streambed were placed in an unimpaired stream for 40 days to
allow colonization by the indigenous benthic macroinvertebrate assemblage. In the mesocosm
experiment, the colonized trays were placed in a stream impaired by the contaminants of interest.
In the microcosm experiment, the colonized trays were placed in artificial streams dosed with
one or mixtures of the contaminants of interest. After a 10-day exposure period, the trays were
removed and the macroinvertebrates processed for identification and counting. The results from
the mesocosm experiment verify that the effects observed in field samples occur as a result of
exposure to the contaminants of interest, whereas the results of the microcosm would quantify
the levels of the contaminants of interest that cause those effects.
8.3.6. Further Investigate Selenium and Sediments
Aqueous selenium concentrations and concentrations in fish fall within a range that can
cause effects on fish and fish-eating birds. Additional analyses, including possibly a study of
stream-based food webs, could better define the extent of this problem, and reproductive tests of
fish collected in high-selenium streams could better define the nature of the problem. To
confirm effects of Se on fish reproduction, fish would be collected from high-Se streams and
spawned in the laboratory. This would be required because Se acts by bioaccumulation in the
females and transfer to the eggs. Little is known about the effects of selenium on stream
invertebrates. Analyses of invertebrates from high-selenium streams and reproductive tests
could determine whether selenium is contributing to observed effects.
Few data are available concerning the effects of MTM-VF on the chemical quality of
sediments in streams below valley fills. While dissolved trace metals in effluent waters below
valley fills appear to be low, there is evidence along with geochemical theory that paniculate
metals should be produced within valley fills and could, under some conditions, be flushed
downstream. Also, there are some metals (i.e., Mn, Ni) whose solubilities are not affected by pH
and whose dissolved concentrations could be somewhat elevated in effluent waters. Therefore,
data on sediment concentrations of metals could be used to assess whether sediment
contamination is a concern associated with MTM-VF. Observations could also determine if
effects associated with the deposition of particulate metals occur. These effects could be similar
to those observed with iron hydroxides in more acidic situations. To completely assess this
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exposure pathway, such sampling could include measurement of pore water concentrations of the
dissolved metals and ammonia or use techniques such as simultaneously extracted metals and
acid volatile sulfide.
8.3.7. Quantify Cumulative Effects
Cumulative impact assessment can be approached from many perspectives. The impacts
of MTM-VF can be evaluated as more land is mined over time. Impacts can be approached from
a watershed perspective, evaluating the overall impact of all human activities that influence
aquatic ecosystems. Or, the influence of stressors from MTM-VF can be traced through different
components of the food web, evaluating effects on ecosystem functions or movement through the
food web. Another perspective might evaluate impacts of upstream alterations from MTM-VF
on downstream ecosystem functions. Among other species, long-lived unionid mussels might be
used to monitor effects on ecosystem functions and food webs.
We found little published literature that evaluated cumulative impacts from any of these
perspectives in our study region. Johnson et al. (2010) successfully modeled changes in
conductivity levels as tributaries combine but did not link these changes to biological endpoints.
Petty et al. (2010) evaluated the impact of increasing mining intensity in watersheds influenced
primarily by acid mine drainage. Impacts increased with intensity of mining but were also
influenced by underlying coal geology, and the spatial configuration of disturbance. Several
papers have documented that exposure to one stressor can make ecological systems more
vulnerable to the impacts from subsequent stressors or disturbances, potentially preventing
recovery (e.g., Brooks et al., 2007; Clements et al., 2008; Paine et al., 1998). However, none of
the studies have investigated the stressors or aquatic ecosystems associated with MTM-VF. The
movement of selenium through ecosystems has been reviewed (Chapman et al., 2010), but no
studies were identified from the aquatic ecosystems of the central Appalachian coalfields.
Additional studies explicitly designed to quantify the cumulative effects of MTM-VF
would help differentiate those effects from the other land uses in the central Appalachian
coalfields, such as abandoned mines, oil and gas development, and residential development.
Additional water chemistry sampling, combined with spatial analyses of the number and volume
of valley fills, could reveal how specific conductivity and other measures of the dissolved ions
increase as the percentage of the watershed in valley fills increases and how export of dissolved
ions changes with time after the creation of a valley fill. Concurrent samples of the biological
assemblages could be used to develop models to predict the temporal and spatial extent of
impacts. Currently, little is known about the cumulative effects of incremental loss of headwater
streams and naturally occurring mountain aquifers on the region's hydrology and water supplies.
Given the extensive scales at which landscape disturbances above and below ground occur in
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coal mining areas, groundwater sampling, tracer studies, and surface flow-groundwater
interaction monitoring, in addition to water quality sampling, might be needed.
8.3.8. Quantify Longitudinal Effects
With the exception of Johnson et al. (2010), which only measured conductivity
r\
longitudinally in a single 118-km catchment, no studies have quantified the change in
conductivity, individual ion concentrations, pH, or precipitates as the water progresses
downstream in a catchment with valley fills. This is needed to quantify the potential longitudinal
effects of these elevated ions downstream from valley fills and understand how the relative
concentrations of the ions change over space. Moreover, if a catchment has more than one valley
fill, as did Johnson et al. (2010), such studies will contribute to our understanding of cumulative
effects on downstream water chemistry. This could also include sampling upstream, in, and
downstream of the sedimentation ponds to further quantify any separate effects of the
sedimentation ponds.
8.3.9. Quantify Effects on Stream Hyporheic Zones
No studies have investigated the effects of MTM-VF on the hyporheic zones of streams
downstream from valley fills or even other groundwater resources. To fully understand the
longitudinal and cumulative effects of MTM-VF, such data are needed. Questions include, is the
chemistry of interstitial waters similar to that of the surface water and how might interaction with
groundwaters affect this chemistry? Are there similar adverse effects to invertebrates and other
organisms living in downstream hyporheic zones?
8.3.10. Quantify Functionality of Constructed Streams and Mitigation Efforts
Finally, although there is a large body of literature on stream restoration ecology in urban
and agricultural streams, we found there is a lack of evidence on the biota and ecosystem
functioning associated with the constructed sediment and flow control channels on valley fills. If
these streams are argued to mitigate the effects of stream burial, the type and degree of
mitigation should be quantified.
We limited our review of reclamation and mitigation activities to evidence of their effects
on on-site water quality, quantity, and aquatic ecosystems. Many research and development
needs remain: methods for decreasing concentrations of ions and improving water quality;
research on the long-term downstream impacts from disturbance, burial and loss of headwater
streams, including physical impacts on sediment supply, hydrology, and geomorphology and
their implications for stream stability and channel adjustment during and post-mining; and
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research on improving stream channels enhancements for areas that have no reference
streamflow curves or gauged streams.
Off-site mitigation of a wide variety of stream impacts, for example, from agriculture or
development, can be used to offset impacts from MTM-VF. The quantification of the
effectiveness of these efforts and the degree to which they compensate for the losses from
MTM-VF is an area of active work and would form the basis of a useful review.
8.3.11. Expand the Scope of Review to Include Evidence from Non-Peer-Reviewed Sources
and Terrestrial Impacts
We limited our source material for the current report to the published peer-reviewed
literature. Evidence reported in theses, dissertations, non-peer-reviewed symposia and reports,
could, after equivalent documentation and quality review of methods and analyses, contribute
additional insights.
The scope of our report is limited to impacts on aquatic ecosystems. An assessment of
impacts of MTM-VF on terrestrial ecosystems would provide a useful companion document.
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LITERATURE CITED
Ahrens, MJ; Morrissey, DJ. (2005) Biological effects of unburnt coal in the marine environment. Oceanogr Marine
Biol Ann Rev 43:69-122.
Aladin, NV; Potts, WTW. (1995) Osmoregulatory capacity of the Cladocera. J Comp Physiol B Biochem Syst
Environ Physiol 164(8):671-683.
Alexander, LC; Delion, M; Hawthorne, DJ; et al. (2009) Mitochondrial lineages and DNA barcoding of closely
related species in the mayfly genus Ephemerella (Ephemeroptera: Ephemerellidae). J N Am Benthol Soc 28:584-
595.
Alexander, LC; Hawthorne, DJ; Palmer, MA; et al. (2011) Loss of genetic diversity in the North American mayfly
Ephemerella invaria associated with deforestation of headwater streams. Freshwater Biol. E-pub Feb
2011:http://dx.doi.org/10.1111/j.l365-2427.2010.02566.x.
Alexander, RB; Smith, RA; Schwarz, GE. (2000) Effect of stream channel size on the delivery of nitrogen to the
Gulf of Mexico. Nature 403(6771):758-761.
Alexander, RB: Boyer, EW; Smith, RA; et al. (2007) The role of headwater streams in downstream water quality. J
Am Water Res Assoc 43(l):41-59.
Altschuler, ZS; Schnepfe, MM; Silber, CC; et al. (1983) Sulfur diagenesis in Everglades peat and origin of pyrite in
coal. Science 221(4607):221-227.
Andrews, JA; Johnson, JE; Torbert, JL; et al. (1998) Minesoil and site properties associated with early height growth
of eastern white pine. J Environ Qual 27:192-199.
Angermeier, PL: Smogor, RA; Stauffer, JR. (2000) Regional frameworks and candidate metrics for assessing biotic
integrity in mid-Atlantic highland streams. Trans Am Fish Soc 129:962-981.
Angradi, TR. (1997) Hydrologic context and macroinvertebrate community response to floods in an Appalachian
headwater stream. Am Midi Nat 138:371-386.
Angradi, T; Hood, R; Tarter, D. (2001) Vertical, longitudinal and temporal variation in the macrobenthos of an
Appalachian headwater stream system. Am Midi Nat 146(2):223-242.
APHA (American Public Health Association), American Water Works Association and Water Environment
Federation. (1998) Standard methods for the Examination of Water and Wastewater. Twentieth Edition. American
Public Health Association, Washington, DC.
Armstead, MY; Yeager-Seagle, JL; Emerson, L. (2004) Benthic macroinvertebrate studies conducted in
mountaintop mining/valley fill influenced streams in conjunction with the USEPA environmental impact study. In:
Barnheisel, RI, editor. 21st Annual Meetings of the American Society of Mining and Reclamation and the 25th West
Virginia Surface Mine Drainage Task Force. Morgantown, WV: American Society of Mining and Reclamation, pp.
87-93.
Arscott, DB; Lamed. S; M.R Scarsbrook. MR; et al. (2010) Aquatic invertebrate community structure along an
intermittence gradient: Selwyn River, New Zealand. Journal of the North American Benthological Society
29(2):530-545.
Ashby, WC. (1997) Soil ripping and herbicides enhance tree and shrub restoration on stripmines. Restor Ecol
5:169-177.
105
-------
Banks, D; Younger, PL; Arnesen, RT; et al. (1997) Mine-water chemistry: the good, the bad and the ugly. Environ
Geol32(3):157-174.
Barbour, MT; Gerritsen, J; Snyder, BD; et al. (1999) Rapid bioassessment protocols for use in streams and wadeable
rivers: periphyton, benthic macroinvertebrates, and fish. Second Edition., U.S. Environmental Protection Agency,
Office of Water, Washington, DC; EPA/841/B-99/002.
Barnes, RF; Nelson, CJ; Collins, M; et al.; eds. (2003) Forages: an introduction to grassland agriculture. 6th edition,
Vol 1. Ames, IA: Iowa State Press.
Barros, AP; Lettenmaier, DP. (1994) Dynamic modeling of orographically induced precipitation. Rev Geophys
32:221-227.
Benda, L; Poff, NL; Miller, D; et al. (2004) The network dynamics hypothesis: How channel networks structure
riverine habitats. Bioscience 54(5):413-427.
Benda, L; Hassan, MA; Church, M: et al. (2005) Geomorphology of steepland headwaters: the transition from
hillslopes to channels. J Am Water Resour Assoc 41(4):835-851.
Benson, MA; Dalrymple, T. (1967) General field and office procedures for indirect discharge measurements.
Chapter Al. In: Techniques of water resources investigations, book 3: applications of hydraulics. Department of
Interior, U.S. Geological Survey, Alexandria, VA. Available online at http://pubs.usgs.gov/twri/.
Berendzen, PB; Simons, AM; Wood, RM; et al. (2008) Recovering cryptic diversity and ancient drainage patterns in
eastern North America: Historical biogeography of the Notropis rubellus species group (Teleostei: Cypriniformes).
Mol Phylogenet Evol 46:721-737.
Bernhardt, ES; Likens, GE; Hall, RO; et al. (2005) Can't see the forest for the stream? In-stream processing and
terrestrial nitrogen exports. Bioscience. 55(3):219-230.
Bernhardt, ES; Palmer, MA. (2007) Restoring streams in an urbanizing world. Freshwater Biol 52:738-751.
Biesinger, KE; Christensen, GM. (1972) Effects of various metals on survival, growth, reproduction, and
metabolism ofDaphnia magna. J Fisheries Res Board, Canada 29:1691-1700.
Blakely, TJ; Harding, JS; Mclntosh, AR; etal. (2006). Barriers to the recovery of aquatic insect communities in
urban streams. Freshwater Biology 51:1634-1645.
Bodkin, R; Kern, J; McClellan, P; et al. (2007) Limiting total dissolved solids to protect aquatic life. J Soil Water
Conserv62(3):57A-61A.
Boulton, AJ; Findlay, S; Marmonier, P; et al. (1998) The functional significance of the hyporheic zone in streams
and rivers. Annu Rev Ecol Syst 29:59-81.
Braatne, JH; Sullivan, SMP; Chamberlain, E. (2007) Leaf decomposition and stream macroinvertebrate colonisation
of Japanese knotweed, an invasive plant species. International Review of Hydrobiology 92:656-665.
Bradley, TJ. (2009) Animal Osmoregulation. Oxford University Press, Inc., New York, NY.
Bradshaw, A. (1997) Restoration of mined lands—using natural processes. Ecological Engineering 8:255-269.
Brinks, JS; Lhotka, JM; Barton, CD; et al. (2011) Effects of fertilization and irrigation on American sycamore and
black locust planted on a reclaimed surface mine in Appalachia. Forest Ecol Manag 261:640-648.
Brittain, JE; Eikeland, TJ. (1988) Invertebrate drift - a review. Hydrobiologia 166:77-93.
106
-------
Brooks, ML; McKnight, DM; Clements, WH. (2007) Photochemical control of copper complexation by dissolved
organic matter in Rocky Mountain streams, Colorado. Limnol Oceanogr 52:766-779.
Brown, JH; Kodric-Brown, A. (1977) Turnover rates in insular biogeography - effect of immigration on extinction.
Ecology 58(2):445-449.
Bryant, G; McPhilliamy, S; Childers, H. (2002) A survey of the water quality of streams in the primary region of
mountaintop/valley fill coal mining, October 1999 to January 2001. In: Draft programmatic environmental impact
statement on mountaintop mining/valley fills in Appalachia - 2003. Appendix D. U.S. Environmental Protection
Agency, Region 3, Philadelphia, PA. Available online at
http://www.epa.gOv/Region3/mtntop/pdf/appendices/d/stream-chemistry/MTMVFChemistryPartl.pdf.
Bullock, A; Acreman, M. (2003) The role of wetlands in the hydrological cycle. Hydrol Earth System Sc 7:358-389.
Cairns, MA; Lackey RT. (1992) Biodiversity and management of natural resources: The issue. Fisheries 17(3):6-10.
Callaghan, T; Brady,K; Chisholm, W; Sanies, G. (2000) Hydrology of the Appalachian bituminous coal basin. In:
Kleinmann, RLP, Ed. Prediction of Water Quality at Surface Coal Mines. The National Mine Land Reclamation
Center. Morgantown, WV: West Virginia University, USA; pp. 36-67.
Carlson, RM; Oyler, AR; Gerhart, EH; et al. (1979) Implications to the aquatic environment of polynuclear aromatic
hydrocarbons liberated from northern Great Plains coal. U.S. Environmental Protection Agency, Office of Research
and Development, Environmental Research Laboratory, Duluth, MN: EPA 600/3-79/093. Available online from the
National Technical Information Service, Springfield, VA, PB 80-115967.
Caruccio, FT; Perm, JC; Home, J; et al. (1977) Paleoenvironment of coal and its relation to drainage quality. U.S.
Environmental Protection Agency, Office of Research and Development, Industrial Environmental Laboratory,
Cincinnati, OH. EPA 600/7-77/067. Available online from the National Technical Information Service, Springfield,
Va, PB-270 080/5.
Casagrande, DJ. (1987) Sulphur in peat and coal. Geol Soc SP 32:87-105.
Chapman, PM; Downie, J; Maynard, A; et al. (1996) Coal and deodorizer residues in marine sediments:
contaminants or pollutants? Environ Toxicol Chem 15(5):638-642.
Chapman, PM; Bailey, H; Canaria, E. (2000) Toxicity of total dissolved solids associated with two mine effluents to
chironomid larvae and early life stages of rainbow trout. Environ Toxicol Chem 19:210-214.
Chapman, PM; Adams, WJ; Brooks, ML; et al. (2010) Ecological assessment of selenium in the aquatic
environment. Pensacola, FL: Society of Environmental Toxicology and Chemistry (SETAC).
Clark, J; Clements, WH. (2006) The use of in situ and stream microcosm experiments to assess population- and
community-level responses to metals. Environ Toxicol Chem 25(9):2306-2312.
Clarke, A; Mac Nally, R; Bond, N; et al. (2008) Macroinvertebrate diversity in headwater streams: a review.
Freshwater Biol 53:1707-1721.
Clements, WH. (2004) Small-scale experiments support causal relationships between metal contamination and
macroinvertebrate community responses. Ecol Appl 14:954-967.
Clements, WH; Brooks, ML; Kashian, DR; et al. (2008) Changes in dissolved organic material determine exposure
of stream benthic communities to UV-B radiation and heavy metals: implications for climate change. Glob Change
Biol 14:2201-2214.
107
-------
Coleman, L; Bragg, LJ; Findelman, RB. (1993) Distribution and mode of occurrence of selenium in US coals.
Environ Geochem Health 15:215-227.
Collins, BM; Sobczak, WV; Colburn, EA. (2007) Subsurface flowpaths in a forested headwater stream harbor a
diverse macroinvertebrate community. Wetlands 27(2):319-325.
Conley, JM; Funk, DH; Buchwalter, DB. (2009) Selenium bioaccumulation and maternal transfer in the mayfly
Centroptilum triangulifer in a life-cycle, periphyton-biofilm trophic assay. Environ Sci Technol 43:7952-7957.
Cooper, DM; Jenkins, A; Keffington, R; et al. (2000) Catchment-scale simulation of stream water chemistry by
spatial mixing: Theory and application. J Hydrol 233:121-137.
Cooper, PD. (1994) Mechanisms of hemolymph acid-base regulation in aquatic insects. Physiol Zool 67(l):29-53.
Courtemanch, DL; Davies, SP. (1987) A coefficient of community loss to assess detrimental change in aquatic
communities. Water Res 21:217-222.
Cross, WF; Wallace, JB; Rosemond, AD; et al. (2006) Whole-system nutrient enrichment increases secondary
production in a detritus-based ecosystem. Ecology 87(6): 1556-1565.
Cummins, KW; Wilzbach, MA; Gates, DM; et al. (1989) Shredders and riparian vegetation - leaf litter that falls into
streams influences communities of stream invertebrates. Bioscience 39:24-30.
Davic, R; Welsh, HH. (2004) On the ecological roles of salamanders. Annu Rev Ecol Evol Syst 35:405-434.
Davies, SP; Jackson, SK. (2006) The biological condition gradient: A descriptive model for interpreting change in
aquatic ecosystems. Ecol Appl 16(4): 1251-1266.
DeBruyn, AMH; Chapman, PM. (2007) Selenium toxicity to invertebrates: will proposed thresholds for toxicity to
fish and birds also protect their prey? Environ Sci Technol 41(5):1766—1770.
DeNicola, DM; Stapleton, MG. (2002) Impact of acid mine drainage onbenthic communities in streams: the relative
roles of substratum vs. aqueous effects. Environ Pollut 119:303-315.
Detenbeck, NE; DeVore, PW; Niemi, GJ; et al. 1990. Recovery of temperate-stream fish communities from
disturbance: A review of case studies and synthesis of theory. EnvironManag 16(l):33-53.
Di Toro, DM; Allen, HE; Bergman, HL; et al. (2001) Biotic ligand model of the acute toxicity of metals. 1.
Technical basis. Environ Toxicol Chem 20(10):2383-2396.
Dickens, PS; Minear, RA; Tschantz, B A. (1989) Hydrologic alteration of mountain watersheds from surface mining.
J Water Pollut ContFed 61:1249-1260.
Dietz, TH. (1979) Uptake of sodium and chloride by freshwater mussels. Can J Zool 57:156-160.
Dietz, TH; Branton, WB. (1975) Ionic regulation in the freshwater mussel, Liguma subrostrata (Say). J Comp
Physiol 104:19-26.
Dittman, EK; Buchwalter, DB. (2010) Manganese bioconcentration in aquatic insects: Mn Oxide coatings, molting
loss, and Mn(II) thiol scavenging. Environ Sci Technol 44:9182-9188.
Doig, LE; Liber, K. (2006) Nickel partitioning in formulated and natural freshwater sediments. Chemosphere
62:968-979.
108
-------
Downing-Kunz, MA; Unthank, MD; Grain, AS. (2005) Compilation of concentrations of total selenium in water,
coal in bottom material, and field measurement data for selected streams in eastern Kentucky, July 1980 Open-File
Report 2005-1354. Department of Interior, U.S. Geological Survey, Reston, VA. Available online at
http://www.dtic.mil/cgi-bin/GetTRDoc? AD=AD A442142&Location=U2&doc=GetTRDoc.pdf.
Dudgeon, DD; Arthington, AH; Gessner, MO; et al. (2006) Freshwater biodiversity: Importance, threats, status and
conservation challenges. BiolRev 81:163-182.
Eble, CF; Hower, JC, (1997) Coal quality trends and distribution of potentially hazardous trace elements in eastern
Kentucky coals. Fuel 76:711-715.
Echols, BS; Currie, RJ; Cherry, DS. (2010) Preliminary results of laboratory toxicity tests with the mayfly,
Isonychia bicolor (Ephemeroptera: Isonychiidae) for development as a standard test organism for evaluating streams
in the Appalachian coalfields of Virginia and West Virginia. Environ Monit Assess 169:487-500.
Ehrenfeld, J; Klein, U. (1997) The key role of the H+ V-ATPase in acid-base balance and Na+ transport processes in
frog skin. J Exp Biol 200:247-256.
Evans, DH. (1975) Ionic exchange mechanisms in fish gills. Comp Biochem Physiol Part A Physiol 51:491-495.
Pagan, WF. (2002) Connectivity, fragmentation, and extinction risk in dendritic metapopulations. Ecology 83:3243-
3249.
Feminella, JW. (1996) Comparison of benthic macroinvertebrate assemblages in small streams along a gradient of
flow permanence. JN AmBenthol Soc 15(4):651-669.
Ferrari, JR; Lookingbill, TR; McCormick, et al. (2009) Surface mining and reclamation effects on flood response of
watersheds in the central Appalachian Plateau region. Water Resour Res 45(4):W04407.
Ferreri, CP; Stauffer, JR; Stecko, TD. (2004) Evaluating impacts of mountain top removal/valley fill coal mining on
stream fish populations. Pages 576-592. In: Barnhisel, R.I. (ed.), Proceedings of the Joint Conference of 21st
Annual Meetings of the American Society of Mining and Reclamation and 25th West Virginia Surface Mine
Drainage Task Force Symposium, Morgantown, WV. American Society of Mining and Reclamation, Lexington,
KY.
Finkelman, RB. (1994) Modes of occurrence of potentially hazardous elements in coal: Levels of confidence. Fuel
Process Tech 39:21-34.
Fox, JR. (2009) Identification of sediment sources in forested watersheds with surface coal mining disturbance using
carbon and nitrogen isotopes. J. Am. Water Resour. Assoc. 45(5): 1273-1289.
Frankham, R. (2005) Genetics and extinction. Biol Conserv 126:131-140.
Fritz, KM; Dodds, WK. (2004) Resistance and resilience of macroinvertebrate assemblages to drying and flood in a
tallgrass prairie stream system. Hydrobiologia 527:99-112.
Freeman, MC; Pringle, CM; Jackson, CR. (2007) Hydrologic connectivity and the contribution of stream headwaters
to ecological integrity at regional scales. J Am Water Resour Assoc 43:5-14.
Fritz, KM; Johnson, BR; Walters, DM. (2006) Field operations manual for assessing the hydrologic permanence and
ecological condition of headwater streams. In. Cincinnati, OH, USA: U.S. Environmental Protection Agency, Office
of Research and Development, National Exposure Research Laboratory.
Fritz, KM; Fulton, S; Johnson, BR; et al. (2010) Structural and functional characteristics of natural and constructed
channels draining a reclaimed mountaintop removal and valley fill coal mine. J N Am Benthol Soc 29(2):673-689.
109
-------
Fulk, F; Autrey, B; Hutchens, J; et al. (2003) Ecological assessment of streams in the coal mining region of West
Virginia using data collected by the U.S. EPA and environmental consulting firms. In: Mountaintop mining/valley
fills in Appalachia. Final programmatic environmental impact statement. U.S. Environmental Protection Agency,
Region 3, Philadelphia, PA. Appendix D. Available online at http://www.epa.gov/Region3/mtntop/pdf/mtm-
vf_fpeis_full-document.pdf.
FWS (U.S. Fish and Wildlife Service) (2003) Federally listed Threatened and Endangered, Candidate and Species of
Concern. Appendix F, in Draft programmatic environmental impact statement on mountaintop mining/valley fills in
Appalachia, U.S. Environmental Protection Agency, Region 3, Philadelphia, PA. Available online at
http://www.epa.gov/Region3/mtntop/eis2003appendices.htarfappf.
FWS (U.S. Fish and Wildlife Service) (2004) Recovery plan for Cumberland elktoe, oyster mussel, Cumberlandian
combshell, purple bean, and rough rabbitsfoot. Southeast Region, U.S. Fish and Wildlife Service, Atlanta, GA.
Available online at http://ecos.fws.gov/docs/recovery_plans/2004/040524.pdf.
Gerhardt, A. (1992) Effects of subacute doses of iron (Fe) onLeptophlebia-marginata (Insecta, Ephemeroptera).
Freshwater Biol 27:79-84.
Gerhardt, A. (1994) Short-term toxicity of iron (Fe) and lead (Pb) to the mayfly Leptophlebia-marginata (L)
(Insecta) in relation to fresh-water acidification. Hydrobiologia 284:157-168.
Gilpin, ME; Soule, ME. (1986) Minimum viable populations: processes of species extinction. In: Soule, ME, editor.
Conservation Biology: The Science of Scarcity and Diversity. Sunderland, MA, USA: Sinauer Associates, Inc.
pp 19-34.
Gomi, T; Sidle, RC; Richardson, JS. (2002) Understanding processes and downstream linkages of headwater
systems. Bioscience 52(10):905-916.
Google maps. (2009) Satellite images of Hobet 21 mine complex and Washington DC. Available online at
http://maps.google.com (accessed 11/19/2009).
Goss, GG; Perry, SF; Wood, CM; et al. (1992) Mechanisms of ion and acid-base regulation at the gills of freshwater
fish. JExp Biol 263:143-159.
Grant, EHC; Green, LE; Lowe, WH. (2009) Salamander occupancy in headwater stream networks. Freshwater Biol
54(6):1370-1378.
Grant, EHC; Nichols, JD; Lowe, WH; et al. (2010) Use of multiple dispersal pathways facilitates amphibian
persistence in stream networks. Proceedings of the National Academy of Sciences of the United States of America
107:6936-6940.
Green, J; Passmore, M; Childers, H. (2000) A survey of the conditions of stream in the primary region of
mountaintop mining/valley fill coal mining. Mountaintop mining/valley fills in Appalachia. Final programmatic
environmental impact statement. U.S. Environmental Protection Agency, Region 3, Philadelphia, PA. Appendix D.
Available online at http://www.epa.gOv/Region3/mtntop/pdf/appendices/d/streams-invertebrate-study/FINAL.pdf.
Green, NB; Pauley, TK. (1987) Amphibians and reptiles in West Virginia. Pittsburgh, PA: University of Pittsburg
Press.
Greenwood, JL; Rosemond. AD. (2005) Periphyton response to long-term nutrient enrichment in a shaded
headwater stream. Can J Fish Aqua Sci 62(9):2033-2045.
Griffith, MB; Perry, SA; Perry, WB. (1995) Macroinvertebrate communities in headwater streams affected by acidic
precipitation in the Central Appalachians. J Environ Qual 24(2):233-238.
110
-------
Grosell, M; Nielsen, C; Bianchini, A. (2002) Sodium turnover rate determines sensitivity to acute copper and silver
exposure in freshwater animals. Comp Biochem Physiol C: Toxicol Pharmacol 133: 287-303.
Gulis, V; Suberkropp, K. (2004) Effects of whole-stream nutrient enrichment on the concentration and abundance of
aquatic Hypohomycete conidia in transport. Mycologia 96:57-65.
Gulis, V; Kuehn, KA Suberkropp, K. (2006) The role of fungi in carbon and nitrogen cycles in freshwater
ecosystems. In: Gadd; GM; ed. Fungi in biogeochemical cycles. Cambridge University Press: Cambridge, UK. pp
404-435.Hakala, JP; Hartman, KJ. (2004) Drought effect on stream morphology and brook trout (Salvelinus
fontinalis) populations in forested headwater streams. Hydrobiologia 515:203-213.
Hall, RO Jr.; Wallace, JB; Eggert, SL. (2000). Organic matter flow in stream food webs with reduced detrital
resource base. Ecology 81:12, 3445-3463.
Hansen, WF. (2001) Identifying stream types and management implications. Forest Ecol Manag 143:39-46.
Harding, JS; Benfield, EF; Bolstad, PV; et al. (1998) Stream biodiversity: The ghost of land use past. Proceedings
of the National Academy of Sciences 95:14843-14847.
Harding, LW; Graham, M; Paton, D. (2005) Accumulation of selenium and lack of severe effects on productivity of
American dippers (Cinclus mexicanus) and spotted sandpipers (Actitis macularia). Arch Environ Contam Toxicol
48:414-423.
Harman, WA; Unger, SJ; Fortney, RH. (2004) A natural channel design approach to stream restoration on reclaimed
surface mine lands. In. Barnhisel, R.I. (ed.), Proceedings of the Joint Conference of 21st Annual Meetings of the
American Society of Mining and Reclamation and 25th West Virginia Surface Mine Drainage Task Force
Symposium, Morgantown, WV. American Society of Mining and Reclamation, Lexington, KY.
Harrison, SSC; Pretty, JL; Shepherd, D; et al. (2004) The effect of instream rehabilitation structures on
macroinvertebrates in lowland rivers. J App Ecol 41:1140-1154.
Hartman, KJ; Kaller, MD; Howell, JW; et al. (2004) How much do valley fills influence headwater streams? Pages
822-846. In: Barnhisel, R.I. (ed.), Proceedings of the Joint Conference of 21st Annual Meetings of the American
Society of Mining and Reclamation and 25th West Virginia Surface Mine Drainage Task Force Symposium,
Morgantown, WV. American Society of Mining and Reclamation, Lexington, KY.
Hartman, KJ; Kaller, MD; Howell, JW; et al. (2005) How much do valley fills influence headwater streams?
Hydrobiologia 532:91-102.
Harvey, BJ. (1992) Energization of sodium absorption by the H+-ATPase pump in mitochondria-rich cells of frog
skin. J Exp Biol 172:289-309.
Hawkins, JW. (2004) Some hydrologic properties of surface mine spoil in the Appalachian Plateau. Pages 860-878.
In: Barnhisel, R.I. (ed.), Proceedings of the Joint Conference of 21st Annual Meetings of the American Society of
Mining and Reclamation and 25th West Virginia Surface Mine Drainage Task Force Symposium, Morgantown, WV.
American Society of Mining and Reclamation, Lexington, KY.
Hawkins, JW; Brady, KBC; Barnes, S; et al. (1996) Shallow ground water flow inunmined regions of the northern
Appalachian Plateau: Part 1. Physical characteristics. In: Annual Meeting of the American Society for Surface
Mining and Reclamation. Knoxville, TN, USA. p 42-51.
Hench, KR; Bissonnette, GK; Sexstone, AJ; et al. (2003) Fate of physical, chemical, and microbial contaminants in
domestic wastewater following treatment by small constructed wetlands. Water Res 37:921-927.
Ill
-------
Henry, RP; Wheatly, MG. (1992) Interaction of respiration, ion regulation, and acid-base balance in the everyday
life of aquatic crustaceans. AmerZool 32:407-416.
Hewlett, JD. (1982) Principles of forest hydrology. Athens, GA: The University of Georgia Press.
Hocutt, CH; Denoncourt, RF; Stauffer, JR, Jr. (1978) Fishes of the Greenbriar River, West Virginia, with drainage
history of the Central Appalachians. J Biogeogr 5(1):59-80.
Holl, KD; Cairns. J. (1994) Vegetational community-development on reclaimed coal surface mines in Virginia. Bull
Torrey Bot Club 121:327-337.
Hooke, RL. (1999) Spatial distribution of human geomorphic activity in the United States: Comparison with rivers.
Earth Surf Proc Land 24(8):687-692.
Houser, JN; Mulholland, PJ; Maloney, KO. (2005) Catchment disturbance and stream metabolism: patterns in
ecosystem respiration and gross primary production along a gradient of upland soil and vegetation disturbance. J N
AmBenthol Soc 24(8):538-552.
Howard, HS; Berrang, B; Flexner, M; et al. (2001) Kentucky mountaintop mining benthic macroinvertebrate survey:
Central Appalachian Ecoregion, Kentucky. In: Draft programmatic environmental impact statement on mountaintop
mining/valley fills in Appalachia - 2003. Appendix D. U.S. Environmental Protection Agency, Region 3,
Philadelphia, PA. Available online at http://www.epa.gov/Region3/mtntop/pdf/appendices/d/kentucky-
macroinvertebrate-study/report.pdf.
Hower, JC; Robertson, JD (2003) Clausthalite in coal. Int J Coal Geol. 53:219-225.
Hughes, JM; Schmidt, DJ; Finn, DS. (2009) Genes in streams: using DNA to understand the movement of
freshwater fauna and their riverine habitat. Bioscience 59:573-583.
Hughes, RM; Noss, RF. (1992) Biological diversity and biological integrity: Current concerns for lakes and streams.
Fishereies 17(2): 11-19.
Johnson, BR; Fritz, KM; Blocksom, KA; et al. (2009) Larval salamanders and channel geomorphology are
indicators of hydrologic permanence in forested headwater streams. Ecol Indica 9:150-159.
Johnson, BR; Haas, A; Fritz, KM. (2010) Use of spatially explicit physiochemical data to measure downstream
impacts of headwater stream disturbance. Water Res Res 46: W09526.
Kaplan, LA; Wiegner, TN; Newbold, JD; et al. (2008) Untangling the complex issue of dissolved organic carbon
uptake: A stable isotope approach. Freshw Biol 53:855-864.
Kapoor, NN. (1979) Osmotic regulation and salinity tolerance of the stonefly nymph Paragnetina media. Journal
of Insect Physiology 25:17-20.
Karr, JR. (1981) Assessment of biotic integrity using fish communities. Fisheries 6:21-27.
Kaufmann, PR; Robinson, EG. (1998) Physical habitat characterization. In: Surface waters: field operations and
methods for measuring the ecological condition of wadeable streams. U.S. Environmental Protection Agency,
Environmental Monitoring and Assessment Program, Washington, DC. EPA/620/R-94/004.
Kefford, BJ; Papas, PJ; Nugegoda, D. (2003) Relative salinity tolerance of macroinvertebrates from the Barwon
River, Victoria, Australia. Mar Freshwater Res 54(6):755-765.
Kefford, BJ; Papas, PJ; Metzeling, L; et al. (2004) Do laboratory salinity tolerances of freshwater animals
correspond with their field salinity? Environ Pollut 129(3):355-362.
112
-------
Kennedy, AJ; Cherry, DS; Currie, RJ. (2003) Field and laboratory assessment of a coal processing effluent in the
Leading Creek Watershed, Meigs County, Ohio. Arch Environ Contam Toxicol 44(3):324-331.
Kennedy, AJ; Cherry, DS; Currie, RJ. (2004) Evaluation of ecologically relevant bioassays for a lotic system
impacted by a coal-mine effluent, using Isonychia. Environ Monit Assess 95:37-55.
Kennedy, AJ; Cherry, DS; Zipper, CE. (2005) Evaluation of ionic contribution to the toxicity of a coal-mine effluent
using Ceriodaphnia dubia. Arch Environ Contam Toxicol 49:155-162.
Ketola, HG; Longacre, D; Greulich, A; et al. (1988) High calcium-concentration in water increases mortality of
salmon and trout eggs. Prog Fish-Cult 50:129-135.
Kiffney, PM; Clements, WH. (1996) Size-dependent response of macroinvertebrates to metals in experimental
streams. Environ Toxicol Chem 15(8):1352-1356.
Kimball, BA; Callender, E; Axtmann, EV. (1995) Effects of colloids on metal transport in a river receiving acid
mine drainage, upper Arkansas River, Colorado, U.S.A. Appl Geochem 10(3):285-306.
Kimmel, WG; Argent, DG. (2006) Development and application of an index of biotic integrity for fish communities
of wadeable Monongahela River tributaries. J Freshwater Ecol 21 (2): 183-190.
Kimmel, WG; Argent, DG. (2010) Stream fish community responses to a gradient of specific conductance. Water
Air Soil Pollut 206:49-56.
Kirchner, F; Stout, B; Wallace, JB. (2003) A survey of eight major aquatic insect orders associated with small
headwater streams subject to valley fills from mountaintop mining. In: Appendix D in Programmatic Environmental
Impact Statement on Mountaintop Mining/Valley Fills in Appalachia (2003, 2005). Philadelphia, PA: U.S.
Environmental Protection Agency, Region 3.
Kirk, E. (1999) An evaluation of the aquatic habitats provided by sediment control ponds and other aquatic
enhancement structures located on mine permitted areas in southern West Virginia. In: Draft programmatic
environmental impact statement on mountaintop mining/valley fills in Appalachia - 2003. Appendix D. U.S.
Environmental Protection Agency, Region 3, Philadelphia, PA. Available online at
http://www.epa.gOv/region03/mtntop/pdf/appendices/d/aquatic-ecosy stem-enhanc-
symp/Proceedings/support/Maggard/PondStudy2.pdf.
Kirk, EJ; Maggard, R. (2004) Long-term downstream impacts of surface mining and valley fill construction to
benthic macroinvertebrates and water quality. In: Barnhisel, RI, editor. Joint Confererence of 21st Annual Meetins
of the American Society of Mining and Reclamation and 25th West Virginia Surface Mine Drainage Task Force
Symposium. Morgantown, WV: American Society of Mining and Reclamation.
Kirschner, LB. (1970) The study of NaCl transport in aquatic animals. Am. Zool. 10:365-376.
Kirschner, LB. (2004) The mechanism of sodium chloride uptake in hyperregulating aquatic animals. J Exp Biol
207:1439-1452.
Kominick, H. (1977) Chloride cells and chloride epithelia of aquatic insects. Int J Cytol 49:285-328.
Kondratieff, BC; Voshell, JR. (1984) The North and Central American species of Isonychia (Ephemeroptera:
Oligoneuriidae). Trans Am Entomol Soc 110:129-144.
Kozak, KH; Elaine, RA; Larson, A. (2006) Gene lineages and eastern North American palaeodrainage basins:
phylogeography and speciation in salamanders of the Eurycea bislineata species complex. Mol Ecol 15:191-207.
113
-------
Larsen, D; R. Mann, R. (2005) Origin of high manganese concentrations in coal mine drainage, eastern Tennessee.
J GeochemExplor 86:143-163.
Lasier, PJ; Hardin, IR. (2010) Observed and predicted reproduction of Ceriodaphnia dubia exposed to chloride,
sulfate, and bicarbonate. Environ Toxicol Chem 29:347-358.
Lasier, PJ; Winger, PV; Reinert, RE. (1997) Toxicity of alkalinity to Hyalella azteca. Bull Environ Contam Toxicol
59(5):807-814.
Lasier, PJ; Winger, PV; Bogenrieder, KJ. (2000) Toxicity of manganese to Ceriodaphnia dubia and Hyalella azteca.
Arch Environ Contam Toxicol 38(3):298-304.
Layzer, JB; Gordon, ME; Anderson, RM. (1993) Mussels: The forgotten fauna of regulated rivers. A case study of
the Caney Fork River. Reg Rivers: Res Manag 8:63-71.
Lee, G; Bigham, JM; Faure, G. (2002) Removal of trace metals by coprecipitation with Fe, Al, and Mn from natural
waters contaminated with acid mine drainage in the Ducktown Mining District, Tennessee. Appl Geochem
17(5):569-581.
Lemly, AD. (1993) Guidelines for evaluating selenium data from aquatic monitoring and assessment. Environ
Monit Assess 28:83-100.
Leopold, LB. (1964) Fluvial processes in geomorphology. San Francisco, CA: W.H. Freeman and Company.
Leopold, LB. (1994) A View of the River. Cambridge, MA, USA: Harvard University Press.
Lepori, F; Palm, D; Malmqvist, B. (2005) Effects of stream restoration on ecosystem functioning: detritus
retentiveness and decomposition. Journal of Applied Ecology 42:228-238.
Letcher, BH; Nislow, KH; Coombs, JA; et al. (2007) Population Response to Habitat Fragmentation in a Stream-
Dwelling Brook Trout Population. PLoS ONE 2:el 139.
Linton, TK; Pacheco, MAW; Mclntyre, DO; et al. (2007) Development of bioassessment-based benchmarks for
iron. Environ Toxicol Chem 26(6): 1291-1298.
Loveland, T; Gutman, G; Buford, M; et al. (2003) Land-use/land cover change. In: Mahoney, JR; ed. Strategic plan
for the U.S. climate change science program. U.S. Climate Change Science Program, Washington, DC; pp 63-70.
Lowe, WH; Likens, GE; Power, ME. (2006) Linking scales in stream ecology. Bioscience 56(7):591-597.
Luoma, SN; Cain, DJ; Rainbow, PS. 2010. Calibrating biomonitoris to ecological disturbance: A new technique for
explaining metal effects in natural waters. Integ Environ Assess Manag 6(2): 199-209.
Luoma, SN; Presser, TS. (2009) Emerging opportunities in management of selenium contamination. Environ Sci
Technol 43:8483-8487.
MacDonald, DD; Ingersoll, CG; Berger, TA. (2000) Development and evaluation of consensus-based sediment
quality guidelines for freshwater ecosystems. Arch Environ Contam Toxicol 39(l):20-31.
Maetz, J; Garcia Romeu, F. (1964) The mechanism of sodium and chloride uptake by the gills of a freshwater fish,
Carassius auratus. II. Evidence forNH4+/Na+ and HCO37Cr exchanges. J. Gen. Physiol. 47:1209-1227.
114
-------
Maggard, R; Kirk, E. (1998) Downstream impacts of surface mining and valley fill construction. Report for Pen
Coal Corporation. In: Draft programmatic environmental impact statement on mountaintop mining/valley fills in
Appalachia - 2003. Appendix D. U.S. Environmental Protection Agency, Region 3, Philadelphia, PA. Available
online at http://www.epa.gOv/Region3/mtntop/pdf/appendices/d/aquatic-ecosystem-enhanc-
symp/Proceedings/support/Maggard/do wnstream-impacts.pdf.
Martinez-Jeronimo, F; Martinez-Jeronimo, L. (2007) Chronic effect of NaCl salinity on a freshwater strain of
Daphnia magna Straus (Crustacea: Cladocera): A demographic study. Ecotoxicol Environ Saf 67(3):411-416.
Mayer, PM; Reynolds, SK; McCutchen, MD. (2007) Meta-analysis of nitrogen removal in riparian buffers. J
Environ Qua! 36(4): 1172-1180.
McGuire, KJ; McDonnell, JJ. (2010) Hydrological connectivity of hillslopes and streams: Characteristic time scales
and nonlinearities. Water Resour Res 46:17.
McGuire, KJ; McDonnell, JJ; Weiler, M; et al. (2005) The role of topography on catchment-scale water residence
time. Water Resour Res 41(5):W05002.
McKeown, PE; Hocutt, CH; Morgan, RP II; et al. (1984) An eletrophoretic analysis of the Etheostoma variatum
complex (Percidae: Etheostomatini), with associated zoogeographic considerations. Environ Biol Fish 11(2):85-95.
McMahon, TE; Zale, AV; Orth, DJ. (1996) Aquatic habitat measurements. In: Murphy, BR; Willis, DW; eds.
Fisheries techniques, 2nd edition. Chapter 4. Bethesda, MD: American Fisheries Society.
Merricks, TC; Cherry, DS; Zipper, CE; et al. (2007) Coal-mine hollow fill and settling pond influences on
headwater streams in southern West Virginia, USA. Environ Monit Assess 129(l-3):359-378.
Merritt, RW; K. W. Cummins, KW. eds. (1996) An introduction to the aquatic insects of North America. 3rd ed.
Dubuque, IA: Kendall/Hunt Publishing Company.
Messinger, T. (2003) Comparison of storm response of streams in small, unmined and valley-filled watersheds,
1999-2001, Ballard Fork, West Virginia. Water-Resources Investigations Report 02-4303, Department of the
Interior, U.S. Geological Survey, Charleston, WV.
Messinger, T; Paybins, KS. (2003) Relations between precipitation and daily and monthly mean flows in gaged,
unmined and valley-filled watersheds, Ballard Fork, West Virginia, 1999-2001. Water-Resources Investigations
Report 03-4113, Department of the Interior, U.S. Geological Survey, Charleston, WV.
Meyer, JL; Wallace, JB. (2001) Lost linkages and lotic ecology: rediscovering small streams. In: Press, MC;
Huntly, NJ; Levin, S: eds. Ecology: achievement and challenge. Maiden, MA: Blackwell Science; pp 295-317.
Meyer, JL; Strayer, DL; Wallace JB; et al. (2007) The contribution of headwater streams to biodiversity in river
networks. J Am Water Res Assoc 43(1):87-103.Meyer, JL; Strayer, DL; Wallace, JB; et al. (2007) The contribution
of headwater streams to biodiversity in river networks. J Am Water Resour Assoc 43:86-103.
Meyer, JS; Sanchez, DA; Brookman, JA; et al. (1985) Chemistry and aquatic toxicity of raw oil shale leachates from
Piceance Basin, Colorado. Environ Toxicol Chem 4(4):559-572.
Minshall, GW. (1967) Role of allochthonous detritus in the tropic structure of a woodland springbrook community.
Ecology 48:139-149.
Morse, JC; Stark, BP; McCafferty, WP. (1993) Southern Appalachian streams at risk - implications for mayflies,
stoneflies, caddisflies, and other aquatic biota. Aqua Conserv Marine Freshw Ecosys 3:293-303.
115
-------
Morse, JC; Stark, BP; McCafferty, WP; et al. (1997) Southern Appalachian and other southeastern streams at risk:
Implications for mayflies, dragonflies, stoneflies, and caddisflies. In: Benz, GW; Collins, DE; eds. Aquatic fauna in
peril: The southeastern perspective. Special publication 1. Decatur, GA: Southeast Aquatic Research Institute. Lenz
Design and Communications, pp 17-42.
Mount, DR; Gulley, DD; Hockett, R; et al. (1997) Statistical models to predict the toxicity of major ions to
Ceriodaphnia dubia, Daphnia magna, andPimephalespmmelas (fathead minnows). Environmental Toxicology and
Chemistry 16(10):2009-2019.
Nadeau, TL; Rains, MC. (2007) Hydrological connectivity between headwater streams and downstream waters:
How science can inform policy. J Am Water Resour Assoc 43:118-133.
Naiman, RJ; Decamps, H. (1997) The ecology of interfaces: Riparian zones. Annu Rev Ecol Systematics
28:621-658.
National Geographic Maps and NatureServe. (2008) Hot spots of rarity-weighted richness for critically imperiled
and imperiled species. National Geographic and NatureServe, Washington, DC. Available online at
http://www.landscope.org/search/, keywords: rarity-weighted species richness map (accessed 12/16/2009).
Negley, TL; Eshleman, KN. (2006) Comparison of stormflow responses of surface-mined and forested watersheds
in the Appalachian Mountains, USA. Hydrol Proc 20(16):3467-3483.
Neuzel, SG; Dulong, FT; Cecil, CB; et al. (2007) Selenium concentrations in Middle Pennsylvanian coal-bearing
strata in the Central Appalachian Basin. Open-File Report 2007-1090. Department of Interior, U.S. Geological
Survey, Reston, VA.
Neuzil, SG. (2001) Chapter I—Summary report on the coal resources, coal production, and coal quality of the
Allegheny Group No. 5 Block, and the Pottsville Group Stockton and Coalburg, Winifrede/Hazard,
Williamson/Amburgy, Campbell Creek/Upper Elkhorn No. 3, and Upper Elkhorn Nos. 1 and 2/Powellton Coal
Zones, Central Appalachian Basin Coal Region, in Northern and Central Appalachian Basin Coal Regions
Assessment Team, 2000 resource assessment of selected coal beds and zones in the northern and central
Appalachian Basin coal regions: U.S. Geological Survey Professional Paper 1625-C, CD-ROM, version 1.0.
Neuzil, SG; Dulong, FT; Cecil, CB. (2005) Spatial trends in ash yield, sulfur, selenium, and other selected trace
element concentrations in coal beds of the Appalachian Plateau region, U.S.A. Open-File Report 2005-1330.
Department of Interior, U.S. Geological Survey, Reston, VA.
Neves, RJ. and J.C. Widlak. 1987. Habitat ecology of juvenile freshwater mussels (Bivalvia: Unionidae) in a
headwater stream in Virginia. Am Malacol Bull 5(1): 1-7.
Niemi, GJ; DeVore, P; Detenbeck, N; et al. (1990) Overview of case studies on recovery of aquatic systems from
disturbance. Environmental Management 14:571-587
Nordstrom, DK; Ball, JW. (1986) The geochemical behavior of aluminum in acidified surface waters. Science
232:54-56.
Northington, RM; Hershey, AE. (2006) Effects of stream restoration and wastewater treatment plant effluent on fish
communities in urban streams. Freshwater Biol 51:1959-1973.
Ohio EPA (Environmental Protection Agency). (2002) Amphibian index of biotic integrity (AmphilBI) for
wetlands. Ohio Environmental Protection Agency, Division of Surface Water, Wetland Ecology Group, Columbus,
OH. Available online at http://www.epa.ohio.gov/portal/35/wetlands/2002_Amphibian_report_final_rev.pdf.
Ohlendorf, HA; Hoffman, DJ; Rattner, BA; et al. (2003) Ecotoxicology of selenium. In: Handbook of
Ecotoxicology, 2nd ed. Boca Raton, FL: Lewis Publishers, pp 465-500.
116
-------
Orr, PL; Guiguer, KR; Russel, CK. (2005) Food chain transfer of selenium in lentic and lotic habitats of a western
Canadian watershed. Ecotoxicol Environ Saf 63:175-188.
Paine, RT; Tegner, MJ; Johnson, EA. (1998) Compounded perturbations yield ecological surprises. Ecosystems
1:535-545.
Palmer, MA. (2009) Reforming watershed restoration: science in need of application and applications in need of
science. Estuar Coast 32:1-17.
Palmer, MA; Bernhardt, ES; Schlesinger, WH; et al. (2010) Mountaintop mining consequences. Science 327:148-
149.
Pan, YD; Stevenson, RJ; Hill, BH; et al. (2000) Ecoregions and benthic diatom assemblages in Mid-Atlantic
Highlands streams, USA. J N AmBenthol Soc 19:518-540.
Pawlowicz, R. (2008) Calculating the conductivity of natural waters. Limnol OceanogrMeth 6:489-501.
Paybins, KS. (2003) Flow origin, drainage area, and hydrologic characteristics for headwater streams in the
mountaintop coal-mining region of southern West Virginia, 2000-01. Water-Resources Investigations Report
02-4300, Department of the Interior, U.S. Geological Survey, Charleston, WV. Available online at
http://pubs.usgs.gov/wri/wri02-4300/pdf/wri02-4300.book.pdf.
Paybins, KS; Messinger, T; Eychaner, JH; et al. (2000) Water quality in the Kanawha-New River Basin: West
Virginia, Virginia, and North Carolina, 1996-98. Circular 1204. Department of the Interior, U.S. Geological Survey,
Charleston, WV. Available online at http://pubs.usgs.gov/circ/circl204/#pdf.
Payne, WJ. (1981) Denitrification. New York: Wiley.
Pierce, SK (1982) Invertebrate cell volume control mechanisms: A coordinated use of intracellular amino acids and
inorganic ions as an osmotic solute. The Biological Bulletin 163:405-419.
Perry, SF. (1997) The chloride cell: Structure and function in the gills of freshwater fishes. Ann Rev Physiol
59:325-347.
Peterman, WE; Crawford, JA; Semilitsch, RD. 2008. Productivity and significance of headwater streams:
population structure and bio mass of the black-bellied salamander (Desmognathus quadramaculatus) Freshwater
Biology 53: 347-357.
Peterson, BJ; Wollheim, WM; Mulholland, PJ; et al. (2001) Control of nitrogen export from watersheds by
headwater streams. Science 292:86-90.
Petranka, JW. (1983) Fish predation: A factor affecting the spatial distribution of a stream-breeding salamander.
Copeia 1983(3):624-628.
Petranka, JW. (1998) Salamanders of the United States and Canada. Washington, DC: Smithsonian Institution
Press.
Petranka, JW; Eldridge, ME; Haley, KE. (1993) Effects of timber harvesting on southern Appalachian salamanders.
Conserv Biol 7(2):363-370.
Petty, JT; Lamothe, PJ; Mazik, PM. (2005) Spatial and seasonal dynamics of brook trout populations inhabiting a
central Appalachian watershed. Trans Am Fish Soc 134:572-587.
Petty, JT; Fulton, JB; Strager, MP; et al. (2010) Landscape indicators and thresholds of stream ecological
impairment in an intensively mined Appalachian watershed. J N Am Benthol Soc 29:1292-1309.
117
-------
Phippen, B; Horvath, C; Nordin, R; et al. (2008) Ambient water quality guidelines for iron. Science and Information
Branch, Ministry of Environment, Province of British Columbia; 46 p. Available online at
http://www.env.gov .be.ca/wat/wq/BCguidelines/iron/iron_tech.pdf.
Pollard, AI; Yuan, L. (2009) Assessing the consistency of response metrics of the invertebrate benthos: a
comparison of trait- and identity-based measures. Freshwater Biol doi:10.1111/j.1365-2427.2009.02235.x.
Ponader, KC; Potapova, MG. (2007) Diatoms from the genus Achnanthidium in flowing waters of the Appalachian
Mountains (North America): ecology, distribution and taxonomic notes. Limnologica 37:227-241.
Pond, GJ. (2004) Effects of surface mining and residential land use on headwater stream biotic integrity in the
eastern Kentucky coalfield region. Kentucky Department of Environmental Protection, Division of Water,
Frankfort, KY. Available online at http://www.water.ky.gov/NR/rdonlyres/ED76CE4E-F46A-4509-8937-
lA5DA40F3838/0/coal_miningl.pdf.
Pond GJ. (2010) Patterns of Ephemeroptera taxa loss in Appalachian headwater streams (Kentucky, USA).
Hydrobiologia 641(1):185-201.
Pond, GJ; McMurray, SE. (2002) A macroinvertebrate bioassessment index for headwater streams of the eastern
coalfield region, Kentucky. Kentucky Department of Environmental Protection, Division of Water, Frankfort, KY.
http://www.water.ky.gOv/NR/rdonlyres/4CA8D7C4-309B-4175-ACC4-lCBDDDF73798/0/EKyMBI.pdf.
Pond, GJ; Passmore, ME; Borsuk, FA; et al. (2008) Downstream effects of mountaintop coal mining: comparing
biological conditions using family- and genus-level macroinvertebrate bioassessment tools. J N Am Benthol Soc
27:717-737.
Potter, KM, Frampton, J; Josserand, SA; et al. (2010) Evolutionary history of two endemic Appalachian conifers
revealed using microsatellite markers. Conserv Genet 11:1499-1513.
Presser, TS; Luoma, SN, (2009) Modeling of selenium for the San Diego Creek watershed and Newport Bay,
California. U.S. Geological Survey Open-File Report 2009-1114, 48 p. Available online at
http://pubs.usgs.gov/of/2009/! 114/.
Pusch, M; Fiebig, D; Brettar, I; et al. (1998) The role of micro-organisms in the ecological connectivity of running
waters. Freshwater Biol 40(3):453-495.
Richardson, DM; Holmes, PM; Esler, KJ; et al. (2007) Riparian vegetation: degradation, alien plant invasions, and
restoration projects. Diversity and Distributions 13:126-139.
Roe, GH. (2005) Orographic precipitation. Annu Rev Earth Plan Sci 33:645-671.
Rose, AW; Cravotta, CA, III. (1998) Geochemistry of coal mine drainage. In: Brady, KBC; Smith, MW; Schueck, J;
eds. Coal mine drainage prediction and pollution prevention in Pennsylvania. Harrisburg, PA, Pennsylvania
Department of Environmental Protection: 1.1-1.22. Available online at
http://www.techtransfer.osmre.gov/nttmainsite/Library/pub/cmdpppp/chapterl.pdf.
Rosgen, DL. (1994) A classification of natural rivers. Catena 22:169-199.
Routman, E; Wu, R; Templeton, AR. (1994) Parsimony, molecular evolution, and biogeography - the case of the
North-American Giant Salamander. Evolution 48:1799-1809.
Rudolph, BL; Andreller, I; Kennedy, CJ. (2008) Reproductive success, early life stage development, and survival of
westslope cutthroat trout (Oncorhynchus clarki lewisi) exposed to elevated selenium in an area of active coal
mining. Environ Sci Technol 42(8):3109-3114.
118
-------
Scholes, L; Shutes, RBE; Revitt, DM; et al. (1998) The treatment of metals in urban runoff by constructed wetlands.
Sci Total Environ 214:211-219.
Schorer, M; Symader, W. (1998) Biofilms as dynamic components for the sorption of inorganic and organic
pollutants in fluvial systems. In: Haigh, MJ; Krecek, J; Rajwar, GS; Kilmartin, MP; eds. Headwaters: water
resources and soil conservation. Rotterdam, Netherlands: A.A. Balkema; pp 187-196.
Shank, M. (2004) Development of a mining fill inventory from multi-date elevation data. West Virginia
Department of Environmental Protection, Charleston, WV. Available online at
http://gis.wvdep.org/tagis/projects/valley_fill_paper.pdf (accessed 11/5/2009).
Sheoran, AS; Sheoran, V. (2006) Heavy metal removal mechanism of acid mine drainage in wetlands: A critical
review. Minerals Engineer 19:105-116.
Skaar, D; Farag, A; Harper, D. (2006) Toxicity of sodium bicarbonate to fish from coal-bed natural gas production
in the Tongue and Powder River drainages, Montana and Wyoming. Fact Sheet 2006-3092 U.S. Geological Survey,
U.S. Department of the the Interior.
Skousen, J; Renton, J; Brown, H; et al. (1997) Neutralization potential of overburden samples containing siderite. J
Environ Qua! 26:673-681.
Skouson, J.G., Sexstone, A; Ziemkiewicz, PF. (2000) Acid mine drainage control and treatment. In: Barnhisel, RI;
Darmody, RG; Daniels, WL; eds. Reclamation of drastically disturbed lands. Agron. Monogr. 41. Madison, WI:
American Society of Agronomy; pp 131-168.
Skousen, J; Ziemkiewicz, P; Venable, C. (2006) Tree recruitment and growth on 20-year-old, unreclaimed surface
mined lands in West Virginia. Int J Min Reclam Environ 20(2): 142-154.
Skousen, J; Gorman, J; Pena-Yewtukhiw, E; et al. (2009) Hardwood tree survival in heavy ground cover on
reclaimed land in West Virginia: Mowing and ripping effects. J Environ Qual 38(4): 1400-1409.
Smith, ME; Lazorchak, JM; Herrin, LE; et al. (1997) A reformulated, reconstituted water for testing the freshwater
amphipod, Hyalella azteca. Environ Toxicol Chem 16(6):1229-1233.
Smith, SL; MacDonald, DD; Keenleyside, KA; et al. (1996) A preliminary evaluation of sediment quality
assessment values for freshwater ecosystems. J Great Lakes Res 22(3):624-638.
Sobek, A; Schuller, W; Freeman, J; et al. (1978) Field and laboratory methods applicable to overburdens and
minesoils. U.S. Environmental Protection Agency, Office of Research and Development, Industrial Environmental
Research Laboratory, Cincinnati, OH. EPA-600/2-78-054. Available online at
http://www.techtransfer.osmre.gov/nttmainsite/Library/hbmanual/fieldlab/front.pdf.
Soltis, DE; Morris, AB; McLachlan, JS; et al. (2006) Comparative phylogeography of unglaciated eastern North
America. MolEcol 18:4261-4293.
Soucek, DJ. (2007a) Bioenergetic effects of sodium sulfate on the freshwater crustacean, Ceriodaphnia dubia.
Ecotoxicology 16(3):317-325.
Soucek, DJ. (2007b) Comparison of hardness- and chloride-regulated acute effects of sodium sulfate on two
freshwater crustaceans. Environ Toxicol Chem 26(4):773-779.
Soucek, DJ; Kennedy, AJ. (2005) Effects of hardness, chloride, and acclimation on the acute toxicity of sulfate to
freshwater invertebrates. Environ Toxicol Chem 24(5):1204-1210.
119
-------
Soucek, DJ; Cherry, DS; Currie, RJ; et al. (2000) Laboratory to field validation in an integrative assessment of an
acid mine drainage-impacted watershed. Environ Toxicol Chemistry 19(4): 1036-1043.
Southerland, MT (1986) The effects of variation in streamside habitats on the composition of mountain salamander
communities. Copeia: American Society of Ichthyologists andHerpetologists.3:731-741.
Stauffer, JR; Ferreri, CP. 2002 Characterization of stream fish assemblages in selected regions of mountain top
removal/valley fill coal mining. In: Draft programmatic environmental impact statement on mountaintop mining/
valley fills in Appalachia - 2003. Appendix D. U.S. Environmental Protection Agency, Region 3, Philadelphia, PA.
Available online at http://www.epa.gov/Region3/mtntop/pdf/appendices/d/fisheries-study/staufferferreri-
oct2002.pdf.
Strahler, AN. (1957) Quantitative analysis of watershed geomorphology. Am Geophys Union Transact 38:913-920.
Stubblefield, WA; Brinkman, SE; Davies, PH; et al. (1997) Effects of water hardness on the toxicity of manganese
to developing brown trout (Salmo trutta). Environ Toxicol Chem 16:2082-2089.
Stumm, W; Morgan, JJ. (1996) Aquatic chemistry: chemical equilibria and rates in natural waters. New York, NY:
John Wiley & Sons, Inc.
Sudduth, EB; Meyer, JL. (2006) Effects of bioengineered streambank stabilization on bank habitat and
macroinvertebrates in urban streams. Environ Manag 38:218-226.
Sweeney, BW; Funk, DH; Standley, LJ. (1993) Use of the stream mayfly, Cloeon triangulifer, as a bioassay
organism: Life history response and body burden following exposure to technical Chlordane. Environ Toxicol
Chem 12:115-125.
Swift, MC. (2002) Stream ecosystem response to, and recovery from, experimental exposure to selenium. J Aqua
Ecosyst Stress Rec 9:159-184.
Talbot, JDR; House, WA; Pethybridge, AD. (1990) Prediction of the temperature dependence of electrical
conductance for river waters. Wat. Res. 24:1295-1304.
Tank, JL; Rosi-Marshall, EJ; Griffiths, NA; et al. (2010) A review of allochthonous organic matter dynamics and
metabolism in streams. J N Am Benthol Soc 29:118-146.
Taylor, TJ; Agouridis, CT; Warner, RC; et al. (2009a) Hydrological characteristics of Appalachian loose-dumped
spoil in the Cumberland Plateau of eastern Kentucky. Hydrol Proc 23:3372-3381.
Taylor, TJ; Agouridis, CT; Warner, RC; et al. (2009b) Runoff curve numbers for loose-dumped spoil in the
Cumberland Plateau of eastern Kentucky. Int J Mining Reclam Enviornm 23(2): 103-120.
Tewalt, SJ; Ruppert, LF; Bragg, LJ; et al. (2001) Chapter F—A digital resource model of the Middle Pennsylvanian
Fire Clay coal zone, Pottsville Group, central Appalachian Basin coal region, in Northern and Central Appalachian
Basin Coal Regions Assessment Team, 2000 resource assessment of selected coal beds and zones in the northern
and central Appalachian Basin coal regions. U.S. Geological Survey Professional Paper 1625-C. CD-ROM, version
1.0.
Thorp, JH; Covich, AP; eds. (1991) Ecology and classification of North American freshwater invertebrates. San
Diego: Academic Press, Inc.
Tietge, JE; Hockett, JR; Evans, JM. (1997) Major ion toxicity of six produced waters to three freshwater species:
Application of ion toxicity models and tie procedures. Environ Toxicol Chem 16(10):2002-2008.
120
-------
Timpano, AJ; Schoenholtz, SH; Zipper, CE; et al. (2010) Isolating effects of total dissolved solids on aquatic life in
central Appalachian coalfield streams. Proc Am Soc Mining Reclam 27:1284-1302.
Townsend, PA; Helmers, DP; Kingdon, CC; et al. (2009) Changes in the extent of surface mining and reclamation in
the Central Appalachians detected using a 1976-2006 Landsat time series. Rem Sens Environ 113:62-72.
Tsuboyama, Y; Sidle, RC; Noguchi, S; et al. (1994) Flow and solute transport through the soil matrix and
macropores of a hillslope segment. Water Resour Res 30:879-890.
U.S. EPA (Environmental Protection Agency). (1979) Environmental assessment of surface mining methods: head-
of-hollow fill and mountaintop removal. Industrial Environmental Research Laboratory, Cincinnati, OH;
EPA/600/7-79/062.
U.S. EPA (Environmental Protection Agency). (2002) Landscape Scale Cumulative Impact Study. In: Draft
programmatic environmental impact statement on mountaintop mining/valley fills in Appalachia - 2003. Appendix I.
U.S. Environmental Protection Agency, Region 3, Philadelphia, PA. Available online at
http ://www. epa. gov/Region3/mtntop/eis2003 appendices. htm#appd.
U.S. EPA (Environmental Protection Agency). (2003) Draft programmatic environmental impact statement on
mountaintop mining/valley fills in Appalachia. U.S. Environmental Protection Agency, Region 3, Philadelphia, PA.
Available online at http://www.epa.gov/Region3/mtntop/eis2003.htm.
U.S. EPA (Environmental Protection Agency). (2004) Draft ambient aquatic life criteria for selenium - 2004.,
Office of Water, Washington, DC; EPA/822/R-04/001. Available online at
http ://www. epa. gov/waterscience/criteria/selenium/pdfs/complete .pdf.
U.S. EPA (Environmental Protection Agency). (2005) Mountaintop mining/valley fills in Appalachia. Final
programmatic environmental impact statement. U.S. Environmental Protection Agency, Region 3, Philadelphia, PA.
Available online at http://www.epa.gov/region3/mtntop/pdf/mtm-vf_fpeis_full-document.pdf.
U.S. EPA (Environmental Protection Agency). (2011) Review of EPA's draft report on aquatic ecosystem effects of
mountaintop mining and valley fills. Science Advisory Board, Washington, DC; EPA-SAB-11-005. Available
online at http://www.epa.gov/sab.
van Dam, RA; Hogan, AC; McCullough, CD; et al. (2010) Aquatic toxicity of magnesium sulfate, and the influence
of calcium, in very low ionic concentration water. Environ Toxicol Chem 29:410-421.
Vannote, RL; Minshall, GE; Cummins, KW; et al. (1980) River continuum concept. Can J Fish Aqua Sci
37:130-137.
Vesper, DJ; Roy, M; Rhoads, CJ. (2008) Selenium distribution and mode of occurrence in the Kanawha Formation,
southern West Virginia, U.S.A. Int J Coal Geol 73(3-4):237-249.
Vinson, MR; Hawkins, CP. (1998) Biodiversity of Stream Insects: Variation at Local, Basin, and Regional Scales.
Annual Review of Entomology 43: 271-293.
Vuori, KM. (1995) Direct and indirect effects of iron on river ecosystems. Annales Zoologici Fennici 32:317-329.
Vymazal, J. (2007) Removal of nutrients in various types of constructed wetlands. Sci Total Environ 380:48-65.
Wallace, JB. 1990. Recovery of lotic macroinvertebrate communities from disturbance. Environ Manag
14(5):605-620.
Wallace, JB; Eggert, SL; Meyer, JL; et al. (1997) Multiple trophic levels of a forest stream linked to terrestrial litter
inputs. Science 277:102-104.
121
-------
Wallace, JB; Eggert, SL; Meyer, JL; et al. (1999) Effects of resource limitation on a detrital-based ecosystem. Ecol
Monogr 69:409-442.
Warnick, SL; Bell, HL. (1969) The acute toxicity of some heavy metals to different species of aquatic insects. J
Water Pollut Con Fed 41:280-284.
Webster, JR; Benfield, EF; Ehrman, TP; et al. (1999) What happens to allochthonous material that falls into
streams? A synthesis of new and published information from Coweeta. Freshwater Biol 41(4):687-705.
Webster, JR; Tank, JL; Wallace, JB; et al. (2000) Effects of litter exclusion and wood removal on phosphorus and
nitrogen retention in a forest stream. Verh Internal Verein Limnol 27:1337-1340.
Welsh, HH, Jr; Droege S. (2001) A case for using plethodontid salamanders for monitoring biodiversity and
ecosystem integrity of North American forests. ConservBiol 15(3):558-569.
Welsh, HH, Jr; Ollivier, LM. (1998) Stream amphibians as indicators of ecosystem stress: A case study from
California's redwoods. Ecol Appl 8(4): 1118-1132.
Wheatly, MG; Gannon, AT. (1995) Ion regulation in crayfish: Freshwater adaptations and the problem of molting.
AmerZool 35:49-59.
Wiley, JB; Brogan, FD. (2003) Comparison of peak discharges among sites with and without valley fills for the July
8-9, 2001, flood in the headwaters of Clear Fork, Coal River basin, mountaintop coal-mining region, southern West
Virginia. Report 03-133, Department of the Interior, U.S. Geological Survey, Charleston, WV. Available online at
http://pubs.usgs.gov/of/2003/ofr03-133/pdf/ofr03133.pdf.
Wiley, JB; Evaldi, RD; Eychaner, JH; et al. (2001) Reconnaissance of stream geomorphology, low streamflow, and
stream temperature in the mountaintop coal-mining region, southern West Virginia, 1999-2000. U.S. Department of
the Interior, U.S. Geological Survey, Charleston, WV. Available online at
http://pubs.usgs.gov/wri/wri014092/pdf/wri01-4092.book_new.pdf.
Williams, DD. (1996) Environmental constraints in temporary fresh waters and their consequences for the insect
fauna. J N AmBenthol Soc 15(4):634-650.
Williams, JM; Wood, PB. (2004) Streamside salamanders in valley fill and reference streams in southern West
Virginia. Pages 2027-2041. In: Barnhisel, RI (ed.), Proceedings of the Joint Conference of 21st Annual Meetings of
the American Society of Mining and Reclamation and 25th West Virginia Surface Mine Drainage Task Force
Symposium, Morgantown, WV. American Society of Mining and Reclamation, Lexington, KY.
Winter, TC. (2007) The role of groundwater in generating streamflow in headwater areas and in maintaining base
flow. J Am Water Resour Assoc 43:15-25.
Wipfli, MS; Gregovich, DP. (2002) Export of invertebrates and detritus from fishless headwater streams in
southeastern Alaska: implications for downstream salmonid production. Freshwater Biol 47(5):957-969.
Wipfli, MS; Richardson, JS; Naiman, RJ. (2007) Ecological linkages between headwaters and downstream
ecosystems: Transport of organic matter, invertebrates, and wood down headwater channels. J Am Water Resour
Assoc 43:72-85.
Wolman, MG. (1954) A method of sampling coarse river-bed material. Trans Am Geophys Union 35(6):951-956.
122
-------
Wood, PB; Edwards, JW; Weakland, CA; Balcerzak, MJ; Chamblin, HD. 2001. Mountaintop removal mining/valley
fill environmental impact statement technical study project report for terrestrial studies. Terrestrial vertebrate
(breeding songbird, raptor, small mammal, herpetofaunal) populations of forested and reclaimed sites. Appendix E
in Draft Programmatic Environmental Impact Statement on Mountaintop Mining / Valley Fills in Appalachia - 2003.
U.S. Environmental Protection Agency, Philadelphia PA. Available online at
http://www.epa.gOv/region03/mtntop/pdf/appendices/e/vertebrate-study/vertebratestudy.pdf
Wunsch, DR; Dinger, JS; Graham, CDR. (1999) Predicting ground-water movement in large mine spoil areas in the
Appalachian Plateau. Int J Coal Geol 41:73-106.
WVDEP (West Virginia Department of Environmental Protection). (2009a) Environmap Explorer. West Virginia
Department of Environmental Protection, Charleston, WV. Available online at
http://gis.wvdep.org/imap/index.html (accessed 10/27/2009).
WVDEP (West Virginia Department of Environmental Protection). (2009b) West Virginia surface mining
reclamation rule. 38 CSR § 2. State of West Virginia, Charleston, WV.
Yang, Y; Hofmann, T; Pies, C;. (2008) Sorptionof polycyclic aromatic hydrocarbons (PAHs) to carbonaceous
materials in a river floodplain soil. Environ Pollut 156(3):1357-1363.
Younger, PL. (2004) Environmental impacts of coal mining and associated wastes: A geochemical perspective. A
Geochemical Perspective. Geol Soc 236:169-209.
Yuan, LL; Norton, SB. (2003) Comparing responses of macroinvertebrate metrics to increasing stress. J N Am
Benth Soc 22:308-322.
Yudovich, YE; Ketris, MP. (2006a) Chlorine in coal: A review. Int J Coal Geol 67:127-144.
Yudovich, YE; Ketris, MP. (2006b) Selenium in coal: A review. Int J Coal Geol 67:112-126.
Zeisset, I; Beebee, TJC. (2008) Amphibian phylogeography: A model for understanding historical aspects of species
distributions. Heredity 101:109-119.
Zotin, AI. (1958) The mechanism of hardening of the salmonid egg membrane after fertilization or spontaneous
activation. J Embryol Exp Morphol 6(4):546-568.
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APPENDIX A
LITERATURE SEARCHES
The peer-reviewed journal articles and reports reviewed in this paper were identified
using a variety of search methods. The report authors identified papers using ISI Web of
Knowledge™ and Google™ Scholar and references that either cited, or were cited by the
Programmatic Environmental Impact Statement or other relevant papers. This search was
supplemented by two more systematic searches described below. Additional sources were
suggested by the Science Advisory Board and public commenters. In total, over 500 sources of
information were identified, including books, conference proceedings, journal articles,
government reports, and theses and dissertations.
All potential sources were evaluated for peer-review status. Suggested conference
proceedings, in general, did not meet this criterion, except for the Proceedings of the American
Society of Mining and Reclamation, which has used a review process since 2002 (Richard
Barnhisel, personal communication), and the 41st Symposium of the British Ecological Society,
Ecology: Achievement and Challenge (Lindsay Haddon, personal communication).
Peer-reviewed sources were classified by region and relevance (see Table A-l). The region of
interest was defined as the central Appalachian coalfields (see Figure 1). Laboratory studies
were included in the category as "stressors in streams from other regions." Most of the sources
judged to be not relevant focused on acid mine drainage, rather than the alkaline discharges that
are typical of mountaintop mines and valley fills.
Table A-l. Categorization of literature retrievals by region and relevance
Description
MTM-VF in region of interest
MTM-VF in other region
Stressors in streams of interest
Stressors in streams from other regions
Review article of stressor of interest
Number of
citations
18
0
24
83
51
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A.l. KEYWORD SEARCH OF ISI WEB OF KNOWLEDGE™ AND GOOGLE™
SCHOLAR
Publications were identified using ISI Web of KnowledgeSM and Google™ Scholar based
on keywords (see Table A-2). The search covered publication dates up to December 2010. ISI
Web of KnowledgeSM searches journal articles dating from 1970. Google™ Scholar does not
specify a date range but generally sorts the search returns so that more recent references are
listed first.
Google™ Scholar generally returned more results than ISI Web of KnowledgeSM.
Google™ Scholar searches the Web across multiple disciplines for journal articles, Web
documents, government reports, other papers, theses/dissertations, books, and abstracts.
Searches are performed in such a manner that the most relevant documents appear on the first
page. Relevancy is determined by "weighing the full text of each article, the author, the
publication in which the article appears and how often the piece has been cited in other scholarly
literature." When searching Google™ Scholar, at minimum, the first five pages were checked
for relevant papers. Search terms were then refined if necessary. ISI Web of KnowledgeSM
returned journal articles that were very specific to the keywords that were entered, which often
resulted in fewer or no returns.
A.2. ECOTOXICOLOGICAL SEARCHES
Searches for ecotoxicological studies on the major ions, and iron, aluminum, and
manganese were supplemented by keyword and Chemical Abstracts Service (CAS) number
searches using BIOSIS, CAS, TOXLINE, Cambridge Scientific Abstracts, and U.S.
Environmental Protection Agency's (EPA) COTOX reference files.
Of the ecotoxicological searches, the one conducted for sulfate compounds calcium
sulfate (CaSO4), MgSO4, potassium persulphate (KSO/O, sodium sulfate (NaSO4), and ferrous
sulfate (FexSO4) was completed in time for inclusion in this appendix. Citations were reviewed
for applicability based on criteria such as the subject of the paper, species group studied and
analytical methods. Of the citations identified, 193 were considered to be applicable and
relevant to organism groups of interest (see Table A-3). Most of the citations judged to be
nonapplicable studied fate and transport rather than effects. The relevant citations were further
reviewed for relevance to the ion mixture typically observed in discharges from MTM-VF.
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Table A-2. Keywords used for ISI Web of Knowledge and Google™
Scholar searches
Keywords
Algae
Alkaline
Amphibian
Anuran
Appalachian streams
aquatic biota
aquatic insects
aquatic toxicity
Arsenic
bank stability
Bivalve
Caddisflies
Caddisfly
Calcium
coal mine
coal mine
overburden
coal mine spoil
Conductivity
Diatom
Discharge
DO
DO
electrical
conductivity
Ephemeroptera
fertilizer
fish
frog
herpetofauna
hollow fill
hydrologic
alteration
leachate
macroinvertebrate
macroinvertebrates
macroinvertebrates
macrophyte
magnesium
manganese
mayflies
mayfly
metals
Mg
mine reclamation
Minnow
Mollusk
Mollusca
Mollusk
mountain top mining
mountaintop mining
mountaintop removal mining
Mussels
Nickel
Nutrients
Overburden
Periphyton
PH
Plecoptera
Potassium
Riparian
Salamander
Salinity
sediment transport
Sediments
Selenium
Snail
Sodium
sodium chloride
specific conductance
Stoneflies
Stonefly
stream temperature
Streams
Sulfate
IDS
Temperature
Thermal
thermal regime
Toad
total dissolved solids
Trichoptera
valley fill
DO = dissolved oxygen; Mg = magnesium.
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Table A-3. Breakdown of sulfate ecotoxicological search results by organism group
Organism Group
Fish
Herpetofauna
Insects
Invertebrates
Plants
Number of citations
62
3
5
73
50
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APPENDIX B
REGULATORY ISSUES RELATED TO MTM-VF OPERATIONS
Mountaintop mines and valley fills (MTM-VF) operations are permitted by state and
federal surface mining and environmental protection authorities. While regulations for
individual mines exist under the Surface Mining Control and Reclamation Act, which is
implemented by the Office of Surface Mining Reclamation and Enforcement (OSMRE) and
delegated states with OSMRE oversight, there are several sections of the Clean Water Act
(CWA) that apply. These are implemented by the U.S. Environmental Protection Agency (EPA)
and the U.S. Army Corps of Engineers (USAGE) and individual states authorized to implement
portions of the CWA. A complete listing and interpretation of regulations that affect MTM-VF
operations is beyond the scope of this report. The following is a general discussion of how the
CWA, and particularly water quality standards, are implemented in the context of MTM-VF.15
B.I. IN GENERAL
Section 301 of the CWA prohibits the discharge of any pollutant by any person except in
compliance with, inter alia, a permit: 33 U.S.C. § 131 l(a) "Except in compliance with this
section and sections... 1342 and 1344 of this title, the discharge of any pollutant by any person
shall be unlawful." For purposes of MTM-VF, there are two relevant CWA permits. The
USACE issues a permit pursuant to Section 404 of the CWA (33 U.S.C. § 1344) for the
discharge of dredged and/or fill material. This permit includes construction of the valley fill
itself and the fill necessary to create an impounded sediment pond downstream of the toe of the
valley fill. The second permit is issued by either the EPA or an authorized state pursuant to
Section 402 of the CWA (33 U.S.C. § 1342). The Section 402 program is also known as the
"National Pollutant Discharge Elimination System" or "NPDES" program. The NPDES permit
includes the discharge from the sediment pond and any stormwater associated with the mining
activity.
NPDES permits must include technology-based effluent limitations. For purposes of
MTM-VF, the applicable technology-based effluent limitations are set forth at 40 C.F.R.
Part 434. In addition to industry sector-specific technology-based effluent limitations,
15While beyond the scope of this paper, it is worth noting that the Surface Mining Control and Reclamation Act
(SMCRA) and its implementing regulations also state that water quality must be maintained and that water quality
standards should not be violated. See, e.g., 30 U.S.C. § 1258(a)(9); 30 U.S.C. § 1265(b)(8)(C); 30 U.S.C. §
1265(b)(10); 30 C.F.R. § 810.2(g); 30 C.F.R. § 816.42; and 30 C.F.R. § 816.57(a)(2). SMCRA also specifically
states that it does not supersede the Clean Water Act and other laws related to preserving water quality. See
30 U.S.C. § 1292(a)(3).
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Section 301(b)(l)(C) of the CWA requires permits to include limits necessary to achieve water
quality standards 33 U.S.C. § 1311(b)(l)(C).
B.2. WATER QUALITY STANDARDS
Water quality standards are the foundation of the water quality-based control program
mandated by the CWA. Water quality standards define the goals for a waterbody by designating
its uses, setting narrative and numeric criteria to protect those uses and establishing provisions to
protect water quality from pollutants. See 40 C.F.R. § 130.3. A water quality standard consists
of four basic elements:
(1) Designated uses of the waterbody (e.g., recreation, water supply, aquatic life,
agriculture)
(2) Water quality criteria to protect designated uses (numeric pollutant concentrations and
narrative requirements)
(3) An antidegradation policy to maintain and protect existing uses and high quality waters
and
(4) General policies addressing implementation issues (e.g., low flows, variances, mixing
zones).
B.2.1. Designated Uses
The water quality standards regulation requires that states and authorized Indian Tribes
specify appropriate water uses to be achieved and protected. Appropriate uses are identified by
taking into consideration the use and value of the water body for public water supply, for
protection offish, shellfish, and wildlife, and for recreational, agricultural, industrial, and
navigational purposes. In designating uses for a water body, states and Tribes examine the
suitability of a water body for the uses based on the physical, chemical, and biological
characteristics of the waterbody, its geographical setting, and scenic qualities and economic
considerations. Each water body does not necessarily require a unique set of uses. Instead, the
characteristics necessary to support a use can be identified so that water bodies having those
characteristics can be grouped together as supporting particular uses.
West Virginia has designated all waters of the state with an aquatic life use (ALU):
§47-2-6. Water Use Categories.
6.1. These rules establish general Water Use Categories and Water Quality Standards for the
waters of the State. Unless otherwise designated by these rules, at a minimum, all waters of the
State are designated for the Propagation and Maintenance of Fish and Other Aquatic Life
(Category B) and for Water Contact Recreation (Category C) consistent with Federal Act goals.
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Incidental utilization for whatever purpose may or may not constitute a justification for
assignment of a water use category to a particular stream segment.
In addition, West Virginia Department of Environmental Protection (WVDEP) water
quality standards specify that waste assimilation and transport are not recognized as a designated
use:
§47-2-6. Water Use Categories.
6. La. Waste assimilation and transport are not recognized as designated uses. The classification
of the waters must take into consideration the use and value of water for public water supplies,
protection and propagation offish, shellfish and wildlife, recreation in and on the water,
agricultural, industrial and other purposes including navigation.
B.2.2. Water Quality Criteria
States establish criteria necessary to protect the designated use. Water quality criteria
may take the form of either specific numeric criteria, such as concentrations of a particular
pollutant, or narrative description of water quality conditions.
B.2.2.1. Numeric Criteria
Section 304(a)(l) of the Clean Water Act requires us to develop numeric criteria for
water quality that accurately reflect the latest scientific knowledge. These criteria are based
solely on data and scientific judgments on pollutant concentrations and ecological or human
health effects. Section 304(a) also provides guidance to states and tribes in adopting water
quality standards. Numeric criteria are developed for the protection of aquatic life as well as for
human health.
Numeric aquatic life criteria are generally pollutant-specific and reflect numeric limits on
the amount of a pollutant that can be present in a water body without harm to indigenous aquatic
life. Aquatic life criteria are designed to provide protection for aquatic organisms from the
effects of acute (short-term) and chronic (long-term) exposure to potentially harmful chemicals.
Human health criteria set allowable concentrations based on human exposure to water
pollutants when humans drink untreated surface water or eat fish, shellfish, or wildlife that have
been contaminated by pollutants in surface waters. To reduce the risk to humans from these
sources, EPA scientists research information to determine the levels at which specific chemicals
are not likely to adversely affect human health.
In making water quality management decisions, a state or tribe should independently
apply each criterion that has been adopted into its water quality standards. If a water body has
multiple designated uses with different criteria for the same pollutant, states/tribes should use the
criterion protective of the most sensitive use.
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B.2.2.2. Narrative Criteria
While numeric criteria help protect a water body from the effects of specific chemicals,
narrative criteria protect a water body from the effects of pollutants that are not easily measured,
or for pollutants that do not yet have numeric criteria, such as chemical mixtures, suspended and
bedded sediments, and floatable debris.
West Virginia's narrative water quality criteria are set forth in a portion of the West
Virginia regulations known as "Conditions Not Allowed":
WV §47-2-3. Conditions Not Allowable in State Waters.
3.2.i. Any other condition, including radiological exposure, which adversely alters the integrity of
the waters of the State including wetlands; no significant adverse impact to the chemical, physical,
hydrologic, or biological components of aquatic ecosystems shall be allowed.
Other examples presented here include excerpts from Kentucky surface water standards
(Chapter 10) and the narrative standards in 401 KAR 10:026-031, which state in part
001 Definitions 401 KAR Chapter 10
(5) "Adversely affect" or "adversely change" means to alter or change the community structure or
function, to reduce the number or proportion of sensitive species, or to increase the number or
proportion of pollution tolerant aquatic species so that aquatic life use support or aquatic habitat is
impaired.
(38) "Impact" means a change in the chemical, physical, or biological quality or condition of
surface water.
(39) "Impairment" means, a detrimental impact to surface water that prevents attainment of a
designated use.
401 KAR 10:031, Section 2: Minimum Criteria Applicable to All Surface Waters.
(1) The following minimum water quality criteria shall be applicable to all surface waters
including mixing zones, with the exception that toxicity to aquatic life in mixing zones shall be
subject to the provisions of 401 KAR 10:029, Section 4. Surface waters shall not be aesthetically
or otherwise degraded by substances that
(a) Settle to form objectionable deposits;
(b) Float as debris, scum, oil, or other matter to form a nuisance;
(c) Produce objectionable color, odor, taste, or turbidity;
(d) Injure, are chronically or acutely toxic to or produce adverse physiological or behavioral
responses in humans, animals, fish and other aquatic life;
(e) Produce undesirable aquatic life or result in the dominance of nuisance species;
(f) Cause fish flesh tainting.
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A narrative chemical criterion for total dissolved solids and specific conductance reads
401 KAR 10:031, Section 4: Aquatic Life.
(f) Total dissolved solids or specific conductance. Total dissolved solids or specific conductance
shall not be changed to the extent that the indigenous aquatic community is adversely affected.
B.2.2.2. Establishing Impairment
Section 303(d) of the CWA requires states to periodically identify those waters that are
not expected to achieve water quality standards even after application of technology-based
effluent limitations to NPDES-permitted point sources (33 U.S.C. § 1313(d)). This identification
is commonly referred to as the state's "Section 303(d) list." By regulation, states must submit
their Section 303(d) lists to EPA for approval every even-numbered year (40 C.F.R. § 130.7(d)).
In establishing its Section 303(d) list, states must consider all existing and readily available
information, including predictive models (40 C.F.R. § 130.7(b)(5)).
In July 1991, EPA transmitted final national policy on the integration of biological,
chemical and toxicological data in water quality assessments. According to this policy, referred
to as "Independent Application," indication of impairment of water quality standards by any one
of the three types of monitoring data (biological, chemical, or toxicological) should be taken as
evidence of impairment regardless of the findings of the other types of data. This policy
continues to the present. See, e.g., Guidance for 2006 Assessment, Listing and Reporting
Requirements Pursuant to Sections 303(d), 305(b) and 314 of the Clean Water Act.
EPA supports use of biological assessments as a direct measure of whether the water
body is achieving the designated use and relevant narrative criteria. A water body in its natural
condition is free from the harmful effects of pollution, habitat loss, and other negative stressors.
It is characterized by a particular biological diversity and abundance of organisms. This
biological integrity—or natural structure and function of aquatic life—can be dramatically
different in various types of water bodies in different parts of the country. EPA recognizes that
biological assessments are a direct measure of the aquatic life use. Because of the natural
variability in ecosystems and aquatic life around the country, EPA could not develop national
biocriteria. Instead, EPA developed methodologies that states can use to assess the biological
integrity of their waters and, in so doing, set protective water quality standards. These
methodologies describe scientific methods for determining a particular aquatic community's
health and for maintaining optimal conditions in various bodies of water. States use these
standard methods to develop their own bioassessment methods and tools. Bioassessment results
are used to support many programs under the CWA (see Figure B-l).
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Wet Weather
Discharge (CSOs,
Storm water)
Listing of Impaired
Waters
(CWA §303d)
Nonpoint Source
Assessment
(CWA §319)
Point Source
Discharge
Permitting
(CWA §402)
Water Quality
Standards and
Criteria
(CWA §303c)
Aquatic Life Use
Assessments
(CWA §305b)
Comprehensive
Watershed
Assessments
Bioassessment
Data
Hazardous Waste
Site Assessments
(CWA§104e)
Marine Point
Source
Discharge
(CWA §403c)
^/
Sewage
Treatment
Plant
Discharges in
Waters
(CWA§301h)
/
p
V
Marine Protection,
Research and
Sanctuaries Act —
Ocean Dumping
(MPRSA)
^^^ Evaluation and Permitting
^^A of Habitat Modifications
\ (CWA §404)
\^
•4
Comprehensive
Risk
Assessment
Figure B-l. Simple representation of CWA programs that rely on biological
assessment data for program implementation. Coal mining activities sections
highlighted in red.
The states have increasingly relied upon biological monitoring in lieu of ambient water
chemistry monitoring because biological monitoring allows the states to maximize monitoring
resources and to assess a larger percentage of their waters. Since 2004, West Virginia has
utilized standard field collection, laboratory, and data analysis methods in its biological
assessment program. This has resulted in West Virginia's use of a family-level benthic metric
developed jointly by EPA and West Virginia Department of Environmental Protection, called the
West Virginia Stream Condition Index (WVSCI), to identify impairment of the aquatic life use.
See http://www.wvdep.org/Docs/536_WV-Index.pdf West Virginia also developed an
assessment methodology for using the WVSCI to interpret its narrative criterion and to make
aquatic life use-attainment decisions. For an example, see WVDEP's 2008 Integrated Water
Quality Monitoring and Assessment Report available at
http://www.wvdep.org/Docs/16495_WV_2008_IR_Supplements_Complete_Version_EPA_Appr
oved.pdf
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In Kentucky, the Kentucky Division of Water assessment methodologies for ALU
attainment are similar, where the state uses biological monitoring data and statistical-based
multimetric index analyses to assess waterbody attainment. For macroinvertebrates, the KY
Macroinvertebrate Bioassessment Index is used to evaluate ALU:
http://www.water.ky.gov/NR/rdonlyres/7F189804-4322-4C3E-B267-
5A58E48AAD3F/0/Statewide_MBI.pdf
In nonheadwater streams, KY uses fish communities as other indicators of ALU with the
KY Index of Biotic integrity, a similarly constructed multimetric index:
http ://www. water.ky.gov/NR/rdonlyres/04C65101 -AF1C-4751-809B-
4F5D09B7269A/0/KffiI_paper.pdf.
Section 303(d) also requires the states to establish total maximum daily loads (TMDLs)
for their impaired waters. Essentially, a TMDL is a measure of the assimilative capacity of a
waterbody considering seasonal variability and critical conditions, allocated among point sources
and nonpoint sources and incorporating a margin of safety. See 33 U.S.C. § 1313(d);
40 C.F.R. § 130.2(i); and!30.7(c).
B.3. IMPLEMENTATION OF WATER QUALITY STANDARDS THROUGH NPDES
PERMITS
As set forth above, Section 301 of the CWA requires NPDES permits to contain both
technology-based effluent limitations and water quality-based effluent limitations. For the
industry sector, that includes surface coal mining with valley fills, the applicable
technology-based effluent limitations are set forth at 40 C.F.R. Part 434. These effluent
limitations include limitations on discharges from coal preparation plants, acid and alkaline mine
drainage, postmining areas, remining and western alkaline mining. For example, effluent
limitations on discharges from a new source of alkaline mine drainage include limits on iron,
total suspended solids and pH. See 40 C.F.R. § 434.45.
The NPDES regulations implement the water quality-based effluent limitations
requirement as set forth in CWA Section 301(b)(l)(C) through the following regulatory
requirements:
No permit may be issued ... (d) When the imposition of conditions cannot ensure compliance with
the applicable water quality requirements of all affected states... (40 C.F.R. § 122.4(d)).
[EJach NPDES permit shall include conditions meeting the following requirements when
applicable .... [A]ny requirements in addition to or more stringent than promulgated effluent
limitations guidelines.. .necessary to: achieve water quality standards under Section 303 of the
CWA, including state narrative criteria for water quality... (40 C.F.R. § 122.44(d)(l)).
No permit may be issued ... (i) To a new source or a new discharger, if the discharge from its
construction or operation will cause or contribute to the violation of water quality standards (40
C.F.R. § 122.4(1)).
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Most states, including West Virginia and Kentucky, have been authorized to issue
NPDES permits for discharges to waters within their borders. EPA retains the ability to review,
object to, and if necessary, take over issuance of a particular NPDES permit. See 33 U.S.C. §
1342(d); 40 C.F.R. § 123.44. The scope of EPA's NPDES permit review in a particular state is
generally spelled out in a Memorandum of Agreement with that state (40 C.F.R. § 123.44). EPA
also retains the ability to enforce discharges without or in violation of an NPDES permit
(33 U.S.C. § 1319).
B.4. IMPLEMENTATION OF WATER QUALITY STANDARDS THROUGH SECTION
404 PERMITS
Section 404(b)(l) directs the EPA in conjunction with the Secretary of the Army to
establish guidelines to be applied by the USAGE when considering an application for a permit to
discharge dredged and/or fill material pursuant to Section 404 of the CWA. This instruction has
resulted in the Section 404(b)(l) Guidelines (40 C.F.R. Part 230), which provide the substantive
environmental criteria that must be applied by the USAGE when considering a Section 404
permit application. Among other things, the USAGE may issue a permit only if it determines
that the project represents the least damaging practicable alternative:
[N]o discharge of dredged or fill material shall be permitted if there is a practicable alternative to
the proposed discharge which would have less adverse impact on the aquatic ecosystem, so long
as the alternative does not have other significant adverse environmental consequences (40 C.F.R. §
230.10(a)).
The USAGE also must ensure that the project proponent has taken "all appropriate and
practicable steps to avoid and minimize adverse impacts to waters of the United States"
(33 C.F.R. §332.1(c)); see also 40 C.F.R. § 230.10(a)(l)(i); 40 C.F.R. §230.10(d); and 40
C.F.R. §§230.70-.77.
In addition, the Section 404(b)(l) Guidelines prohibit the issuance of a permit "if it:
(1) Causes or contributes, after consideration of disposal site dilution and dispersion, to
violations of any applicable state water quality standard," (40 C.F.R. § 230.10(b)(l)), or if it
"will cause or contribute to significant degradation of the waters of the United States, ...
[including] (1) Significantly adverse effects of the discharge of pollutants on human health or
welfare, including but not limited to effects on municipal water supplies, plankton, fish, shellfish,
wildlife and special aquatic sites. (2) Significantly adverse effects of the discharge of pollutants
on life stages of aquatic life and other wildlife dependent on aquatic ecosystems, including the
transfer, concentration and spread of pollutants or their by-products outside of the disposal site
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through biological, physical and chemical processes; (3) Significantly adverse effects of the
discharge of pollutants on aquatic ecosystem diversity, productivity and stability...." (40 C.F.R. §
230.10(c)). The USAGE also must consider the effect of the discharge on fish, crustaceans,
molluscs and other aquatic organisms in the food web (40 C.F.R. § 230.31), the effect on benthos
(40 C.F.R. § 230.61(b)(3)), and the suitability of water bodies for populations of aquatic
organisms (40 C.F.R. § 230.22).
Before issuing a federal permit or license, federal agencies, including the USAGE, must
obtain a certification from the state in which the discharge will originate that the discharge will
comply with applicable provisions of 33 U.S.C. § 1311, 1312, 1313, 1316, and 1317. Among
other things, therefore, the USAGE must obtain a certification that the discharge will comply
with applicable water quality standards, which are established pursuant to 33 U.S.C. § 1313. In
considering the potential of a discharge to cause or contribute to an excursion from water quality
standards, the USAGE generally will consider conclusive the state's CWA Section 401 water
quality certification, unless EPA advises of other water quality aspects to be taken into
consideration (33 C.F.R. § 320.4(d)).
While the USAGE is the permit-issuing authority for Section 404, EPA retains significant
authorities, including the authority to prohibit, deny, or restrict the use of any defined area for
specification as a disposal site pursuant to Section 404(c) (33 U.S.C. § 1344(c)), the ability to
request consideration of particular permits by the USAGE at the Headquarters level pursuant to
the Memorandum of Agreement described in Section 404(q) (33 U.S.C. § 1344(q)), the ability to
identify waters that are within the scope of the CWA and to determine the applicability of
exemptions pursuant to a Memorandum of Agreement with the USAGE under Section 404(f)
(33 U.S.C. § 1344(f)), and the ability to enforce discharges without a permit (33 U.S.C. § 1319).
Figure B-2 depicts the sequence of actions necessary to address impaired streams under
the CWA.
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Determine
Protection Level
(EPA Criteria/State WQS)
Measure Progress
Conduct WQ
Assessment
(Identify Impaired Waters)
Monitor and Enforce
Compliance
(Self Monitoring, Agency Monitoring)
\
^L
\
Set Priorities
(Rank/Target Waterbodies)
Establish Source
Controls
(Point Source, NPS)
Evaluate Appropriateness
of WQS for Specific Waters
(Reaffirm WQS)
Define and Allocate
Control Responsibilities
(TMDL/WLA/LA)
Figure B-2. Water quality-based approach to pollution control.
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