United States
Environmental Protection
Agency
Anaerobic Compost
Constructed Wetlands System
(CWS) Technology
Innovative Technology
Evaluation Report
SUPERFUND INNOVATIVE
TECHNOLOGY EVALUATION
XXX
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EPA/540/R-02/506
December 2002
Anaerobic Compost
Constructed Wetlands System
(CWS) Technology
Innovative Technology Evaluation Report
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
Recycled/Recyclable
Printed with vegetable-based Ink on
paper that contains a minimum of
50% post-consumer fiber content
processed chlorine free.
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Notice
The information in this document has been funded by the U. S. Environmental Protection Agency (EPA) under Contract No.
68-C5-0037 to Terra Tech EM Inc. It has been subjected to the Agency's peer and administrative reviews and has been
approved for publication as an EPA document. Mention of trade names or commercial products does not constitute an
endorsement or recommendation for use.
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Foreword
The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation's land, air, and water
resources. Under a mandate of national environmental laws, the Agency strives to formulate and implement actions leading
to a compatible balance between human activities and the ability of natural systems to support and nurture life. To meet this
mandate, EPA's research program is providing data and technical support for solving environmental problems today and
building a science knowledge base necessary to manage our ecological resources wisely, understand how pollutants affect our
health, and prevent or reduce environmental risks in the future.
The National Risk Management Research Laboratory is the Agency's center for investigation of technological and
management approaches for reducing risks from threats to human health and the environment. The focus of the Laboratory's
research program is on methods for the prevention and control of pollution to air, land, water and subsurface resources;
protection of water quality in public water systems; remediation of contaminated sites and ground water; and prevention and
control of indoor air pollution. The goal of this research effort is to catalyze development and implementation of innovative,
cost-effective environmental technologies; develop scientific and engineering information needed by EPA to support
regulatory and policy decisions; and provide technical support and information transfer to ensure effective implementation of
environmental regulations and strategies.
This publication has been produced as part of the Laboratory's strategic long-term research plan. It is published and made
available by EPA's Office of Research and Development to assist the user community and to link researchers with their clients.
Hugh W. McKinnon, Director
National Risk Management Research Laboratory
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Abstract
As part of the Superfund Innovative Technology Evaluation (SITE) Program, the U.S. Environmental Protection Agency (EPA)
evaluated constructed wetlands systems (CWS) for removing high concentrations of zinc from mine drainage at the Burleigh
Tunnel in Silver Plume, Colorado.
Exploration geologists have known for many years that metals, most commonly copper, iron, manganese, uranium, and
zinc, frequently accumulate in swamps and bogs located in mineralized areas. This understanding forms the basis for the
design of CWS—essentially excavated pits filled with organic matter—that have been developed and constructed over the
past 15 years to treat drainage from abandoned coal mines in the eastern United States. Mine drainage is routed through
the organic material, where metals are removed through a combination of physical, chemical, and biological processes.
In fall 1994, anaerobic compost wetlands in both upflow and downflow configurations were constructed adjacent to and
received drainage from the Burleigh Tunnel, which forms part of the Clear Creek/Central City Superfund site. The
systems were operated over a 3-year period. The effectiveness of treatment by the CWS was evaluated by comparing the
concentration of zinc and other metals from corresponding influent and effluent analyses. By far the dominant toxic metal
present in the drainage was zinc. The upflow CWS removed an average of 93 percent of the zinc during the first year of
operation, and 49 and 43 percent during the second and third years. The downflow CWS removed an average of 77
percent of zinc during the first year and 70 percent during the second year. (Flow was discontinued to the downflow
system in the third year.)
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Contents
List of Figures and Tables viii
Acronyms, Abbreviations, and Symbols ix
Conversion Factors xi
Acknowledgments xii
Executive Summary 1
1 Introduction 5
1.1 Brief Description of the SITE Program and Reports 5
1.2 Purpose of the Innovative Technology Evaluation Report 6
1.3 Technology Description 6
1.3.1 Treatment Technology 8
1.3.2 System Components and Function 8
1.3.3 Key Features of the CWS Technology 9
1.4 Key Contacts 11
2 Technology Application Analysis 12
2.1 Applicable Wastes 12
2.2 Factors Affecting Performance 12
2.2.1 Mine Drainage Characteristics 12
2.2.2 Operating Parameters 13
2.2.3 Compost Performance 13
2.3 Site Characteristics 13
2.3.1 Support Systems 13
2.3.2 Site Area, Preparation, and Access 15
2.3.3 Climate 15
2.3.4 Utilities 15
2.3.5 Services and Supplies 15
2.4 Availability, Adaptability, and Transportability of Equipment 15
2.5 Material Handling Requirements 16
2.6 Personnel Requirements 16
2,7 Potential Community Exposures 16
2.8 Evaluation of Technology Against RI/FS Criteria 16
2.9 Potential Regulatory Requirements 18
v
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Contents (continued)
2.9.1 Comprehensive Environmental Response, Compensation, and Liability Act 18
2.9.2 Resource Conservation and Recovery Act 18
2.9.3 Clean Water Act 19
2.9.4 Occupational Safety and Health Act 19
2.10Limitations of the Technology 19
3 Treatment Effectiveness 22
3.1 Background 22
3.2 Review of SITE Demonstration 22
3.2.1 Treatability Study 22
3.2,2 Technology Demonstration 23
3.2.3 Operational and Sampling Problems and Variations from the Work Plan 23
3.2.4 Site Demobilization 24
3.3 Demonstration Methodology 24
3.3.1 Testing Approach 25
3.3.2 Sampling, Analysis, and Measurement Procedures 25
3.4 Site Demonstration Results 27
3.4.1 Burleigh Mine Drainage Chemistry 27
3.4.2 DownflowCWS 27
3.4.3 UpflowCWS 36
3.4.4 Clear Creek 40
3.4.5 Toxicity Testing Results 40
3.4.6 Microbial Toxicity Testing 42
3.5 Attainment of Demonstration Objectives 43
3.6 Design Effectiveness 44
3.6.1 DownflowCell 44
3.6.2 UpllowCell 45
4 Data Quality Review 46
4.1 Zinc Data Quality Review 46
4.1.1 Quality Assurance Results for Field Sampling Activities 46
4.1.2 Quality Assurance Results for Sample Analysis 47
4.2 Acute Toxicity Data Quality Review 48
4.2.1 Analytical Quality Assurance 48
4.3 Noncritical Parameters Data Quality Review 50
5 Economic Analysis 52
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Contents (continued)
5.1 Basis of Economic Analysis 52
5.2 Cost Categories 53
5.2.1 Site Preparation Costs 53
5.2.2 Permitting and Regulatory Requirements 53
5.2.3 Capital Equipment 53
5.2.4 Startup 55
5.2.5 Labor 55
5.2.6 Consumables and Supplies 55
5.2.7 Utilities 55
5.2.8 Residual Waste Shipping and Handling 55
5.2.9 Analytical Services 55
5.2.10 Maintenance and Modifications 55
5.2.11 Demobilization 56
Technology Status 57
References 58
Appendix
A Analytical Results Summary Tables
B Case Study
VII
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Figures
1 Site Location 3
2 Schematic Cross-Section of an Anaerobic CWS Upflow Cell 5
3 Flow Rates Measured for Effluent Cells 17
4 CWS Zinc Concentration by Month 21
5 CWS Cadmium Concentration by Month 23
6 CWS Lead Removal by Month 24
7 CWS Manganese Removal by Month 25
8 Sulfate-Reducing Bacteria, Downflow CWS Substrate 27
9 Monthly Zinc Loading, Downflow CWS 29
10 Sulfate-Reducing Bacteria, Upflow CWS Substrate 32
11 Monthly Zinc Loading, Upflow CWS 33
Tables
1 Evaluation of CWS Treatment Versus Rl/FS Criteria 11
2 Treatment Standards and Influent concentrations for the CWS SITE Demonstration 13
3 Summary of Standard Methods and Procedures 20
4 Average Downflow CWS Substrate Results 26
5 Average Upflow CWS Substrate Results 31
6 Clear Creek Upstream 35
7 Clear Creek Downstream 35
8 CWS Demonstration Toxicity (LC,U) Results 36
9 CWS Costs for Different Treatment Flow Rates 47
viii
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Acronyms, Abbreviations, and Symbols
°c
°F
%C
%R
AA
ARAR
ASTM
AVS
BOD
CDPHE
CDM
CPU
CERCLA
CFR
cws
DQO
Eh
EPA
FS
gpm
H2S
HOPE
HSWA
ICP
ITER
LC50
MCAWW
MCL
MS
Degrees Celsius
Degrees Fahrenheit
Percent completeness
Percent recovery
Atomic absorption
Applicable or relevant and appropriate requirement
American Society for Testing and Materials
Acid volatile sulfide
Biochemical oxygen demand
Colorado Department of Public Health and Environment
Camp, Dresser, & McKee, Inc.
Colony forming units
Comprehensive Environmental Response, Compensation, and Liability Act
Code of Federal Regulations
Constructed wetlands system
Data quality objective
Oxidation reduction potential
U.S. Environmental Protection Agency
Feasibility study
Gallons per minute
Hydrogen sulfide
High-density polyethylene
Hazardous and Solid Waste Amendments of 1984
Inductively coupled plasma
Innovative technology evaluation report
Lethal concentration for 50 percent of the test organisms
Methods for Chemical Analysis of Water and Wastes
Maximum contaminant level
Micrograms
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Acronyms, Abbreviations, and Symbols (continued)
u.s
mg/kg
mg/L
MS
NCP
NIST
NPDES
NRMRL
O&M
ORD
ORP
OSHA
OSWER
PPE
ppm
PRC
PVC
QAPP
QA/QC
RCRA
RI
RPD
SARA
SITE
SDWA
SOP
SRM
SWDA
TCLP
TOO
IDS
TSS
yd3
Microsiemens
Milligrams per kilogram
Milligrams per liter
Matrix spike
National Oil and Hazardous Substances Pollution Contingency Plan
National Institute of Standards and Technology
National Pollutant Discharge Elimination System
National Risk Management Research Laboratory
Operation and maintenance
Office of Research and Development
Oxidation/reduction potential
Occupational Safety and Health Administration
Office of Solid Waste and Emergency Response
Personal protective equipment
Parts per million
PRC Environmental Management, Inc.
Polyvinyl chloride
Quality assurance project plan
Quality assurance/quality control
Resource Conservation and Recovery Act
Remedial investigation
Relative percent difference
Superfund Amendments and Reauthorization Act
Superfimd Innovative Technology Evaluation
Safe Drinking Water Act
Standard operating procedure
Standard reference material
Solid Waste Disposal Act
Toxicity characteristic leaching procedure
Total organic carbon
Total dissolved solids
Total suspended solids
Cubic yards
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Conversion Factors
To Convert From
To
Multiply By
Length
Area:
Volume:
inch
foot
mile
square foot
acre
gallon
cubic foot
centimeter
meter
kilometer
square meter
square meter
liter
cubic meter
2.54
0.305
1.61
0.0929
4,047
3.78
0.0283
Mass:
pound
kilogram
0.454
Energy:
kilowatt-hour
megajoule
3.60
Power:
kilowatt
horsepower
1.34
Temperature: (°Fahrenheit - 32) "Celsius
0.556
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Acknowledgments
This report was prepared under the direction of Mr. Edward Bates, the U.S. Environmental Protection Agency (EPA) Superfund
Innovative Technology Evaluation (SITE) project manager at the National Risk Management Research Laboratory (NRMRL)
in Cincinnati, Ohio; Ms. Dana Allen, U.S. EPA Region VIII; and Mr. James Lewis, Colorado Department of Public Health
and Environment. This report was prepared by Mr. Gary Miller, Mr. Garry Farmer, Mr. Jon Bridges, and Ms. Shaleigh
Whitesell of Tetra Tech EM Inc. (Tetra Tech) and Mr. Mark Kadnuck of the Colorado Department of Public Health and
Environment (formerly of Tetra Tech). This report was typed by Ms. Robin Richey and Ms. June Diller, edited by Mr. Butch
Fries, and reviewed by Dr. Kenneth Partymiller of Tetra Tech.
This project consisted of a demonstration conducted under the SITE program to evaluate the anaerobic compost Con-
structed Wetland System (CWS) technology developed by the Colorado Department of Public Health and Environment
(CDPHE). The technology demonstration was conducted on mineral mine drainage at the Burleigh Tunnel in Silver
Plume, Colorado, which is included in the Clear Creek/Central City Superfund site, Passive treatment was selected as the
preferred treatment alternative for the Burleigh Tunnel drainage in a 1991 Record of Decision (ROD). This Innovative
Technology Evaluation Report (ITER) interprets the data that was collected during the nearly four-year demonstration and
discusses the potential applicability of the technology.
The cooperation and participation of the following people are gratefully acknowledged: Mr. Vincent Gallardo, Ms. Ann
Vega, and Dr. James Lazorchek of NRMRL; Ms. Holly Fliniau of EPA Region VIII and Mr. Rick Brown of CDPHE.
xli
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Executive Summary
This executive summary of the Constructed Wetlands
System (CWS) technology demonstration discusses
technology applications, describes system effectiveness,
and presents an evaluation of the costs associated with the
system and lessons learned during the field demonstration.
Introduction
The anaerobic compost CWS technology was evaluated
under the Superfund Innovative Technology Evaluation
(SITE) program. The SITE program was developed by
the U.S. Environmental Protection Agency (EPA) in
response to the mandate of the Superfund Amendments
and Reauthorization Act (SARA) of 1986. The primary
purpose of the program is to maximize the use of alternative
treatmenttechnologies. Tothisend,reliableperformance
and cost data on innovative technologies are developed
during demonstrations where the technology is used to
treat a specific waste.
After the demonstration, EPA publishes an Innovative
Technology Evaluation Report (ITER) designed to aid
decision makers in evaluating the technology for further
consideration as an appropriate cleanup option. This
ITER includes a review of the technology application, an
economic analysis of treatment costs, and the results of
the demonstration.
For this demonstration, wetlands were designed and
constructed to treat mine drainage through a combination
ofphysical, chemical, and biological processes. The mine
drainage, containing primarily zinc contamination, flowed
into the constructed wetlands where metals were removed
by sorption, precipitation, and biological sulfate reduction.
The demonstration included the evaluation of two CWS
treatment cells (pits) filled with an organic-rich compost
(96 percent) and alfalfa hay (4 percent) mixture. Both
treatment cells were constructed adjacent to, and received
drainage from, the Burleigh Tunnel in Silver Plume,
Colorado. The Burleigh Tunnel is part of the Clear Creek/
Central City Superfund site. Passive wetlands treatment
was identified by the Colorado Department of Public
Health and Environment (CDPHE) as the preferred
remedial alternative for the Burleigh Tunnel drainage.
Each treatment cell covered 0.05 acres and differed in
flow configuration. One cell was constructed in an upflow
configuration, in which water entered from the base of the
cell and was forced upward to discharge; the other was
constructed in a downflow configuration, in which water
entered from the top of the cell and flowed by gravity to
discharge. The compost and hay mixture was 4 feet deep
in both cells. Flow rates of water into and out of the cells
were controlled by a series of v-notch weirs; each cell
was designed to treat 7 gallons per minute (gpm).
Technology Applications Analysis
The primary objectives of the CWS technology
demonstration were to (1) measure the reduction of zinc
in Burleigh Tunnel drainage resulting from the CWS
treatment with respect to cell configuration and seasonal
variation (temperature); (2) assess the toxicity of the
Burleigh Tunnel drainage; (3) characterize the toxicity
reduction resulting from treatment of the drainage by the
CWS; and (4) estimate toxicity reductions in the stream
(Clear Creek) receiving the Burleigh Tunnel drainage.
CWS treatment effectiveness was evaluatedby comparing
the concentration of zinc and other metals from
corresponding CWS influent and effluent analyses
(see Section 3.0). The results indicate the concentration
of zinc in the Burleigh Tunnel drainage ranged from 50 to
60 milligrams per liter (mg/L) during the first year of the
demonstration. However, in May and June 1995, agreat
deal of spring snow and rain and a rapid thaw combined
to increase the amount of runoff entering the mine
network drained by the Burleigh Tunnel. At that time,
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flow from the tunnel increased from 45 gpm to more than
300 gpm, and zinc concentrations increased from 55 mg/
L (April 12,1995) to 109 mg/L (August 8,1995). Over
the final 2 years of the demonstration, zinc concentrations
in Burleigh Tunnel mine drainage were lower in the
winter, dropped again in April or May when flow through
the mine workings increased, and rapidly increased in
summer, remaining high throughout the fall. During this
period, Burleigh Tunnel mine drainage zinc concentrations
generally remained between 45 and 84 mg/L, with
increases to more than 100 mg/L noted during the late
summer and fall. The Burleigh Tunnel drainage is also
characterized by moderate pH and alkalinity and low
concentrations of metals other than zinc.
Downflow
In the first year of operation, CDPHE reported the
downflow cell developed flow problems on occasion,
preventing treatment of the intended amount of water.
Remedies, such as fluffing the compost, were tried and
were somewhat successful allowing the system to flow at
4 to 6 gpm during the first two years of operation. During
the third year, the flow in this cell dropped to less than
1 gpm and flow to this cell was discontinued.
The permeability loss is believed to be related to
precipitation of metal oxides, hydroxides, and carbonates,
settling of fine materials in the cell, and compaction of the
substrate material.
In general, the downflow cell was effective in removing
zinc during the first year of operation. Zinc removal by this
cell ranged from 69 to 96 percent with a mean removal of
77 percent. During the second year of operation, zinc
removal ranged from 62 to 79 percent with a mean of 70
percent. During the final 6 months of operation, flow
through the downflow cell continued to decline increasing
the residence time of the mine drainage in the cell. The
increased residence time improved zinc removal. Zinc
removal during this period ranged from 67 to 93 percent
with a mean of 82 percent.
Aqueous geochemical modeling, observations of cell
compost, sulfate-reducing bacteria count results, and acid
volatile sulfide data suggest that biological sulfate reduction
is not the primary zinc removal mechanism within this cell.
Instead, the primary metal removal mechanism is thought
to be the precipitation of zinc oxides, hydroxides, and
carbonates in aerobic sections of the downflow cell.
Upflow
During the first 6 months of operation, upflow cell effluent
samples contained low (less than 1 mg/L) concentrations
of zinc. However, during the later part of 1994 and into
1995, upflow cell effluent zinc concentrations began to
increase. The concentrations of zinc ranged from 0.13 mg/
L in early 1994 to 60.1 mg/L in May 1997.
In the spring of 1995, heavy spring runoff overwhelmed
the CWS system, channeling 20 gpm of aerobic water
(nearly three times the design flow) through the upflow
cell. This high runoff also apparently mobilized more zinc
from the mine workings or mine waters and substantially
increased the concentration of zinc in the mine drainage.
The large flows created aerobic conditions and the
increased zinc loading had a detrimental effect on the
upflow cell. These new conditions apparently initiated a
change in the cell's microbial ecology. After the high flow
event, the upflow cell removed only 50 to 60 percent of the
zinc in the mine drainage. Prior to the high flow event, the
upflow cell removed greater than 90 percent of the zinc
contamination (year 1 mean removal was 93 percent).
The loss of substrate hydraulic conductivity also affected
the upflow CWS. During the demonstration, the height of
the influent wier was periodically raised to increase the
hydraulic pressure to maintain flow through the upflow
CWS. The water level was raised approximately 1 foot
over the 4-year demonstration. In 1997, this cell developed
a visibly obvious preferential pathway in the southeast
corner, adjacenttothebermedsidewall. This preferential
pathway was eliminated by terminating flow to this section
of the wetland through excavating of the wetland substrate
to allow installation of a cap on the influent line.
The high initial zinc removal rates in the upflow cell were
likely the result of absorption of metals and biological
sulfate reduction. The decline in metal removal by the
upflow cell after the high flow event is likely related to the
decline in sulfate reducing bacteria in this cell. There are
several possible reasons for the decline of the sulfate-
reducing bacteria including toxicity produced by high zinc
concentrations for the bacteria, prolonged exposure to
aerobic conditions allowing other wetland bacteria to
outcompete the sulfate-reducing bacteria, or the utilization
of all the most readily metabolized growth materials by the
sulfate reducing bacteria leading to lower activity and
eventually lowerpopulations of these bacteria. Ultimately,
the primary metal removal mechanism over the last
several years of the demonstration was likely chemical
precipitation.
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Economic Analysis
An economic analysis was conducted to examine 11 cost
categories for the CWS technology. The 11 categories
include (1) site preparation; (2) permitting and regulatory
requirements; (3) capital equipment and construction;
(4) startup; (5) labor; (6) consumables and supplies;
(7) utilities; (8) residual and waste shipping and handling;
(9) analytical services; (10) maintenance and
modifications; and(ll) demobilization.
A number of factors affect the estimated costs of treating
mine drainage with the CWS technology. These factors
generally include flow rate, type and concentration of
contaminants, water chemistry, physical site conditions,
site location, and treatment goals. In addition, the
characteristics of the spent compost produced by a CWS
will affect disposal costs since the compost may require
treatment for off-site disposal.
Based on the criteria evaluated in the cost analysis, the
average estimated cost for a constructed wetland at
50 gallons per minute (gpm) over a 15-year period is
$ 1,744,100 million or $0.0045 per gallon of water treated.
Treatment Effectiveness
Based on this demonstration, the following conclusions
may be drawn about the effectiveness of the anaerobic
compost CWS technology.
• The upflow CWS removed an average (arithmetic
mean) of 53 mg/L (93 percent) of zinc during the
first year of operation.
• Upflow cell zinc removal averaged 41 mg/L
(49 percent) dining the second year and 30 mg/L
(43 percent) during the third year of operation.
• During the first year of operation, the upflow cell
effluent was not toxic to Ceriodaphnia dubia or
Pimephales promelas.
• The downflow CWS removed an average of 44.2 mg/
L (77.4 percent) of zinc during the first year and 58
mg/L (70 percent) during the second year of
operation.
• The CWS is relatively easy to construct with readily
available materials.
In summary, results from this SITE demonstration and
additional tests of the CWS technology suggest that the
CWS is capable of reducing the toxicity of contaminated
mine drainage by removing metals such as zinc, cadmium,
iron, lead, nickel, and silver.
However, the results of this demonstration also clearly
show that an anaerobic compost CWS using sulfate
reduction may have difficulty in recovering from upset
conditions such as the high flow event that occurred
during this demonstration.
In addition, application of this technology to mine drainage
containing high concentrations of iron may require
pretreatment to remove the iron. If not removed, the iron
could precipitate in the wetland and could lead to loss of
wetland permeability.
Lessons Learned
The following items highlight lessons learned during the
CDPHE constructed wetlands demonstration. The list is
partitioned among five categories of considerations (or
concerns): theory, design, construction, operation and
maintenance, and analytical.
Theory
• An upflow CWS using biological sulfate reduction is
capable of reducing the concentration of several
metals including zinc, cadmium, nickel, lead, iron, and
silver. The extent of metal reduction depends on the
concentration of the metal and sulfate in the mine
drainage, and the performance of the CWS.
• The primary metal removal process in the downflow
CWS did not appear to be biological sulfate reduction.
Zinc in the demonstration CWS downflow cell
appeared to be primarily removed by chemical
precipitation. Generally, zinc removal by the
demonstration downflow cell ranged between 70 and
80 percent. However, the accumulation of zinc
carbonate in the cell compost may have attributed to
a loss of cell permeability during the demonstration.
Design
• A hydraulic residence time of 50 hours (estimated)
provided good metal removal in the upflow cell
during the first 8 months of the demonstration.
However, the decline in metal removal after this
initial period and inability to re-establish the sulfate-
reducing bacteria in the upflow cell suggest this
residence time may be a lower limit for mine drainages
containing high metal concentrations.
• Hydraulically, the upflow cell performed well with
4 feet of compost. However, some short circuiting
was observed after 3 years of operation.
• The mixture of fresh compost (96 percent) and
hay (4 percent) used as a substrate during the
demonstration was a superior environment for sulfate-
reducing bacteria. However, the compost contains
high levels of ammonia that is readily leached during
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wetland startup, resulting in elevated levels of
ammonia in the discharge. The addition of wood
products to the substrate can reduce the amount of
ammonia generated. Land treatment has been used
at some sites to dispose of wetland startup discharge.
• Each wetland cell should have an easily adjustable
influent conveyance with the capability of bypassing
200 to 300 percent of typical peak flows.
Construction
• Bermed sidewalls lined with high-density polyethylene
(HOPE) is a suitable construction technique for cold
region applications. However, the use of a geonet
on the wetland surface to allow animals and people
to walk on the wetland is not recommended. The
geonet did not allow additional compost or hay to be
added to the wetland. In addition, the use of geofabric
to separate the piping networks from the compost is
not recommended.
• Effluent collection pipes (polyvinyl chloride [PVC])
should be larger than 1 inch in diameter to prevent
clogging from precipitated material. In addition, the
effluent collection structure should include cleanouts
that allow precipitated material to be periodically
removed without driving the precipitate back into the
wetland compost.
• Lining a downflow cell with HOPE above the level
of the ponded water allows this water to short circuit
the wetland compost. Short circuits are most
noticeable during the winter when the compost
becomes frozen and contracts from the liner, creating
a gap between the compost and liner.
Operation and Maintenance
• Constructed wetlands can require frequent inspections
to ensure that proper flows are maintained within the
treatment cells. However, properly designed and
constructed influent distribution and effluent collection
networks may reduce inspection frequency.
• Treatment system downtime with CWS treatment is
not high. Effluent piping networks should be cleaned
out periodically (once or twice a year was appropriate
for the Burleigh Tunnel CWS). The frequency of
compost removal and replacement will depend on
contaminant loading, metal removal efficiencies, and
the desired performance level of the CWS. Compost
removal and replacement frequency for the
demonstration CWS upflow cell is estimated to be
once every 4 to 5 years.
• Straw bales covered with insulated construction
blankets (used to cure concrete in cold weather) arc
an effective insulator for an upflow CWS during
winter operation. However, their use requires an
upper support structure such as a geonet. An
equally effective insulation system could include
6 inches of fresh compost and hay covered by
construction blankets, although this system has not
been tested.
• Straw bales used for winter insulation must not be
allowed to become saturated by water. Their
combined weight will compress the wetland compost,
making it impermeable,
Analytical
• Routine (monthly) total metals analysis in
conjunction with quarterly dissolved metals analysis
were useful in evaluating the performance of the
CWS. The mine drainage and effluents were
sampled and analyzed every 2 weeks during the first
2 years of the demonstration; however, monthly
sampling (conducted over the final year of the
demonstration) is adequate to track treatment
performance.
• Routine aquatic toxicity testing of the mine drainage
and CWS effluent also provides useful water quality
information. During the CWS demonstration, these
analyses were conducted every 3 to 4 months, but
semi-annual analyses could also be used.
Demonstration aquatic toxicity testing used two test
organisms, fathead minnows (Pimephalus promelas)
and water fleas (Ceriodaphnia dubia); however, other
test organisms including trout fry could also be used.
• Sulfate-reducing bacteria analyses of wetland
compost were conducted monthly during the first 2
years of the CWS demonstration. These analyses,
while useful, did not show much variation until the
high flow event, and their frequency could easily be
reduced to every other month or even a quarterly.
Acid volatile sulfide analysis can indicate the
accumulation of metal sulfides within the CWS
compost; however, the compost sample must be
collected from the area of metal filtration. The acid
volatile sulfide analysis procedure is not routine for
most laboratories, and meaningful results may not be
achievable.
• All aqueous field analyses conducted during the
CWS demonstration including pH, Eh (effluent),
dissolved oxygen (influent), conductivity, and
temperature were useful measurements. It should
be noted that the platinum element of the Eh probe is
prone to poisoning, requiring periodic replacement.
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Section 1
Introduction
This section provides background information about the
SITE program, discusses the purpose of this ITER, and
describes the CWS technology. Key contacts for additional
information about the SITE program, this technology, and
the demonstration site are listed at the end of this section.
1.1 Brief Description of the SITE
Program and Reports
SARA mandates that EPA select, to the maximum extent
practicable, remedial actions at Superfund sites that create
permanent solutions (as opposed to land-based disposal)
for contamination that affects human health and the
environment. In response to this mandate, the SITE
program was established by EP A's Office of Solid Waste
and Emergency Response (OSWER) and Office of
Research and Development (ORD). The SITE program
promotes the development, demonstration, and use of
new or innovative technologies to clean up Superfund
sites across the country.
The SITE program's primary purpose is to maximize the
use of alternatives in cleaning up hazardous waste sites by
encouraging the development and demonstration
of innovative treatment and monitoring technologies. It
consists of the Demonstration Program, the Emerging
Technology Program, the Monitoring and Measurement
Technologies Program, and the Technology Transfer
Program. These programs are discussed in more detail
below.
The objective of the Demonstration Program is to develop
reliable performance and cost data on innovative treatment
technologies so that potential users may assess specific
technologies. Technologies evaluated either are currently
or will soon be available for remediation of Superfund
sites. SITE demonstrations are conducted at hazardous
waste sites under conditions that closely simulate full-
scale remediation, thus assuring the usefulness and
reliability of information collected. Data collected are
used to assess the performance of the technology, the
potential need for pre- and post-treatment processing of
wastes, potential operating problems, and approximate
costs. The demonstrations also allow evaluation of long-
term risks and operating and maintenance (O&M) costs.
The Emerging Technology Program focuses on
successfully proven, bench-scale technologies that are in
an early stage of development involving pilot-scale
or laboratory testing. Successful technologies are
encouraged to advance to the Demonstration Program.
The constructed wetlands is an example of a successful
graduate of the Emerging Technology Program that was
evaluated in the Demonstration Program.
Existing technologies that improve field monitoring and
site characterization are identified in the Monitoring and
Measurement Technologies Program. New technologies
that provide faster, more cost-effective contamination
and site assessment data are supported by this program.
The Monitoring and Measurement Technologies Program
also formulates the protocols and standard operating
procedures for demonstrating methods and equipment.
The Technology Transfer Program disseminates technical
information on innovative technologies in the
Demonstration, Emerging Technology, and Monitoring
and Measurement Technologies Programs through various
activities. These activities increase the awareness and
promote the use of innovative technologies for assessment
andremediationofSuperfundsites. The goal of technology
transfer is to promote communication among remedial
managers requiring up-to-date technical information.
Technologies are selected for the SITE Demonstration
Program through annual requests for proposals. ORD
staff review the proposals, including any unsolicited
proposals that may be submitted throughout the year, to
determine which technologies show the most promise for
use at Superfund sites. Technologies chosen must be at
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the pilot- or full-scale stage, must be innovative, and must
have some advantage over existing technologies. Mobile
technologies are of particular interest. Once EPA has
accepted a proposal, cooperative agreements between
EPA and tiie technology developer establish responsibilities
for conducting the demonstrations and evaluating the
technology. The developer is responsible for demonstrating
the technology at the selected site and is expected to pay
any costs for transportation, operation, and removal of
equipment, EPA is responsible for project planning, site
preparation, sampling and analysis, quality assurance and
quality control (QA/QC), and for preparing reports,
disseminating information, and transporting and disposing
of untreated and treated waste material. For the CWS
evaluation, CDPHE (the lead agency of the Burleigh
Tunnel site) identified passive wetlands treatment as the
preferred treatment alternative with agreement by EPA
and the division of responsibilities was essentially as
described.
The results of the CWS technology demonstration are
published in two documents: the SITE technology capsule
and the present ITER. The SITE technology capsule
provides relevant information on the technology,
emphasizing key features of the results of the SITE field
demonstration. The ITER is discussed in the following
section. Both the SITE technology capsule and the ITER
are intended for use by remedial managers making a
detailed evaluation of the technology for a specific site and
waste.
1.2 Purpose of the Innovative
Technology Evaluation Report
The ITER provides information on the CWS technology
and includes a comprehensive description of the
demonstration and its results. The ITER is intended for
use by EPA remedial project managers, EPA on-scene
coordinators, contractors, and other decision makers for
implementing specific remedial actions. The ITER is
designed to aid decision makers in evaluating specific
technologies for further consideration as an option in a
particular cleanup operation. This report represents a
critical step in the development and commercialization of
a treatment technology. To encourage the general use of
demonstration technologies, EPA provides information
regarding the applicability of each technology to specific
sites and wastes. Therefore, the ITER includes information
on cost and site-specific characteristics. It also discusses
advantages, disadvantages, and limitations of the
technology. Each SITE demonstration evaluates the
performance of a technology in treating a specific waste.
The waste characteristics at other sites may differ from
the characteristics of the treated waste. Therefore,
successful field demonstration of a technology at one site
does not necessarily ensure that it will be applicable at
other sites. Data from the field demonstration may
require extrapolation for estimating the operating ranges
in which the technology will perform satisfactorily. Only
limited conclusions can be drawn from a single field
demonstration.
1.3 Technology Description
The Colorado Department of Public Health and
Environment submitted a proposal to the SITE program
for demonstrating the anaerobic compost CWS technology.
This technology was selected for a SITE demonstration at
the Burleigh Tunnel in Silver Plume, Colorado. The
demonstration was carried out under a cooperative
agreement involving the EPA National Risk Management
Research Laboratory (NRMRL), CDPHE, and EPA
Region 8.
The Burleigh Tunnel is located approximately 50 miles
west of Denver in the Silver Plume - Georgetown mining
district (Figure 1), within the Clear Creek/Central City
Superfund site. The Silver Plume - Georgetown mining
district occupies an area of about 25 square miles
surrounding the towns of Silver Plume and Georgetown.
The tunnel entrance is at an elevation of 9,152 feet, about
400 feet north of Clear Creek, on the western side of the
town of Silver Plume. The area immediately surrounding
the tunnel entrance is littered with mill tailings and waste
rock dumps. Dilapidated buildings and equipment
from previous milling operations are also present.
No mining operations are active in the immediate area.
The water draining from the Burleigh Tunnel is of near-
neutral pH (ranging from 6.9 to 7.9) and has high zinc
concentrations (ranging from 44.8 to 109 mg/L). The
drainage also contains moderate alkalinity and low levels
of metals other than zinc.
A treatability study was conducted at the Burleigh Tunnel
between June 18, 1993 and August 12, 1993. The
treatability study involved the construction, operation, and
sampling of two upflow compost and hay bioreactors that
treated mine drainage from the Burleigh Tunnel. The
treatability study (PRC 1993) showed that low levels of
sulfate in the mine drainage would not limit biological
sulfate reduction, thereby permitting the removal of zinc
and other metals by the bioreactors or the demonstration
scale treatment cells. Construction of the CWS
demonstration cells began in August 1993 and was
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Grand
County
Burleigh Tunnel
Constructed
Wetlands Site
5 Ml 0 5 Ml 10
SCALE: 1" = 10 MILES
Figure 1. Site location.
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completed in November 1993, The demonstration began
in January 1994 and continued for a 46-month period
through November 1997. Evaluation of the CWS
technology is based on results of the treatability study and
the SITE demonstration at the Burleigh Tunnel site.
1.3.1 Treatment Technology
There are generally three types of constructed wetlands:
free-water surface systems, subsurface flow systems,
and aquatic plant systems (EPA 1988). A free-water
system typically consists of shallow basins or channels
with slow- flowing water and plant life. A subsurface
flow wetland consists of basins or channels filled with
permeable substrate material; the water flows through,
rather than over, this substrate. An aquatic plant system
is essentially a free water surface system with deeper
channels containing floating or suspended plants. In
general, free-water surface and aquatic plant systems are
aerobic wetlands that remove metals primarily by aerobic
oxidation of iron followedby precipitation of iron hydroxides,
that leads to the precipitation or adsorption of other
metals. Aerobic wetlands are most successful in removing
iron, arsenic, selenium and, to some extent, manganese
from moderately low to neutral pH mine waters (Gusek
and others 1994).
Anaerobic compost wetlands are designed to treat mine
drainage through a combination of physical, chemical, and
biological processes. Mine drainage is directed into
constructed wetlands that contain an organic-rich compost
substrate. Initially, sorption to the CWS substrate is the
primary metal removal mechanism active within the
system. Sorption includes adsorption of metals to organic
and inorganic wetlands materials and absorption of metals
into wetlands microorganisms and plants.
• Adsorption refers to the binding of positively charged
ions to mineral surfaces by metal cations in solution.
The sorption of inorganic ions is largely determined
by complex chemical equilibria involving the charge
and size of the element or complex ion, the nature of
the sorbing material, and the pH of the aqueous
solution. The properties of the surface that influence
inorganic sorption include net surface charge and the
presence, configuration, and pH dependence of
binding sites. The structure of the solid may also
affect adsorption reactions.
• Absorption refers to the incorporation of ions
or compounds into the cell structure of
microorganisms or plants. Metals may also be
incorporated into the structure of complex humic
substances formed during the degradation of the
substrate.
After several months, the sorption capacity of the wetlands
is exhausted and metal removal efficiencies by this
mechanism decline.
Once the sorption capacity of the CWS substrate is
expended, the formation, precipitation, and filtration of
metal sulfides become the primary metal removal
mechanism in the CWS. The process is believed to be
biologically mediated by sulfate-reducing bacteria present
in anaerobic zones within the CWS.
The bacteria oxidize organic matter provided by the
wetland with the simultaneous reduction of sulfate to
hydrogen sulfide. The hydrogen sulfide reacts with
dissolved metals to produce metal sulfides. The metal
sulfides, with low aqueous solubilities, precipitate and
become trapped in the wetlands substrate by filtration.
The following reactions illustrate the overall oxidation/
sulfate reduction reactions and subsequent formation of
metal sulfides.
SO;2 + 2CH20 —> HS- + 2HCO3' + IP
M+2 + H2S or HS- —> MS(s) + 2H+
where: M is a metal such as zinc (Zn+2), iron (Fe+2), nickel
(Ni+2), and (s) indicates a solid.
In addition, other reactions within the wetlands may
contribute to observed metal removal, including mineral
precipitation and chelation (binding) to suspended organic
material. In general, mine drainage contains low levels of
dissolved oxygen that, when exposed to air, will take up
oxygen and become aerobic. This process can lead to
geochemical disequilibrium where the metal is no longer
soluble at this concentration and may initiate metal
precipitation. Zinc carbonate (Smithsonite) is an example
of a mineral that may precipitate in the demonstration
downflow CWS. In addition, the decay of wetland
compost and biomass will produce dissolved and suspended
organic material in the wetland pore water. These
materials can chelate metals in solution. Although chelated
metals may not be effectively removed (filtered) by the
wetland, they may not be available biochemically to
aquatic plants and organisms exposed to the effluent.
1.3.2 System Components and Function
Two CWS treatment cells were located adjacent to the
Burleigh Tunnel between a compressor building and an
old mill. Each cell covered 0.05 acre; the two cells
differed in flow configuration. The cell nearest the mine
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adit was an upflow system, in which water entered the cell
under pressure from the bottom and flowed upward
through the substrate material to discharge. The second
cell was a downflow system, in which the water entered
the cell from the top and flowed by gravity to the bottom
for discharge. The demonstration C WS cells were highly
engineered systems compared to many of the previously
tested constructed wetlands, including the Big 5 wetlands
evaluated in the Emerging Technology Program (EPA/
540/R-93/523). Figure 2 shows a cross-section schematic
of the upflow CWS treatment cell. The downflow cell
was identical except the direction of mine drainage flow
in the compost is reversed.
Both CWS treatment cells were installed below grade to
reduce freezing of the cells during winter. Both had
bermed earthen side walls. The base of each cell was
made up of a gravel subgrade, a 16-ounce geofabric, a
sand layer, a clay liner, and a high density polyethylene
liner. The base was separated from the influent or
effluent piping by a geonet. A 7-ounce geofabric separated
theperforatedPVCpipingfromthecompost. The compost
was held inplace with a combination of 7-ounce geofabric
and geogrid in the upflow cell. The perforated effluent
piping was also supported by the geogrid in the upflow cell.
Up to 6 inches of dry substrate material was located above
the perforated piping. The geonet and the perforated
piping ensured even distribution of the influent water into
the treatment cells and prevented short circuiting of water
through the cells. The influent and effluent distribution
piping were also staggered horizontally as an additional
precaution against short circuiting.
Existing construction near the Burleigh Tunnel entrance
required that the upflow cell be 10 percent smaller by
volume than the downflow cell. The dimensions of the
cells are as follows;
• Upflow cell - 69 feet long, 25.5 feet wide, and 4 feet
deep, with an estimated total substrate volume of
198 cubic yards
• Downflow cell - 62 feet long, 33 feet wide, and 4
feet deep, with an estimated total substrate volume
at 218 cubic yards
Note: The dimensions listed are at the top of the cell
wall The volumes listed take into account the sloped
walls of the cells.
The organic-rich compost substrate was composed of a
mixture of 95 to 96 percent manure compost and 4 to
5 percent hay. The compost was produced from cattle
manure and unidentified paper products. The compost
and hay mixture had been identified as the most effective
medium in removing zinc from the drainage during the
previous bench-scale test (Camp, Dresser and McKee
1993). Wood based substrates have also been used in
constructed wetland systems.
The flow to the CWS cells was regulated by a series of
concrete v-notch weirs, one for the influent and one for
the effluent of each cell. The effluent weir controlled the
flow and the hydraulic residence time of the mine drainage
through both CWS cells. Each cell was designed for a
flow of 7 gpm with a total flow capacity for the two cells
of 14 gpm. The remaining flow from the Burleigh Tunnel
drainage was diverted to Clear Creek (untreated) via the
influent weir. A drainage collection structure was
constructed within the Burleigh Tunnel to build sufficient
hydraulic head to drive the flow through the two CWS.
1.3.3 Key Features of the CWS
Technology
Certain features of the CWS technology allow it to be
adapted to a variety of settings:
• The hardware components (geosynthetic materials,
PVC piping, and flow control units) of the CWS are
readily available,
• Compost materials can be composed of readily
available materials. However, the actual composition
of a substrate material for a site-specific constructed
wetland is best determined through pilot studies.
Composted manure was used during this study.
• Operation and maintenance costs are low since the
systems are generally self-contained, requiring only
periodic changes of the compost depending on site-
specific conditions.
Other features that shouldbe thoroughly evaluated before
constructing a CWS include the following:
• Properties of the drainage to be treated. Some
drainages may need some type of pretreatment
before entering the CWS. For example, drainage
with high iron or aluminum content might prematurely
clog the CWS if not pretreated to remove some of
the metal.
• Climate conditions must be evaluated to assess the
potential for reduced efficiency of the system during
different seasons of the year.
• Contingencies if the system does not perform as
expected.
-------
7oz. Geofabric
Geogrid
7oz. Geofabric
Perf. Effluent
Piping
Perf. Influent
Piping
7oz. Geofabric
Geonet
HOPE Liner
Geosynthetic
Clay Liner
16oz. Geofabric
Figure 2. Schematic cross-section of an anaerobic CWS upflow cell.
10
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• Proximity to a populated area—odors generally are
associated with CWS treatment.
• Land availability near the source of the contaminated
water to avoid extended transport. The CWS
typically requires more land than a conventional
treatment system. Consequently, locations with
steep slopes and drainages would make construction
more difficult and costly.
• Cost of constructing the system if substrate and
other materials are not readily available.
• Possible use of concrete basins to eliminate
replacement costs for liners.
• Potential for vandalism of the CWS, which could
result in increased costs.
• Seasonal fluctuation of water flow or chemistry and
the potential impact to the CWS.
• Production and release of nutrients from substrate
and stream standard requirements for discharge of
produced nutrients
1.4 Key Contacts
Additional information on the CWS technology, the SITE
program, and the demonstration site can be obtained from
the following sources:
The CWS Technology
James Lewis
Colorado Department of Public Health and Environment
HMWMD-RP-82
4300 Cherry Creek Drive South
Denver, Colorado 80222-1530
Telephone: (303)692-3390
Fax: (303)759-5355
The SITE Program
Edward Bates, Project Manager
U.S. Environmental Protection Agency
National Risk Management Research Laboratory
26 West Martin Luther King Drive
Cincinnati, Ohio 45268
Telephone: (513)569-7774
Fax: (513)569-7676
The Clear Creek/Central City Superfund Site
Michael Holmes, Remedial Project Manager
U.S. Environmental Protection Agency
Region 8
999 18th Street, Suite 300
Denver, Colorado 80202
Telephone: (303)312-6607
11
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Section 2
Technology Applications Analysis
This section of the ITER describes the general applicability
of the CWS technology to contaminated waste sites. The
analysis is based primarily on the SITE CWS treatability
study and demonstration results. A detailed discussion
of the treatability study and demonstration results is
presented in Section 3.0 of this report. An article containing
a constructed wetlands case study is presented in
Appendix B.
2.1 Applicable Wastes
Constructed wetlands have been demonstrated to be
effective in removing organic, metal, and nutrient elements
including nitrogen and phosphorus from municipal
wastewaters, mine drainage, industrial effluents, and
agricultural run-off. The technology is waste-stream
specific, requiring characterization of all organic and
inorganic constituents.
Because constructed wetlands can treat a wide variety of
wastes, they vary considerably in their design. Constructed
wetlands can be simple, single-cell systems, such as the
two cells evaluated during this demonstration, or complex
multicell ormulticomponent systems. Complex constructed
wetlands may include multiple wetland cells in series,
anoxic limestone drains, marshes, ponds, and rock filters.
Constructed wetlands tested in the eastern U.S. to
remediate slightly acidic coal mine drainage have
incorporated an anoxic limestone drain to provide alkalinity,
followed by a holding pond, a constructed wetland, a
shallow marsh, and finally a rock filter. The holding pond
and wetland promote precipitation of iron hydroxides,
while the marsh and rock filter remove manganese and
suspended solids. Constructed wetlands design criteria
are discussed in detail in an article by Gusek and Wildeman
(1995).
The results of the CWS demonstration (see Section 3.0)
suggest the primary metals removal mechanisms are not
identical within the upflow and downflow wetland cells.
In the upflow cell, biological sulfate reduction appeared
to be the primary zinc removal mechanism. Metals
shown to be removed by this process include cadmium,
copper, iron, lead, nickel, silver, and zinc (PRC 1995). In
addition, biological sulfate reduction may also remove
cobalt, mercury, and molybdenum contamination. In the
downflow cell, chemical precipitation appeared to be the
primary zinc removal mechanism. Because of the
numerous geochemical species and complex equilibria
involved in wetlands treatment of mine drainage, it is often
difficult to predict which metals will precipitate.
An equilibrium aqueous geochemical wetlands model
(MINTEQ.AK) has been developed to help predict metal
removal by constructed wetlands (Klusman 1993),
2.2 Factors Affecting Performance
Because CWS designs are so diverse, the number of
parameters affecting their operation is also large. In the
discussion that follows, the performance factors described
pertain to this demonstration CWS (anaerobic compost)
or to similar systems treating metal-contaminated mine
drainage. These performance factors may or may not be
relevant to constructed wetlands designed to treat organic
or inorganic (nonmetal) contamination. Several factors
influenced the performance of the two demonstration
CWS. These factors can be grouped into three categories;
(1) mine drainage characteristics, (2) operating parameters,
and (3) compost degradation.
2.2.1 Mine Drainage Characteristics
The CWS technology is capable of treating a range of
contaminated waters containing heavymetals. However,
the effectiveness of a CWS can be reduced as
contaminants in high concentrations precipitate and clog
the system prematurely. Often, contaminated coal mine
drainages in the eastern U.S. contain elevated
concentrations of iron or aluminum. When the pH of these
drainages is raised during wetland treatment, iron and
12
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aluminum hydroxides can form and precipitate (Hedin and
others 1994).
These precipitates can lead to a loss of permeability or a
gradual filling of the wetland. Because sulfate-reducing
bacteria cannot survive in low pH environments, low pH
mine drainage can also affect the ability of the biological
sulfate reduction wetland to remove contaminants. The
oxidation/reduction potential (ORP) of the mine drainage
may also affect the performance of the constructed
wetland. However, the extent of the ORP effect is
unknown.
2.2.2 Operating Parameters
The operating parameters that can be adjusted during the
treatment process include the flow rate and hydraulic
residence time of water within the wetland. In general,
the selection and design for the hydraulic residence time
is a function of the rate of metal loading. A hydraulic
residence time of 50 to 100 hours was found to work well
in the biological sulfate reduction reactors used during the
short-term CWS treatability study (Figure 3).
The residence time in the upflow and downflow cells
during the demonstration was calculated at between 50
and 60 hours. The calculation was based on the substrate
volume of the wetlands, the percent moisture of the
substrate (generally, 50 to 65 percent with 50 percentused
in the calculation), and a flow rate of 7 gpm.
Maintaining proper hydraulic residence times is one of the
most important factors in successful wetlands treatment.
In biological-based systems, a short residence time may
not allow metals to precipitate and be filtered out by the
wetland or may expose the bacteria to inhibitory levels of
metal contaminants. Both may result in lower metal
removal rates. In chemical precipitation systems,
compounds that precipitate slowly may notbe removed to
the same extent as rapidly precipitating compounds.
Chemical amendments, such as alkalinity or nutrients, are
also examples of parameters that can be adjusted during
the wetland treatment process. Alkalinity may be added
via an anoxic limestone drain or directly to the mine
drainage as lime. Nutrients could also be added directly
to the mine drainage or applied to the ponded surface
water of downflow cells. Neither alkalinity nor nutrients
was added to the SITE demonstration CWS.
2.2.3 Compost Performance
Compost performance depends on the compost materials
used and the characteristics of the mine drainage. When
using manure compost, the metals concentrations of the
drainage, the nutrient concentrations in the compost, and
gradual breakdown and compaction of the compost
materials are the most important factors controlling compost
effectiveness. Of these factors, substrate breakdown
and compaction that leads to a loss ofhydraulic conductivity
is probably the most important factor. The breakdown of
the complex biological polymers to smaller compounds by
fermentative bacteria gradually destroys the structural
intensity of the compost and leads to compaction. One
way to extend substrate lifetime is to include materials that
are degraded at a moderate rate. Based on the loss of
nutrients and hydraulic conductivity in the upflow CWS,
the wetland compost material is expected to last 4 to
5 years before becoming ineffective.
The accumulation of metals within the constructed wetlands
may eventually cause the compost material to become a
hazardous waste, substantially decreasing the number of
compost disposal options and increasing treatment costs.
However, after 4 years of near-continuous operation of
the demonstration CWS, neither cell's compost material
developed hazardous characteristics based on thresholds
defined in 40 Code of Federal Regulations (CFR)
Part 261.24. However, the primary contaminant in the
Burleigh Mine Drainage, zinc is not a TCLP analysis
parameter.
2.3 Site Characteristics
Site characteristics are important when considering CWS
technology because they can affect system application.
All characteristics should be considered before selecting
the technology to remediate a specific site. Site-specific
factors include support systems, site area and preparation,
site access, climate, hydrology, utilities, and the availability
of services and supplies.
2.3.1 Support Systems
If on-site facilities are not already available, a small
storage building equipped with electricity maybe desirable
near the treatment system. The on-site building could be
used for storing operating and sampling equipment (tools,
field instrumentation, and health- and safety-related gear)
and providing shelter for sampling personnel during
inclement weather. The building may also be used for
calibrating field equipment for system monitoring.
13
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2,3.2 Site Area, Preparation, and Access
Constructed wetlands typically require a larger level
area compared to other treatment options. The results
of this investigation suggest that a 50-60 hour hydraulic
residence time is near the lower limit required of these
systems to provide consistent metal removal. Researchers
in this field have suggested that longer residence times
ranging from 75 to 150 hours may be required for long-
term metal removal (Dr. Ronald Klusman and Dr. Richard
Gammons, personal communications) The depth of the
compost in the demonstration CWS cells was 4 feet. The
maximum depth of compost that can be used while
maintaining treatment effectiveness is unknown.
Consequently, some sites may require extensive grading
and leveling to allow construction of a CWS. Depending
on the site, grading and leveling may be cost prohibitive.
Piping or other mechanisms for conveying mine drainage
to the wetlands is also necessary. In addition, a relatively
constant rate of flow is desired to keep the wetlands
active. Thus, site conditions may require a mine
drainage collection, storage, and distribution structure.
Furthermore, an upflo w constructed wetland may require
that the mine drainage distribution network include a dam
or pump to maintain sufficient hydraulic head to force
the mine drainage through the compost. Also, piping is
required to bypass flow around the wetland. This bypass
piping or conveyance should be oversized to manage
200 to 300 percent of the predicted maximum mine
drainage discharge.
Access roads for heavy equipment (excavation and
hauling) are required to install, operate, and maintain a
CWS.
2.3.3 Climate
The climate at potential constructed wetland sites can be
a limiting factor. Extended periods of severe cold,
extreme hot and arid conditions, and frequent severe
storms or flooding will affect system performance.
Extreme cold can freeze portions of the wetland resulting
in channeling of the mine drainage through the substrate,
thus, reducing the hydraulic residence time. In addition,
cold temperatures may reduce microbial activity or
populations. Reductions in hydraulic residence time and
microbial activity will both lessen the ability of the
constructed wetland to remove metals and may require it
to be oversized. The large water surface areas and plant
life associated with wetlands enhance evaporation and
evapotransportation. A constructed wetland in a hot and
arid climate may periodically dry up at a site with low
water flow rates. If the wetland design does not consider
cyclical periods of wet and dry, it may be less effective
during the wet periods. Constructing wetlands in areas
with frequent flooding or severe storms can lead to
hydraulic overloading or washout of substrate materials.
The engineering controls required to overcome these
climatic or geographic limitations may eliminate the low
cost and low maintenance advantages that make
constructed wetlands appealing.
2.3.4 Utilities
The CWS is a passive treatment technology, so utilities
are not required to operate the system. However, in some
situations electricity for pumps or on-site analytical
instrumentation may be desirable. In remote areas, an on-
site storage building should be provided if possible. A
telephone connection or cellular phone is required for
operating and sampling personnel to contact emergency
services if needed and for routine communications.
2.3.5 Services and Supplies
The main services required by the CWS are periodic
adjustment of system flow rates, cleanout of effluent
piping, and the removal and replacement of compost
materials. During the CWS demonstration, flow rate
adjustments were required every 3 to 6 months, and effluent
piping cleanout was conducted once. However, both
CWS demonstration cells were operated from a single v-
notch weir and the flow diverted to the cells. The
frequency of flow adjustment would be lower if each cell
had been constructed with its own weir. The time
between changeout of wetland compost depends on the
chemical constituents of the influent water, the
configuration and capacity of the constructed wetland,
and the preferred method of disposal. The compost
lifetime, estimated from nutrient loss and the development
of short circuiting during this demonstration is estimated to
be 4 to 5 years.
2.4 Availability, Adaptability, and
Transportability of Equipment
The components of a simple CWS are generally available
locally. The components include standard construction
materials for the structure of the wetland cells, liner
materials available from several sources, and compost
materials, the type of which will depend on the contaminants
in the mine drainage. The most suitable compost for a
given application can be identified during a treatability
study using materials available locally.
15
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2.5 Material Handling Requirements
The CWS generates spent compost material. Substrate
material will require testing to evaluate disposal options.
Depending on the disposal option, dewatering or other
pretreatment may be necessary prior to shipment for off-
site disposal. Depending on regulatory requirements, the
effluent water generated during dewatering may also
require additional treatment prior to discharge.
Some CWS compost materials may contain high levels of
water-soluble nitrogen or phosphorus compounds. These
compounds can be readily leached from the fresh compost
during startup of the constructed wetland. Thus, the CWS
effluent at startup may require treatment to reduce or
remove excess nitrogen or phosphorous. Treatment may
include land application, if permitted, or effluent collection
for subsequent recycling through the CWS.
2.6 Personnel Requirements
Wetlands construction and compost replacement require
heavy equipment operators, laborers, and a construction
supervisor. After the CWS is installed, personnel
requirements include a sampling team and personnel to
adjust system flow rates. Sampling personnel should be
able to collect water and substrate samples for laboratory
analysis and measure field parameters using standard
instrumentation.
All personnel should have completed an Occupational
Safety and Health Administration (OSHA)initial 40-hour
health and safety training course with annual 8-hour
refresher courses, if applicable, before constructing,
sampling, replacing compost, or removing a constructed
wetland at hazardous waste sites. They should also
participate in a medical monitoring program as specified
under OSHA requirements.
2.7 Potential Community Exposures
Fencing and signs should be installed around a CWS to
restrict access to the system for both humans and wildlife.
The potential routes of exposure include the mine drainage
or waste stream, the compost material, and the CWS
effluent. The actual exposure risk depends on the
constituents of the specific waste being treated and the
effectiveness of the treatment.
The CWS may also generate low concentrations of
hydrogen sulfide gas, depending on the time of year and
the biological activity of the CWS. Odors caused by
hydrogen sulfide and volatile fatty acids from the decaying
manure may be a nuisance to a local community.
2.8 Evaluation of Technology Against
RI/FS Criteria
EPA has developed nine evaluation criteria to fulfill
the requirements of the Comprehensive Environmental
Response, Compensation, and Liability Act (CERCLA),
as well as additional technical and policy considerations
that have proven important for selecting potential remedial
alternatives. These criteria serve as the basis for
conducting bench-scale testing during the remedial
investigation (RI) at a hazardous waste site, for conducting
the detailed analysis during the feasibility study (FS), and
for subsequently selecting an appropriate remedial action.
Each SITE technology is evaluated against the nine EPA
criteria because these technologies may be considered as
potential remedial alternatives. The nine evaluation criteria
are;
• Overall protection of human health and the
environment
• Compliance with applicable or relevant and
appropriate requirements (ARAR)
• Long-term effectiveness and permanence
• Reduction of toxicity, mobility, or volume
• Short-term effectiveness
• Implementability
• Cost
• State acceptance
• Community acceptance
Table 1 presents the results of this evaluation for the
CWS. The demonstration results indicate the upflow
CWS canprovideshort-termprotectionofthe environment;
reduces contaminant mobility, toxicity, and volume; is cost
effective; implementable, and is an acceptable remedy to
the community and state regulators. However, neither
CWS cell tested in this demonstration, provided long-term
effectiveness. This in part is the result of low zinc
discharge requirements (200 ug/L) at the demonstration
site. Other sites may have less strict discharge
requirements. In addition, the upset condition resulting
from the high flow event also contributed to the lack of
long-effectiveness particularly in regards to the upflow
cell.
16
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Table 1, Evaluation of CWS Treatment Versus RI/FS Criteria
Criterion
Discussion
1. Overall Protection of Human Health
and the Environment
2. Compliance with Applicable or
Relevant and Appropriate
Requirements (ARAR)
3. Long-Term Effectiveness and
Permanence
4. Short-term Effectiveness
5. Reduction of Toxicity, Mobility, or
Volume of contaminates through
Treatment
6. Implementability
7. Cost
8. Community Acceptance
9. State Acceptance
As tested, the CWS provided only short-term
effectiveness. Indifferentcircumstances.theCWSmay
provide short- and bng-term protection by removing
mine drainage contaminants.
Substrate is a recycled product, not mined or
manufactured.
Wetland effluentdischarge may require compliance with
Clean Water Act regulations.
Substrate disposal may require compliance with RCRA
regulations.
CWS treatment removes contamination from mine
drainage, but may not meet low-level discharge
requirements.
Use of CWS treatment with other technologies may be
effective in meeting low-level discharge requirements.
Presents few short-term risks to workers, community, or
wildlife.
Minimal personal protective equipment required for
operators.
CWS treatment reduces contaminant mobility, toxicity,
and volume.
Generally a passive treatment system, but can be
active.
Construction uses standard material and practices
common in the industry.
Construction cost of full-scale (SOgpm) system is
estimated at approximately $290,000.
O&M of full-scale CWS system is estimated to be
$57,000 per year.
The public usually views the technology as a natural
approach to treatment; therefore, the public generally
accepts this technology.
CDPHE found the technology shows promise for
treating AMD; however, based on constraints at the
Burleigh site, including the cold climate and proximity to
town, CDPHE recommended not implementing a full-
scale, permanent system at the site.
Colorado Division of Minerals has built several CWSsto
treat AMD.
17
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2.9 Potential Regulatory Requirements
This section discusses specific environmental regulations
pertinent to operation of a CWS, including the transport,
treatment, storage, and disposal of wastes and treatment
residuals, and analyzes these regulations in view of the
demonstration results. State and local regulatory
requirements, which may be more stringent, must also be
addressed by remedial managers.
ARARs include the following: (1) CERCLA; (2) the
Resource Conservation and Recovery Act (RCRA);
(3) the Clean Water Act; and (4) OSHA regulations.
These four general ARARs are discussed below; specific
ARARs mustbe identified by remedial managers for each
site.
2.9.1 Comprehensive Environmental
Response, Compensation, and
Liability Act
CERCLA, as amended by SARA, authorizes the federal
government to respond to releases or potential releases of
any hazardous substance into the environment, as well as
to releases of pollutants or contaminants that may present
an imminent or significant danger to public health and
welfare or the environment.
As part of the requirements of CERCLA, EPA has
prepared the National Oil and Hazardous Substances
Pollution Contingency Plan (NCP) for hazardous substance
response. The NCP, codified at 40 CFR Part 300,
delineates methods and criteria used to determine the
appropriate extent of removal and cleanup for hazardous
waste contamination.
SARA amended CERCLA and directed EPA to:
• Use remedial alternatives that permanently and
significantly reduce the volume, toxicity, or mobility
of hazardous substances, pollutants, or contaminants.
• Select remedial actions that protect human health
and the environment, are cost-effective, and involve
permanent solutions and alternative treatment or
resource recovery technologies to the maximum
extent possible.
• Avoid off-site transport and disposal of untreated
hazardous substances or contaminated materials when
practicable treatment technologies exist (Section
121 [b]).
In general, two types of responses are possible under
CERCLA: removals and remedial actions. The CWS
technology is likely to be part of a CERCLA remedial
action. Remedial actions are governed by CERCLA as
amended by SARA. As stated above, these amendments
promote remedies that permanently reduce the volume,
toxicity, and mobility of hazardous substances, pollutants,
or contaminants.
On-site remedial actions must comply with federal and
state ARARs. ARARs are identified on a site-by-site
basis and may be waived under six conditions: (1) the
action is an interim measure, and the ARAR will be met
at completion; (2) compliance with the ARAR would pose
a greater risk to human health and the environment than
noncompliance; (3) it is technically impracticable to meet
the ARAR; (4) the standard of performance of an ARAR
can be met by an equivalent method; (5) a state ARAR
has not been consistently applied elsewhere; and (6)
ARAR compliance would not provide a balance between
the protection achieved at a particular site and demands
on the Superfund for other sites. These waiver options
apply only to Superfund actions taken on site, and
justification for the waiver must be clearly demonstrated.
2.9.2 Resource Conservation and
Recovery Act
RCRA, an amendment to the Solid Waste Disposal Act
(SWDA), was enacted in 1976 to address the problem of
safe disposal of the enormous volume of municipal and
industrial solid waste generated annually. RCRA
specifically addressed the identification and management
of hazardous wastes. The Hazardous and Solid Waste
Amendments of 1984 (HSWA) greatly expanded the
scope and requirements of RCRA.
The presence of RCRA-defined hazardous waste
determines whether RCRA regulations apply to the
CWS technology. RCRA regulations define andregulate
hazardous waste transport, treatment, storage, and disposal.
Wastes defined as hazardous under RCRA include
characteristic and listed wastes. Criteria for identifying
characteristic hazardous wastes are included in 40 CFR
Part 261 Subpart C. Listed wastes from nonspecific and
specific industrial sources, off-specification products, spill
cleanups, and other industrial sources are itemized in 40
CFR Part 261, Subpart D.
The CWS demonstration treated mine discharge water
from the Burleigh Tunnel, which is included in the Clear
Creek/Central City Superfund site. The manure compost
was tested regularly to determine whether it would become
a hazardous waste during the demonstration. The concern
18
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was that sorption and precipitation of metals could cause
the substrate to become a hazardous waste, thus restricting
options and increasing cost for material disposal. The
substrate did not exhibit the characteristics of hazardous
waste after nearly 4 years of operation.
2.9.3 Clean Water Act
The objective of the Clean Water Act is to restore and
maintain the chemical, physical, and biological integrity of
the .nation's waters. To achieve this objective, effluent
limitations of toxic pollutants from point sources were
established. Wastewater discharges are most commonly
controlled through effluent standards and discharge permits
administered through the National Pollutant Discharge
Elimination System (NPDES) by individual states with
input from the federal EPA. Under this system, discharge
permits are issued with limits on the quantity and quality
of effluents. These limits are based on a case-by-case
evaluation of potential environmental impacts and on
wasteload allocation studies aimedat distributing discharge
allowances fairly. Discharge permits are designed as an
enforcement tool with the ultimate goal of achieving
ambient water quality standards (Metcalf and Eddy 1979).
NPDES permit requirements must be evaluated for each
CWS when the effluent water is discharged into a
waterway or water body. The requirements and standards
that must be met in the effluent for each CWS will be
based on the waterway or water body into which the CWS
discharges. The effluent limits will be established through
the NPDES permitting process by the state in which the
CWS is constructed and by EPA.
CDPHE has identified stream standards for Clear Creek
at the Burleigh Tunnel discharge. Table 2 provides these
standards for both low- and high-flow conditions. The
zinc standard for both low- and high-flow conditions is 200
jig/L in the receiving stream (Clear Creek). In order to
met this standard, the discharge from Burleigh Tunnel
must contain less than 13,650 n-g/L zinc under low-flow
conditions and 65,700 ug/L under high-flow conditions.
2.9.4 Occupational Safety and Health Act
CERCLA remedial actions and RCRA corrective actions
must be conducted in accordance with OSHA requirements
detailed in 29 CFR Parts 1900 through 1926, especially
Part 1910.120, which provides for health and safety of
workers at hazardous waste sites. On-site construction at
Superfund or RCRA corrective action sites must be
conducted in accordance with 29 CFR Part 1926, which
provides safety and health regulations for construction
sites. State OSHA requirements, which may be
significantly stricter than federal standards, must also be
met.
Construction and maintenance personnel and sampling
teams for the Burleigh Tunnel CWS demonstration all
met the OSHA requirements for hazardous waste sites.
For most sites, the minimum personal protective equipment
(PPE) required would include gloves, hard hats (during
construction), steel toed boots, and eye protection.
Additional PPE may be required during summer or winter
months to protect against extreme temperatures.
2.10 Limitations of the Technology
Land required for constructed wetland systems is typically
extensive compared to conventional treatment systems.
Thus, in areas with high land values, a constructed
wetland treatment system may not be appropriate. Land
availability relatively close to the source of contaminated
water is preferred to avoid extended transport.
The climate at potential constructed wetland sites can also
be a limiting factor. Extended periods of severe cold,
extreme heat, arid conditions, and frequent severe storms
or flooding can result in performance problems.
Contaminant levels in treated and discharged water can
vary in response to variations of influent volumes and
chemistry. This may also be a limiting factor if there is no
tolerance in contaminant level discharge requirements.
19
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Table 2. Treatment Standards and Influent Concentrations of the CWS SITE Demonstration
ro
o
Element
Average Influent Concentration
(M9/L)
Colorado Department of Public Health and
Environment Chronic Water Quality Standards
(Clear Creek)
Low Flow (pg/L)
Colorado Department of Public Health and
Environment Chronic Water Quality Standards
(Clear Creek)
High Flow (pg/L)
Aluminum
Arsenic
Cadmium
Copper
Iron
Lead
Magnesium
Manganese
Nickel
Potassium
Silver
Sodium
Zinc
Sulfate
Fluoride
Chloride
Phosphorus (total)
Orthophosphate
Nitrate plus Nitrite (as
N)
Nitrite asN
Nitrate as N
Ammonia
20
6 150
89 0.84
<10 8.5
302 1.000
16 2,25
46,000
2,360 1,000
47 0.1
3,080
0.2 0.039
14,000
57,000 200
383,000
102
20,000
ND
66
245
ND
245
ND 0.02
-
150
0.49
4.7
1,000
0.84
-
1,000
42.09
-
0.0117
-
200
-
-
-
-
-
-
-
-
0.02
-------
Table 2 (continued). Treatment Standards and Influent Concentrations of the CWS SITE Demonstration
Element
Average Influent Concentration
(M9/L)
Colorado Department of Public Health and
Environment Chronic WaterQuality Standards
(dear Creek)
Low Flow (pg/L)
Secondary Maximum Contaminant Level
No standards exist
Colorado Department of Public Heal'
Environment Chronic Water Quality Ste
(Clear Creek)
High Flow (pg/L)
TSS
IDS
Alkalinity
(total as CaCO3)
Alkalinity
(bicarbonate as CaCO3)
Dissolved Oxygen
(mg/L)
pH
Conductivity
Temperature
11.300
680,000
104,000 -
104,000
7.9
7.11 pH units 6.5-8.5"
723 pS
8.6 °C
-
-
-
-
-
6.5-8.5*
-
-
Notes:
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Section 3
Treatment Effectiveness
The following sections discuss the treatment effectiveness
of the CWS demonstration in Silver Plume, Colorado.
The discussion includes a background section, a review of
the demonstration, demonstration methodology, site
demonstration results, and demonstration conclusions.
3.1 Background
The Burleigh Tunnel is located approximately 50 miles
west of Denver in the Georgetown-Silver Plume mining
district (Figure 1). The Georgetown-Silver Plume mining
district occupies an area of about 25 square miles
surrounding the towns of Silver Plume and Georgetown.
In general, the period of significant silver production in the
area commenced in 1872, reached a peak in 1894, and
gradually declined after. Mining in the district increased
briefly during World Wars I and II, when many old mines
were reopened and considerable amounts of lead and zinc
were mined from old slopes, dumps, and wastes left from
the silver mining boom.
The Burleigh Tunnel drains a group of mines on Sherman
and Republican mountains. Many of these mines intercept
shallow groundwater migrating through fractures in the
rock or surface water collected by slopes. The intercepted
waters are transported through the mines and are eventually
discharged through the Burleigh Tunnel. The Burleigh
Tunnel discharge contains elevated levels of zinc, typically
between 45 and 65 mg/L. However, greater than normal
precipitation during the spring of 1995 mobilized a large
amount of zinc and increased zinc concentrations within
the drainage to 109 mg/L. Burleigh Tunnel discharge
rates are generally between 40 to 60 gpm and increase to
100 to 140 gpm during spring runoff. The elevated levels
of zinc and significant flow rates combine to make the
Burleigh Tunnel a major source of zinc to Clear Creek.
Because of the large amount of zinc being discharged to
Clear Creek and the potential impact of the zinc on the
Clear Creek fishery, the drainage from the Burleigh
Tunnel was included in the Clear Creek/Central City
Superfund site.
The elevation of the Burleigh Tunnel is 9,15 2 feet, and the
climate is typical ofmountainous alpine regions in Colorado.
Summers are short and cool and winters are long and cold.
Strong eastward, down-valley winds are typical during the
winter months. Winds are lighter during the summer
months and occasionally blow westward, up the valley.
Snow accumulation during the winter months in the
immediate area of the tunnel is usually not significant due
to the open, south-facing exposure of the hillside and high
winds. Snow accumulation at higher elevations in more
sheltered areas is significant, with some snow fields
persisting until late summer. The average annual
temperature is approximately 43.5 degrees Fahrenheit
(°F), with a mean minimum of 31 °F and a mean maximum
of 55.9°F. The average annual precipitation is 15.14
inches.
3.2 Review of SITE Demonstration
The SITE demonstration was divided into three phases:
(1) CWS treatability study; (2) CWS technology
demonstration;and(3) site demobilization. These activities
are reviewed in the following sections, which also discuss
variations from the work plan and the CWS performance
during the technology demonstration phase.
3.2.1 Treatability Study
A treatability study was conducted at the Burleigh Tunnel
between June 18,1993, and August 12,1993. The goal of
the treatability study was to show that bacterial sulfate
reduction could remove zinc from the low-sulfate mine
drainage from the Burleigh Tunnel and to estimate levels
of zinc reduction that could be expectedby CWS treatment.
The treatability study involved the construction, operation,
and sampling of two bioreactors. Each bioreactor was
22
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filled with a mixture of composted manure (96 percent)
and alfalfa hay (4 percent), the same substrate that was
to be used in the CWS demonstration treatment cells.
Both reactors used an upflow configuration, in which
Burleigh Tunnel drainage entered the bioreactors from
thebottom and was forced to flow up through the substrate.
The small bioreactor was 4 feet tall and 22 inches in
diameter and held approximately 60 gallons of compost
and water. The large bioreactor was 8 feet tall and 22
inches in diameter and held approximately 130 gallons of
compostand water. Thelower6 inches of each bioreactor
was filled with gravel to support inlet piping and minimize
channeling. Peristaltic pumps were used to establish a
flow rate of 20 to 30 milliliters per minute for the small
bioreactor and 50 to 60 milliliters per minute for the large
bioreactor. The flow rates for the bioreactors were set to
provide an estimated hydraulic residence time of 50 to
100 hours.
The results of the treatability study indicated that after
8 weeks of operation, both bioreactors achieved removal
efficiencies of 99 percent for zinc and similar efficiencies
for cadmium and manganese. Zinc was the major metal
of concern for the Burleigh Tunnel drainage. Sorption of
metals in the substrate is believed to be the dominant
removal process during the first 1 to 2 weeks of bioreactor
operation. After this brief period of sorption, biological
sulfate reduction apparently became the primary metal
removal process in the bioreactors. Results of sulfate-
reducing bacteria counts and sulfate and sulfide analyses
indicated that a large population of sulfate-reducing
microorganisms was active in the system. The results
supported the theory that the bacteria reduce sulfate in the
water to hydrogen sulfide ions, which react with dissolved
metals to produce insoluble metal sulfides. The results
indicated that the Burleigh Tunnel drainage contains a
sufficient concentration of sulfate to promote metal removal
by microbial sulfate reduction. Compost sample results
from both bioreactors indicated that the compost
accumulated metals and sulfide but did not become a
reactive or hazardous waste after 8 weeks of operation.
3.2.2 Technology Demonstration
Site preparation requirements for the CWS demonstration
were minimal because of previous mining and treatability
study activities. Moreover, the area surrounding the
Burleigh Tunnel adit is level and required only minor
gradingto install the twoCWStreatmentcells. Construction
of the CWS treatment cells and all drainage conveyances
was the responsibility of the developer (CDPHE).
The demonstration evaluated two treatment cells that
differed only in flow configuration, one upward and the
other downward. The demonstration evaluated the ability
of each cell to remove zinc and other metals from the
Burleigh Tunnel mine drainage without pretreatment.
Efforts were made to maintain constant flow rates;
however, flow rates did vary. In addition, several events
resulted in brief interruptions of flow to the cells.
Approximately 12.7 million gallons of water from the
Burleigh Tunnel were passively treated by the upflow
constructed wetland cell and 11 million gallons by the
downflow CWS over the 46-month demonstration.
Figure 3 shows the flow rates measured for both wetland
cell effluents during the demonstration.
Throughout the demonstration, mine drainage influent and
wetlands system effluent samples were collected for
analysis of total metals, anions, total suspended solids
(TSS), and total organic carbon (TOC). In addition,
wetlands substrate samples were collected monthly for
sulfate-reducing bacteria analysis and quarterly for analysis
of total metals, acid-volatile sulfides (AVS), and toxicity
characteristic leaching procedure (TCLP) metals. The
substrate samples were analyzed to evaluate the
effectiveness of the treatment system in sequestering
zinc, to assess the tendency of the substrate to become a
hazardous waste, and to estimate the role of sulfate-
reducing bacteria within the wetlands substrate.
3.2.3 Operational and Sampling Problems
and Variations from the Work Plan
The CWS experienced several operational problems during
the demonstration. Some of these problems resulted in
changes to the schedule and sampling events. Problems
encountered and resolutions effected during the
demonstration are described below.
• The upflow cell froze in December 1993 and remained
frozen until the middle of February 1994. The cell
froze because flow to the cells was interrupted when
the dike within the Burleigh Tunnel collapsed. The
dike was quickly repaired; however, as a result of
the cold conditions and the lack of flow to the cells,
the upflow cell froze to a depth of 18 inches. A
livestock water heater and a steam cleaner were
used to thaw the cell so that flow through the cell
could be maintained. The freezing of the upflow cell
delayed the start of the demonstration by 1 month.
In order to prevent the upflow cell from freezing
during the winter of 1995, straw bales were placed
on top of the cell to provide insulation from the cold.
• The insulation provided by the straw bales maintained
the wetland water temperatures consistent with
23
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influent values and the upflow cell effluent piping did
not freeze.
• The 1995 spring runoff was exceptionally high, and
more flow was channeled to the CWS than the
wetlands were designed to handle. More than
20 gpm were flowing through the upflow cell for a 2-
week period in early June 1995. CDPHE responded
to the flooding by installing a 6-inch bypass pipe to
carry overflow from the influent weir around the
wetlands Once installed, the bypass allowed flow
rates to be returned to 7 gpm for each cell. However,
CDPHE had not removed the straw bales insulating
the upflow cell before the spring runoff began, and
the straw bales became saturated. The weight of
the saturated straw compressed the substrate,
reducing the flow within the upflow cell to less than
1 gpm. The straw bales were removed from the
upflow cell, and flow was restored to the cell within
a week.
• In late November 1994, a large block of rock,
roughly 10 feet by 10 feet, fell from the hillside and
rolled onto a corner of the upflow CWS cell. The
rock appeared to have depressed the effluent
accumulation network and created a high spot in the
piping at the collection point to the effluent weir.
The high point in the piping may have resulted in the
collection of precipitated metal sulfides in the piping,
causing a flow restriction.
• During the summer and fall of 1994 and 1995, the
effluent flowrate from the downflow cell could not
be maintained at 7 gpm. It was not clear if biological
surface growth, chemical precipitation in the cell, or
settling and compaction of fine particles in the
substrate was responsible for the decreased cell
permeability.
• Several substrate sampling techniques were proposed
for the demonstration, including polyethylene dipper
and sediment core samplers. Both techniques
appeared to be equally effective; however, the dippers
were determined to be preferable. The dippers
were selected because they were inexpensive and
could be dedicated to each sampling cell, reducing
the number of equipment blank samples required
during the demonstration.
3.2.4 Site Demobilization
The demonstration-scale wetland was removed by
CDPHE at the end of the demonstration. Wetland removal
entailed:
• Removal and disposal of the wetland substrate
Filling the wetland cells with site materials
Filling or removal of wetland weirs
• The CWS demonstration substrate was not a
hazardous material, and potential disposal options
included:
Disposal at a municipal landfill
- Disposal in landfill bipbeds (compost piles)
- Mixing with site mining waste rock and soil to
provide needed organic matter
Reuse in an interim ponded wetland
• The CWS Demonstration substrate was disposed of
in a nearby municipal landfill
3.3 Demonstration Methodology
The primary objectives of the CWS technology
demonstration were to (1) measure the reduction of zinc
in Burleigh Tunnel drainage resulting from the CWS
treatment with respect to cell configuration and seasonal
variation (temperature); (2) assess the toxicity of the
Burleigh Tunnel drainage; (3) characterize the toxicity
reduction resulting from treatment of the drainage by the
CWS; and (4) estimate toxicity reductions in the stream
(Clear Creek) receiving the Burleigh Tunnel drainage. In
addition, secondary objectives of the demonstration
included:
• Estimating the metal removal capacity (lifetime) of
the substrate, including the effect of treatment cell
flow configuration. The results of influent and
effluent metal analyses, CWS flow rate data, and
TCLP metal analysis were compared to substrate
metal accumulation estimates to evaluate the removal
capacities of each CWS treatment cell. The TCLP
metals analysis was used because the substrate
could become a hazardous waste before its metal
removal capabilities were exhausted. Replacing the
substrate before it becomes a hazardous waste was
determined to be the most cost-effective solution.
• Estimating the extent to which sulfate-reduction
processes within the CWS are responsible for the
removal of zinc from the drainage. Substrate was
analyzed for sulfate-reducing bacteria and acid-
volatile sulfides to estimate the extent to which sulfate-
reduction processes are removing zinc from the
drainage. The approximate number of sulfate-
reducing bacteria was correlated to metal removal
efficiencies as part of the determination. In addition,
the accumulation of AVS in the substrate was
compared to metal loading in the treatment cells to
determine trends. Furthermore, the AVS analyses
included an analysis of zinc to verify that the metal
sulfides accumulating in the CWS were zinc sulfides.
Previous investigations suggested that AVS analyses
were indicative of metal sulfide accumulation
attributed to sulfate-reducing bacteria (Reynolds
1991).
24
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• Evaluating the impact of the CWS effluent on Clear
Creek. Clear Creek samples were analyzed for total
metals, TSS, total dissolved solids (TDS), TOC,
nitrate, and phosphate. Results of the stream analyses
were compared to CWS effluent analyses to assess
the effect of CWS effluent on Clear Creek. Clear
Creek samples were collected upstream and
downstream of the CWS outfall.
• Estimating the capital and operating costs of the
CWS.
Critical parameters are the data required to meet the
primary objectives. The primary critical parameters were
influent andeffluent analyses for zinc (total), and toxicity
testing with fathead minnows (Pimephalus promelas) and
water fleas (Ceriodaphnia dubia).
Noncritical parameters are data required to address
secondary objectives of the demonstration. Secondary
objectives provide useful information topotential technology
users but are not critical to evaluate the technology. The
noncritical parameters of the CWS demonstration included:
• Total metals, nitrate and phosphate analysis of the
Burleigh Tunnel drainage and CWS effluents
• Metal loading, metal accumulation, and TCLP metals
in CWS substrate samples
• Sulfate-reducing bacteria counts and AVS
accumulation in CWS substrate samples
• Clear Creek samples for total metals, TDS, TSS,
TOC, biochemical oxygen demand (BOD), and
aquatic toxicity
• Construction, operation, maintenance, substrate
disposal, and miscellaneous costs
3.3.1 Testing Approach
In general, the testing approach of the demonstration
incorporated the collection and analysis of wetland influent
and effluent samples every 2 weeks for a period of
20 months. Monthly sampling was conducted for the
remainder of the nearly 4-year demonstration. The
effluent zinc results for each sampling event were
compared to influent data and a removal efficiency
calculated. An initial 2-week interval was selected
because it provided for 3 to 7 pore volumes of water to be
passed through the CWS, assuming a hydraulic residence
time of between 50 and 100 hours. In addition, the 2-week
interval was chosen because several factors, such as
precipitation or evaporation, could cause variation in the
measured concentration ofzinc in wetland effluent samples.
By increasing the number of influent and effluent water
samples, performance trends display better continuity, the
effects of weather are reduced, and calculated removal
efficiencies are expected to more closely reflect true
values. Also, sampling intervals shorter than 2 weeks
were not economically feasible considering the length of
the demonstration. The initial 20-month schedule was the
maximum tune allowable for the demonstration. This time
frame is allowed because the CWS is a biological
technology and performance depended, in part, on primary
substances andnutrients within the substrate. By allowing
the system to operate for an extended period, results were
expected to show a relationship (positive or negative)
between declining nutrient concentrations in the substrate
and CWS performance.
The frequency of demonstration toxicity testing was
limited to every 3 to 4 months due to budget considerations.
Essentially, the sample collection and testing schedule
was designed to evaluate toxicity reduction during periods
of widely different zinc removal (different seasons) and
critical periods for the receiving stream.
3.3.2 Sampling, Analysis, and
Measurement Procedures
Mine drainage samples were collected from the influent
weir, and CWS effluent samples were collected from the
effluent weirs. Clear Creek samples were collected
above and below the CWS outfall. Influent and effluent
samples were analyzed for total recoverable zinc and
toxicity (critical analyses), other metals, anions, TDS,
TSS, and TOC (effluent only). These samples were
collected at the frequency discussed in the previous
section.
Two substrate sampling points were located in each cell.
Initially, substrate samples were collected monthly for
sulfate-reducing bacteria analysis and quarterly for total
metals, AVS, and TCLP metals analyses for a period of
20 months. Quarterly and semi-annual sampling was
conducted for the remainder of the demonstration.
Substrate samples were collected from two locations
within each cell, at approximately I to 2 feet below the
wetland surface.
Mine drainage, wetlands effluent, and substrate were
analyzed for critical and noncritical parameters using the
methods listed in Table 3.
Field analyses included measurement of pH and
conductivity for all aqueous samples, Eh for wetlands
effluent samples, and dissolved oxygen for mine drainage
25
-------
Table 3. CWS Demonstration Summary of Standard Analytical Methods and Procedures
Parameter
Metals
Sulfate
Fluoride
Nitrate/Nitrite
Chloride
Total and
Orthophosphate
pH
TSS
TDS
TOG
Ammonia
Alkalinity
Sulfide
Aquatic Toxicity
Acid Volatile Sulfide
(AVS)
Sulfate reducing bacteria
count
Toxicity leaching
procedure
Reactive sulfide
Orthophosphate
Sulfate
Physical parameters
Residence time
PH
Temperature
Dissolved oxygen
Conductivity
Notes:
1 Test Methods for Evaluat
Sample Type
Aqueous and
Substrate
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
Substrate
Substrate
Substrate
Substrate
Substrate
Substrate
Substrate
Aqueous
Aqueous
Aqueous
Aqueous
Aqueous
ina Solid Wastes. V<
Method Number
6010A, 6020, 7470
300.0
9056
353.2 and 354.1
300.0
365.3
9040
160.2
160.1
9060
350.1
310.1
376.2
EPA SOPs3
EPA Method
None
1311
EPA4
365.3
300.0
Various 3
ND
SOP3 12
SOP3 11
SOP3 62
SOP3 99
Dlumes IA-IC: Laboraton
Method Title
ICP, ICP/MS, orAA
Ion chromatography
Ion chromatography
Various
Ion chromatography
Various
Electrometric
Gravimetric
Gravimetric
Various
Various
Various
Various
Acid volatile sulfide
Anaerobic deep tube
ICP, ICP-MSorAA
Titration
Various
Various
Variouss
ND
/ Manual. Phvsical/Chemi
Source
SW-8461
MCAWW2
SW-846
MCAWW2
MCAWW2
MCAWW
MCAWW
MCAWW
MCAWW
SW-846
MCAWW2
MCAWW2
MCAWW2
EPA5
EPA 1991
CSM3
SW-846
SW-846
MCAWW
MCAWW
ASTM
ND
Tetra Tech6
Tetra Tech6
Tetra Tech6
Tetra Teche
cal Methods: at
Volume II Field Manual. Physical/Chemical Methods, SW-846. 3d Edition. Office of Solid Waste and Emergency
Response. U.S. Environmental Protection Agency (EPA). 1986.
2 Methods for Chemical Analysis of Water and Wastes (MCAWW). EPA 600/4-79-020. Environmental Monitoring and
Support Laboratory, Cincinnati, Ohio. EPA. 1983 and subsequent EPA-600/4.
3 The analytical methods selected for the analysis of critical and noncritical parameters, and the rationale used in their
selection, are discussed in Section 4.2,
4 Interim Guidance for Reactive Sulfide. Section 7.3.4.2, SW-846.
5 Methods for Measuring the Acute Toxicity of Effluents and Receiving Waters to Freshwater and Marine Organisms.
EPA/600/4-90/027F. EPA 1993.
These are field measurements made by Tetra Tech.
26
-------
and Clear Creek samples. All field measurements were
made in accordance with standard operating procedures.
3.4 Site Demonstration Results
This section presents the results of the CWS demonstration
conductedfromJanuary 1994 to November 1997. Initially,
aqueous chemistry data for the Burleigh Tunnel mine
drainage are presented, followed by the demonstration
results for the two CWS cells (Sections 3.4.1 through
3.4.3).
Section 3.4.4 presents data for the receiving stream, Clear
Creek, and Sections 3.4.5 and 3.4.6 present toxicity
results. Tables summarizing analytical results for
the Burleigh Tunnel mine drainage are included in Appendix
A. An evaluation of demonstration data quality parameters
for critical analyses is contained in Section 4.
The data discussed in this section were generally collected
using demonstration sampling and analysis techniques.
However, influentandeffluentdataformuchof 1996 were
collected and analyzed by the CDPHE laboratory
(Analytica, in Broomfield, Colorado). In addition, data
was not collected by Tetra Tech or CDPHE for 3 months
(September through November) in 1996. Tetra Tech
discontinued CWS sampling at the end of its initial SITE
contract and the resumption of sampling was slowed by
contractual delays.
3.4.1 Burleigh Mine Drainage Chemistry
The Burleigh Tunnel drains a network of interconnected
mines on Republican Mountain and Sherman Mountain.
Unlike many metal mine drainages, the Burleigh Tunnel
effluent has near-neutral pH and carbonate alkalinity of
approximately 100 mg/L.
The mine drainage contains high levels of zinc that
typically range from 45 to 65 mg/L. However, in May and
June 1995, agreat deal of spring snow and rain and arapid
thaw combined to increase the amount of runoff entering
the mine network drainedby the Burleigh Tunnel. At that
time, flow from the tunnel increased from 45 gpm to more
than 300 gpm, and zinc concentrations increased from 55
mg/L (April 12,1995) to 109 mg/L (August 8,1995).
Over the final 2 years of the demonstration, zinc
concentrations in Burleigh Tunnel mine drainage were
lower in the winter, dropped again in April or May when
flow through the mine workings increased, and rapidly
increased in summer, remaining high throughout the fall.
During this period, Burleigh Tunnel mine drainage
zinc concentrations generally remained between 45 and
84 mg/L, with increases to more than 100 mg/L noted
during the late summer and fall. Zinc concentrations in
Burleigh Tunnel mine drainage between September and
November 1996 are assumed to be similar to zinc
concentrations measured during the same period in 1995.
Figure 4 shows zinc concentrations for the Burleigh
Tunnel mine drainage measured during the demonstration.
In addition to zinc, cadmium, lead, nickel, and manganese
are also demonstration metals of interest. Cadmium, lead,
and nickel readily form sulfides and are expected to be
removed by the CWS. Manganese does not form a stable
sulfide but was shown to be removed in a short-
term treatability study conducted prior to the demonstration
(PRC 1993). Cadmium, lead, and nickel levels were
generally less than 0.1 mg/L in the Burleigh Tunnel mine
drainage. After the high flow event in 1995, cadmium
levels increased to concentrations ranging from 0.11 to
0.26 mg/L. Lead and nickel levels were generally much
lower than cadmium and did not increase to the same
extent after the high flow event.
Anion concentrations also increased during the
demonstration. Sulfate concentrations in the Burleigh
Tunnel drainage ranged from 279 to 652 mg/L and also
increased after the high flow event. Carbonate (total
alkalinity) concentrations were measured over arelatively
narrow range of 82.4 to 125 mg/L. The highest carbonate
concentrations were measured during a 1-month period
in June and July 1995, corresponding to the period of
highest flow from the Burleigh Tunnel. The simultaneous
increases in zinc, sulfate, carbonate, and calcium without
an increase inpH suggest these mine drainage constituents
originate from mineral dissolution. Calcite (CaCO3) is
commonly found in hydrothermal vein deposits in
association with lead-silver-zinc formations (Correns 1969)
and is also reported in the Silver Plume mining district.
The high concentration of both zinc and carbonate at near
neutral pH suggests the Burleigh Tunnel mine drainage is
a combination of waters from multiple sources.
3.4.2 Downflow CWS
The downflow cell was operated for approximately
2 '/a years during the demonstration. Over this period, the
system removed 60 to 95 percent of the zinc contamination
from the Burleigh Tunnel mine drainage.
Figure 4 shows zinc concentrations in the Burleigh Tunnel
mine drainage (influent), and the effluents of both CWS
27
-------
100
80 -
E 60 -
2Q H
Oi
I
-*•
O)
I
o>
I
Q.
0>
tn
O)
I
H-H-
en
s
-i — i — i — i — i — i — i — i — i — i — r—i — f— i — i — r
m m m m m
0> CR 0) 0)
c Ji I i I
0
-i — i — i — i — i — i — \ — i — i — i — i — i — i — i — i — i — r-
m m to to to O> OJ CO O) O>
111)11
0. O >- C Q. O
a> o> o -3 Ul Q
~i — i — r-
r--
OJ
L
a
2
T — 1 — 1 — 1
r-.
Ol
1
c
-3
I — f-_
r-
01
O-
co
Month
Influent Zinc
Downflow Effluent Zinc
Upflow Effluent Zinc
Inferred Estimated
Figure 4. CWS zinc concentrations by month.
-------
cells. During the first year of operation, influent zinc
concentrations ranged from 45 to 63 mg/L (average of
57.1 mg/L) and the amount of zinc removed by the
downflow cell ranged from 35 to 54 mg/L (average of
44.2 mg/L). Zinc removal efficiency during the first year
averaged 77.4 percent. During the second year, zinc
levels in mine drainage ranged from 53 to 109 mg/L
(average of 83 mg/L) and downflow zinc removal ranged
from 41 to 78 mg/L (average of 58 mg/L). Zinc removal
efficiency during the second year averaged 70 percent.
Over the final 6 months this cell operated, influent zinc
levels ranged from 46 to 84 mg/L, while downflow CWS
zinc removal ranged from 31 to 78 mg/L. In general,
greatest zinc removal corresponded to times with the
highest influent zinc concentrations, and the lowest zinc
removal was observed during periods of lesser zinc in the
mine drainage suggesting metal removal was effected by
a physical process.
Although present only in low levels in the influent water,
cadmium, lead, and nickel were removed to a great extent
by the downflow CWS treatment. Influent cadmium
concentrations ranged from 0.071 to 0.10 mg/L, while
effluent levels ranged from 0.0007 to 0.003 mg/L during
the first year. During the second year, cadmium
concentrations increased in the influent, ranging from
0.057 to 0.26 mg/L, and downflow effluent levels ranged
from 0.0001 to 0.007 mg/L with few detections. Figure 5
shows cadmium concentrations for the influent and both
effluents during the first 2 years of the demonstration
Substantial cadmium removal continued over the final
6 months by the downflow cell, with the exception of the
April 1996 sample.
Samples were not regularly analyzed for lead or nickel
during the demonstration. Figure 6 shows lead
concentrations for the influent and both effluents during
the first 2 years of the demonstration. During the first
year, influent lead concentrations ranged from 0.013 to
0.020 mg/L, while downflow effluent concentrations
ranged from 0.00065 to 0.0054 mg/L. Throughout the
remainder of 1995, influent levels of lead increased
slightly while effluent levels remained very low with few
detections.
Nickel was also removed by the downflow cell; however,
the extent of removal declined when influent nickel
concentrations increased after the high flow event.
Nickel levels in the influent ranged from 0.033 to 0.68 mg/
L, and downflow effluent ranged from 0.0073 to 0.020
mg/L in thefirstyear, Throughouttheremainderof 1995,
influent nickel levels ranged from 0.045 to 0.093 mg/L,
and downflow effluent levels ranged from 0.014 to
0.040 mg/L.
Manganese concentrations in the mine drainage were
initially between 1 to 2 mg/L. Manganese removal by the
downflow CWS was low during the demonstration. Figure
7 shows manganese concentrations for the influent and
both effluents.
The extended residence time of the influent within the
downflow cell substrate caused by low flow rates may be
one reason the downflow CWS was effective in removing
metals from the mine drainage. Both wetland cells were
designed to treat 7 gpm; however, the permeability of the
downflow cell declined during the first year of operation,
and flow through the cell dropped to 4 gpm particularly
during the summer months. Although attempts were
made to increase its permeability by fluffing the substrate
with compressed air, these procedures resulted in only
temporary improvements. Flow through the downflow
cell improved during winter months when the substrate
froze and contracted from the liner allowing the influent to
flow down the sides of the interior cell. Flow through the
downflow cell averaged 6.5 gpm during the first year; 5.8
gpm in the second year; and 6 gpm over the final 6 months
of operation.
Analytical results for the downflow substrate (Table 4)
showed a substantial increase in zinc levels over the
period of the demonstration. Substrate zinc levels ranged
from a low of 59.7 milligrams per kilogram (mg/kg) to a
high of 5,630 mg/kg. Substrate samples were generally
collected from between 1 to 2 feet below the surface of
the CWS. Downflow substrate samples contained little
visible evidence of sulfate reduction and low concentrations
of AVS. Sulfate-reducing bacteria counts showed much
variability (FigureS).
After the first 6 months of operation, the downflow cell
was removing more zinc from the mine drainage compared
with the upflow cell. However, the reason for the greater
removal was likely the higher residence time of the mine
drainage within the downflow wetland. The increasing
residence time was a function of mine drainage flow
through the cell, that was generally lower in the summer
compared to winter. A reduction of flow from 7 to 5 gpm
increases residence time by 19 hours nearly a 40 percent
increase. The loss of permeability is believed to be related
to the loss of permeability in the downflow cell resulting
from biological surface growth, chemical precipitation of
29
-------
oe
I
»
1/1
n
Cadmium (mg/L)
ooooopppppppp
b b b b '-* '-* *_» 1* L* io k) ki k>
•q o
=2 *
o 5
* =5
o
D
a
3
c
3
o
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a
3
E
3
May-97
Sep-97
-------
0.06
0.05
0.04
65 0.03
E
S 0.02
0.01
-4-
o>
I
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I
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JD
01
U.
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a.
0)
Month
Influent Lead
Downflow Effluent Lead
Upflow Effluent Lead
Figure 6. CWS lead concentrations by month.
-------
o
LJ
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LJ
CO
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•si-
en
I
QL
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POPULATION COUNTS PER GRAM
CJI
o
b
o
o
Ol O O1
o o o
(ji UJ •£• 4* Cn
O tn o ui O
p o p o o
b b b b b b b b b
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MAR-94
APR-94
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3 FEB-95 --
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-------
Table 4. Average Downflow CWS Substrate Results
0-6 months
6- 12 months
12-18 months
Cadmium
(mg/kg)
2.7
8.0
23
Lead
(mg/kg)
18
31
74
Nickel
(mg/kg)
3.1
6.1
7.0
Znc
(mg/kg)
1,100
3,400
5,200
Acid Volatile
Sulfides
(mg/kg)
180
120
460
Sulfate-
Reducing
Bacteria
(count)
8.5 x104
1.1 x105
3.3 x 10s
Ortho-
phosphate
(mg/kg)
34
12
2.6
Notes:
mg/kg Milligram per kilogram
Average Arithmetic Mean
Substrate samples collected from 1-2 feet below wetland surface
zinc compounds, microbial breakdown of the substrate to
finer particulates, and the settling of these particles into
substrate pore spaces. The increase of flow during winter
is believed to result from freezing of the wetland substrate
at the edge of the cell causing the substrate to contract
from the liner. The contraction allowed ponded water at
the surface of the wetland to flow between the frozen
substrate and liner to the base of the cell forming a
preferential pathway.
Loading is the amount of metals retained by the wetland
over time. It is a function of the flowrate through the
wetland, the concentration of metals in the mine drainage,
and the removal efficiency of the treatment. For this
discussion, monthly loading of each wetland was calculated
from measured flow rates and simultaneously collected
samples of the mine drainage and the wetland effluent.
Figure 9 shows the monthly zinc loading to the downflow
CWS over the demonstration. The graph indicates that
loading was initially high (maximum of 60 kg/month) but
dropped as the downflow cell flow rate declined in the Fall
of 1994. In winter, loading also increased as flow
improved. The greatest loading to the downflow CWS
occurred during the high flow event in the late spring and
early summer of 1995. After the high flow event, loading
in this cell declined dramatically and eventually dropped to
less than 5 kg/month in May 1996.
The primary metal removal mechanism active in this cell
did not appear to be sulfate reduction. Substrate analyses
indicate a significant portion of the zinc removal in this
CWS occurred in the upper 1 to 2 feet of substrate, where
few AVS or sulfate-reducing bacteria were found. Pockets
of sulfide-rich substrate were observed in this CWS cell
at depths of 3 to 4 feet below the wetland surface,
suggesting some sulfate reduction contributes to metal
removal in this wetland. Aqueous geochemical modeling
of the mine drainage suggests gypsum is oversaturated;
however, visual observations of Burleigh Tunnel mine
drainage precipitate and historical mine reports suggest
the material is a zinc carbonate, probably smithsonite or
hydrozincite.
The following can be concluded from the evaluation of the
downflow CWS:
• As tested, the downflow CWS did not retain sufficient
permeability to be considered a reasonable long-
term treatment option.
• Chemical precipitation (suspected to be mineral
carbonate accumulations) may have been the primary
metal removal process in this CWS treating Burleigh
Tunnel mine drainage.
• A 2-foot substrate depth should be adequate, as
most metal removal occurred at between 1 to 2 feet
below the wetland surface. A thinner substrate
should decrease the flow resistence of the downflow
CWS and increase the effectiveness of the system.
• A 2-foot downflow CWS may be a good pretreatment
for an upflow CWS treating the Burleigh Tunnel
mine drainage allowing some physical precipitation
of the zinc.
The concentration of orthophosphate in the substrate also
decreased after the high flow event in 1995. The high
orthophosphate concentration, measured at the beginning
of the demonstration, was 114 mg/kg; the low, 1 to 2 mg/
kg, was measured in August 1995.
34
-------
Date
Figure 9. Monthly zinc loading, downflow CWS.
-------
3.4.3 UpflowCWS
The upflow cell was demonstrated for nearly 4 years and,
during this period, removed zinc and other metals initially
by adsorption, later by sulfate reduction, and eventually by
chemical precipitation (presumed). Theadsorptionperiod;
appeared to last roughly 4 to 5 months as indicated by
manganese removal. After the adsorption phase, sulfate
reduction appeared to be the primary metal removal
process; however, oxidation/reduction (ORP)
measurements suggested the activity of the sulfate-
reducing bacteria appeared to drop in late fall and through
the winter of 1994. Counts of sulfate-reducing bacteria
declined coincidentally with the decline in ORP. The drop
may have been caused by lower winter temperatures, or
an increase in flow through the cell that occurred in
September through October 1994, or may result from the
use of all the most easily metabolized materials in the
compostsubstratebythe bacteria. During this period, the
concentration of zinc in the upflow effluent increased
from3.2mg/L(Octoberl2,1994)tol8mg/L(Marchl5,
1995).
By May 1995, zinc levels were approaching levels that are
inhibitory to sulfate-reducing bacteria at the observed
area loading of 250 square feet per gallon. During May
and June of that year, the high flow event exposed the
wetland sulfate-reducing bacteria to elevated levels of
zinc, and the high influent flow probably created aerobic
conditions within the cell. The periodic high zinc
concentrations observed in influent waters during the
summer and fall of 1996 and 1997 likely prevented the
sulfate-reducing bacteria from reestablishing activity to
previous levels. The flow was halted to the upflow cell in
the summer of 1997 for approximately one month for
repairs. At that time, much of the water was removed
from the cell, allowing wetland sulfate-reducing bacteria
an opportunity to become reestablished.
However, there was no indication thatthebacteriabecame
re-established during the final 4 to 5 months of the
demonstration. One of the repairs involved plugging a
short section of the influent piping in the upflow cell.
Visible observation of this influent pipe noted a black
coating on the inside of approximately 1/16 inch and
accumulations of black precipitate nearly filling the holes
in the perforated pipe. Overlying the black material in the
piping was a layer of cream colored to yellow material up
to 1/8 of an inch thick.
Analytical results for influent and effluent samples from
the upflow system showed that zinc was nearly completely
removed by this system during the first 8 months of the
demonstration (Figure 4). After this period, zinc
concentrations in the upflow effluent gradually increased
from 1.4 mg/L (September 19,1994) to 18.5 mg/L in the
spring of 1995 corresponding to zinc removal efficiencies
of97.6and66.8, respectively. InMayandJune 1995,high
flow from the Burleigh Tunnel increased flow through the
upflow cell to 20 gpm and zinc concentrations nearly
doubled. Over the next 6 months, as flow decreased from
the tunnel, influent zinc concentrations rose to a high of
109 mg/L. From May to November 1995, effluent zinc
levels increased from 26.7 to 73.6 mg/L. The amount of
zinc removed by the upflow cell averaged 41 mg/L (49.3
percent) during the second year.
During the third year of operation, zinc levels in the
influent ranged from 56 to 84 mg/L; however, data were
not collected between September and November 1996.
Zinc concentrations in the upflow effluent over the third
year ranged from 30 to 49 mg/L with an average removal
of 30 mg/L (39,6 percent). In the final year of operation,
zinc influent concentrations ranged from 42 to 104 mg/L
and effluent levels ranged from 15 to 60 mg/L with an
average removal efficiency of 65.1 percent. Effluent
levels were greater in the May 28,1997 sample (60 mg/
L) compared to the influent sample (56 mg/L). Over the
final 6 months, the upflow cell removed greater amounts
of zinc as flow through the cell decreased. Flow through
the upflow cell at this time ranged from 2 to 5 gpm.
Cadmium removal by the upflow cell followed a pattern
similar to zinc removal (Figure 5). Initially, cadmium was
removed to nondetect levels; however, cadmium
concentrations increased two and a half times after the
high flow event. After this period, cadmium removal
remained high for 4 months but declined in the latter part
of 1995 and remained low through 1996 and 1997.
Lead (Figure 6) and nickel were also removed to lower
concentrations by the upflow CWS. Influent lead and
nickel concentrations were approximately 0.015 mg/L
and 0.043 mg/L, respectively. During the first year, lead
was removed to nondetect levels and nickel effluent
concentrations ranged from 0.0005 to 0.019 mg/L. Unlike
zinc and cadmium, lead and nickel concentrations did not
increase significantly after the high flow event; however,
the removal of both decreased somewhat until flow values
through the cell declined in the final months of the
demonstration.
Manganese was initially present in the mine drainage at
concentrations ranging from 1 to 3 mg/L. Manganese
36
-------
was removed by the upflow cell for the first 4 months of
operation but was not removed throughout the remainder
of the demonstration.
Analytical results for the upflow substrate showed an
increase in zinc levels over the period of the demonstration.
Table 5 summarizes mean annual results for selected
analysis from upflow cell substrate samples collected
during the demonstration. Zinc levels ranged from a low
of 40 mg/kg to a high of 4,800 mg/kg. The zinc content is
expected to be higher in the removal zone of the upflow
cell (deeper in the substrate of the cell). In general, upflow
substrate samples were collected approximately 2 feet
below the wetland surface, above the removal zone.
Counts of sulfate-reducing bacteria in the upflow cell
were generally very high between April 1994, through
July 1995. However, counts were 1 to 2 orders of
magnitude lower in upflow cell samples collected in
April 1996 through September 1997. The final substrate
sample analyzed for sulfate-reducing bacteria
contained approximately 250,000 CFU/gram substrate.
Figure 10 shows the results of sulfate-reducing bacteria
counts conducted on upflow cell substrate samples
collected during the demonstration.
The change from strongly reducing to slightly reducing
conditions in the fall of 1994 may have made previously
removed metal sulfides less stable within the wetland
substrate. Substrate observations in the summer of 1997
indicated there were fewer sulfides present compared to
substrate samples collected in 1994 and 1995. If half of
the zinc removed in the first year of operation were
released over the subsequent 2 years, the resulting zinc
increase in the effluent would have been 33 mg/L. The
higher zinc concentration measured in the May 28,1997
effluent sample compared to the corresponding influent
sample suggests some previously removed zinc was
released.
Between March and December 1994, metals loading to
the upflow CWS ranged from 53 to 97 kg/month but
dropped to 26 kg/month in February 1995. This drop in
loading corresponded with the increase of zinc in the
effluent, an increase in ORP, and a decrease in flow rate
through the cell. Flow through the cell increased in March
andApril I995,leadingtohigherloading. Themaximum
loading to the upflow CWS (107 kg/month) occurred in
May 1995 during the high flow event. Throughout the
remainder of the demonstration, loading to this cell declined
as the zinc removal efficiency decreased to 40 to 50
percent; eventually, flow through the cell ended in 1997.
Figure 11 shows zinc loading to the upflow CWS over the
demonstration.
The effect of the high flow event on the performance of
the upfiow CWS reveals the major shortcoming of passive
systems, the inability to adapt to rapidly changing conditions.
In this demonstration, the upflow CWS could not adjust to
the increased influx of zinc or the change in environmental
conditions.
As several constructed wetlands have successfully treated
mine drainage with much higher concentrations of zinc, it
may be concluded that the bacteria are somehow able to
protect themselves from the high metals concentration. If
this mechanism is sulfate reduction, the rate of sulfate
Table 5. Average Upfiow CWS Substrate Results
Yearl
Year 2
Year 3
Year 4
Cadmium
(mg/kg)
0.17
0.18
5.0
9.6
Lead
(mg/kg)
9.9
13
40.0
NR
Nickel
(mg/kg)
1.9
2.0
4.1
6.2
Zinc
(mg/kg)
40
71
1,500
4,800
Acid Volatile
Sulfides
(mg/kg)
210
460
1,300
1,000
Sulfate-
Reducing
Bacteria (count)
7.2 x106
3.2 x106
2.2 x105
6.2x10"
Ortho-
phosphate
(mg/kg)
55
54
6.3
6.9
Notes:
mg/kg Milligram per kilogram
NR Not Reported
Average = Arithmetic Mean
Substrate samples collected from 1-2 feet below wetland surface
37
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POPULATION COUNTS PER GRAM
O
O
O
b
o
o
o
o
o
b
o
o
o
o
o
b
o
o
10
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APR-94
MAY-94
-------
120
100
M
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in
01
01
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Figure 11. Monthly zinc loading, upflow CWS.
-------
reduction must be great enough to reduce zinc
concentrations in the substrate to below inhibitory levels.
This hypothesis suggests that the effectiveness of an
anaerobic compost CWS is a function of the rate of sulfate
reduction, residence time of the mine drainage in the
wetland substrate, and the concentration of zinc (or other
inhibitory metals) in the mine drainage. Low temperature
is also a factor that will affect the activity of sulfate-
reducing bacteria in the wetland.
The following can be concluded from the evaluation of the
upflowcell:
• The upflow CWS is effective in removing many
metal contaminants from mine drainage; however,
the CWS may have difficulty recovering from rapidly
increasing metals loading conditions. Reinnoculation
and incubation of sulfate-reducing bacteria may
improve recovery of these systems.
• Control of mine drainage flow to the constructed
wetland is critical to ensure that residence time and
operational conditions are maintained.
• The operational lifetime of an upflow CWS (with a
compost substrate depth of 4 feet) is roughly 4 to
5 years.
• The upflow cell had superior hydraulic performance
throughout most of the demonstration.
• Winter freezing can be prevented by covering the
wetland surface with hay or blankets used in curing
concrete.
• Piping cleanouts should allow all piping networks to
be easily cleaned.
3.4.4 Clear Creek
The untreated Burleigh Tunnel mine drainage and the
effluents of both CWS cells discharge to Clear Creek. To
assess the impact of treatment on the receiving stream,
upstream and downstream Camples collected from Clear
Creek were also analyzed for total metals and aquatic
toxicity. The metals results indicated that although the
wetlands may be removing metals from the mine drainage,
the demonstration-scale CWS treated only a small portion
of the total discharge from the Burleigh Tunnel, not
enough to show a measurable decrease in the metals
content of the stream. The demonstration-scale CWS
treated approximately 30 percent of the total flow from
the Burleigh Tunnel, and during high flow treated only
about 5 percent of the flow. A full-scale system could
show a more significant decrease in the metals content of
Clear Creek downstream of the system.
The stream results for upstream versus downstream
samples are presented in Tables 6 and 7, The results show
that Burleigh Tunnel mine drainage is a significant source
of zinc to Clear Creek. However, CDPHE reports there
are also additional nonpoint sources of zinc-contaminated
water received by the creek.
3.4.5 Toxicity Testing Results
Constructed wetland treatment is a complex
biogeochemical process involving adsorption, chemical
precipitation, and microbial interactions with contaminants.
The primary metal removal mechanisms in the CWS are
chemical precipitation and microbial sulfate reduction;
however, treatment may also complex metal contaminants,
making them unavailable to receptor organisms. Thus,
aquatic toxicity analyses were conducted by the EPA
National Exposure Research Laboratory - Aquatic Toxicity
during the demonstration to evaluate the reduction in
toxicity resulting from CWS treatment. Two test organisms
were used in the toxicity testing: water fleas (Ceriodaphnia
dubia) and fathead minnows (Pimephales promelas). A
total of eight rounds of aquatic toxicity testing were
conducted during the demonstration. Initially, toxicity
samples were collected and analyzed every 3 to 4 months
until late 1995, when demonstration activities were
temporarily suspended. When demonstration monitoring
resumed, toxicity testing was conducted every 4 to
6 months. In 1997, a microbial toxicity test was conducted
on wetland sulfate-reducing bacteria with Burleigh Tunnel
mine drainage. The results of the microbial toxicity test
are presented in Section 3.4.6.
Aquatic toxicity testing results correlated well with
increasing zinc concentrations observed in the effluents of
the treatment cells during the first 2 years of the
demonstration. Results of testing conducted during the
first 8 months of the demonstration indicate the effluents
from both cells were not toxic to either the C. dubia or the
P. promelas. The Burleigh Tunnel mine drainage was
toxic to both test organisms at low concentration (dilution)
throughout the demonstration. Table 8 provides influent
and effluent concentrations resulting in the death of
50 percent of the test organisms (LC50) in each round of
testing. As zinc concentrations increased in the effluents
of both cells through 1995, so did the toxicity to the test
organisms.
The first test conducted that year (February 1995) indicated
that effluent from the upflow cell had become toxic to
C. dubia at a concentration of 8.4 percent. The high
runoff event that occurred in the spring of 1995 and
40
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Table 6. Clear Creek Upstream
Average
Maximum
Minimum
Cadmium
(mg/L)
0.0022
0.0094
0.0
Lead
(mg/L)
0.0034
0.013
0.0
Nickel
(mg/L)
0.0047
0.015
0.0
Zinc
(mg/L)
0.126
0.56
0.11
PH
7.8
8.1
7.6
Conductivity
(MS)
155.7
167.5
144.0
Temperature
CO
5.4
9.7
0.9
Notes:
°C
mg/L
MS
ND
PH
Degrees Celsius
Milligrams per liter
MicroSiemens
Not Detected
Standard units
Average sArithmetic Mean
Table 7. Clear Creek Downstream
Cadmium
(mg/L)
Average 0.00075
Maximum 0.0017
Minimum ND
Lead
(mg/L)
0.0013
0.0024
ND
Nickel
(mg/L)
0.0068
0.026
ND
Zinc
(mg/L)
0.512
0.56
0.14
pH
7.6
8.1
6.5
Conductivity
(MS)
132.8
173.3
80.0
Temperature
4.3
9.7
Notes:
°C Degrees Celsius
mg/L Milligrams per liter
pS MicroSiemens
ND Not Detected
pH Standard units
Average = Arithmetic Mean
associated increases in flow through the CWS cells and
elevated zinc concentrations resulted in higher zinc levels
in the CWS effluents. At that time, the effluent from both
cells became toxic to the test organisms. The upflow cell
effluent was toxic to C. dubia at a concentration of
0.1 percent and to P. promelas at concentrations ranging
from 1.2 to 2.3 percent. The downflow cell effluent was
toxic to C. dubia at concentrations ranging from 0.31 to
0.51 percent and to P. promelas at concentrations ranging
from 2.6 to 30 percent.
Over the final 2 years of the demonstration, the upflow cell
effluent continued to be toxic to C. dubia at concentrations
below 1 percent and to P. promelas at a concentration of
14 percent. Toxicity samples were not collected from the
downflow cell: operation of this cell was discontinued in
September 1996.
Demonstration toxicity testing results indicate that the
ability of the wetlands to reduce toxicity to aquatic
organisms gradually declined over the first 2 years. In
addition, the high flow event in 1995 had a significant
impact on zinc and toxicity removal by the upflow cell over
the final 2 years of the demonstration.
Water samples for toxicity testing were collected from
Clear Creek above and below the CWS discharge three
times during the demonstration. As mentioned, the
constructed wetlands treated only 30 percent of the mine
drainage; thus, the impact of treatment on the receiving
stream was minor. One set of samples contained higher
toxicity in the upstream sample while samples collected
after June 1995 indicated that there was no acute toxicity
in the upstream samples but that addition of the mine
drainage to the stream resulted in an increase in toxicity.
41
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Table 8. CWS Demonstration Toxicity (LC,n) Results
Indicator Species
Fathead Minnows
(Pimephalus
promotes)
Water Fleas
(Ceriodaphnia dubia)
Date
Collected
08/24/94
09/19/94
02/22/95
06/12/95
09/05/95
12/10/96
06/24/97
1 0/29/97
10/29/971
08/24/94
09/19/94
02/22/951
02/22/95
06/12/95
12/10/96
06/24/97
09/05/95
1 0/29/97
10/29/971
Influent
1.1
0.73
1.6
1.0
0.62
0.62
0.69
1.4
0.46
0.31
1.0
0.10
0.09
0.43
0.10
0.15
Upflow
Effluent
No toxic ity
No toxic ity
No toxicity
2.3
1.2
1.6
24
14
11
No toxicity
No tox icily
8.4
0.43
0.22
0.41
<0.19
0.13
0.19
Downflow
Effluent
NA2
No toxicity
No toxicity
2.6
30
NA
NA
NA
NA
No toxicity
No toxicity
No toxicity
0.51
NA
NA
0.31
NA
NA
Clear Creek
Upstream
No toxic ity
No toxicity
No toxic ity
No toxicity
No toxicity
No toxicity
Clear Creek
Downstream
No toxicity
No toxicity
No toxicity
No toxicity
No toxicity
No toxicity
Notes:
1 Duplicate Sample
2 NA-Not analyzed
3.4.6 Microbial Toxicity Testing
Microbial toxicity testing was undertaken when repairs to
the upflow cell indicated that there were few metal
sulfides in the wetland substrate compared with
observations conducted in previous years. The lack
of metal sulfide deposits in the substrate suggested
that the sulfate-reducing bacteria were not actively
producing sulfide. Thus, Burleigh Tunnel mine drainage
was tested at the Colorado School of Mines for toxicity to
sulfate-reducing bacteria isolated from the upflow cell.
The tests indicated that the mine drainage is inhibitory to
sulfate-reducing bacteria at low concentrations (dilution)
corresponding to a zinc concentration of 17.5 mg/L.
In addition, zinc sulfate (ZnSO4-7 H2O) was used to
show that the zinc was the toxic constituent (positive
control) in the mine drainage. The zinc sulfate was also
toxic to the sulfate-reducing bacteria at a similar zinc
concentration (18.8 mg/L). The concentration of zinc in
the Burleigh Tunnel mine drainage typically exceeds the
inhibitory level measured in this study. A similar study
conducted using Desulfovibrio desulfricans also found a
zinc concentration of 13 mg/L resulted in inhibition to the
bacteria. (Paulson and others 1997).
Evidence that sulfate reduction was important to the
removal of zinc in the upflow CWS include the large
population of sulfate-reducing bacteria observed when
zinc removal was also high (first year of demonstration),
the accumulation of AVS, primarily zinc sulfide, in the
substrate of this cell, and the decline of sulfate-reducing
bacteria populations after the high flow event that
corresponded with lower zinc removal by the upflow cell.
42
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Visible observations of the upflow cell substrate observed
blackening of the substrate during the first year of operation
suggesting metal sulfides were accumulating, however,
observations of wetland substrate conducted three years
later, showed little blackening of the substrate. These
results suggest sulfate-reduction was not as an important
metal removal mechanism and was occurring to a much
lesser extent during the latter portion of the demonstration.
These observations also suggest that previously formed
metal sulfides are not stable when environmental conditions
within the wetland changes.
3.5 Attainment of Demonstration
Objectives
This section discusses the results of the CWS demonstration
in regard to the attainment of primary and secondary
demonstration objectives. In addition, metal removal
mechanisms, some of the causes for poor performance,
and substrate lifetimes are discussed for each cell.
The results of the demonstration were able to achieve
many but not all of the primary objectives outlined in
Section 3.3. The first primary objective was the
measurement of wetland effectiveness with respect to
cell flow configuration and seasonal variation. This
primary obj active was achieved in part. The demonstration
zinc results indicate zinc removal is greater with an upflow
configured wetland; however, the technology as tested is
not capable of meeting low metal discharge requirements
for extended periods.
The better zinc removal and flow of the mine drainage
through the upflow CWS compared to the downflow
CWS indicate the upflow configuration is superior.
Unfoitunately.itwasnotpossibleduringthis demonstration
to determine the effect of season variation on the
performance of the upflow CWS. The downflow CWS
actually performed better during the winter. The reason
for the improved winter performance is discussed in
Section 3.4.2.
The second primary objective was to determine the
toxicity of the Burleigh Tunnel mine drainage. This
primary objective was achieved. The Burleigh Tunnel
mine drainage is toxic to both the C. dubia and P.
promelas. Measured LC50 values for the P. promelas
(fathead minnows) ranged from 0.62 to 1.6 percent (mine
drainage) and for the C. dubia (water fleas) ranged from
0.10 to 1.0 percent.
The third primary objective was the characterization of
toxicity reduction resulting from CWS treatment. This
primary objective was also achieved. The demonstration
toxicity results indicate the ability of the wetlands to
reduce toxicity to aquatic organisms declined over the first
two years of operation. Further, the high flow event had
a significant impact on toxicity removal in both wetland
cells.
The final primary objective was to estimate the toxicity
reduction to the mine drainage receiving stream (Clear
Creek). This primary objective was not achieved as none
of the demonstration stream samples were toxic to either
test organism.
The most significant primary objective not achieved is the
inability to determine the seasonal variability of the upflow
CWS. During winter, constructed wetlands located in
cold climates may be less effective as a result of lower
microbial activity. This may require pretreatment of the
mine drainage during winter, oversizing the CWS or
retaining a portion of the flow until warmer conditions
return.
The first secondary objective of the demonstration was
to estimate the lifetime of the substrate material. The
lifetime of substrate material is estimated to be 4 to 5
years. The estimate is based on the breakdown of the
substrate material resulting in settling and compaction of
the substrate that leads to flow restrictions. In addition,
demonstration substrate data for nutrients indicate
elements such as phosphate (orthophosphate) have been
depleted in the substrate by this time. If low discharge
limits must be met then demonstration re suits suggest the
substrate lifetime is approximately one year (taking into
account the demonstration starting time and freezing of
the upflow cell during the first year). However, in this
situation it would likely be more cost effective to pretreat
the mine drainage or amend it with an electron donor such
as ethanol to extend the lifetime of the substrate material.
The second, noncritical or secondary objective was to
estimate metal removal by sulfate reducing bacterial.
This evaluation was expected to be qualitative as the
bacteria counts and acid-volatile sulfide analyses are not
highly precise and the metal removal may not be uniform
throughout the treatment cells. As discussed in Section
3.4.2, the downflow cell data did not indicate the primary
metal removal mechanism tobe sulfate reduction. Section
3.4.3 discusses the upflow cell results for sulfate-reducing
bacteria removal of metals. Data indicated an initial high
43
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rate of removal with a longer term reduction in this
mechanism of metals removal.
The third noncritical, secondary objective was to evaluate
the impact of the systems effluent on Clear Creek. These
data are discussed in Section 3.4.4, and indicate that
although the treatment was effective in removing metals
from the Burleigh Tunnel drainage, the relatively small
portion of the discharge being treated did not produce a
measureable decrease in the metals content of Clear
Creek.
The fourth and final noncritical objective was to evaluate
capital operating costs for the CWS. Section 5.0 of this
report provides a detailed economic analysis and
successfully provides data useful for estimating costs for
application of this technology at other sites.
3.6 Design Effectiveness
The following sections discuss the effectiveness of the
upflow and downflow CWS tested during the Burleigh
Tunnel demonstration. The basic design of each wetland
cell is discussed in Section 1.3.2 of this report. This
discussion focuses on general design parameters and
factors that affected each cell.
The basic design of the CWS demonstration system
consisted of a dam inside the Burleigh Tunnel, piping from
the dam to the influent weir, the two wetland cells, an
effluent weir, and a bypass pipe. The dam collected the
mine drainage and provided adequate hydraulic head to
drive the mine drainage through the upflow cell. The
influent weir partitioned the mine drainage to the CWS
cells and channeled the excess water to the bypass piping.
From the influent weir, the mine drainage was channeled
to a ball valve that separated flow to the CWS cells. Water
collected from the cells was piped to the effluent weir and
was discharged to Clear Creek. The purpose of the
effluent weir was to regulate flow through the wetland
cells.
Construction materials associated with this design were
generally inexpensive, readily available, and easily
transported to remote areas. Installation techniques were
also straightforward.
The major drawbacks of this design observed during the
demonstration centered on the flow control valves and
the inability of the effluent weir to regulate flow through
the cells. Because flow through the cells could not be
controlled with the effluent weir, flow through the cells
was regulated at the influent weir and control valve.
Unfortunately, this design meant that any adjustment in
flow to one cell affected flow to the other cell. Future
systems should use easily controlled flow structures such
as weirs to regulate flow to both cells independently.
In addition, the capacity of the initial 4-inch bypass line
was insufficient to accommodate the large water volume
during spring runoff. Eventually, a 6-inch bypass line was
installed. Piping connecting the influent control structure
and the cells should be direct and accessible for routine
cleanout.
A drawback associated with the use of compost substrates
is the high concentration of nitrate in the effluent water
during startup. During this demonstration, no attempt was
made to remove the nitrate from the water prior to
discharge. In a similar wetland evaluation, startup effluents
were applied to surface soils, Alternatively, the startup
effluent could be stored on site in a pond or tank and fed
back into the CWS.
3.6.1 Downflow Cell
The downflow cell consisted of 4 feet of a compost (95 to
96 percent) and hay (4 to 5 percent) substrate. The mine
drainage flowed from the top to a PVC piping collection
network at the base of the cell. The influent and effluent
distribution networks were staggered within the cell to
minimize short-circuiting of the mine drainage in the
substrate.
The design of the downflow cell is discussed in
Section 1.3.2; Figure 2 shows a cross section of the
anaerobic CWS in an upflow configuration. The downflow
configuration is only a reversal of the influent and effluent
flows, not the construction of the cell.
For the most part, the materials used in the construction of
the cells-HDPE liner, geonets, and PVC piping were
acceptable. However, the geofabric was found to fill with
fine material and lose permeability over the 2'/2-year
demonstration. In addition, the cell piping networks did not
include cleanouts. Cleanouts shouldbe included in future
CWS designs. Finally, the influent piping network did not
evenly distribute the mine drainage in this cell. An
additional row of perforated piping in this cell would more
evenly distribute the mine drainage.
The cell was designed to treat 7 gpm. However, during
the demonstration, the downflow cell became less
permeable. The permeability loss is believed to be related
44
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to precipitation of metal oxides, hydroxides, and carbonates,
settling of fine materials in the cell, and compaction of the
substrate material. In winter months, flow through the
downflow cell improved; presumably, the contraction of
frozen substrate allowed water to flow between the liner
and the substrate. However, this short circuiting did not
substantially affect metal removal by the cell.
In an attempt to restore flow through the downflow cell,
air was injected into the substrate to fluff the material.
Although this technique improved flow, the effect was
typically short lived. The results of this demonstration
indicate that substrates with high concentrations of compost
will not retain permeability in a downflow configuration
and are not recommended. However, some recent
downflow wetlands have used substrate mixtures of 50
percent limestone with sawdust and compost to improve
hydraulic characteristics.
3.6.2 Upflow Cell
The design of the upflow C WS is identical to the downflow
cell except that the mine drainage is channeled up though
the compost substrate. Figure 2 shows a cross section of
the demonstration anaerobic compost CWS. The design
of the demonstration wetlands is discussed in Section
1.3.2.
In general, the upflow cell retained permeability throughout
the demonstration. However, some hydraulic restriction
developed during the later half of the demonstration
resulting in a preferential flow pathway. In addition, gas
buildup produced by fermenative bacteria within the
upflow cell may have restricted flow to the effluent lines
in the wetland during the last year of the demonstration.
Gas was released from the cell by periodically puncturing
the upper geofabric with a pitch folk. Replacing the
geofabric with a fine mesh geonet could eliminate gas
buildup. Also, the decline of sulfate-reducing bacteria and
apparent increases in the population of fermentative
bacteria likely exacerbated the problem*
The upflow cell was prone to freezing during winter.
During startup, the dike within the Burleigh Tunnel gave
way, stopping flow to the upflow cell. Flow was restored
by thawing the ice around the effluent line with a steam
cleaner and water tank heater. The following winter, hay
bales were placed over the substrate followed by insulated
blankets (identical to insulated blankets used for curing
concrete), and the system was operational throughout the
winter. However, the straw bales became saturated with
water and the combined weight compressed the substrate
so that all flow ceased through the cell. Flow through
the cell was restored once the hay bales were removed.
During year three, the insulated blankets were used alone
to insulate the cell and there were no interruptions in flow
during this period. In the final year, the ponded water in
the upflow cell was allowed to freeze and did so to a depth
of approximately 6 inches. There were no interruptions in
flow during that winter.
Residence time is an important factor in anaerobic
constructed wetlands that use sulfate-reducing bacteria.
Decreasing residence times may overload the wetland,
exposing the bacteria to inhibitory concentrations of zinc.
Based on the size of the wetlands and substrate water
volumes (percent moisture results of 50 percent) the
calculated residence time for a flow rate of 7 gpm is
48 hours,and67hoursataflowrateof5 gpm. Verification
of residence times was one of the more difficult
measurements undertaken during the demonstration. Both
a chloride tracer (treatability study) and an organic dye
test (demonstration) were unsuccessful in measuring
residence time. The chloride couldnot be readily measured
as background levels of dissolved salts was somewhat
high during the treatability study and the organic dye likely
absorbedto the wetland substrate during this demonstration
test.
During the final year of the demonstration, flow through
the upflow cell began to short circuit in an area adjacent
to the southeastern bermed sidewall. An excavation was
made into the wetland to the influent line feeding this
section of the cell and the line was capped. De watering
the excavation was somewhat difficult and would have
been aided by a sump within the cell. Inspection of the
influent line found precipitates coating the piping walls and
in the piping perforations. The amount of material in the
perforations and the pressure on the piping against the
geofabric would have caused a notable restriction in flow.
Replacing the geofabric with a fine mesh geonet should
alleviate the problem.
45
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Section 4
Data Quality Review
This section presents the summarized results of QA
procedures established to ensure the validity of the zinc
and acute toxicity data collected during the demonstration.
Section 4.1 discusses zinc data quality, and Section 4.2
discusses acute toxicity data quality. A comprehensive
discussion for both zinc and acute toxicity, along with
supporting summary tables, is presented in the Technical
Evaluation Report.
4.1 Zinc Data Quality Review
This section discusses the results of the QA procedures
established to ensure the validity of the zinc data collected
during the demonstration. The QA procedures were
established prior to the demonstration and were recorded
in the quality assurance project plan (Q APP) as part of the
demonstration plan. Both field and analytical QA
procedures were specified to ensure sample integrity and
the generation of data of known quality.
4.1.1 Quality Assurance Results for Field
Sampling Activities
The procedures followed during field activities to maintain
sample integrity and quality are discussed below. They
include specifications for sample collection, labeling,
containerization, preservation, holding times, and chain of
custody.
Sample Containerization, Preservation, and Holding
Times
This section describes sample labeling, shipment, chain-
of-custody, and laboratory receipt procedures for zinc
samples. Conformance with and documentation of these
procedures provide a definitive record of sample integrity
from origin to analysis.
Each sample container was labeled with a unique sample
identification number. The label identified the sampling
location, date, time of collection, and analysis to be
performed. All chain-of-custody forms included the
project number, project name, sampler's name, station
number, date, time, sampling location, number of containers,
and analytical parameters. Samples were hand-delivered
to Quanterra Environmental Services in Arvada, Colorado.
Chain-of-custody forms gathered during the demonstration
were reviewed for content and completeness and appeared
in good order.
All samples analyzed for critical parameters arrived at the
laboratory intact. Several of the coolers used for shipping
the samples arrived with inside temperatures greater than
4 degrees Celsius as specified in the QAPP. However,
the results of associated QA samples suggest that the
elevated temperature did not affect sample integrity. All
samples were analyzed within their designated holding
times (6 months); the majority were analyzed within
1 month of sample collection.
Equipment and Field Blanks
Equipment blanks were collected during the demonstration
to assess sample contamination resulting from sampling
equipment. Throughout the demonstration, dedicated
sampling equipment was used for sample collection to
reduce sample cross contamination. As a result, few
equipment blanks or field blanks were collected during the
demonstration. The data quality objective (DQO) for
equipment and field blanks was results below reporting
limits for all analytes.
Twoequipmentblanks(WEV090794EBandEB012197)
were collected with a polyethelene dipper by pouring
deionized water into the dipper and decanting the water
into an appropriate sample container. The equipment
blank collected in September 1994, contained an estimated
zinc concentration of 0.019 mg/L, which is below the
0.020 mg/L reporting limit. The equipment blank collected
in January 1997, contained 0.052 mg/L zinc, above the
0.020 mg/L reporting limit.
46
-------
Field blanks were used to assess whether zinc
contamination was introduced during the handling,
presentation, or transport of aqueous samples. The field
blank was prepared by adding deionized water into an
appropriate sample container in place of a real sample.
One field blank was collected during the demonstration
(FB060194). Zinc was found in this field blank at a
concentration of 0.034 mg/L, slightly above the reporting
limit of 0.020 mg/L.
The level of contamination in the equipment and field
blanks qualifies data near the reporting limit for accuracy.
The source of the contamination is unknown; however,
the commercial distilled water is suspected. All of the
CWS performance data contained zinc concentrations
at least one order of magnitude greater than the
reporting limit and in most cases two or three orders of
magnitude above the reporting limit. Consequently, the
demonstration zinc data are considered acceptable for
their intended use.
Method Blanks
Method blanks verify that laboratory extraction and sample
cleanup and concentration procedures used do not introduce
contaminants that compromise the analytical results.
Method blanks were prepared and analyzed with each
batch of laboratory analysis. The method blank DQO was
for results to be below reporting limits for all analytes of
interest.
Five out of the 40 batches analyzed during this
demonstration contained reportable quantities of zinc in
the method blanks. Values ranged from 0.020 mg/L to
0.046 mg/L. All samples corresponding to these five
analytical batches were qualified for blank contamination
(B). All of the sample results were greater than five times
the associated blank contamination; thus, no zinc results
were qualified as nondetected due to blank contamination
(UB).
4.1.2 Quality Assurance Results for
Sample Analysis
Analytical QA includes methods and procedures used to
ensure data reliability. Thisprocessinvolvesestablishing
dataquality objectives for the project data and developing
data quality indicators (quanitative or qualitative measures
of precision, accuracy, completeness, representativeness,
and comparability) that can be used to determine whether
the data meet the project's QA objectives.
The QA objective for the CWS demonstration data were
established in the QAPP with specific performance goals
for precision, accuracy, representativeness, completeness,
and comparability. The following sections evaluate the
demonstration data with respect to these performance
goals.
Precision and Accuracy
Precision is a measure of the reproducibility of
measurements under a given set of conditions. Accuracy
is the degree of agreement between an analytical
measurement and the true value. The overall precision for
zinc concentrations was a function of both sampling and
laboratory precision. Overall precision was evaluated
using data from field duplicates, and laboratory precision
was evaluated using data from laboratory duplicates.
Relative percent difference (RPD) between duplicate
samples was used to evaluate precision using the following
formula:
RPD =
- B)
0.5 (A + B)
X100
where: A = first duplicate concentration
B = second duplicate concentration or
Fifteen field duplicate samples were collected during
this demonstration, yielding RPDs ranging from 0 to
3.7 percent. Laboratory duplicate control sampling were
analyzedfor51 rounds of sampling activities. Alllaboratory
RPDs were within the established DQO of 20 percent
with the exception of one, of 28 percent. Overall, the
precision objectives for zinc analyses were achieved.
The accuracy of a measurement is affected by errors
introduced through the sampling process and in handling,
sample matrix, sample preservation, and analytical
techniques. A program of sample spiking at the laboratory
and analysis of standard reference materials (SRMs) was
also used to evaluate laboratory accuracy.
Accuracy for zinc measurements was estimated as percent
recovery (%R) of the true analyte level from SRMs and
by evaluation of matrix spike (MS) recoveries. The
following formula was used to calculate MS percent
recovery:
%R = (S-C)/TX100
47
-------
where; S - measured spike concentration
C = sample concentration
T = true or actual concentration of the spike or
MS spiking recoveries were all within the DQO limits with
one exception. One MS sample analyzed (collected on
July 27,1994) yielded a recovery of 134 percent, slightly
above the DQO. When the data were rechecked by the
laboratory, the deviations were not found to bias the
results sufficiently to affect data use. The laboratory
concluded that the magnitude of the errors was too small
relative to the zinc concentrations to have a significant
effect on the zinc values.
Reported results for the SRM indicate that the analytical
method measured larger concentrations of zinc than
reported in National Institutes of Standards and Testing
(NIST) standard reference material 1643c. The higher
recoveries were considered to be the result of matrix
interferences and the low level of zinc in the SRM. The
DQO for accuracy is 75 to 125 percent recovery. SRM
recoveries were 123 and 149 percent. Quanterra was
immediately notified of the problem, and the laboratory
control samples were checked to confirm that all other
analytical controls were within acceptable parameters.
Tetra Tech determined that some demonstration results
with very low levels of zinc may be positively biased. The
zinc results affected are from the upflow cell effluent
during the first 6 months of operation.
Overall laboratory accuracy for the demonstration data
was acceptable.
Representativeness
Representativeness expresses the degree to which sample
data accurately and precisely represent the characteristics
of a population, parameter variations at a sampling point,
or an environmental condition they are intended to
represent. For the CWS demonstration, the low RPDs
associated with field duplicate results suggest the data
collected are representative of the CWS system for the
environmental and physical conditions at the Burleigh
Tunnel site.
Completeness
Completeness is a measure of the amount of acceptable
data obtained compared to the amount of data needed
to achieve a particular level of confidence in the results.
Acceptable data are obtained when (1) samples are
collected and analyzed in accordance with the
QC procedures outlined in the demonstration plan, and
(2) criteria that affect data quality are not exceeded.
CWS percent project completeness (%C) was calculated
using the following equation;
%C = (V/T) X 100
where: %C = percent completeness
V = number of measurements judged
acceptable
T = total number of measurements planned
The QA objective for degree of completeness was
90 percent for the critical parameter zinc. All data
collected are considered usable for the intended purpose;
therefore, the QA objective for completeness was
achieved.
Comparability
The comparability parameter is designed to identify
deviations in the data that may result from inconsistencies
in field conditions, sampling methods, or laboratory analysis.
During this demonstration, changes in sampling techniques
and laboratory analysis were minimized to ensure
comparability of results. However, the end of the first
SITE contract and delays in restarting the new SITE
contract required the use of data collected by CDPHE.
The results of a laboratory intercalibration exercise with
Quanterra, the CDPHE laboratory (Analytica), and a
referee laboratory suggest that the data are comparable.
4.2 Acute Toxlclty Data Quality Review
This section discusses the results of QA data collected to
document the validity of the acute toxicity data. The QA
procedures were established prior to the demonstration
and recorded in the QAPP as part of the demonstration
plan. Both field and analytical QA procedures were
specified to ensure sample integrity and the generation of
data of known quality.
4.2.1 Analytical Quality Assurance
Analytical QA is the process of ensuring and confirming
data reliability. This process includes establishing
DQOs for the project data and developing data quality
indicators (quantitative or qualitative measures of precision,
accuracy, completeness, representativeness, and
comparability) that can be used to evaluate whether the
data met the project's QA objectives. The QA objectives
for acute toxicity testing during the CWS demonstration
48
-------
were established in the QAPP and are summarized in the
following discussions.
Water Chemistry Results for Environmental
Samples and Reference Toxicant Tests
To ensure that laboratory water quality conditions did not
adversely affect the reference toxicant or environmental
sample results, water quality parameters were documented
throughout all test series. The water chemistry results
indicate that the water quality conditions for testing were
appropriate for the test organisms during all test dates and
that no abnormal water conditions were documented that
could influence the survivability results.
Precision and Accuracy
Precision and accuracy in toxicity tests are controlled and
evaluated through documentation of reference toxicant
responses of indicator species against inter- and intra-
laboratory historical records; and by carefully controlling
and documenting the environmental conditions tested.
The following discussion documents the laboratory testing
conditions for growth, feeding, and maintenance of indicator
species during the tests; and documents the results of
indicator species survivability results against laboratory
historical records for identical tests.
Acute toxicity and metal concentration in the mine drainage
were used to infer a response relationship between the
mostprevalenttoxiccomponentpresent(zinc)and indicator
species survival. Preliminary chemical analysis had
identified zinc in various forms as the most predominant
metal contaminant.
Zinc sulfate was used as a reference toxicant to simulate
the population response of the indicator species to a
soluble zinc compound present in the mine drainage
matrix. Potassium chloride was used as a laboratory
reference test for population viability and toxic response
of the indicator species.
Pimephales promelus and Ceriodaphnia dubia were used
as the test organism populations in the 48-hour static-
renewal acute toxicity tests. Indicator species survival
rates (LC50) at the 95 percent confidence level (EPA
1993a) in a static series of potassium chloride and zinc
sulfate concentration dilutions were calculated and
compared with laboratory historical records. The
comparison provided a control on the viability of the test
species and the testing methodology.
The quantitative precision and accuracy requirements for
acute toxicity for Pimephales promelus and Ceriodaphnia
dubia when exposed to zinc sulfate were established by
toxicant equivalent concentration values generated from
both external and internal laboratory records of earlier
tests. The quantitative precision and accuracy objectives
for acute toxicity for Pimephales promelus and
Ceriodaphnia dubia when exposed to potassium chloride
were established bymonthly cumulative laboratory toxicant
equivalent concentration values.
All reference toxicant results fell within the prescribed
ranges, indicating that the response of the indicator
species response to test conditions was appropriate for
evaluating the toxin present. Therefore, the quantitative
results of acute toxicity to the environmental samples are
comparable to other tests under identical conditions.
Sample Duplicates
The results of sample (field) duplicates is another indicator
of overall precision. The sample duplicate was collected
on February 27, 1995 from the treated effluent from the
downflow cell (samples designated WED and WED II).
Generally, the analysis of duplicate acute toxicity values
for sampling and analytical precision is a numerical
comparison of the difference in reported acute toxicity
values to the magnitude of the values themselves.
However, sample WED for February 27, 1995 was not
toxic enough to generate an LC50 value, which is the
normal endpoint for acute toxicity analysis. Consequently,
the analysis of test sampling and analytical precision
presented is a subjective comparison of the sample and
duplicate routine chemistry and intermediate toxicity
results.
The chemistry for duplicate samples WED and WEDII
shows no significant difference, with less than 10 percent
variation in all measured parameters. Those variables
having the greatest difference - inpH, DO, and temperature
- were consistently lower for WEDII than for WED. The
values, however, do not strongly indicate a difference in
water quality conditions. The initial and final chemistry for
both species tests also show slight differences, but no
consistent variability in an individual parameter.
Qualitatively, the survival rates for C. dubia of the individual
sample dilutions for duplicate samples WED and WEDII
both show very slight toxicity, especially noting that both
controls had survival rates of 20/20. Quantitatively, the
100 percent WEDII sample yields a survival ratio
49
-------
Section 5
Economic Analysis
This section presents cost estimates for using an anaerobic
compost CWS system to treat mine drainage with water
chemistry similar to the Burleigh Tunnel. The baseline
scenario used for developing this cost estimate was a 50
gpm flowrate, the total flow from the Burleigh Tunnel, and
a 15-year system life, The baseline costs were then
adjusted for flowrates of 25 gpm and 100 gpm to develop
cost estimates for other cases.
Cost estimates presented in this section are based primarily
on data compiled during the SITE demonstration at the
Burleigh Tunnel (CDPHE 1995). Additional cost data
were obtained from standard engineering cost reference
manuals (Means 1992). Costs have been assigned to
11 categories applicable to typical cleanup activities at
Superrund and RCRA sites (Evans 1990). Costs are
presented in year 1995 dollars and are considered estimates,
with an accuracy ofplus 50 percent and minusBO percent.
5.1 Basis of Economic Analysis
A number of factors affect the costs of treating mine
drainage with an anaerobic compost CWS system. These
factors generally include flow rate, type and concentration
of contaminants, physical site conditions, geographical
site location, and treatment goals. The characteristics of
spent substrate produced by a CWS system will also
affect disposal costs. Spent substrate will require off-site
disposal. Mine drainage containing cadmium at 0.05 parts
per million (ppm), iron at 50 ppm, nickel at 0.5 ppm,
and zincatSO ppm was selected forthis economic analysis.
The following presents additional assumptions and
conditions as they apply to each case.
For each case, this analysis assumes that an upflow CWS
system will treat contaminated mine drainage continuously,
24 hours per day, 7 days per week. An average metals
removal efficiency of 96 percent was assumed for all
cases. Based on these assumptions, the CWS system will
treat about 26.3 million gallons of water per year of
operation at the baseline flowrate of 50 gpm.
• Further assumptions about constructed wetlands
treatment for each case include the following:
• A residence time of 75 to 150 hours is recommended
for adequate metals removal.
• A porosity of 50 percent is assumed for the substrate
material.
• Two baseline wetlands, size of 90 feet by 90 feet by
4 feet (2,300 cubic yards [yd3]), will provide a 78
hour residence time at a flowrate of 50 gpm (wetland
size is directly proportional to flowrate). Square
wetlands were used for the cost estimation; however,
other shapes may be preferable.
• Substrate material will require removal and
replacement once every 5 years.
• The spent substrate is not a RCRA hazardous waste:
thus, it will be dewatered on site and can be recycled
or disposed of at an industrial landfill.
• An aerobic polishing pond to increase displaced
oxygen is not required.
This analysis assumes that aquatic-based standards are
most appropriate; and the attainment of these standards
depends on the affected organisms, receiving waters and
volume of mine drainage. Attainment may not be feasible
in all cases for the technology as tested during this
demonstration.
The following assumptions were also made for each case
in this analysis:
• The site is located within 200 miles of the disposal
location.
• The site is located within 100 miles of a moderate-
sized city.
52
-------
• The site will allow for gravity flow of the mine
drainage through the wetland.
• A staging area is available for dewatering spent
substrate.
• Access roads exist at the site.
* Utilities, such as electricity and telephone lines, are
available on site.
* The treatment goal for the site will be to reduce zinc
contaminant levels by 90 percent.
* Spent substrate will be dewatered and disposed of
off site.
• One influent water sample and two effluent water
samples will be collected monthly and two composite
substrate samples will be collected quarterly to
monitor system performance.
• One part-time operator will be required to inspect
the system, collect all required samples, and conduct
minor maintenance and repairs.
5.2 Cost Categories
Cost data associated with the CWS technology have been
assigned to one of the following 11 categories: (1) site
preparation; (2) permitting and regulatory requirements;
(3) capital equipment and construction; (4) startup;
(5) labor; (6) consumables and supplies; (7) utilities;
(8) residual and waste shipping andhandling; (9) analytical
services; (10) maintenance and modifications; and
(11) demobilization. Costs associated with each category
are presented in the sections that follow. Some sections
end with a summary of significant costs within the category.
Table 9 presents the cost breakdown for the flow variant
cases. This table also presents total one-time, fixed costs,
and total variable O&M costs; the total project costs; and
the costs per gallon of water treated.
5.2.7 Site Preparation Costs
Site preparation includes administration, pilot-scale testing,
mobilization costs. This analysis assumes a total area of
about 65 acres will be needed to accommodate the
wetland and staging area, construction equipment, and
sampling and maintenance equipment storage areas. A
solid gravel (or ground) surface is preferred for any
remote treatment project. Pavement is not necessary, but
the surface must be able to support construction equipment.
This analysis assumes adequate surface areas exist at the
site and that only moderate modifications will be required
for wetland construction.
Administrative costs, such as legal searches and access
rights, are estimated to be an additional $ 10,000.
Mobilization involves transporting all construction
equipment and materials to the site. For this analysis, it is
assumed that the site is located within 100 miles of a city
where construction equipment is available. The total
estimatedmobilization cost will be $5,000.
For each case, total site preparation costs are estimated
to be $15,000.
5.2.2 Permitting and Regulatory
Requirements
Permitting and regulatory costs vary depending on whether
treatment occurs at a Superfund site and on the disposal
method selected for treated effluent and any solid wastes
generated. At Superfund sites, remedial actions must be
consistent with ARARs, environmental laws, ordinances,
and regulations, including federal, state, and local standards
and criteria. In general, ARARs must be identified on a
site-specific basis. At an active mining site, a NPDES
permit will likely be required and may require additional
monitoring records and sampling protocols, which can
increase permitting andregulatory costs. For this analysis,
total permitting and regulatory costs are estimated to be
$5,000.
5.2.3 Capital Equipment
Capital costs include all wetland construction and
construction materials and a site building for housing
sampling, monitoring, and maintenance equipment.
Construction materials include sand, synthetic liners,
geotextile liners, PVC piping, valves, concrete vaults or
sumps, weirs, and other miscellaneous materials. Capital
costs for the baseline wetland of 50 gpm are presented
below. Site preparation and excavation include clearing
the site of brush and trees, excavation of the wetland cell,
grading the cell, and construction of the earthen berms.
The total cost of site preparation and excavation is $ 19,500
for the 50 gpm system.
Construction of the wetland cell itself involves system
design, subgrade preparation and installation of a sand
layer, liner, piping distribution and collection systems, and
the substrate. Also included is piping to and from the cell
as well as system bypass piping, and concrete sumps with
weirs at the influent of the wetland to control flow through
53
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Table 9. CWS Costs for Different Treatment Flow Rates*
Cost Categories
25gpm
System Life 15 Years
SQgpm
100 gpm
Fixed Costs
Site Preparation
Administrative
Mobilization
Permitting and Regulatory
Requirements
Capital Equipment
System Design
Excavation and Site
Preparation
Wetland Cell Construction
Piping and Valves
Storage Building
Startup
Demobifization
Excavation and Backfilling
Substrate Disposal
Total Fixed Costs
Variable Costs
Labor
Operations Staff
Consumables and Supplies
Personal Protective
Equipment
Utilities
Residual and Waste Shipping and
Handling
Substrate Disposal
Analytical Services
Maintenance and Modifications
Annual Maintenance
Substrate Removal and
Replacement
Total Variable Costs
Total Costs
Total Cost Per Gallon Treated
$15,000
$10,000
5,000
$5,000
$215,300
$50,000
9,800
120,000
25,500
10,000
$1,500
$52,250
$10,000
42,250
$316,000
$153,000
$39,000
$153,000
$39,000
NA
$120,000
40,000(3)
$360,000
$247,550
$5,000
80,850(3}
$919,550
$1,235,500
$0.0063
$15,000
$10,000
5,000
$5,000
$345,000
$50,000
19,500
240,000
25,500
10,000
$1,500
$104,500
$20,000
84,500
$492,000
$153,000
$39,000
$153,000
$39,000
NA
$240,000
80,000 (3)
$360,000
$490,100
$5,000
161,700(3}
$1,282,100
$1,774,100
$0.0045
$15,000
$10,000
5,000
$5,000
$604,500
$50,000
39,000
480,000
25,500
10,000
$1,500
$209,000
$40,000
169,000
$844,000
$153,000
$39,000
$153,000
$39,000
NA
$480,000
160,000(3}
$360,000
$975,200
$5,000
323,400 (3)
$2,007,200
$2,851,200
$0.0036
*Costs are based on July 1995 dollars, rounded to the nearest $100.
Substrate removal and replacement estimated to be necessary every 5 years.
(3) Number of removais anticipated
NA Not applicable
54
-------
the system. The total cost for wetland cell construction of
a 50 gpm system is $335.000.
A small building is required for storing sampling equipment
and providing work space for the system operator. The
cost for a simple building with electricity has been estimated
at $10,000.
The total capital cost for a 50 gpm wetland system is
$345,000.
5.2.4 Startup
Startup requirements are minimal for a wetland system.
System startup involves introducing flow to the wetland
with frequent inspections to verity proper hydraulic
operation. Operators are assumed to be trained in health
and safety procedures. Therefore, training costs are not
incurred as a direct startup cost. The only costs directly
related to system startup are labor costs associated with
more frequent system inspection. Startup costs are
estimated at $1,500.
5.2.5 Labor-
Labor costs include a part-time technician to sample,
operate, and maintain the system. Once the system is
functioning, it is assumed to operate continuously at the
design flow rate. One technician will monitor the system
on a weekly basis. Weekly monitoring will require several
hours 2 to 3 times per week to check flowrate and overall
system operation. Sampling is assumed to be conducted
once a month and will require two technicians for 2 hours.
These requirements equate to 175 hours annually for
general O&M. An additional 80 hours of labor are
included for miscellaneous O&M and review of data.
Based on $40 per hour for a technician, the annual cost for
general labor O&M is $10,200.
5.2.6 Consumables and Supplies
The only consumables and supplies used during wetland
operations are disposable PPE. Disposable PPE includes
Tyvek coveralls, gloves, and bootcovers. The treatment
system operator will wear PPE when required by health
and safety plans during system operation. PPE will cost
about $25 per day per person on site. Based on the
assumed labor required above and an additional 22 days
for miscellaneous O&M, PPE will be required 100 days
annually, for an annual PPE cost of about $2,500.
5.2.7 Utilities
Utilities used by the wetland system are negligible. The
wetland system requires no utilities for operation. The
only utility required is for electricity for lights in the on-site
storage building and for charging monitoring equipment.
For this analysis, utility costs are assumed to be zero.
5.2.8 Residual Waste Shipping and
Handling
The residual waste for the wetland is assumed to be spent
substrate. This analysis assumes that substrate will
require removal and replacement once every 5 years. It
is assumed that spent substrate will be dewatered on site
and disposed of at a recycler or landfill. Substrate removal
and replacement costs are covered in Section 5.2.11,
maintenance and modifications. Loading dewatered
substrate into 20 yd3 haul trucks is estimated to cost
$14,500. Hauling the substrate to a recycler or landfill
is estimated to cost $28,000; disposal of substrate at
the landfill costs $42,000. Oversightof substrate removal,
hauling and replacement is expected to cost $3,200 (10 8-
hour days at $40/hr). Loading of the new substrate is
expected to cost $12,000 and the cost of the substrate is
$65,200. The total waste shipping and handling cost per
substrate replacement is $161,700. Costs for residual
waste shipping and handling are based solely on substrate
volume. Costs for different sized wetlands are proportional
to the 50 gpm baseline system described here.
5.2.9 Analytical Services
Analytical costs associated with a wetlands system include
laboratory analysis, data reduction and tabulation, QA/
QC, and reporting. For each case, this analysis assumes
that one influent sample and two effluent samples will be
collected once a month and that two substrate samples
will be collected quarterly. The substrate samples will be
analyzed for total metals. Influent and effluent samples
will be analyzed for total metals, ammonia, nitrate,
phosphate, BOD, TSS, and TDS. Monthly laboratory
analysis will cost about $1,050, and substrate analysis
$3,500 per year. Data reduction, tabulation, QA/QC, and
reporting are estimated to cost about $660 per month.
Total annual analytical services for each case are estimated
to cost about $24,000 per year.
5.2.10 Maintenance and Modifications
Annual repair and maintenance costs are expected to be
minimal and forthis analysis are assumed to be $5,000 for
each case. No modification costs are assumed to be
55
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incurred. The major maintenance cost will be removal
and replacement of the substrate every 5 years. Excavation
of substrate material has been estimated to cost $14,500
for the 50 gpm scenario. Replacement of the distribution
and collection piping was estimated to cost $14,300.
Purchase and transport of new substrate was estimated
to cost $65,400. The total estimated cost of substrate
removal and replacement is $ 161,700. The removal and
replacement cost will vary proportionally with the wetland
size.
5.2.11 Demobilization
Site demobilization costs include excavation of the substrate
and concrete vaults and weirs, disposal of substrate, and
backfilling the wetland. For the 50 gpm scenario, excavation
costs are estimated at $10,000. Substrate disposal costs
are $80,000. Backfilling of the wetland is expected to cost
$ 10,000, assumingnative material fromthe original wetland
excavation was left on site. The total demobilization cost
is estimated to be $104,500. This cost will vary
proportionally with wetland size.
56
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Section 6
Technology Status
Currently, several hundred constructed and natural
wetlands are treating coal mine drainage in the eastern
United States. The effectiveness of these systems is
discussed in several publications including Hammer 1989,
Moshiri 1993, and the proceedings of .annual meetings of
the American Society for Surface Mining and Reclamation,
and several U.S. Bureau of Mines papers (U.S. Bureau
of Mines Special Publication SP066-4 and Information
Circular 1C 9389) (see Appendix B).
In addition, any constructed wetlands designed to treat
metal mine drainages have been constructed and tested or
are being tested by EPA, various state agencies, and
industry. In Colorado, the state Division of Minerals has
constructed several wetland systems to treat metal mine
drainage. Constructed wetlands treatment is also being
considered for the full-scale remedy of the Burleigh
Tunnel drainage.
57
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Section 7
References
Camp, Dresser, and McKee (COM). 1993. Clear Creek
Remedial Design Passive Treatment at Burleigh
Tunnel, Draft Preliminary Design at Burleigh Tunnel.
June.
Colorado Department of Public Health and Environment
(CDPHE). 1995. Facsimile Communication with
Garry Farmer, Tetra Tech. February, 1995.
Correns, C.W. 1969. Introduction to Mineralogy.
Springer-Verlag. New York. Berlin.
Environmental Restoration Unit Cost Book (ECHOS).
1995. ECHOS, Los Angeles, California.
Evans, G. 1990. Estimating Innovative Technology Costs
for the SITE program. Journal of Air and Waste
Management Association. 40:7:1047-1051.
Gusek, J.J., and Wildeman, Dr. T. R.. 1995. New
Developments in Passive Treatment of Acid Rock
Drainage. Paper presented at Engineering Foundation
Conference on Technological Solutions for Pollution
Prevention in the Mining and Mineral Processing
Industries, Palm Coast Florida, January 23,1995.
Gusek, J.J., J.T. Gormley, and J.W. Sheetz. 1994.
Design and construction aspects of pilot-scale passive
treatment systems for acid rock drainage at metal
mines. Proc. Society of Chemical Industry
Symposium. Chapman and Hall, London.
Hammer, D.A. 1989. Constructed Wetlands
for Wastewater Treatment. Lewis Publishers.
Chelsea, Michigan.
Hedin, R.S., R.W. Narin, and R.L.P. Kleinmann. 1994.
Passive Treatment of Coal Mine Drainage. United
States Bureau of Mines Information Circular 9389.
Klusman, R.W. 1993. Computer Code to Model
Constructed Wetlands for Aid in Engineering Design.
Repoit to United States Bureau of Mines, Contract
J0219003.
Means, R.S. 1992. Means Building Construction
Cost Data. Construction Consultants and Publishers,
Kingston, Massachusetts.
Metcalf and Eddy, Inc. 1979. Wastewater Engineering
Treatment, Disposal, and Reuse. Revised by George
Tchobanoglous and Franklin L. Burton. McGraw-
Hill Publishing Company. New York, New York.
Moshiri, G.A. 1993. Constructed Wetlands for
Water Quality Improvement. Lewis Publishers.
Boca Raton, Florida.
PRC Environmental Management, Inc. (PRC) 1993.
Colorado Department of Public Health and
Environment, Constructed Wetlands System
Treatability Study at the Burleigh Tunnel, Silver
Plume, Colorado, Treatability Study Work Plan,
Denver, Colorado, February 1993.
PRC. 1995. Colorado Department of Public Health and
Environment Constructed Wetlands System
Demonstration Plan. July.
Reynolds, J,S. 1991. Determination of the Rate of
Sulfide Production by Sulfate-reducing Bacteria at
the Big 5 Wetland. Masters Thesis. Colorado
School of Mines, Golden, Colorado.
U.S. Bureau of Mines. 1994b. Proceedings of the
International Land Reclamation and Mine Drainage
Conference and Third International Conference on
the Abatement of Acidic Drainage. Pittsburgh,
Pennsylvania, April 24-29, 1994, Bureau of Mines
Special Publication SP 066-4.
U.S. Environmental Protection Agency (EPA). 1988.
Constructed Wetlands and Aquatic Plant Systems
for Municipal Wastewater Office of Research and
Development. Washington, D.C. EPA/625/1-88/
022. September.
EPA. 1993a, Methods for Measuring the
Acute Toxicology of Effluents and Receiving Waters
to Freshwater and Marine Organisms. Office of
Research and Development. Washington, D.C
EPA/600/4-90/027F. 4th Edition. September.
EPA. 1993b. Handbook for Constructed
Wetlands Receiving Acid Mine Drainage. Office of
Research and Development. Cincinnati, OH.
September.
58
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Appendix A
Analytical Results Summary Tables
59
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Table A-l. Influent Results
INFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUSNITRITEASN
NIT RITE AS N
NIT RATE AS N
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaC03
ALKALINITY, BICARB
AS CAC03
DISSOLVED OXYGEN (mg/L)
pH
CONDUCTIVITY (jjS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354,1
350.1
160.2
160.1
9060
310.1
310.1
-
--
--
-
WI030994
03/09/94
mg/L
ND
ND
0.10
84.8
0.31
0.014
41.8
2.3
0.045
2,6
0.0011
10.3
55.0
386
NA
1.0
19.9
ND
ND
ND
ND
ND
ND
16.8
732
1.1
100
100
8.1
7.4
730
6.9
WM32394
03/23/94
mg/L
ND
0.0041
0.099
88,0
0.33
0.015
43.1
2.4
0.039
2.9
0.00012
9.3
56.1
374
NA
1.2
21.8
ND
0.30
ND
ND
ND
ND
8.8
655
NA
107
107
8.3
7.5
745
7.3
W 1041)694
04/06/94
mg/L
ND
0,0068
0.10
91.7
0,33
0.014
44.2
2,5
0.042
3.0
0.000066
10.9
60.1
387
NA
I.I
22.3
ND
ND
0.060
ND
0.060
ND
20,4
640
NA
105
105
6.8"
7.5
745
7.3
W 1042094
04/20/94
mg/L
ND
0.020
0.10
96.9
0,34
0.016
46,5
2.6
0,047
3.1
0,000070
9,1
64.0
384
NA
I.I
21,9
ND
ND
0.11
ND
0.11
ND
15,2
663
ND
107
107
W 1050594
05/05/94
mg/L
ND
0.060
0.098
89.9
0.32
0.016
47.1
2.3
0.043
3.6
0.000098
14.0
56.1
317
NA
0,98
19.0
ND
ND
ND
ND
ND
ND
7.4
641
NA
104
104
NA
7.4
699
8.9
W 105 1994
05/19/94
mg/L
0.045
0.052
0.081
83.2
0.21
0,014
49.1
1.8
0.035
3,2
0.00019
10.5
44.8
314
NA
1.0
15.0
ND
0.40
ND
ND
ND
ND
8.4
622
NA
107
107
NA
7.5
698
9.4
•-•Not applicable
fjS - MicroSiemens
mg/L - Milligrams per liter
NA *• Not analyzed
ND-Not detected
60
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANALY1E
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE ASN
NITRITE ASN
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCOS
ALKALINITY, BICARB
ASCAC03
DISSOLVED OXYffiN (mg/L)
pH
CONDUCT I VITY(nS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310,1
..
--
--
--
WI060194
06/01/94
mg/L
ND
ND
0.092
89.6
0.25
0.020
50.6
1.9
0.033
3.6
0.00019
13.2
49.1
357
NA
1.0
16.9
ND
ND
ND
ND
ND
ND
4.4
657
NA
109
109
8.7
7.6
775
9.4
WIO 62994
06/29/94
mg/L
0.068
ND
0.089
86.1
0,23
0.017
45.4
2.1
0.045
3.0
ND
12.8
54.2
378
NA
1.0
17.9
ND
0.44
ND
ND
ND
ND
11.2
680
NA
107
107
NA
7.57
980
9.5
WI071394
07/13/94
mg/L
ND
ND
0.086
94.5
0.23
0.013
48.3
2.2
0.044
3.1
0.00013
13.00
56.8
377
NA
0.90
17.5
ND
ND
ND
ND
ND
ND
9.2
685
NA
109
109
8.2
7.5
950
9.4
WI072894
07/28/94
mg/L
ND
ND
0.098
91.2
0.30
0.017
46.4
2.2
0.043
2.9
0.00015
12.0
59.1
397
NA
1.1
18.7
ND
0.077
2.0
ND
2.0
ND
9.6
707
NA
103
103
NA
NA
927
9.5
WI081594
08/15/94
mg/L
ND
ND
0.10
92.5
0.24
0.016
47.7
2.3
0.042
2.9
0.00017
14.4
54.7
374
NA
1.1
18.6
ND
ND
1.7
ND
1.7
ND
2.4
759
NA
105
105
NA
7.5
948
9.4
WI082494
08/24/94
mg/L
ND
ND
0.0952
94.6
0.25
0.014
48.1
2.4
0.046
3.2
ND
15.3
57.5
403
NA
1.1
19.6
ND
ND
1.9
ND
1.9
ND
18.4
703
NA
102
102
7.6
7.4
920
9.4
** = Degrees Farenheit
-- = Not applicable
jjS = microSiemens
mg/L= Milligrams per liter
NA = Not analyzed
ND = Not detected
61
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANALVTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS;
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUSNITRITE AS N
NIT RITE ASN
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
DISSOLVED OXYGEN (mgO.)
PH
CONDUCTIVITY OiS)
TEMPERATURE (degrees Q
ANALYTICAL
MEIHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354,1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
--
--
--
--
W 1090794
09/07/94
mg/L
ND
ND
0.098
90.2
0.29
0.017
46.5
2,3
0.047
3.9
0.00040*
12.1
56.4
416
NA
1.0
20-2
ND
ND
ND
ND
ND
ND
17.6
711
NA
102
102
9.5
7.41
922
9.3
WI091994
09/19/94
mg/L
ND
ND
0.085
89.7
0.29
0.015
46.6
2.3
0.042
3.1
0.00041
12.5
57.6
404
NA
1.0
19.6
ND
ND
ND
ND
ND
ND
8.4
723
NA
101
101
7.8
7.4
930
9,3
WJ100494
10/04/94
mg/L
ND
ND
0.089
92.6
0.31
0,014
47.3
2.3
0.052
3.0
0.00050
11.6
59.7
400
NA
1.0
19.8
ND
ND
ND
ND
ND
ND
18.8
695
NA
112
112
NA
7.4
935
9.1
WII01994
10/19/1994
mg/L
ND
ND
.10
92.4
0.25
0.014
46,7
2.4
0.046
3.0
ND
13
57.6
409
NA
ND
19.5
ND
ND
ND
ND
ND
ND
18.8
695
NA
102
102
NA
7.1
750
8.5
VV 11 10294
11/02/94
mg/L
0.030
ND
0.10
89.2
0.28
0.014
46.2
2,2
0.051
2.9
ND
14.8
56.5
410
NA
1.0
20.1
ND
0.13
ND
ND
ND
ND
8.0
709
NA
82.4
82.4
NA
6.9
900
8.7
WI1 12094
11/20/94
mg/L
ND
ND
0.091
93.5
0,32
0.016
47.3
2.3
0.050
3,1
0.00030
14.4
58.2
407
NA
1.1
21.3
ND
ND
ND
ND
ND
ND
18.0
711
NA
101
101
NA
6.9
NA
8.1
-- = Not applicable
^8= MicroSiemcns
mg/L = Milligrams per liter
NA = Not detected
ND = Not detected
62
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORT HOP HOSP HATE
NITRATE PLUS NITRITE ASN
NITRITE AS N
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaC03
ALKALINITY, BICARB
AS CAC03
DISSOLVED OXYGEN (mg/L)
pH
CONDUCT I VlTY(nS)
TEMPERATURE (degrees C)
ANALYTICAL
MEIHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353,2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
-
--
-
--
WI113094
11/30/94
mg/L
ND
ND
0.086
95.4
0.34
0.014
47.7
2.5
0.044
2.8
0.00036
14.2
62.8
411
NA
1.1
21.4
ND
0.13
ND
ND
ND
ND
16.4
711
NA
99.6
99.6
NA
6.9
605
7.9
WI121494
12/14/94
mg/L
0.036
ND
0.092
98.1
0.37
0.018
48.9
2.5
0.050
3.3
ND
19.5
63.0
413
NA
1.0
21.2
ND
0.36
ND
ND
ND
ND
10.4
687
NA
103
103
8.0
7.54
600
8.0
W 10 10495
01/04/95
mg/L
0.032
ND
0.82
87.7
0.31
0.016
46.5
2.3
0.048
2.9
0.00037
15.0
55.5
395
NA
1.1
21.6
ND
ND
ND
ND
ND
ND
5.2
689
NA
104
104
8.5
7.5
610
6.5
WI011895
01/18/95
mg/L
0.038
ND
0.076
90.8
ND
0.015
45.4
2.4
0.046
3.0
0.00021
15.9
57.1
386
NA
1.1
21.7
ND
ND
ND
ND
ND
ND
12.0
693
NA
106
106
7.3
7.5
600
9.0
WI020195
02/01/95
mg/L
0,047
ND
0.089
90.1
0.34
0.016
44.1
2.4
0.052
2.8
ND
14.1
56.6
402
NA
1.1
22.5
ND
ND
1.7
ND
1.7
ND
12.8
694
NA
106
106
7.6
7.9
610
7.9
WI02I595
02/15/95
mg/L
0.043
ND
0.084
100.0
0.39
0.015
49.4
2.7
0.048
3.5
ND
20.4
58.9
390
NA
I.I
22.8
ND
0.10
ND
ND
ND
ND
12,8
656
NA
106
106
NA
7.0
NA
8.1
* = Dissolved metals
- =Not applicable
US = Microsiemens
mj^L = Milligrams per lilcr
NA = Not analyzed
ND = Not detected
63
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL,
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NIT RATE PLUS NITRITE ASM
NIT RITE AS N
NIT RATE AS N
AMMONIA
TOTAL SOLIDS;
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaC03
ALKALINITY, BICARB
ASCAC03
DISSOLVED OXYGEN (mg/L)
PH
CONDUCT IVITY(jiS)
TEMPERATURE (degrees C)
ANALYTICAL
MErraoo
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300,0
376.2
340.2
300.0
365,3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
-
--
--
-
WI022795
02/27/95
mg/L
0.024
ND
0.071
92.6
0.33
0,014
45.1
2.5
0,068
2.9
ND
16.2
58,6
384,0
NA
l.l
22.6
ND
ND
ND
ND
ND
ND
11.2
692
NA
107
107
7.8
7.4
630
8.6
W 103 1595
03/15/95
mg/L
0.049
ND
0.076
91.4
0.36
0.016
44.4
2.5
0.045
2.9
ND
15.8
57.0
384.0
NA
1.1
22.4
ND
ND
ND
ND
ND
ND
9.2
672
NA
104
104
NA
7.5
620
9.3
WI032995
03/29/95
mg/L
ND
ND
0.074
85.2
0,33
0.014
41.9
2.3
0.045
2.8
ND
16.4
53.1
368,0
NA
1,0
23.1
ND
ND
ND
ND
ND
ND
12.8
655
NA
107
107
7,5
7,7
600
8,1
WI041295
04/12/95
mg/L
ND
ND
0.057
90.9
0,32
0.015
42.9
2.4
0.048
3.0
ND
16.1
55.0
376.0
NA
1.0
22.4
ND
0,11
ND
ND
ND
ND
14.4
656
NA
107
107
8.6
7.5
620
8.4
WI042695
04/26/95
mg/L
0.060
ND
0.095
88.2
0.41
0.022
41.2
2.6
0,071
2.9
ND
14,2
55.7
370.0
NA
1.1
23.8
ND
ND
0.14
ND
0.14
ND
7,2
575
NA
104
104
7.5
NA
600
9.0
WI051095
05/10/95
mg/L
0.15
ND
0.095
92.0
0.48
0.026
41.9
3.0
0.054
3.1
ND
14.8
61.4
374
NA
1.1
20.5
ND
ND
ND
ND
ND
ND
2.8
689
ND
103
103
* = Dissolved metals
-• -Not applicable
nS"Microsiemens
mg/L = Milligrams per liter
NA - Not analyzed
ND - Not detected
64
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE ASN
NITRITE ASN
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL;
ASCaCOS
ALKALINITY, BICARB
AS CAC03
DISSOLVED OXYGEN (mg/L)
PH
CONDUCTIVITY (US)
TEMPERATURE (degrees Q
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376,2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160,1
9060
310.1
310.1
--
--
--
--
W 106 1295
6/12/1995
mg/L
0.065
ND
0.25
94.4
0.12
0.058
58.3
3.9
0.061
4.1
ND
9.9
75.5
499
0.8
6.9
ND
ND
0.13
ND
0.13
ND
20.4
838
120
120
NA
7.4
NA
10.2
WI06289S
6/28/1995
mg/L
ND
ND
0.26
111
0.11
0.051
61.4
4.4
0.073
ND
ND
14.2
86.8
502
0.89
ND
ND
0.10
ND
ND
ND
20.4
967
125
125
7.1
7.2
700
10.3
W 1071 095
7/1 0/1 995
mg/L
ND
ND
0.25
119
0.10
0.050
64.0
5.0
0.081
3.6
ND
14.8
99.8
582
0.96
8.8
ND
ND
ND
ND
ND
ND
24.8
1010
118
118
NA .
7.4
NA
10.3
WI072695
7/26/1995
mg/L
ND
ND
0.24
129
ND
0.038
64.2
5.5
0.084
3.7
ND
13.2
105
596
0.88
10.2
ND
ND
0.63
ND
0.63
ND
22.4
999
107
107
NA
NA
NA
NA
WI080895
8/8/1995
mg/L
ND
ND
0.26
123
0.15
0.043
61.7
5.2
0.093
3.5
ND
14.1
109
638
0.87
11.7
ND
0.095
ND
ND
ND
ND
18.8
10.0
107
107
NA
NA
750
10.4
WI082395
8/23/1995
mg/L
0.079
ND
0.240
125
0.19
0.039
61.3
5.2
0.086
3.2
ND
15.2
108
630
0.95
13.1
0.093
ND
ND
ND
ND
ND
32.0
1050
107
107
NA
NA
NA
NA
* = Dissolved metals
— = Not applicable
jjS = Microsiemens
mg'L = Milligrams per liter
NA = Not analyzed
ND=Not detected
65
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORT HOP HOSP HATE
NITRATE PLUS NITRITE ASN
NITRITE AS N
NIT RATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaC03
ALKALINITY. BICARB
ASCAC03
DISSOLVED OXYGEN (mg/L)
pH
CONDUCTIVITY (jfi)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
-
--
--
--
WI090595
9/5/1995
mg/L
ND
ND
0.24
123
0.28
0.038
60.2
5.2
0.087
ND
ND
12.4
107
652
NA
0.88
NA
0,067
ND
ND
ND
ND
ND
18.4
1050
107
107
W 11 10995
11/9/1995
mg/L
ND
ND
0.20
113
0.18
0.027
56.2
5.2
0.082
3.2
ND
15.6
105
591
NA
0.97
17.7
0.060
0.20
ND
ND
ND
ND
14.4
956
95.7
95.7
CDPHE
1/29/1996
mg/L
NA
NA
0.160
NA
0.24
NA
NA
3.60
NA
NA
NA
NA
73
490
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
2/29/1996
mg/L
NA
NA
0.200
NA
0.26
NA
NA
3.50
NA
NA
NA
NA
69
450
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
4/25/1996
mg/L
NA
NA
0.12
NA
0.18
NA
NA
2.4
NA
NA
NA
NA
46
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
5/31/1996
mg/L
NA
NA
0.14
NA
0.17
NA
NA
2.7
NA
NA
NA
NA
56
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
* = Dissolved metals
— = Nol applicable
j£ = Micro-Siemens
mg/L = Milligrams per liter
NA = Not analyzed
ND = Not detected
66
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANION&
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORT HOP HOSP HATE
NITRATE PLUSNITRITE ASN
NITRITE ASN
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
DISSOLVED OXYGEN (mg/L)
pH
CONDUCT IWTYOfi)
TEMPERATURE (degrees Q
ANALYTICAL
MEIHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300,0
376.2
340.2
300,0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160,2
160.1
9060
310.1
310.1
--
--
-
-
CDPHE
6/14/1996
mg/L
NA
NA
0.16
NA
0.18
NA
NA
2.9
NA
NA
NA
NA
60
430
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
7/19/1996
mg/L
NA
NA
0.19
NA
0.20
NA
NA
3.5
NA
NA
NA
NA
71
490
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
8/31/1996
mg/L
NA
NA
0.20
NA
0.24
NA
NA
4.1
NA
NA
NA
NA
84
520
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
WI120996
12/9/1996
mg/L
ND
NA
0.15
104
0.30
NA
52.8
3.7
0.07
3.1 J
NA
17.4
78
488
NA
NA
17.8
NA
0.31
ND
ND
ND
ND
NA
849
0.8J
97.6
97.6
7.4
7,2
NA
10.0
WI012197
1/21/1997
mg/L
ND
NA
0.12
100.0
0.30
NA
51.2
3.5
0.06
3.0 J
NA
16.4
74
491
NA
NA
18.2
NA
0.17
ND
ND
ND
ND
8.4
796
1.1
94.9
94.9
8.8
5.1
NA
8.2
W 1022097
2/20/1997
mg/L
ND
NA
0.11
105
0.33
NA
52
3.7
0.06
3.0 J
NA
17.0
78
471
NA
NA
18.3
NA
0.22
ND
ND
ND
ND
3.2
809
1.8
101
101
8.6
7.5
NA
3.2
* = Dissolved metals
- = Not applicable
jjS=Microsiemens
mg/L= Milligrams per liter
NA = Not analyzed
ND = Not detected
67
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANAL VIE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORT HOP HOSP HATE
NITRATE PLUS NITRITE ASN
NITRITE AS N
NIT RATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
DISSOLVED OXYGEN (mg/L)
PH
CONDUCTIVITY (US)
TEMPERATURE (degrees C)
ANALYTICAL
MEIHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300,0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310,1
310.1
--
-
-
•-
W 103209 7
3/20/1997
mg/L
ND
NA
0.14
97,5
0.34
NA
48.8
3.6
0.07
ND
NA
15.6
75
476
NA
NA
18.7
NA
0.15
ND
0.002 1J
ND
ND
7.6
751
0.20 J
96.3
96.3
7,8
6,9
NA
8.6
WI042297
4/22/1997
mg/L
0.17
NA
0.07
67.2
0,34
NA
37.3
2.0
0.034 J
2.7 J
NA
ND
42
279
NA
NA
9.3
NA
ND
0.14
0.0046 J
ND
ND
1.6 J
507
1.30
99.7
99.7
7.3
7,4
NA
9.7
WI052897
5/28/1997
mg/L
ND
NA
0,11
86.4
0,24
NA
53.8
2,7
0.042
3.3 J
NA
14.9
56
358
NA
NA
7.2
NA
ND
ND
0.0024 J
ND
ND
12.4
653
1.4
107
107
7.3
7.4
NA
10.5
WI062397
6/23/1997
mg/L
ND
NA
0.19
95.6
0.26
NA
52.3
3.3
0,030 J
3.5 J
NA
ND
72
428
NA
NA
9.2
NA
0.10
0.14
0.0028 J
ND
ND
14.4
765
0.98 J
121
121
8
7.5
NA
9.7
W 1082 897
8/28/1997
mg/L
ND
NA
0.22
121
0.3
NA
61.9
4.9
0.090
4.8 J
NA
ND
104
541
NA
NA
13.8
NA
ND
ND
0.0037J
ND
ND
927
0,80 J
8.7
6.9
NA
9.6
WI093097
9/30/1997
mg/L
ND
NA
0.200
119
0.33
NA
58.4
4.9
0.098
3.4 J
NA
18.3
104
568
NA
NA
16
NA
ND
0.19
ND
ND
ND
16.4
940
0.58 J
102
102
NA
6.9
NA
9.4
* = Dissolved metals
— = Not applicable
nS = Microsiemens
mg'L = Milligrams per liter
NA -Not analyzed
ND= Not detected
68
-------
Table A-l (continued). Influent Results
INFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANION&
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE ASN
NITRITE ASN
NIT RATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
DISSOLVED OXYGEN (mg/L)
pH
CONDUCTIVITY (nS)
TEMPERATURE (degrees C)
ANALYTICAL
MEfflOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
-
--
--
--
WI102997
10/29/1997
mg/L
ND
NA
0.19
113
0.37
NA
58.8
4.9
0.079
3.4 J
NA
18.3
95
571
NA
NA
17.5
NA
ND
0.11
0.002J
NA
ND
10.4
940
0.71J
84
84
10.3
7.2
NA
9.2
WI112597
11/25/1997
mg/L
ND
NA
0.22
103
0.39
NA
50.4
4.2
0.065
ND
NA
16.5
86
548
NA
NA
17.8
NA
0.15
ND
0.0025J
NA
ND
14.8
869.0
1.8
102
102
7.5
7.2
NA
8.9
* = Dissolved metals
-- = Not applicable
|jS= Microsiemens
mg/L= Milligrams per liter
NA = Not analyzed
ND = Not detected
69
-------
Table A-2. Downftow Effluent Results
DOWNFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSHNIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORT HO PHOSPHATE
NITRATE PLUS NITRITE AS N
NIT RITE AS N
NITRATE AS N
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaCO3
ALKALINITY, BICARB
AS CAC03
ORP (mV)
pH
CONDUCTIVITY fciS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
--
-
--
-
WED030994
03/09/94
mg/L
0,021
ND
0.00034
105.0
1.5
0.0015
56.7
1.6
0.0073
55.8
0.0015
19.0
14.2
350
4.1
0.82
15.6
9.9
10.6
0.24
ND
0.24
5.4
51.0
864
60,4
193
193
-77.0
7.3
845
4.1
WED 032394
03/23/94
mg/L
0.021
0.00056
0.00025
107.0
1.2
0.0012
56.9
1.5
0.0081
56.6
0.00012
17.1
14.9
357
5.2
0.93
28.4
10.6
12.4
ND
ND
ND
6.2
27,0
781
20.6
209
209
WED040694
04/06/94
mg/L
0.027
0.029
0.00028
110.0
1,1
0.00065
58.6
1.5
0,0086
54.0
0.000060
18.1
15.6
338
5.7
0.88
27.2
11.0
10.7
ND
ND
ND
5,9
47.0
766
29
200
200
-180
7.2
889
5.2
WED042094
04/20/94
mg/L
0,029
0.016
0.00053
113.0
1.0
0.0015
58.3
1.4
0.010
50.6
0.000089
15.3
15,3
337
2.1
0.90
28
10.8
11.1
ND
ND
ND
5.8
39.2
783
28.2
213
213
WED 050594
OS/05/94
mg/L
0.033
0.076
0.00072
113.0
1.1
0.0017
58.9
1.4
0.0090
48.3
0.0051
18.6
13.1
280
0.74
0.87
22
10.4
11.1
ND
ND
ND
4.6
3.8
753
20.8
193
193
-184
7.6
803
8.8
-- = Not applicable
H$ = M icroSiemens
mg/L = Milligrams per liter
mV= Millivolts
NA = Not analyzed
ND = Not detected
70
-------
Table A-2 (continued). Downflow Effluent Results
DOXVNFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE AS N
NITRITE AS N
NITRATE AS N
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaCO3
ALKALINITY, BICARB
ASCAC03
ORP(mV)
PH
CONDUCTIVITY (uS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
--
-
--
-
WED051994
OS/19/94
mg/L
0.024
0.066
0.001 1
107.0
1.0
0.0013
57.1
1.3
0.0088
39.5
0.000063
15.4
9.9
270
3.2
0.91
17.4
11.4
10.6
ND
ND
ND
4.4
ND
739
26.3
196
196
-271
7.28
812
12.2
WED060194
06/01/94
mg/L
0.030
0.0013
0.00073
112.0
1.1
0.0011
60.8
1.4
0.015
29.2
ND
15.2
10.3
319
2.4
0.95
18.4
10.1
9.2
ND
ND
ND
3.2
3.6
741
35.6
208
208
WED062994
06/29/94
mg/L
0.017
0,0011
ND
106.0
1.0
ND
55.2
1.5
0.014
19.8
0.00010
13.8
12.6
338
2.1
0.80
19.6
8.9
8.6
ND
ND
ND
2.3
33.6
709
17.8
188
188
-253
7.10
1040
12.3
WED07I394
07/13/94
mg/L
0.012
0.0010
ND
118.0
I.I
ND
57.9
1.8
0.0089
20.8
0.00025
14.7
15.3
337
1.3
0.90
17.8
9.5
8.6
ND
ND
ND
3.1
43
722
15.9
190
190
-250
7
1010
11.6
WED072894
07/28/94
mg/L
0.017
0.0012
ND
116.0
1.1
ND
55.9
1.8
0.013
17.8
ND
14.5
16.5
354
6.9
1.1
19.8
7.8
7.5
2.3
ND
2,3
2.9
45.6
747
15.4
188
188
NA
NA
996
11.8
WED081S94
08/15/94
mg/L
0.016
0.0011
0.00033
114.0
1.3
ND
56.6
2,1
0.013
23.0
0.00014
15.5
14.5
311
1.5
1.0
19.2
8.7
6.7
1.7
ND
1.7
3.2
43,2
759
15.6
194
194
NA
7.06
1006
12.1
-- = Not applicable
US = MicroSiemens
mgT- = Milligrams per liter
mV= Millivolts
NA = Not analyzed
ND = Not detected
71
-------
Table A-2 (continued). Downflow Effluent Results
DOWNFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE AS N
NITRITE AS N
NITRATE AS N
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL;
ASCaCOS
ALKALINITY, BICARB
AS CAC03
ORP(mV)
PH
CONDUCTI VITY(nS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310,1
--
--
-
--
WED082494
08/24A4
mg/L
0.015
0.001 1
0.00030
117.0
1.7
ND
57.5
2.2
0.014
21.7
ND
15.6
15.3
345
4.5
1.0
21.3
10.4
7.9
1.8
ND
1.8
3.2
48.8
713
13.8
191
191
-125
6.88
973
13.4
WED090794
09/07*4
mg/L
0.053
ND
ND
113.0
1.8
0.0016
55,8
2.0
0.013
25.0
0.00032*
14.5
15.2
349
0.12
0.94
22.3
1.6
8.6
ND
ND
ND
2.6
49.6
741
12.3
194
194
-163
6,91
997
12.4
WED091994
09/19/94
mg/L
0.022
0.001 1
ND
124.0
2.0
0.0023
63.9
2.2
0.020
24.9
0.00034
16.4
17.5
349
5.3
0.96
21.0
9.1
13.8
ND
ND
ND
2.5
47.2
738
10.3
184
184
-216
6,9
1010
10.7
WED 100494
10/04/94
mg/L
0.037
0.0018
0.00038
115.0
1.8
0.0032
57.6
1.9
0.019
21.6
0.0012
14.4
15.5
333
10.7
0.88
21.0
8.8
8.5
ND
ND
ND
2,9
52.0
716
9,7
200
200
-220
6.9
960
9.0
WED 101 994
10/19/1994
mg/L
0.018
ND
0.00048
112.0
1.7
ND
57.7
1.8
0.020
19.5
ND
14.5
14.2
353
4.8
0.85
20.3
9.0
8.4
ND
ND
ND
2.2
45.6
698
8.1
174
174
-331
6.66
750
6.8
WED 110294
11/02/94
mg/L
0.023
ND
0.00041
112.0
1.8
ND
58.0
1.6
0.020
16.8
ND
15.5
12.1
365
7.4
0.87
20.8
8.2
8.8
ND
ND
ND
1.5
40.0
734
5.0
152
152
-149
6.92
890
4.9
--"Not applicable
nS=MicroSiemens
mg/L = Milligrams per liter
mV-Millivolts
NA = Not analyzed
ND = Not detected
72
-------
Table A-2 (continued). Downflow Effluent Results
DO WNFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE AS N
NITRITE AS N
NITRATE AS N
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCOS
ALKALINITY, BICARB
ASCAC03
ORP(mV)
PH
CONDUCTIVITY (yS)
TEMPERATURE (degrees C)
WED 112094
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
--
-
--
--
11/20/94
mg/L
0.018
ND
0.00030
120.0
1.8
0.0054
60.6
1.6
0.019
16.0
ND
14.6
10.9
357
0.11
0.90
21.0
6.5
3.1
ND
ND
ND
2.2
41.0
750
6.9
187
187
-170
7.6
NA
3.7
WED 11309*
11/30/94
mg/L
0.023
ND
0.00030
118.0
2.4
0.0018
58.0
1.7
0.019
13.1
0.00022
14.5
11.7
391
5.8
1.1
22.0
7.2
5.0
ND
ND
ND
2.0
40.5
767
20.4
143
143
' -220
7.12
600
3,0
WED121494
12/14/94
mg/L
0.013
ND
0.00088
120.0
2.0
0.011
56.6
1.5
0.017
11.5
ND
15.0
8.8
391
3.1
0.99
21.2
7.3
6.2
ND
ND
ND
0.41
28.5
744
5.7
152
152
-195
7.46
600
2.9
WED 01 049 5
01/04/95
mg/L
0.013
0.0039
ND
117.0
2.7
ND
57.1
1.9
0.013
9.7
ND
14.3
8.3
386
3.3
1.1
22.1
6.6
5.5
ND
ND
ND
1.6
34.0
729
4.8
146
146
-20.0
7,26
590
3.3
WED011895
01/18/95
mg/L
0.014
0.0035
ND
119.0
3.0
0.0012
54.5
1.9
0.014
9.9
ND
14.9
9.7
386
1.6
1.0
22.1
6.4
4.9
ND
ND
ND
1.5
37,0
718
5.6
141
141
-6.5
7.6
590
3.0
WED02019!
02/01/95
mg/L
0.022
ND
ND
115.0
2.6
ND
50.7
1.8
0.018
8.3
ND
15.0
10.5
380
2.3
1.0
21.9
6.3
6.0
ND
ND
ND
1.3
33.0
721
4.8
129
129
-7.3
7.6
670
4.0
-- = Not applicable
nS = MicroSiemens
mg/L = Milligrams per liter
mV= Millivolts
NA - Not analyzed
ND - Not detected
73
-------
Table A-2 (continued). Downflow Effluent Results
DOWNFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE AS N
NITRITE AS N
NITRATE AS N
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
ORP (mV)
pH
CONDUCTIVITY(nS)
TEMPERATURE(degrees C)
WED021595!
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353,2/354.1
350.1
160.2
160.1
9060
310.1
310.1
»
--
--
-
02/15/95
mg/L
0.018
0.001 1
0.00033
116.0
2.4
0.0010
51.2
1.9
0.016
8.4
ND
15.9
10.7
359
1.9
1.1
22.1
17.5
5.1
ND
ND
ND
1.2
32.4
679
4.3
140
140
59.0
8.8
NA
2.8
WED022795;
02/27*5
mg/L
0.011
0.0019
ND
121.0
2.1
ND
52,5
1.9
0.015
8.5
ND
15.1
11.7
346
1.9
1.0
22.7
5.9
5.8
ND
ND
ND
1.2
34.0
723
5.5
152
152
-82.0
7,1
620
5.6
WED031595
03/15*5
mg/L
0.011
ND
ND
126.0
2.2
ND
54.3
2.1
0.018
9.0
ND
16.5
13.0
370
3.1
1.1
24.4
5.7
5.7
ND
ND
ND
1.5
33.0
707
5.4
152
152
-65.0
7.1
680
6.8
WED032993
03/29/95
mg/L
ND
ND
ND
103,0
1.8
ND
46.0
1.8
0.019
6,9
ND
14.7
12.2
341
3.1
1.1
22.5
5.2
3.8
ND
ND
ND
1.3
31.0
662
5.8
141
141
-81.1
7.3
580
5.6
WED041295
04/12/95
mg/L
0.014
0.0021
ND
113.0
1.8
ND
48.1
1.9
0.014
6.7
ND
14.1
12.6
338
0.099
1.0
21.8
4.7
5,4
ND
ND
ND
1,1
35.0
655
6,9
143
143
35.0
7.2
580
4.8
WED04269;
04/26/95
mg/L
ND*
ND
ND
109.0
1.7
ND
46.6
1.9
0,014
6.9
ND
14.1
11.9
341
1.6
1.1
23.8
4.7
2.4
ND
ND
ND
1.1
31.2
651
6.8
141
141
NA
NA
600
7,0
— = Not applicable
^8= MicroSiemens
mg/L = Milligrams per liter
mV= Millivolts
NA = Not analyzed
ND = Not detected
74
-------
Table A-2 (continued). Downflow Effluent Results
DOWNFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE AS N
NITRITE AS N
NITRATE AS N
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaCO3
ALKALINITY, BICARB
ASCAC03
ORP(mV)
pH
CONDUCTIVITY (nS)
TEMPERATURE (degrees C)
\VED05109S
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353,2^54.1
330.1
160.2
160.1
9060
310.1
310.1
05/10/95
mg/L
ND*
ND
ND
121.0
2.1
ND
47.8
2.4
0.016
6.5
ND
14.1
13.3
348.0
0.38
1.1
22.6
4.3
4.1
ND
ND
ND
0.96
29.2
707
4.4
137
137
WED061295
6/12/1995
mg/L
ND
ND
ND
125
4.2
ND
52.7
3.9
0.017
6.8
ND
8.7
26.5
425
0.054
0.87
7.0
3.7
2.2
ND
ND
ND
0.90
43.0
763
6.6
129
129
-80
6.8
NA
11.7
WED062895
6/28/1 995
mg/L
ND
ND
ND
142
3.9
ND
61.9
4.4
0.020
7.1
ND
10.6
31.2
453
6.9
0.80
7.2
4.7
1.5
ND
ND
ND
0.94
53.6
918
11.4
195
195
-68
6.6
720
12.3
WED071095
7/10/1995
mg/L
ND
ND
ND
144
3.9
ND
68.7
4.1
0.021
8.2
ND
12.8
30.8
525
5.7
0.96
8.6
3.5
3.7
ND
ND
ND
1.0
48.0
946
5.4
146
146
-52
6.7
NA
13.8
WED072695
7/26/1995
mg/L
ND
ND
ND
157
2.9
ND
71.7
4.1
0.020
7.6
ND
12.6
29.7
537
0.83
0.86
10.1
2.6
2.0
ND
ND
ND
0.50
28.0
959
7.2
141
141
WED08089!
8/8/1995
mg/L
0.015
ND
ND
148
2.8
ND
68,6
3.8
0.022
6.8
ND
12.5
33.1
535
10.0
0.91
11.1
2.5
1.6
ND
ND
ND
0.64
38.8
1090
4.7
14
7.1
850
14.1
* - Aluminum was re-analyzed 6/2/95 due to blank contamination
-- = Not applicable mV= Millivolts
US = MicroSiemens NA = Not analyzed
m^l = M illigrams p er liter -ND = Not detected
75
-------
Table A-2 (continued). Downflow Effluent Results
DOWNFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE ASN
NITRITE AS N
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaC03
ALKALINITY, BICARB
AS CAC03
ORP(mV)
PH
CONDUCTIVITY (nS)
TEMPERATURE (degrees C)
WED082395
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353,2/354.1
350.1
160.2
160.1
9060
310.1
310.1
8/23/1995
mg/L
ND
ND
ND
155
2.7
ND
70.2
3.9
0.026
6.2
ND
13.7
34.1
539
11.4
0.85
12.2
3.0
3.0
ND
0.0070
ND
0.78
50.0
996
4.2
143
143
WED 090595
9/5/1995
mg/L
0.016
ND
ND
147
2.2
ND
66,3
3.7
0.028
6.2
ND
12.5
29.1
529
5.6
0.82
14
2.8
1.3
ND
ND
ND
0.64
45.6
941
4.9
179
179
WED 11 0995
11/9/1995
mg/L
ND
ND
0.00030
149
2.4
0.0016
66.2
4.0
0.04
5.3
ND
14.5
34.5
535
3.8
0.81
17.3
2.5
1.1
ND
ND
ND
0.39
12.8
957
4.2
152
152
-60
6.7
750
4.7
CDPHE
1/29/1996
mg/L
NA
NA
0.00012
NA
NA
NA
NA
3.2
NA
NA
NA
NA
28
440
NA
NA
NA
NA
NA
NA
NA
NA
1.0
AM
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
2/29/1996
mg/L
NA
NA
0.00072
AM
0.28
AM
NA
3.0
NA
NA
NA
NA
26
430
NA
NA
NA
NA
NA
NA
NA
NA
1.1
AM
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
4/25/1996
mg/L
AM
NA
0.15
AM
1.7
AM
NA
2.2
NA
NA
NA
NA
15
318
NA
NA
NA
NA
NA
NA
NA
NA
1.1
AM
NA
NA
NA
NA
NA
NA
NA
NA
.--Not applicable
nS-MicroSiemens
mg/L - M illigrams per liter
mV-Millivolts
NA - Not analyzed
ND - Not detected
76
-------
Table A-2 (continued). Downflow Effluent Results
DOWNFLO W EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE AS N
NITRITE AS N
NITRATE AS N
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
AS CaC03
ALKALINITY, BICARB
ASCAC03
ORP(mV)
pH
CONDUCTIVITY (uS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
CDPHE
5/31/1996
mg/L
NA
NA
0.00016
NA
0.87
NA
NA
1.8
NA
NA
NA
NA
11
230
NA
NA
NA
NA
NA
NA
NA
NA
0.67
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
6/14/1996
mg/L
NA
NA
ND
NA
0.92
NA
NA
2.00
NA
NA
NA
NA
9.7
82
NA
NA
NA
NA
NA
NA
NA
NA
1.2
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
7/19/1996
mg/L
NA
NA
0.00021
NA
1.10
NA
NA
2.10
NA
NA
NA
NA
8.7
340
NA
NA
NA
NA
NA
NA
NA
NA
0.90
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
8/31/1996
mg/L
NA
NA
0.00013
NA
1.60
NA
NA
2.20
NA
NA
NA
NA
5.8
350
NA
NA
NA
NA
NA
NA
NA
NA
ND
NA
NA
NA
NA
NA
NA
NA
NA
NA
WED01219V
1/21/1997
mg/l
0.098
NA
0.016
115
0.53
57,3
3.3
NA
0.05
0.39
NA
16.6
55
421
0.13
NA
18.6
NA
1.1
ND
0.0025
NA
0.24
7.2
787
7.2
158
158
110
5.3
NA
1.8
WED02209-J
2/20/1997
mg/l
ND
NA
0.034
113
0.72
56.9
5.0
NA
0.035
3.80
NA
16
59.7
322
ND
NA
18.6
NA
0.54
0.2
0.0055
NA
0.20
6.0
752
25.8
259
259
92.0
7.0
NA
1.8
- = Not applicable
|iS = MicroSiemens
mg/L = Milli^ams per liter
mV= Millivolts
NA = Not analyzed
ND = Not detected
77
-------
Table A-.1. Upflow Effluent Results
UPFLOW EFFLUENT
AQUEOUS
ANALY1E
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORT HOP HOSP HATE
NITRATE PLUS NIT RITE ASN
NIT RITE AS N
NIT RATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
ORP(mV)
pH
CONDUCTIVITY (^S)
TEMPERATURE (degrees C)
ANALYTICAL
MEIHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350,1
160.2
160,1
9060
310.1
310.1
-
--
--
--
WEU030994
03/09/94
mg/L
0.077
0,0062
0.00042
75.3
0.48
0.0042
72.7
0.051
0,0054
223.0
0.0014
33.9
0.22
354
0.38
0,30
83.2
24.3
26.8
ND
ND
ND
23.8
6
1390
264
367
367
-377
8
1410
5
WEU032394
03/23/94
mg/L
0.20
0.0071
0.00049
96.2
0.61
0.0030
71.4
0.072
0.0071
188.0
0.00015
31.2
0.22
388
7.9
0.57
76.0
23.2
26.7
ND
ND
ND
19,6
12.0
1200
51.3
347
347
WEU040694
04/06/94
mg/L
0,078
0.036
0.00034
112.0
0.48
0.0038
69.3
0.065
0.0095
150.0
0.000084
27.3
0.13
364
9.4
0,62
59.7
20.5
20.9
0.060
ND
0.060
15.0
6.0
1110
60.0
310
310
-280
7.85
1222
6.0
WEU042094
04/20/94
mg/L
0.39
0.028
0.00036
115.0
0.99
0.020
63.1
0.16
0.0086
108.0
0,00048
21,8
0.43
343
1.9
0.72
50.0
20.8
20.6
ND
ND
ND
12.9
25.2
1010
49.3
308
308
WEUOS0594
05/05/94
mg/L
0.062
0.085
0.00024
123,0
0.27
0.0022
66.0
0.17
0.0086
91.2
0.000071
22
0.14
292
0,47
0.71
35.5
18.3
18.6
ND
ND
ND
10.5
ND
934
35.6
265
265
-269
7,20
954
7.8
WEU05199<
05/19/94
mg/L
0.028
0.067
0.00020
115.0
0.25
0.0015
60.1
0.25
0.0086
49.4
0.000072
16.8
0.32
265
2,4
0,88
21.8
17.6
15.9
ND
ND
ND
6.8
ND
804
23,8
230
230
-271
7.84
893
8.8
-- = Not applicable
li/s = Micro Siemens
mg/L = Milligrams per liter
mV= Millivolts
NA == Not analyzed
ND = Not detected
78
-------
Table A-3 (continued). Upflow Effluent Results
UPFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NIT RITE ASN
NITRITE ASN
NIT RATE ASN
AMMONIA
TOTAL SOLIDS
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
ORP(tnV)
pH
CONDUCTIVITY (nS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
-
--
--
--
WEU060194
06/01/94
mg/L
0.045
ND
ND
117
0.26
0.0030
61.5
0.33
0.014
37.3
ND
15.7
0.20
330
5
0.81
22.2
27.3
14.9
ND
ND
ND
5.6
ND
$08
28.0
244
244
WEU062994
06/29/94
mg/L
0.021
ND
ND
120
0.47
0.0017
61.7
0.79
0.011
24.2
0.00014
15.6
0.35
355
3.2
0.90
20.9
12.8
21.3
ND
ND
ND
3.0
2.4
759
11.4
220
220
-275
7.7
1115
9.7
WEU071394
07/13/95
hig/L
ND
ND
ND
132
0.79
ND
61.4
1.3
0.0052
17.3
0.00015
15
0.18
372
0.59
0.80
18.9
13.3
19.5
ND
ND
ND
3.0
2.0
766
9.0
211
211
-280
7.6
1090
9.4
WEU072894
07/28/95
mg/L
0.38
ND
ND
132
1.4
ND
58.6
1.7
0.0075
13.7
ND
14.2
0.29
356
1.5
1.0
20.2
10.8
10.5
1.9
ND
1.9
2.6
18.8
816
9.6
206
206
NA
NA
1049
9.7
WEU081594
08/15/94
mg/L
0.015
ND
ND
134
2.7
ND
58.3
2.1
0.0089
12.8
0.00021
14.4
0.38
369
0.69
0.96
19.9
10.5
7,8
1.7
0.077
1.7
1.6
7.6
802
8.8
194
194
NA
7.6
1069
9.4
WEU08249*
08/24/94
mg/L
0.023
ND
ND
132
3.3
ND
57.1
2.3
0.0077
11.3
ND
14.4
0.58
392
1.0
1.1
20.5
9.8
9.2
1.8
ND
1,8
1.3
27,2
767
6.0
183
183
-344
7,46
1037
10.0
- = Not applicable
H/s = Micro Siemens
mj^L = Milligrams per liter
mV = Millivolts
NA = Not analyzed
ND = Not detected
79
-------
Table A-3 (continued). Upflow Effluent Results
UPFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOS>HATE
NITRATE PLUS NITRITE ASN
NITRITE AS N
NIT RATE ASN
AMMONIA
TOTAL SOLIDS
TSS
IDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
ORP(mV)
pH
CONDUCT IVITY(nS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
--
--
--
--
WEU090794
09/07/94
mg/L
0,052
ND
ND
126
4.3
0.0011
53.3
2.4
0.0083
10.2
0.00011*
13.6
0.82
395
0.12
1.0
21.1
1.6
8.8
ND
ND
ND
1.0
27.6
787
6.4
175
175
-315
7.39
1007
9.3
VEU090794E
09/07/94
mg/L
ND
ND
ND
0.15
ND
ND
ND
ND
ND
ND
ND*
ND
0.019
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
WEU091994
09/19/94
mg/L
0.023
ND
ND
132
5.1
0.0015
56.6
2.6
0.015
9.0
0.00046
14.2
1.4
391
0.23
1.1
20.4
7.9
9.8
ND
ND
ND
0,87
28.8
790
5.8
164
164
-267
7.3
990
9.2
WEU100494
10/04/94
mg/L
0.017
0.0011
ND
127
5.7
ND
54.5
2.4
0,015
11.9
0.00052
13.8
2.4
369
5.0
0.99
21,4
8.0
7.1
ND
ND
ND
1.0
37.6
750
7.4
182
182
-260
7.3
960
8.7
WB5100494
10/04/94
mg/L
6.1
0.0021
0.024
344
92.7
0,020
139.0
28.6
0.20
7,4
0.00099
46.8
9.4
1760
NS
1.0
6.0
NS
ND
NS
ND
NS
NS
49.6
2520
NS
ND
ND
NA
5.2
NA
15.0
WEU10199'
10/19/94
mg/L
0.015
0.0011
ND
128
5.6
ND
54.1
2.7
0.019
7.7
ND
14.6
3.1
392
1.3
0.95
20.2
6,8
6,8
ND
0.018
ND
0.51
40.8
734
5,3
150
150
-344
6.95
760
7,7
-=Not applicable
jVs = MicroSJemens
mg/L *= Milligrams per liter
mV= Millivolts
NA-Not analyzed
ND " Not detected
BO
-------
Table A-3 (continued). Upflow Effluent Results
UPFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE ASN
NIT RITE AS N
NIT RATE ASN
AMMONIA
TOTAL SOLIDS
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCOS
ALKALINITY, BICARB
ASCAC03
ORP(mV)
pH
CONDUCTIVITY (pS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
--
--
--
-
VEW01994I
10/19/94
mg/L
0.025
ND
ND
130
5.7
ND
54.8
2.7
0.018
7.6
ND
15
3.2
380
1.8
0.97
20.0
6.9
6.2
ND
0.017
ND
0.52
36.8
742
5.6
148
148
-344
6.95
760
7.7
WEU1I0294
11/02/94
mg/L
0.025
0.0011
ND
122
7.0
ND
52.5
2.7
0.018
11.6
ND
14.2
6.8
371
3.8
1.1
23.2
6.2
5.9
ND
0.016
ND
0.38
52.0
121
9.4
141
141
-164
7.01
935
8.5
WEU112094
11/20/94
mg/L
0.012
0.0011
ND
127
6.0
ND
53.8
2.8
0.016
9.6
0.00082
14.7
6.5
360
3.8
1.0
23.0
5.5
2.7
ND
ND
ND
0.74
49.0
745
7.3
185
185
-160
7.2
NA
8.1
WEU1 13094
11/30/94
mg/1
0.013
0.0014
ND
123
7.5
ND
51.4
2.9
0.018
7.6
0.00028
14.5
7,9
379
4.6
1.2
22.2
6.9
2.7
ND
ND
ND
0.55
47.0
729
19.1
142
142
-216
6.8
640
7.1
WEU121494
12/14/94
mg/L
0.015
0.0010
ND
127
6.8
ND
52,4
2.9
0.016
7.6
ND
15.6
9.0
375
3.2
1.1
22.4
5.3
4.7
ND
ND
ND
1.5
44.0
729
6.4
157
157
-196
7.33
670
7.7
WEU01049J
01/04/95
mg/L
0.020
0.0035
ND
116
6.3
ND
53.1
2.7
0.012
15.3
0.00033
14.5
11,7
341
3.3
1.1
25.6
4.8
3.0
ND
ND
ND
0.68
51.0
707
12.5
171
171
-80
7.0
670
7.0
-- = Not applicable
ps = Micro Siemens
mg/L = Milligrams per liter
mV= Millivolts
NA - Not applicable
ND - Not detected
81
-------
Table A-3 (continued). Upflow Effluent Results
UPFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUSNITRITE ASN
NITRITE ASN
NIT RATE ASN
AMMONIA
TOTAL SOLIDS
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaC03
ALKALINITY, BICARB
ASCAC03
ORP(mV)
PH
CONDUCT I VITY(uS)
TEMPERATURE (degrees C)
ANALYTICAL
MEIHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365,3
353.2
354.1
353.2/354.1
350.1
160.2
WEU011895
01/18/95
mg/L
0.026
0.0043
ND
116
5.4
0.0034
49,5
2,6
0.012
10.5
ND
15
12,5
347
3.0
1.1
23.0
4.7
3.0
ND
ND
ND
0.63
51,0
160.1 693
9060
310.1
310.1
-
-
--
-
7.8
168
168
5
7.1
650
8.3
WEU020195
02/01/95
mg/L
0.017
0.0015
ND
119
4.9
ND
49.0
2.5
0.016
9.1
ND
16.7
16.9
330
6.0
1.0
23.4
5.5
5.0
1.4
ND
1.4
0.52
54.0
692
6.8
161
161
-11.7
7.4
610
6.1
WEU021595
02/15/95
mg/L
0.012
0.0020
ND
119
4.3
ND
49.1
2.5
0.016
9.1
NP
16
12.9
308
3,3
1.1
23.4
13.3
3.7
ND
ND
ND
0.51
45.2
682
6.2
191
191
-44.0
7.2
NA
7.6
WEU022795
02/27/95
mg/L
0.014
0.0021
ND
116
4.0
ND
48.2
2.4
0,016
8.9
ND
15.2
17.8
340
4.3
1,0
23,6
3.4
3.0
ND
ND
ND
0.34
47.0
700
5.8
150
150
-65
6.9
680
8.4
WEU031S95
03/15/95
mg/L
0.012
0.0012
ND
116
4.0
ND
48.2
2.4
0.019
7.5
ND
16.0
18,0
335
2.7
1.1
24.3
4.1
2.6
ND
ND
ND
0.38
39.0
671
4.3
151
151
-63
6.9
650
8.8
\VEU03299i
03/29/95
mg/L
0.015
0.0012
ND
105
3.5
ND
44.6
2.2
0.019
5.9
ND
15.7
17.5
317
4.3
1.1
23.0
3.4
1.8
ND
ND
ND
0.31
41.0
667
7.0
154
154
-81.1
7.3
580
5.6
-- =Not applicable
nS=MicroSiemens
mg'L = Milligrams per liter
mV= Millivolts
NA = Not analyzed
ND = Not detected
82
-------
Table A-3 (continued). Upflow Effluent Results
UPFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE ASN
NITRITE AS N
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
IDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
AS CAC03
ORP(mV)
pH
CONDUCT IVITYCuS)
TEMPERATURE (degrees C)
ANALYTICAL
MEIHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365,3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
--
--
--
--
WEU04129S
04/12/95
mg/L
0.013
0.0028
ND
114
3.5
ND
46.5
2.5
0.013
7.1
ND
15.3
15.9
326
0.39
1.0
22.5
3.0
2.7
ND
ND
ND
0.33
41.9
657
8.0
152
152
-7.0
7.1
620
8.5
WEU042695
04/26/95
mg/L
ND'
ND
0.00078
106
2.2
ND
45.3
2.0
0.015
11.7
ND
14.2
18.5
326
2.9
1.1
26.0
3.2
1.6
ND
ND
ND
0.31
31.2
607
8.3
147
147
NA
NA
620
8.0
WEU051095
05/10/95
mg/L
ND*
ND
0.0094
110
2.2
0.0019
44.5
2.5
0.022
18.1
ND
13.3
26.7
355
1.3
1.2
25.9
2.0
2,4
ND
ND
ND
0.42
29.0
724
9.9
138
138
WEU06129S
06/12/95
mg/L
0.028
ND
0.0084
103
4.6
0.0018
45.2
3
0.019
7.5
ND
8.9
33.5
326
0.065
0.90
7.6
2.3
1,5
ND
ND
ND
0.36
47.3
668
9.1
181
181
-57
6,7
NA
10.1
WEU06289S
6/28/1995
mg/L
ND
ND
0.0045
121
3.7
ND
60.2
4.0
0.026
6.0
ND
11.2
47.1
494
1.5
0.90
7.0
1.2
0.34
ND
ND
ND
0.20
18.8
885
4.9
136
136 -
6.9
WEU071095
7/10/1995
mg/L
ND
ND
ND
130
3.8
ND
68.2
4.1
0.026
5.4
ND
13.2
50.8
514
1.5
0.96
8.3
1.5
0.48
ND
ND
ND
0.20
25.6
944
4.8
147
147
6.9
* - Aluminum was re-analyzed 6/2/95 due to blank contamination
~ = Not applicable N/¥>= Not analyzed
uS= MicroSiemens ND =Not detected
mg/L = Milligrams per liter
mV= Millivolts
83
-------
Table A-3 (continued). Upflow Effluent Results
UPFLOW EFFLUENT
AQ UEO US
ANALVTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS TOTAL
ORT HOP HOSP HATE
NITRATE PLUS NITRITE ASN
NIT RITE AS N
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL:
ASCaCO3
ALKALINITY, BICARB
ASCAC03
ORP(mV)
PH
CONDUCT IVITY(nS)
TEMPERATURE (degrees C)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365,3
353.2
354.1
353.2/354.1
350.1
160,2
160.1
9060
310,1
310.1
--
--
--
--
WEU072695
7/26/1995
mg/L
ND
ND
0.0060
144
2,5
ND
68,6
4.1
0.028
5.7
ND
12.2
53.2
549
4.3
0.89
10,0
1.1
1.0
ND
ND
ND
0.11
16.8
961
5.7
135
135
6.8
WEU080895
8/8/1995
mg/L
ND
ND
0.0046
135
2.5
ND
64.4
3.8
0.032
4.9
ND
13.3
56.6
584
3.5
0.88
11.2
1.1
0.9
ND
ND
ND
ND
17.6
999
3.4
138
138
7.1
WEU082395
8/23/1995
mg/L
ND
ND
0.0093
141
2.1
ND
66.1
3.8
0.036
4.5
ND
14.0
59.8
561
5.2
0.90
12.5
1.3
1,2
ND
0.0080
ND
0.21
30.0
1010
3.2
149
149
7.1
WEU090595
9/5/1995
mg/L
ND
ND
0.010
137
1.8
ND
64.3
3.6
0.04
ND
ND
12,2
59.9
569
2.8
0.86
13.7
1.5
0.43
ND
ND
ND
ND
26
978
3.7
160
160
6.9
WEUI10995
11/9/1995
mg/L
ND
ND
0,04400
133
0.93
0.0022
62.1
4.4
0.059
4,3
ND
15.6
73.6
559
0.84
0.96
17.1
0.69
0.80
ND
ND
ND
ND
5.2
932
2.1
115
115
7.0
CDPHE
1/29/1996
mg/L
AM
NA
0.037
AM
1.6
AM
NA
3.3
AM
AM
AM
NA
47
460
AM
AM
NA
NA
AM
AM
AM
AM
0.2
AM
AM
AM
AM
AM
AM
AM
NA
-- =Not applicable
uS = MicroSiemens
mg/L = Milligrams per liter
mV = Millivolts
NA = Not analyzed
NiJ=Not detected
84
-------
Table A-3 (continued). Upftow Effluent Results
UPFLOW EFFLUENT
AQUEOUS
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS;
SULFATE
SULFIDE TOTAL
FLUORIDE
CHLORIDE
PHOSPHORUS TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE ASN
NITRITE ASN
NITRATE ASN
AMMONIA
TOTAL SOLIDS:
TSS
TDS
TOC
ALKALINITY, TOTAL;
ASCaCO3
ALKALINITY, BICARB
ASCAC03
ORP(mV)
pH
CONDUCTIVITY (nS)
TEMPERATURE (degrees C)
ANALYTICAL
MEIHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
376.2
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
160.2
160.1
9060
310.1
310.1
--
-
--
--
CDPHE
2/29/1996
mg/L
AM
NA
0.035
NA
1.3
NA
NA
3.1
NA
NA
NA
NA
42
430
AM
NA
NA
NA
NA
NA
NA
NA
0.4
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
4/25/1996
mg/L
NA
NA
0.030
NA
0.81
NA
NA
2.3
NA
NA
NA
NA
31
329
AM
AM
AW
AM
AM
NA
NA
NA
0.3
NA
NA
NA
NA
NA
NA
NA
NA
NA
CDPHE
5/31/1996
mg/L
NA
NA
0.140
AM
0.17
NA
NA
2.7
AM
AM
NA
NA
56
420
AM
AM
AM
NA
NA
NA
NA
NA
ND
AM
AM
NA
NA
NA
NA
NA
NA
NA
CDPHE
6/14/1996
mg/L
AM
NA
0.031
AM
1.1
NA
NA
2.2
AM
AM
AM
AM
30
310
AM
AM
AM
AM
AM
AM
AM
AM
0.2
AM
AM
AM
AM
AM
AM
AM
AM
AM
CDPHE
7/19/1996
mg/L
AM
NA
0.051
AM
0.87
NA
NA
2.6
AM
AM
AM
AM
41
410
AM
AM
AM
AM
AM
AM
AM
AM
0.2
AM
AM
AM
AM
AM
AM
AM
AM
AM
CDPHE
8/31/1996
mg/L
NA
NA
0.053
NA
0.90
NA
NA
2.5
NA
NA
NA
NA
43.0
45
NA
NA
NA
NA
NA
NA
NA
NA
ND
NA
NA
NA
NA
NA
NA
NA
NA
AM
-- = Not applicable
nS=MicroSiemens
mg'L = Milligrams per liter
mV= Millivolts
NA = Not analyzed
ND = Not detected
85
-------
Table A-4, Substrate Results - Downflow Cell
SUBSTRATE - DOWNFLOW CELL
SEDIMENT
ANALYTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE, REACTIVE
SULFIDE, ACID VOLATILE
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE A S N
NITRITE AS N
NITRATE AS N
AMMONIA
WATER (%)
ANALYTICAL
METHOD
6010
6020
SD2032394
03/23/94
mg/kg
1410.0
2.9
6020 2.2
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
EPA/OSW
EPA (Draft)
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
ILM01.1
7040.0
2250.0
7.4
2140.0
99.2
3.9
890.0
0.061
ND
1560.0
214
0.40
NA
NA
NA
NA
25.8
NA
NA
NA
NA
82
SD2062994
06/29/94
mg/kg
65.6
0.14
0.56
406
88.7
3.1
145
4.1
ND
149.0
0.024
76.3
59.7
56.5
19.1
226
NA
NA
NA
63.4
NA
NA
NA
NA
62
SD5062994
06/29/94
mg/kg
423.0
ND
4.8
2330.0
653.0
53.4
571.0
36.0
1.9
184.0
0.79
ND
1000.0
143.0
18.6
178.0
NA
NA
NA
30.5
NA
NA
NA
NA
70
SD5082594
08/25/94
mg/kg
2580.0
0.59
5.1
7650.0
3650.0
16.2
2120.0
140.0
4.9
1360.0
0.16
ND
2650.0
214
3,2
ND
NA
NA
NA
18.8
NA
NA
NA
NA
75
NA = Not analyzed
ND = Not detected
88
-------
Table A-4 (continued). Substrate Results - Downflow Cell
SUBSTRATE - DOWNFLOW CELL
SEDIMENT
ANALVTE
ALUMINUM
ARSENIC
CADMIUM
CALCIUM
IRON
LEAD
MAGNESIUM
MANGANESE
NICKEL
POTASSIUM
SILVER
SODIUM
ZINC
ANIONS:
SULFATE
SULFIDE, REACTIVE
SULFIDE, ACID VOLATILE
FLUORIDE
CHLORIDE
PHOSPHORUS, TOTAL
ORTHOPHOSPHATE
NITRATE PLUS NITRITE AS N
NITRITE AS N
NITRATE AS N
AMMONIA
WATER (%)
ANALYTICAL
METHOD
6010
6020
6020
6010
6010
6020
6010
6010
6010
6010
6020
6010
6010
300.0
EPA/OSW
EPA (Draft)
340.2
300.0
365.3
365.3
353.2
354.1
353.2/354.1
350.1
ILM01.0
SD2100494
10/04/94
me/kg
2640.0
1.5
4.6
8460.0
3410.0
46.4
2180.0
160.0
3.7
930.0
0.17
ND
1510.0
86.8
103.0
190.0
NA
NA
NA
39.0
NA
NA
NA
NA
62
SD5100494
10/04/94
mg/kg
3200.0
0.97
10.5
4890.0
4640.0
30.8
1800.0
151.0
6.4
1410.0
0.29
108.0
2850.0
187.0
79.3
70.6
NA
NA
NA
3.3
NA
NA
NA
NA
70
SD2110294
11/02/94
mg/kg
3200.0
1.3
4.3
11700.0
4860.0
11.3
2910.0
232.0
7.0
1140.0
0.069
92.8
3170.0
159.0
1.1
171.0
NA
NA
NA
12.6
NA
NA
NA
NA
NA
SD2010495
01/04/95
mg/kg
2430.0
1.5
4.3
8770.0
3460.0
18.2
2190.0
144.0
4,9
729.0
0.28
ND
3250.0
184.0
15.3
117.0
NA
NA
NA
6.4
NA
NA
NA
NA
63
NA = Not analyzed
ND = Not detected
89
-------
BUREAU OF MINES
INFORMATION CIRCULAR/1994
PB94173341
Passive Treatment of Coal Mine
Drainage
By Robert S. Hedln, Robert W. Narln,
and Robert L, p. Klelnmann
UNITED STATES DEPARTMENT OF THE INTERIOR
KPMOUCIDir;
UI,OWlmint*
NMtttf T«*nto*l tnttrmttUfl
92
-------
U.S. Department of the Interior
Mission Statement
As the Nation's principal conservation agency, the Department of
the Interior has responsibility for most of our nationally-owned
public lands and natural resources. This includes fostering
sound use of our land and water resources; protecting our fish,
wildlife, and biological diversity; preserving the environmental
and cultural values of ournational parks and historical places; and
providing for the enjoyment of life through outdoor recreation.
The Department assesses our energy and mineral resources and
works to ensure that their development is in the best interests of
all our people by encouraging stewardship and citizen participa-
tion in their care. The Department also has a major responsibility
for American Indian reservation communities and for people who
live in island territories under U.S. administration.
93
-------
REPORT DOCUMENTATION PAGE
OMB No. OTM-Diea
Public reporting burden/or Ms collection of information Is estimated to average I hour per response, including the time for reviewing Instructions, searching existing
data sources, gathering and maintaining the data needed and completing and reviewing the collection of Information. Send
any osher aspect of this collection of information, including suggestions for reducing Ait bardtn io Washington Headquartt
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Project 70+4188), Washington, DC 20503.
"T
PB94-173341
STITDTAKD »U»TITLE
Passive Treatment of Coal Mine Drainage
Z. RCponT OAT I
October 4, 1993
Information Circular 9389
6. AUTHOR*8]
Hedin, R. S., Nairn, R. W., and Kleinmann, R. L. P.
B, FUNHNQ NUMBERS
l AND AODNEUIE*)
U.S. Bureau of Mines
Pittsburgh Research Center
P.O. Box 18070
Pittsburgh, PA 16236
8. PBWSRMINO
OM1ANIZAT1QN REKtTT NUMBER
1C 9389
NCV INAMHS) AND ADDniMIUI
U.S. Bureau of Mines
Research
8107th Street, NW
Washington, DC 20241
ID. (PON»OIUNUMOMT0RINa
AOINCV REPORT NUMBM
. •UPPlEMENTAHT NOTE*
None
1ZA. DUTRIBUTWNfAVAILABIUTY •TATEMEMT
TZB. DMTfllBUTION CODE
U, AUTflAcT (MubtHB 3M wail)
Passive methods of treating mine water utilize chemical and biological processes that decrease metal concentrations and
neutralize ecldity. Compared to conventional chemical treatment, passive methods generally require more land area, but utilize less
costly reagents and require less operational attention and maintenance. Currently, three types of passive technologies exist:
aerobic wetlands, wetlands that contain an organic substrate, and anoxic limestone drains. Aerobic wetlands promote mixed
oxidation and hydrolysis reactions, and are most effective when the raw mine water is net alkaline. Organic substrate wetlands
promote anaerobic bacterial activity that results in the precipitation of metal sulfides and the generation of bicarbonate alkalinity.
Anoxic limestone drains generate bicarbonate alkalinity and can be useful for the pretreatment of mine water before It flows Into a
wetland.
Rates of metal and acidity removal for passive systems have been developed empirically. Aerobic wetlands remove Fe and
Mn from alkaline water at rates of 10-20 QTrra*d-' and 0.6-1.0 g*rrrJtd-'( respectively. Wetlands with a composted organic
substrate remove acidity from mine water at rates of 3-9 g • m-* * d-1. A model lor the design and sizing of passive treatment
systems is presented in this report.
14. lUBjtOT TCTMt
Acid mine drainage, constructed wetlands, anoxic limestone drains, iron, sulfate, manganese, water
treatment
TS. NUMBDl OF MOM
38
IB. PNORCOOl
T. WTCUBITV 13. IECUNTV CLAMJflGATION W THW IS. IICUMTV CLAMWCATIOW OF ZOrDMITATION OF AMTAACT
ClAIHNCATION OP NEPOHT PAOI AHTRAOT
TIT
96
-------
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-------
CONTENTS
Page
Abstract 1
Introduction , 2
Treatment of mine water 2
Background of passive treatment 3
Acknowledgments 3
Chapter 1. Materials and methods 4
Collection of water samples 4
Analysis of water samples 4
Analytical quality control 4
Flow rate measurements 4
Analysis of surface deposits 4
Chapter 2. Chemical and biological processes in passive treatment systems 5
Acidity , 5
Alkalinity 6
Metal removal processes 7
Metal removal in aerobic environments 7
Iron oxidation and hydrolysis 7
Manganese oxidation and hydrolysis 8
Mine water chemistry in anaerobic environments 10
Limestone dissolution 10
Sulfate reduction 11
Aluminum reactions in mine water 13
Chapter 3. Removal of contaminants by passive treatment systems 14
Evaluation of treatment system performance 14
Dilution adjustments 16
Loading limitations 17
Study sites 18
Effects of treatment systems on contaminant concentrations 19
Dilution factors 19
Removal of metals from alkaline mine water 21
Removal of metals and acidity from acid mine drainage ., 23
Chapter 4, Design and sizing of passive treatment systems 25
Characterization of mine drainage discharges 26
Calculations of contaminant loadings 27
Classification of discharges 27
Passive treatment of net alkaline water 27
Passive treatment of net acid water 28
Pretreatment of acidic water with ALD 28
Treating mine water with compost wetland 30
Operation and maintenance 31
Chapter 5. Summary and conclusions 31
Kinetics of contaminant removal processes ...... 32
Long-term performance , 32
Continually evolving passive technologies 33
References 34
ILLUSTRATIONS
1. Comparison of calculated and measured acidities for water samples collected at Friendship Hill wetland 6
2. Removal of Fe2* from acidic and alkaline mine waters in a laboratory experiment 8
3. Concentrations of Fe3* and Held pH for water samples collected from Emlenton wetland 8
-------
ILLUSTRATIONS-Continued
Page
4. Concentrations of Felot and field pH at two constructed wetlands 9
5. Mean concentrations of Fe, Mn, and Mg at the Morrison wetland 9
6, Changes in concentrations of Fe2+ and Mn2* , 10
7. Concentrations of Ca, Fc, and alkalinity for water as it flows through the Howe Bridge ALD 11
8. Influent and effluent concentations at the Latrobe wetland 14
9. Relationship between mean Fe removal rates and ,4, mean influent pH and B, mean influent alkalinity .. 22
10. Relationship between mean Mn removal rates andX, mean influent pH and B, mean influent alkalinity . 22
11. Measured rates of alkalinity generation and acidity removal at the Friendship Hill wetland 25
12. Flow chart showing chemical determinations necessary for the design of passive treatment systems 26
13. Longitudinal-section and cross-section of the Morrison ALD — 29
TABLES
1. Federal effluent limitations for coal mine drainage 2
2. Calculated and measured acidities for synthetic acidic mine water 6
3. Acidic components of mine drainage influent at three passive treatment systems 6
4. Chemical compositions of mine drainages that contain high concentrations of alkalinity 7
5. Chemistry of mine water flowing through the Howe Bridge ancotic limestone drain, January 23,1992 ... 11
6. Solubility products of some metal sulfldes 12
7. Sinks for HjS in constructed wetlands and their net effect oa mine water acidity and alkalinity 12
8. Surface and pore water chemistry at the Latrobe Wetland 13
9. Hypothetical wetland data and performance evaluations 15
10. Influent and effluent concentrations of Ca, Mg, Na, and sulfate at eight constructed wetlands 16
11. Average concentrations of Fe, Mn, and Mg at the Morrison passive treatment system 17
12. Construction characteristics of the constructed wetlands IS
13. Average chemical characteristics of influent water at the constructed wetlands 19
14. Mean water quality for sampling stations at the constructed wetlands 20
15. Dilution factors for the constructed wetlands 21
16. Fe and Mn removal rates at the constructed wetland 21
17. Fe and SO4 content of ferric oxyhydroxide deposits , 24
18. Average rates of acidity removal, sulfate removal, and calcium addition at sites receiving acidic mine
water , 24
19. Recommended sizing for passive treatment systems 25
-------
UNIT OF MEASURE ABBREVIATIONS USED
cm
°C
ft
g
.-3
g-m"3
centimeter
degree Celsius
foot
gram
gram per cubic centimeter
gram per day
gram per square meter
nr'-d"1 gram per square meter per day
nr'-yr1 gram per square meter per year
gallon per minute
hectare
hour
kilogram (concentration)
kilogram per day
kilogram per cubic meter
liter
gpm
ha
h
kg-d-1
fcg-nr3
L
L-min'1
m
m1
fan
meg
mg
mg'L'1
mg-L^'
mi,
min
nmol
yda
yr
IN THIS REPORT
liter per minute
meter
square meter
micrometer
railliequivaJent
milligram
milligram per liter
milligram per liter per hour
milliliter
minute
nanomole
nanomole per cubic centimeter
per day
square yard
year
-------
PASSIVE TREATMENT OF COAL MINE DRAINAGE
By Robert S. Hedln,1 Robert W. Nairn,2 and Robert L. P. Klelnmann3
ABSTRACT
Passive methods of treating mine water use chemical and biological processes that decrease metal
concentrations and neutralize acidity. Compared with conventional chemical treatment, passive methods
generally require more land area, but use less costly reagents and require less operational attention and
maintenance. Currently, three types of passive technologies exist: aerobic wetlands, organic substrate
wetlands, and anoxic limestone drains. Aerobic wetlands promote mixed oxidation and hydrolysis
reactions, and are most effective when the raw mine water is net alkaline. Organic substrate wetlands
promote anaerobic bacterial activity that results in the precipitation of metal sulfides and the generation
of bicarbonate alkalinity. Anoxic limestone drains generate bicarbonate alkalinity and can be useful for
the pretreatment of mine water before it flows into a wetland.
Rates of metal and acidity removal for passive systems have been developed empirically by the U.S.
Bureau of Mines. Aerobic wetlands remove Fe and Mn from alkaline water at rates of 10-20 and 0.5-
1.0 g»m"2'd-:, respectively. Wetlands with a composted organic substrate remove acidity from mine
water at rates of 3-9 g-m~2-d"1. A model for the design and sizing of passive treatmeut systems is
presented in this report.
Research biologist.
^Research biologist (nowwith The Ohio State University, Columbus, OH).
3Reseaich supervisor,
Pittsburgh Research Center, U.S. Bureau of Mines, Pittsburgh, PA.
-------
INTRODUCTION
TREATMENT OF MINE WATER
The mining of coal in the Eastern and Midwestern
United States can result in drainage that is contaminated
with high concentrations of dissolved iron, manganese,
aluminum, and sulfate. At sites mined since May 4,1984,
drainage chemistry must meet strict effluent quality criteria
(table 1). To meet these criteria, mining companies com-
monly treat contaminated drainage using chemical meth-
ods. In most treatment systems, metal contaminants are
removed through the addition of alkaline chemicals (e.g.,
sodium hydroxide, calcium hydroxide, calcium oxide, sodi-
um carbonate or ammonia). The chemicals used in these
treatment systems can be expensive, especially when re-
quired in large quantities. In addition, there are operation
and maintenance costs associated with aeration and mixing
devices, and additional costs associated with the disposal
of metal-laden sludges that accumulate in settling ponds.
It is not unusual for the water treatment costs to exceed
$10,000 per year at sites that are otherwise successfully
reclaimed. Total water treatment costs for the coal mining
industry are estimated to exceed $1,000,000 per day (I).4
The high costs of water treatment place a serious financial
burden on active mining companies and have contributed
to the bankruptcies of many others.
Table 1 .-Federal affluent limitation* for coaJ mine drainage
Pollutant or
pollutant
property
Fe total
Mn total
Maximum for any
1 day,
6.0
4.0
Average of dally value*
for 30 consecutive
days mfl-L'1
3.0
2.0
pH between 6.0 and 9.0.
The high costs of chemical systems also limit the water
treatment efforts at abandoned sites. Thousands of miles
of streams and rivers in Appalachia are currently polluted
by the input of mine drainage from sites that were mined
and abandoned before enactment of strict effluent regula-
tions (2-3). State and Federal reclamation agencies, local
conservation organizations, and watershed associations ah*
consider the treatment of contaminated coal mine dis-
charges to be a high priority. Unfortunately, insufficient
funds are available for chemical water treatment, except in
a few watersheds of special value.
Natural processes commonly ameliorate mine drainage
pollution. As contaminated coal mine drainage flows into
and through receiving systems (streams, rivers, and lakes),
4Italic numbers in parentheses refer to items in the list of references
at the end of this report
its toxic characteristics decrease naturally as a result of
chemical and biological reactions and by dilution with
uncontaminated water. The low pH that is common to
many mine drainages is raised when the water mixes with
less acidic or alkaline water or through direct contact with
carbonate rocks. Metal contaminants of coal mine
drainage then precipitate as oxides and hydroxides under
the aerobic conditions found in most surface waters. Dis-
solved Fe precipitates as an oxyhydroxide, staining the
bottoms of many streams orange and often accumulating
to sufficient depths to suffocate benthic organisms. Less
commonly, dissolved Mn precipitates as an oxide that
stains rocks and dctrital material black. Dissolved Al
precipitates as a white hydroxide.
During the last decade, the possibility that mine water
might be treated passively has developed from an experi-
mental concept to full-scale field implementation at hun-
dreds of sites. Passive technologies take advantage of
natural chemical and biological processes that ameliorate
contaminated water conditions. Ideally, passive treatment
systems require no input of chemicals and little or no
operation and maintenance requirements. The costs of
passive treatment systems are generally measured in then-
land use requirements. Passive treatment systems use con-
taminant removal processes that are slower than that of
conventional treatment and thus require longer retention
times and larger areas to achieve similar results.
The goal of passive mine drainage treatment systems
is to enhance the natural amelioration processes so that
they occur within the treatment system, not ia the re-
ceiving water body. Two factors that determine whether
this goal can be accomplished are the kinetics of the
contaminant removal processes and the retention time of
the mine water in the treatment system. The retention
time for a particular minesite is often limited by available
land area. However, the kinetics of contaminant removal
processes can often be affected by manipulating the
environmental conditions that exist within the passive
treatment system. Efficient manipulation of contaminant
removal processes requires that the nature of the rate-
limiting aspects of each removal process be understood.
This U.S. Bureau of Mines (USBM) report describes
the chemical and biological processes that underlie the
passive technologies currently used in the eastern United
States for the treatment of contaminated coal mine
drainage. After reviewing the background of passive treat-
ment and the methods used in these studies (Chapter 1),
the chemical behavior of mine drainage contaminants is
reviewed (Chapter 2). This discussion highlights the dif-
ference between alkaline and acidic mine water, and de-
tails the processes in passive treatment systems that
generate alkalinity. In Chapter 3, contaminant removal is
-------
evaluated for 13 passive treatment systems through the
calculation of contaminant removal rates. These rates,
which incorporate the size of the treatment system, the
flow rate of the water, and mine drainage chemistry, are
the only measures of treatment system performance that
can be reliably compared between systems. In Chapter 4,
the chemical background provided in Chapter 2 and the
observed contaminant removal rates presented in Chap-
ter 3 are combined in a model that gives design and sizing
recommendations for future passive treatment systems.
Chapter 5 summarizes the results of this study and iden-
tifies future research needs,
BACKGROUND OF PASSIVE TREATMENT
The current interest in passive treatment technologies
can be traced to two independent research projects that
indicated that natural Sphagnum wetlands caused an
amelioration of mine drainage pollution without incurring
any obvious ecological damage (4-$)- These observations
prompted the idea that wetlands might be constructed for
the intentional treatment of coal mine drainage. Research
efforts were initiated by West Virginia University^ Wright
State University, Pennsylvania State University, and the
USBM to evaluate the feasibility of the idea. As a result
of promising preliminary reports (<$-#), experimental wet-
lands were built by mining companies and reclamation
groups. Initially, most of these wetlands were constructed
to mimic Sphagnum moss wetlands. However, Sphagnum
moss was not readily available, proved difficult to trans-
plant, and tended to accumulate metals to levels that were
toxic to the Sphagnum after several months of exposure to
mine drainage (9-lff). Instead of abandoning the concept,
researchers experimented with different kinds of con-
structed wetlands. Eventually, a wetland design evolved
that proved tolerant to years of exposure to contaminated
mine drainage and was effective at lowering concentrations
of dissolved metals. Most of these treatment systems con-
sist of a series of small wetlands (<1 ha) that are vege-
tated with cattails (Typha latifolia) (11-12). In northern
Appalachia, many wetlands contain a compost and lime-
stone substrate in which the cattails root. In southern
Appalachia, most wetlands have been constructed without
an exogenous organic substrate; emergent plants have been
rooted in whatever soil or spoil substrate was available on
the site when the treatment system was constructed (13).
Recently, treatment technologies have been developed
that do not rely at all on the wetland model that the early
systems were designed to mimic. Ponds, ditches, and rock-
filled basins have been constructed that are not planted
with emergent plants, and in some cases, contain no soil or
organic substrate (14). Pretreatment systems have been
developed where acidic water contacts limestone in an
anoxic environment before flowing into a settling pond or
wetland system (IS}. In these cases, the water is treated
with limestone followed by passive aeration; however, the
low cost and chemical behavior of limestone make possible
the construction of wetland systems that should, theo-
retically, require no maintenance and last for decades.
A wide diversity of opinions exist on the merits of pas-
sive treatment systems for mine drainage. Wieder's anal-
ysis of a survey of constructed wetlands conducted by the
Office of Surface Mining (OSM) indicated no strong re-
lationships between concentration efficiency and wetland
design features, leading him to question the feasibility of
the constructed wetland concept (72). In a separate study
by Wieder and his colleagues, measurements of the Fe
content of Sphagnum peat exposed to synthetic acid mine
drainage were used to calculate that an average wetland
system should cease to remove metals after 11 weeks of
operation (16). These negative reports contrast with many
other studies of successful wetlands. Examples include an
Ohio wetland that is treating Fe-contammated mine
drainage effectively in its 8th year of operation (17) and six
Tennessee Valley Authority (TVA) wetlands that have
produced compliance water for at least 4 years (18). A
vast majority of the passive treatment systems constructed
in the United States during the last decade achieve per-
formance that is better than Wieder and his colleagues
would predict, though not necessarily enough to consist-
ently meet effluent limits. Hundreds of constructed wet-
lands discharge water that contains lower concentrations
of metal contaminants than was contained in the inflow
drainage. These improvements in water quality decrease
the costs of subsequent water treatment at active sites and
decrease deleterious impacts that discharges from aban-
doned sites have on receiving streams and lakes. In gen-
eral, the systems that are not 100% effective were im-
properly designed, were undersized, or both. This report
has been prepared so that designers of future systems can
avoid these errors.
ACKNOWLEDGMENTS
The authors thank Holly Biddle for invaluable assist-
ance with a reorganization of this report that occurred be-
tween draft and final versions. Laboratory analyses were
conducted by Mark Wesolowski, Joyce Swank, and Dennis
Viscusi. Adrian Woods, John Odoski, John Klemhenz,
and Robert Neupert assisted with field work. Partial
funding for research described in this report was provided
by the U.S. Office of Surface Mining.
-------
CHAPTER 1. MATERIALS AND METHODS
COLLECTION OF WATER SAMPLES
Water samples were collected at passive treatment
systems from their influent and effluent points, and, if
applicable, between treatment cells within the system.
Raw and acidified (2 mL of concentrated HCl) water sam-
ples were collected in 250 mL plastic bottles at each sam-
pling point. Measurements of pH and temperature were
made in the field with a calibrated Orion SA 270, SA 250
or SA 290 portable pH/ISE meter.3 Alkalinity was meas-
ured in the field using a pH meter and an Orion Total
Alkalinity Test Kit, At sites where participates were vis-
ible in water samples, an extra sample was collected that
was filtered through a 0.22-pra membrane Miter before
acidification. All samples were immediately placed on
ice in ait insulated cooler and returned to the laboratory
within 36 h of collection. Samples were refrigerated at
4° C until analysis.
Substrate pore water samples were collected using a
dialysis method similar to that described by Wheeler and
Oilier (79). Lengths of 6,000-8,000 molecular weight
dialysis tubing were filled with 250 mL of deionized, de-
oxygenated water and buried 30-45 cm deep in the organic
substrate of the wetland. Three weeks later, the dialysis
tubes were retrieved and the contents immediately filtered
through a 0.45-/ira membrane filter. Laboratory experi-
ments established that the chemistry of water within the
sampling tubes equilibrated with surrounding pore water
within 24 h. The 3-week equilibration period was allowed
so that chemical anomalies caused by the burial process
would dissipate. Portions of the filtered water samples
were preserved with NaOH (for dissolved sulfide deter-
minations), HCl (for cation analysis), or were left unprc-
served (for alkalinity, acidity, and sulfate analyses).
ANALYSIS OF WATER SAMPLES
Concentrations of Fe, Mn, Al, Ca, Mg, and Na were
determined in the acidified samples using Inductively
Coupled Argon Plasma Spectroscopy, ICP (Instrumenta-
tion Laboratory Plasma 100 model). The acidified samples
were first filtered through a 0.45-/*m membrane filter to
prevent clogging of the small diameter tubing in the ICP.
Ferrous iron concentrations were determined on acid-
ified samples by the potassium dichromate method (20).
Sulfate concentrations were determined by reaction with
^Reference to specific product* dot* not imply endonement by the
US. Bureau of Mines.
barium chloride (BaCl) after first passing the raw sample
through a cation exchange resin, Thorin was used as the
end-point indicator. Dissolved sulfide species were deter-
mined using a sulfide-specific electrode.
Acidity was determined by boiling a 50-mL raw sample
with 1 mL of 30% H2Oa (hydrogen peroxide), and then
titrating the solution with 0,1 W NaOH (sodium hydroxide)
to pH 8.3 (21). Acidity and alkalinity are reported as
mg-L"1 CaCO3 equivalents.
ANALYTICAL QUALITY CONTROL
For each set of samples for a particular sice, a dupli-
cate, standard, and spike were analyzed for quality control
purposes. The relative standard deviation for the duplicate
was always at least 95%. Percent recovery for the stand-
ards were within 3% of the original standard. Spike recov-
eries were within 5% of the expected values.
FLOW RATE MEASUREMENTS
Mine water flow rates were determined by several
methods. Whenever possible, flow was determined with a
bucket and stopwatch. In all cases, three to five meas-
urements of the time needed to collect a known volume
of water were made at each sampling location, and the
average flow rate of these measurements was reported. At
two sites where flows were occasionally too high to meas-
ure with a bucket (the Latrobe and Piney Wetlands), 0.50
or 0.75 ft H-type flumes were installed and flows were
determined from the depth of water in the flume. At the
Keystone site, flows were determined by measuring the
depth of water in a drainage pipe and then using the
Manning formula for measurement of gravity flow in open
channels (22).
ANALYSIS OF SURFACE DEPOSITS
The chemical composition of surface deposits collected
from several constructed wetlands were determined by the
following procedure. The samples were rinsed with
deionized water, dried at 100° C, and weighed. The acid-
soluble component was extracted by boiling 5 g of dry
sample in 20 mL of concentrated HCl for 2 min. The acid
extractants were filtered and analyzed for metal content
by ICP Spectroscopy and for sulfate content by liquid
chromatography. The acid-insoluble material was dried at
100° C and weighed. The acid-soluble component was
determined by subtracting the dry weight of the insoluble
material from the original dry weight.
-------
CHAPTER 2. CHEMICAL AND BIOLOGICAL PROCESSES
IN PASSIVE TREATMENT SYSTEMS
Coal mining can promote pyrite oxidation and result
in drainage containing high concentrations of Fe, Mn, and
Al, as well as SO4, Ca, Mg, and Na. The solubilities of Fe,
Mn, and Al arc generally very low (<1 mg«Lrl) in nat-
ural waters because of chemical and biological processes
that cause their precipitation in surface water environ-
ments. The same chemical and biological processes re*
move Fe, Mn, and Al from contaminated coal mine drain-
age, but the metal loadings from abandoned minesites are
often so high that the deleterious effects of these elements
persist long enough to result in the pollution of receiving
waters.
Passive treatment systems function by retaining con-
taminated mine water long enough to decrease contam-
inant concentrations to acceptable levels. The chemical
and biological processes that remove contaminants vary
between metals and are affected by the mine water pH
and oxidation-reduction potential (Eh). Efficient passive
treatment systems create conditions that promote the
processes that most rapidly remove target contaminants.
Thus, the design of passive treatment systems must be
based on a solid understanding of mine drainage chem-
istry and how different passive technologies affect this
chemistry.
This chapter provides the basic chemical and biological
background necessary to efficiently design passive treat-
ment systems. The authors begin with a discussion of
acidity and alkalinity because many of the decisions about
how to treat mine water passively depend on determina-
tions of these parameters. Next, the chemistry of Fe, Ma,
and Al in aerobic and anaerobic aquatic environments is
described. Throughout the discussion, chemical and bio-
logical concepts are illustrated with data collected from
passive treatment systems.
ACIDITY
Acidity is a measurement of the base neutralization
capacity of a volume of water. Three types of acidity exist:
proton acidity associated with pH (a measure of free H*
ions), organic acidity associated with dissolved organic
compounds, and mineral acidity associated with dissolved
metals (23), Mine waters generally have a very low dis-
solved organic carbon content, so organic acidity ts very
low. The acidity of coal mine drainage arises from free
protons (low pH) and the mineral acidity from dissolved
Fe, Mn, and Al. These metals are considered acidic be-
cause they can undergo hydrolysis reactions that produce
H*.
+ 1/4O2 + 3/2H20 - FcOOH + 2H* (A)
Fe3* + 2H2O .-* FeOOH + 3H+ (B)
Al3* + SHjO -* A1(OH)3 «• 3H* (C)
Mnr + 1/402 + 3/2H2O -* MnOOH * 2H* (D)
These reactions can be used to calculate the total
acidity of a mine water sample and to partition the acidity
into its various components. The expected acidity of a
mine water sample is calculated from its pH and the sum
of the millicquivalents of acidic metals. For most coal
mine drainages, the calculation is as follows:
2+
3*
Acidwlc = 50(2Fe+/56 + 3Fe*/56
(1)
3A1/27 + 2Mn/55 + 1000(10"pH»
where all metal concentrations are in milligram per liter
and 50 is the equivalent weight of CaCO3, and thus trans-
forms milliequivalent per liter of acidity into milligram per
liter CaCO3 equivalent. For water samples with pH <4.5
(no alkalinity present), equation 1 calculates a mine water
acidity that corresponds closely with measurements of
acidity made using the standard H2O2 method (21). Using
synthetic mine drainages with a wide range of composi-
tions, It was determined that calculated acidities differed
from measured values by less than 10% (table 2),
Equation 1 accurately characterizes mineral acidity for
samples of actual acid mine drainage as well. At one site
where numerous measurements of metal chemistry and
total acidity were made, the mean acidity of samples with
pH <4J5 was 693 tog*!*'1, while the predicted acidities for
these samples averaged 655 mg-V1, a difference of only
6% (figure I).
Equation 1 can be used to partition total acidity into its
individual constituents. When the total acidities of con-
taminated coal mine drainages are partitioned in this
manner, the importance of mineral acidity becomes ap-
parent. A breakdown of the acidic components of three
mine drainages is shown in table 3. At each site, the acid-
ity arising from protons (pH) was a minor component of
the totat acidity. Mine drainage at the Friendship Hill
wetland had extremely low pH (2.7), but the acidity of the
-------
_ 1,500
i
g 1.200
3 900
§
O 600
td
15
^ 3OO
ALKALINITY
When mine water has pH >4.5, it has acid neutralizing
capacity and is said to contain alkalinity. Alkalinity can
result from hydroxyl ion (OH~), carbonate, silicate, bo-
rate, organic llgands, phosphate, and ammonia (25). The
principal source of alkalinity in mine water is dissolved
carbonate, which can exist in a bicarbonate (HCO3~) or
carbonate form (CO32~). Both can neutralize proton
acidity.
300
600
900
MEASURED ACIDITY, mfl'L
UOO
-1
1,500
Figure 1.-Comparison of calculated and measured acidities
for water sample* collected at Friendship Hill wetland.
mine water resulted primarily from dissolved ferric iron
and Al. The Somerset wetland received water with low
pH (3.7), but the acidity of the water resulted largely from
dissolved ferrous iron and Mn. At the Cedar Grove sys-
tem, where the mine water was circumneutial, ferrous iron
accounted for 98% of the acidity, while the hydrogen ion
accounted for <1% of mine water acidity.
H* + HCO3~ -* H2O + CO2 (E)
2H+ + COj" -* H2O + CO2 (F)
In the pH range of most alkaline mine waters (5 to 8),
bicarbonate is the principal source of alkalinity.
The presence of bicarbonate alkalinity in mine waters
that contain elevated ievels of metals is not unusual.
Table 4 shows the chemical composition of 12 mine waters
in northern Appalachia that contain alkalinity and are also
contaminated with ferrous iron and Mn. None are con-
taminated with dissolved ferric iron or Al because the
solubility of these metals is low in mine waters with pH
greater than 5.5 (23-24).
Table 2,-Calculated and measured acidities for synthetic acidic mine water
Synthetic Mine Water Composition1
PH
3.9
3.9
3.6
3.8
Fejt
98
0
0
13
Fe3t
1
0
0
0
Al
0
106
0
47
Mn
0
0
97
42
Calculated2
181
598
192
370
Acidity
Measured3
184
578
186
335
Diff,4
-2%
+3%
+3%
+9%
Measured values are the average of three tests. Metal concentrations are
mg.f1. Acidities are mg-L1 CaC03 equivalent.
2From reaction 1.
30ata determined by the hot H202 acidity method (27).
4(l.00-mea3/cal)x 100.
Table 3.-AoldIc component* of mine drainage Influent at three passive treatment systems
Param
Fe2* ...
Fea+ ...
Al3* . . . .
Mn2+ . . .
PH
l§w Concen-
tration,
mfl'l'1
7
163
58
9
2.6
Friendship Hill
Acid
equivalent,1
mo-L-1
13
434
317
16
112
%of
total
acidity
1
49
36
1
13
Concen-
tration,
mg-C1
193
9
3
59
3.7
Somerset
Acid
equivalent,1
mg-c1
345
24
17
107
10
%of
total
acidity
69
5
3
21
2
Concen-
tration,
mg-L'1
95
<1
<1
2
6.3
Cedar Grove
Acid
equivalent,1
mg-L4
170
<1
<1
4
<1
%of
total
acidity
98
<1
<1
2
<1
'CaCOj equivalents calculated from the stolchiometry of reactions A-D.
-------
Table 4,~-Ch*mlcal composition* of mine drainages that contain high concentrations of alkalinity
Location
Pennsylvania:
Fallston
Latrobe
Sligo
Somerset
Uniontown
PH
6,1
6.3
6.6
6.2
6.5
6.2
6.1
6.4
5.5
$.3
6.1
6.3
AHolinity, t
mg-L4 mg
1S2
300
214
120
106
204
163
263 <
93
27S
255
220
y, Feu, Fe3
•L4 mg-L4 mg-l
d 119 <•
:1 96 <
;1 39 <•
:1 30 <'
si 37 <
:1 102 <
:1 51 <•
:1 32 <
t1 43 <
c1 2 <
;1 29 <
(1 70 <
+, Mn,
-4 mg-L4
1 2
1 2
i 8
t 3
1 1
1 6
t 28
1 1
1 26
1 6
t 9
1 3
S04, 1
ma-L4
1,325
1,260
S30
390
331
1,200
493
620
1,720
750
250
950
^let alkalinity,1
mg-L4
-50
140
130
66
72
15
51
209
-31
265
203
95
1 Alkalinity minus acidity.
Alkalinity and acidity are not mutually exclusive terms.
All of the mine waters shown in table 4 contain both acid-
ity and alkalinity. When water contains both mineral
acidity and alkalinity, a comparison of the two measure-
ments results in a determination as to whether the water
is net alkaline (alkalinity greater than acidity) or net acidic
(acidity greater than alkalinity). Net alkaline water con-
tains enough alkalinity to neutralize the mineral acidity
represented by dissolved ferrous iron and Mn. As these
metals oxidize and hydrolyze, the proton acidity that is
produced is rapidly neutralized by bicarbonate. For waters
contaminated with Fe2+, the net reaction for the oxidation,
hydrolysis and neutralization reactions is
complex because it differs between metals and also
between abiotic and biotic processes.
METAt REMOVAL IN AEROBIC ENVIRONMENTS
Iron Oxidation and Hydrolysis
The most common contaminant of coal mine drainage
is ferrous iron. In oxidizing environments common to
moat surface waters, ferrous iron is oxidized to ferric iron.
Ferrous iron oxidation occurs both abiotically and as a
result of bacterial activity. The stoichiometry of the reac-
tion is the same for both oxidation processes.
Fe2* + ^03 * 2HC03~ -*• FeOOH * y2H2O +2CO2 (G)
Reaction G indicates that net alkaline waters contain
at least 1.8 mg«L~l alkalinity for each 1.0 mg'lr1 of dis-
solved Fe. Waters that contain a lesser ratio are net
acidic, since the oxidation and hydrolysis of the total dis-
solved iron content results in a net release of protons and
a decrease in the pH.
METAL REMOVAL PROCESSES
Oxidation and hydrolysis reactions already discussed
cause concentrations of Fe2*, Fe3*, Mn, and Al to com-
monly decrease when mine water flows through an aerobic
environment. Whether these reactions occur quickly
enough to lower metal concentrations to an acceptable
level depends on the availability of oxygen for oxidation
reactions, the pH of the water, the activity of microbial
catalysts> and the retention time of water in the treatment
system. The pH is an especially important parameter
because it influences both the solubility of raetal hydrox-
ide precipitates and the kinetics of the oxidation and
hydrolysis processes. The relationship between pH and
metal-removal processes in passive treatment systems is
(H)
The pH of the mine water affects the kinetics of both the
abiotic and biotic processes (25-26), When oxygen is not
limiting, the rate of abiotic Fe oxidation slows 100-fold for
every unit decrease in pH, At pH values >8, the abiotic
process is fast (rates are measured in seconds), while at
pH values <5 the abiotic process is slow (rates are
measured in days). In contrast, bacterial oxidation of
ferrous iron peaks at pH values between 2 and 3, while
less activity occurs at pH values >5 (27). The presence of
bicarbonate alkalinity buffers mine water at a pH of 6 to
7, a range at which abiotic iron oxidation processes should
dominate. Waters containing no alkalinity have a pH <4.5
and the removal of Fe under oxidizing conditions occurs
primarily by bacterial oxidation accompanied by hydrolysis
and precipitation.
The effect that pH can have on the mechanism of iron
oxidation is shown by the data in figure 2. Samples were
collected from two mine drainages that were both con-
taminated with ferrous iron, but had different pH and
alkalinity values. The samples were returned to the lab-
oratory and exposed to aerobic conditions. For the cir-
cumneutral waters, oxidation of ferrous iron occurred at a
-------
40
30
20
10
10 20 30 40 50
80
60
40
20
KEY
a Unfllterwt
Filtered
0 2 4 6 6 tO 12
TIME.h
Flaur* 2,-ftwnoval of f*2' from aoMlo and alkalln* mm*
watara In laboratory •xparlmant. Raw mbM dralnaga was eol-
ktttad from A, aoldio Latrob* tfta; B, alkaUM Cadar Orova tit*.
Spltta of aaoh aamplv waro flKw-ttarllhad {0-22-Ain fitter). Th*
Lattob* aamplaa w*r* tftakan throughout *x0*rimant; air waa
bubbtod through Cadar Orov* aamptot during axpartmont
rate of 18 mg'L~l*h-], while the rate for the raw acidic
samples was only 1.4 mg»V*'h-1. To evaluate the signi-
ficance of bacterial processes in iron oxidation, splits of
both samples were filter-sterilized (0.22-pm membrane
filter) before the experiment was begun. Removal of bac-
teria had ao effect on the oxidation of ferrous iron for the
circumneutral water, but completely inhibited ferrous iron
oxidation for the acidic water,
As ferrous iron is converted to ferric iron, it is sub-
ject to hydrolysis reactions that can precipitate it as a
hydroxide (reaction B). The hydrolysis reaction occurs
abioticaUy; catalysis of the reaction by microorganisms has
not been demonstrated. The solubility of the ferric hy-
droxide solid is such that, under equilibrium conditions,
negligible dissolved ferric iron (<1 ing'L'1) exists unless
the pH of the mine water is <2.5. In actuality, the rate of
the hydrolysis reaction is also pH dependent, and sig-
nificant Fcs* can be found in mine water with a pH above
2.5. Singer and Stumm (25) suggested a fourth-order rela-
tionship with pH, which indicated that ferric iron hydrol-
ysis processes shift from a very rapid rate at pH >3 to a
very slow rate at pH <2.5. Figure 3 shows the relation-
ship between pH and concentrations of Fe3* at a site
where pH varied by almost 3 units. Ferric iron was not
generally indicated unless the pH was <4, and the highest
*tu
35
30
L 25
f 20
+ • 13
n
l£ 10
5
i
— -r 1 r- r-
•
•
Vi
* • »» 1 *
• %nL
•(•II •
•*c
ji^ • *
m 3* •
•* -f**iiJlf*lL--'T--L M •
'34367
FIELD, pH
Figure 3.— Conoontntfone of F*3* and fl*ld pH for w«Mr
•amplM oolltotod from Emlwton wttkmd.
concentrations of ferric iron occurred when the pH was
<3.
The tendency for dissolved iron to oxidize and hydro-
lyze in aerobic environments with pH >3 results in the
precipitation of ferric hydroxide. Because the net result of
the oxidation and hydrolysis process is the production of
protons, the process can decrease pH. Thus, natural or
constructed wetlands receiving circumneutral net acidic
water commonly decrease both Fe concentrations and pH.
An example of this phenomenon is shown in figure 44.
M water flowed through the constructed wetland, iron
concentrations decreased from 95 to 15 mg*L~S and pH
decreased from 55 to 3.2. Figure 4B shows Fe concen-
trations and pH within a wetland that received mine water
with a net alkalinity. Despite the removal of 60 mg« I/1
Fea* and the production of enough protons to theoret-
ically lower the pH to 2.7, the pH did not decrease
because bicarbonate alkalinity neutralized the proton
acidity.
Manganese Oxidation and Hydrolysis
Manganese undergoes oxidation and hydrolysis reac-
tions that result in the precipitation of manganese oxy-
hydrorides. The specific mechanising) by which Mn2'
precipitates from aerobic mine water in the absence of
chemical additions is uncertain. Mn3* may be oxidized to
either a +3 or a +4 valance, either one of which rapidly
precipitates (reaction D). If MnOOH precipitates, over
time it likely oxidizes to the more stable MnOa. In alka-
line environments, Mn1* can precipitate as a carbonate,
which may also be oxidized by oxygen to MnO2 (25).
Mn2* * HC03~ •* MnCO3 * H* (I)
MnCO3 * HO3 -* Mn02 + CO2 (J)
-------
Regardless of the mechanism by which Mn2+ is oxidized
to Mn*% the removal of one mole of Mna* from solution
results in the release of two moles of H* or an equivalent
decrease in alkalinity (HCO3").
The kinetics of Mn2* oxidation reactions are strongly
affected by pH. Abiotic oxidation reactions are very slow
at pH <8 (24}. Microorganisms can catalyze Mn2* oxida-
tion, but their activity is limited to aerobic waters with pH
>6 (29}. '
Although the hydrolysis of Mn produces protons, the
precipitation of MnOOH does not result in large declines
in pH as can happen when FeOOH precipitates. This dif-
ference between Mn and Fe chemistry is because of the
fact that no natural mechanism exists that rapidly oxidizes
Mn2* under acidic conditions. If pH falls below 6, Mn2'
oxidation virtually ceases, the proton-producing hydrolysis
reaction ceases, and pH stabilizes.
The oxidation and precipitation of Mn2* from solution
is accelerated by the presence of MnO2 and FeOOH (24,
30). Both solids reportedly act as adsorption surfaces for
Mn2* and catalyze the oxidation mechanism. While addi-
tions of FeOOH to Mn-contaimng water might accelerate
Mn oxidation, the direct precipitation of FeOOH from
mine water containing Fe1' does not generally stimulate
1ZO
90
60
30
0
r
£
150
120
90
60
30
n
\f \j" \r v- v- Ji-
r ' ' 1
0
* — ""*
KEY
*-*^ A-FeTOT
**•* ,,
N* **pH
N^
s
X
SAMPLING STATION
Figure 4.-Cono«ntritioni of Fe101 and field pH at two con-
structed wetlands. A, Emlanton wetland; 8, Cedar Grove wetland.
Mil-removal processes in passive treatment systems. Fig-
ure 5 shows concentrations of Mn and Fe for mine water
as it flowed through a constructed wetland that markedly
decreased concentrations of both metals. On average, Fc
decreased from 150 to <1 mg»l/S while Mn decreased
from 42 to 11 mg»lA Removal of metals occurred se-
quentially, not simultaneously. Two-thirds of the decrease
in iron concentration occurred between the first and
second sampling stations. The wetland substrate in this
area was covered with precipitated FeOOH and the water
was turbid with suspended FeOOH. Despite the presence
of large quantities of FeOOH, little change in the con-
centration of Mn occurred between the first and second
sampling station. The slight decrease in Mn that occurred
was proportionally similar to the change in Mg, suggesting
that dilution was the most likely cause of the decrease in
Mn concentrations (the use of Mg to estimate dilution is
discussed In detail in chapter 3). Between stations 3 and
5, there was little Fe present in the water and little visual
evidence of FeOOH sludge on the wetland substrate.
Most of the observed removal of Mn occurred in this Fe-
free zone.
The absence of simultaneous precipitation of dissolved
Fe and Mn from aerobic alkaline waters likely results from
the reduction of oxidized forms of Mn by ferrous iron,
2*
2+
MnOj + 2Fe* + 2H2O •* 2FeOOH + Mn+ + 2H* (K)
or
MnOOH + Fe2* •+ FeOOH * Mn2* (L)
Figure 6 shows the results of a laboratory study that
demonstrate the instability of Mn oxides in the presence
of ferrous iron. Water samples and Mn-oxidcs were
150
125
ICO
75
50
25
\ KEY
\ *-^ J«
\ ««a Mn
\ ^5 '
, V .
60
30
16
ALP
ElMjOl
ditch
Pond
W*ilond
Mo. (
SAMPLING STATION
Figure 5.—Mean concentration* of Pe, Mn, and Mg at tne
Morrison Wetland. Mine water flow* linearly from station 1 to
station S. VeiUota bars are one standard error of the mean.
-------
10
bU
501
40
30'
LJ 20
eft
6 10
2
Q 0
S 60
£
UJ 501
| ^O
30'
20
10
0
1 r- (- i
A Mine water only
1 •/ / ^ f"
**x>^ _ £i
1111
9-^^^ Mine water + MnOOH
KEY
I .^ pfiZ-f _
,k * Mn
^.x-> A A ^
20 40 60 80 100
TIME, h
Flfluro 6,—Changw In concentration* of F«2t and Mn2*. A,
abtanct; B, prtMnc* of MnOOH. Mln* wat*r wat collected from
Influent pfp« of Blair watiand. MnOOH wac coltocted from Inild*
of final effluent pipe.
collected from a wetland that removed Fe and Ma in a
sequential manner. The wetland influent was alkaline
(pH 6.2, 162 rag* I/1 alkalinity) and contaminated with
50 mg'L"1 Fe and 32 rag-L'1 Mn. Two flasks of mine
water received MnO2 additions, while the controls did not
receive Mn02. Concentrations of dissolved Fe and Mn
were monitored in each flask over a 73-h period. In all
flasks, concentrations of Fe decreased to <1 mg'L'1. In
the control flasks, concentrations of Fe decreased to
<3 mg"L-1 within 43 h. In flasks that received MnO2,
concentrations of Fe decreased to <3 mg'lr1 in only
22 h. No change in concentrations of Mn occurred in the
control flasks. Concentrations of Mn in the MnO2 flasks
increased by 15 mg-L"1 during the first 22 h and did not
change during the remaining 50 h of the experiment. The
association of accelerated precipitation of Fe with
solubitization of Mna+ suggests that the MnO2 oxidized
Fez* in a manner analogous to reaction K.
The data presented in figures 5 and 6 demonstrate
aspects of Fe and Mn chemistry that are important in
passive treatment systems. Iron oxidizes and precipitates
from alkaline mine water much more rapidly than does
Mn. One reason for the differences in kinetics is that the
oxidized Mn solids, which are presumed to result from
Mn:t oxidation reactions, are not stable in the presence
of Fe2*. Concentrations of ferrous iron must decrease to
very low levels before Mn2* oxidation processes can result
in a stable solid precipitate. In the absence of Fe2t, Mn
removal is still a very slow process under laboratory con-
ditions. Conditions in a wetland may either accelerate
Mn-removal reactions or promote mechanisms that are not
simulated in simple laboratory experiments. However,
both field and laboratory investigations indicate that, under
aerobic conditions, the removal of Mn occurs at a much
slower rate than does the removal of Fe (empirical evi-
dence for this concept is presented in chapter 3).
MINE WATER CHEMISTRY IN ANAEROBIC
ENVIRONMENTS
Chemical and microbial processes in anaerobic envi-
ronments differ from those observed in aerobic envi-
ronments. Because O2 is absent, Fe2* and Mn2* do not
oxidize and oxyhydroxide precipitates do not form. Hy-
droxides of the reduced Fe and Mn ions, Fe(OH)2 and
Mn(OH)2, do not form because of their high solubility
under acidic or circumneulral conditions. In passive treat-
ment systems where mine water flows through anaerobic
environments, its chemistry is affected by chemical and
biological processes that generate bicarbonate and hydro-
gen sulfide.
Limestone Dissolution
A major source of bicarbonate in many anaerobic en-
vironments is the dissolution of carbonate minerals, such
as calcite.
CaCO3
Ca
HCO
(M)
Carbonate dissolution can result in higher concen-
trations of bicarbonate in anaerobic mine water environ-
ments than aerobic environments for two reasons. First,
the absence of Fe3* in most anaerobic environments limits
the formation of FeOOH coatings that armor carbonate
surfaces and inhibit further carbonate dissolution in aero-
bic environments (32). Second, the solubilities of carbon-
ate compounds are directly affected by the partial pressure
of dissolved CO2 (23-24, 32). Anaerobic mine water en-
vironments commonly contain high CO2 partial pressures
because of the decomposition of organic matter and the
neutralization of proton acidity.
The observation that limestone dissolution is enhanced
when contact with mine water occurs in an anaerobic
environment has resulted in the construction of anaerobic
limestone treatment systems. The first demonstration of
-------
11
this technology was by Turner and McCoy (75) who
showed that when anoxic acidic mine water was directed
through a plastic-covered buried bed of limestone, it was
discharged in an alkaline condition. After exposure to the
atmosphere metal contaminants precipitated from this
alkaline discharge much faster than they did from the
original acid discharge.
Since Turner and McCoy described their findings in
1990, dozens of additional limestone treatment systems
have been constructed (33-35). These passive mine water
pretreatment systems have become known as anoxic
limestone drains or ALD's. In an ALD, mine water is
made to flow through a bed of limestone gravel that has
been buried to limit inputs of atmospheric oxygen. The
containment caused by the burial also traps CO2 within the
treatment system, allowing the development of high COZ
partial pressures (5(5).
Water quality data from an ALD in western Penn-
sylvania are shown in table 5 and figure 7. This ALD is a
rectangular bed of limestone gravel that is 37 m long by
6 m wide by 1 m deep. The limestone bed is covered with
filter fabric and 1 m of day. No organic matter was
incorporated into the limestone system. Water samples
were collected from the ALD influent and effluent and at
four locations within the ALD. The influent mine water
contained high concentrations of ferrous iron and Mn and
a small amount of alkalinity. As the mine water flowed
through the ALD, pH and concentrations of calcium and
alkalinity increased while other measured parameters were
unchanged. Between the influent and effluent locations,
changes in concentrations of alkalinity (137 mg-L'1) and
Ca (58 mg-L*1) were in stoichiometric agreement with
those expected from CaCO3 dissolution.
Table s.—Chemistry of mine water flowing through the Howe
Bridge anoxic limestone drain, January 23,1992
Parameter
pH
Alkalinity ...
Ca
Fe2+
Fe3+
Mn
Al
Mg
Na
sex
CO,
In
5.9
39
140
249
<1
34
<1
90
11
1175
6.3
WetM
6.1
75
150
237
<1
33
<1
87
11
1175
4.0
Well 2
6.4
141
183
246
<1
34
<1
91
11
1200
4.7
Well 3
6.5
179
201
246
O
34
<1
91
11
1150
4.3
Well 4
6.5
183
206
245
<1
34
<1
90
11
1200
4.7
Eff
6.3
176
198
?44
<1
34
•Ca
.v\ .v1*- ,V1> v.h
«»N .lie* iitft* .mV
Figure 7.~-Concentrations of Ca, and alkalinity for water as tt
flows through the Howt Bridge ALD. Water flows linearly from
Influent to effluent
Dissolution of CaCO3 within the ALD was greater than
would be expected from an open system in equilibrium
with atmospheric concentrations of CO2 (0.035%). An
equilibrated open system would only produce alkalinity in
the range of 50 to 60 mg*L~'} and increase Ca concen-
trations by 4 to 8 mg'L'1. Observations of elevated CO2
gas concentrations within the ALD, and the higher sol-
ubility of CaCO3 within the ALD indicate that the ALD
acts as a closed system.
Concentrations of alkalinity and Ca changed little be-
tween the third well and the ALD effluent. This obser-
vation suggests that water within the ALD was already in
equilibrium with CaCO3 by the time it reached the third
well location. Thus, the amount of alkalinity that can be
generated by this ALD is limited to a maximum value that
is a function of the CO: partial pressures within the ALD.
Similar observations of solubility-limited alkalinity gen-
eration by an ALD have also been made at a second site
in western Pennsylvania (36).
Sulfate Reduction
When mine water flows through an anaerobic envi-
ronment that contains an organic substrate, the water
chemistry can be affected by bacterial sulfate reduction.
In this process, bacteria oxidize organic compounds using
sulfate as the terminal electron sink and release hydrogen
sulfide and bicarbonate,
2CH20
.2-
2HCO,
(N)
where CH2O is used to represent organic matter. Bac-
terial sulfate reduction is limited to certain environmental
-------
12
conditions (37). The bacteria require the presence of sul-
fatc, suitable concentrations of low-molecular weight car-
bon compounds, pH >4, and the absence of oxidizing
agents such as Oj, Fe3* and Mn<+. These conditions are
commonly satisfied in treatment systems that receive coal
mine drainage and contain organic matter. High concen-
trations of sulfate (>200 mg-L'1) are characteristic of
contaminated coal mine drainage. The oxygen demand of
organic substrates causes the development of anoxic con-
ditions and an absence of oxidized forms of Fe or Mn.
The low-molecular weight compounds that sulfate-reducing
bacteria utilize (tactate, acetate) are common end products
of microbial fermentation processes in anoxic environ-
ments. The pH requirements can be satisfied by alkalinity
generated by microbial activity and carbonate dissolution.
Bacterial sulfate reduction directly affects concentra-
tions of dissolved metals by precipitating them as metal
sulfide solids.
M2+ + H2S + 2HCO3*" •* MS + 2H2O + 2CO2 (O)
For Fe, the formation of pyrite is also possible
2H +
(P)
The removal of dissolved metals as sulfide compounds
depends on pH, the solubility product of the specific metal
sulfide, and the concentrations of the reactants. The sol-
ubilities of various metal sulfides are shown in table 6.
Laboratory studies have verified that metal removal from
mine water subjected to inflows of hydrogen sulfide occurs
in an order consistent with the solubility products shown
in table 6 (39). The first metal sulfide that forms is CuS
followed by PbS, ZnS, and CdS. FeS is one of the last
metal sulfides to form. MnS is the most soluble metal
sulfide shown and is expected lo form only when the con-
centrations of all other metals in the table are very
low(«lrng«L'1).
For coal mine drainage, where metal contamination is
generally limited to Fe, Mn, and Al, the hydrogen sulfide
produced by bacterial sulfate reduction primarily affects
dissolved iron concentrations. Aluminum does not form
any sulfide compounds in wetland environments and the
relatively high solubility of MnS makes its formation
unlikely.
Table 0.-SolublHty products of some metal sulfides
CdS
CuS
FsS
MnS
NJS
PbS
ZnS
Solubility product
1,4 x 10*
4.0 x 1QT*
1.0 x Iff1*
5.6 x Iff"
3.0 X 10'n
1.0x10*
4,5 X 10"M
'See reference 36.
The precipitation of metal sulfides in an organic sub-
strate improves water quality by decreasing the mineral
acidity without causing a parallel increase in proton acidity.
Proton-releasing aspects of the H2S dissociation process
(HaS -» 2H" + S*-) are neutralized by an equal release of
bicarbonate during sulfate reduction. An organic substrate
in which 100% of the HjS produced by sulfate reduction
precipitated as FeS would have no effect on the mine
water pH or alkalinity (although acidity would decrease).
In fact, however, the chemistry of pore water in wetlands
constructed with an organic substrate characteristically
has pH 6 to 8 and is highly alkaline (40-41). These alka-
line conditions result, in part, from reactions involving
hydrogen sulfide that result in the net generation of bicar-
bonate. Hydrogen sulfide is a very reactive compound that
can undergo a variety of reactions in a constructed wet-
land. In most wetlands (constructed and natural), surface
waters are aerobic while the underlying pore waters in
contact with organic substrate are anaerobic. When sul-
fidic pore waters diffuse from the organic substrate into
zones that contain dissolved ferric iron, dissolved oxygen,
or precipitated Fe and Mn oxides, the hydrogen sulfide can
be oxidized (table 7). These reactions affect the mineral
acidity and the alkalinity in various manners.
Table 7.~Slnk* for HjS In constructed wetland* wid their net effect on mint
water acidity and alkalinity
Ha8
H2S
HjS
H7S
HjS
H,S
H^S
+ 2HCCy
+ 2HCCV
+ 2HCCV
+ 2HCCV
+ 2HC
-------
13
Table 8 shows the chemistry of surface water and sub-
strate pore water samples collected from a wetland con-
structed with limestone and spent mushroom compost.
Spent mushroom compost consists of a mixture of spoiled
hay, horse manure, corn cobs, wood chips, and limestone.
At the wetland used in this example, 10 to 15 cm of lime-
stone sand was covered with 20 to SO cm of compost and
planted with cattails. Water flowed through the wetland
primarily by surface paths; no efforts were made to force
the water through the compost. This design is typical
of many compost wetlands constructed in northern
Appalachia during the last 10 years. The data shown in
table 8 were collected 15 months after the wetland was
constructed.
Table 8,—Surface and pore water chemittry
at the Latrobe wetland
Parameter Pore water1 Surface water2
Al
Ca
Fe2+
H2S
MB
Mn
Na
S04
Acidity3
Alkalinity
Net Alkalinity4 . . .
pH
Mean
1
467
215
2
37
175
24
11
1,674
493
885
392
6.8
Std dev
5
188
183
9
75
48
10
10
532
340
296
NAp
.8
Mean
35
308
73
24
166
42
5
1,967
503
o
-503
3.1
Std dev
5
29
39
16
0
9
2
1
115
86
o
NAp
.1
NAp Not applicable.
Std dev Standard deviation.
1 A total of 52 water samples wen collected on July 25 and
August 11, 1988, by the dialysis tube method. Metals were ana*
lyzed for every sample. Reid pH was measured for 29 samples.
Alkalinity was measured for nine samples.
2 Six samples collected In July and August 1988.
3 Calculated from pH, Fe2*, Fe3*, Al, Mn, and HjS for pore
water samples and measured by the H^ method for surface
water samples.
4 Average alkalinity minus average acidity. The nine pore
water samples for which alkalinity was measured had a mean net
alkalinity of 653 mg/L (aid dev - 590).
Surface water at the study site had low pH and high
concentrations of Fe, Al, and Mn (table 8). Compared
with the surface water, the substrate pore water had higher
pH, higher concentrations of alkalinity, ferrous iron,
calcium, and hydrogen sulfide, and lower concentrations of
sulfate, ferric iron, and aluminum. On average, the pore
water had a net alkalinity while the surface water had a
net acidity. The alkalinity of the pore water appeared to
result from a combination of limestone dissolution and
sulfate reduction. The average alkalinity calculated to
result from these processes was 703 mg*L-1, a value that
corresponded reasonably well with the measured difference
in acidity, 895 mg«L'1.6
Compared with surface water, substrate pore water
contained elevated concentrations of ferrous iron. High
concentrations of Fe2+ likely resulted from the dissolution
of ferric oxyhydroxides at the redox boundary. FeOOH
can be reduced by direct heterotropbic bacterial activity
CH2O + 4FeOOH + H2O -* 4Fe2+ + 8OH" + CO2 (Q)
and also by H2S that results from sulfate reduction.
H2S + 2FeOOH -» 2Fe2* + 4OH" + S° (R)
In both cases, the solubilization of ferric hydroxides results
in the release of OH~, which acts to raise pH to cir-
cumneutral levels and also reacts with dissolved CO2 to
form bicarbonate. Reduction of ferric hydroxide has no
effect on the net acidity of the mine water because the
increase in alkalinity is exactly matched by an increase in
mineral acidity. If the Fe-enriched pore water diffuses
into an aerobic zone, the ferrous iron content should
oxidize, hydrolyze, and reprecipitate as ferric oxyhydroxide.
4Fe2* + 8OH' + O2 -* 4FeOOH + 2H2O (S)
Because the pore water has circumneutral pH and is
strongly buffered by bicarbonate, the removal of iron by
oxidation processes from pore water as it diffuses into
aerobic surface waters should occur rapidly. Indeed,
during the summer months, when the data in table 8 were
collected, comparisons of the wetland influent and effluent
indicated that the wetland decreased both concentrations
of iron and total acidity on every sampling day (figure 8).
The decrease hi acidity indicates that alkaline pore water
was mixing with surface water and neutralizing acidity.
The decrease in concentrations of Fe in the surface water
indicates that elevated concentrations of Fe1* observed in
the pore water were rapidly removed in surface water
environments.
ALUMINUM REACTIONS IN MINE WATER
Aluminum has only one oxidation state in aquatic
systems, +3. Oxidation and reduction processes, which
complicate Fe and Mn chemistry, do not directly affect
'The difference between surface and pore water concentrations of
sutfate averaged 293 mg*!*"1, which is equivalent to 305 mg'L"1
CaCO3 alkalinity (reaction N); the difference in calcium concentrations
averaged 159 mg«L~l, which is equivalent to 398 mg-L"1 CaCO3
alkalinity (reaction M).
-------
14
250
200
150
100
50
0
Q
O
1,200
1.000 -
800
600
400
300
0
B
o-'
•O—O*'
KEY
o Influent
• Effluent
- o-
Figure 8.—Influent and effluent concentration* at the Utrobe wetland during the summer of 1988. A, Fa; fl, acidity.
concentrations of dissolved Al. Instead, concentrations of
Al in mine waters are primarily influenced by the solubility
of A1(OH)3 (23, 43), At pH levels between 5 and 8,
A1(OH)3 is highly insoluble and concentrations of dissolved
Al are usually <1 tng-L"1. At pH values <4, Al(OH), is
highly soluble and concentrations >2 mg-L"1 are possible.
The passage of mine water through highly oxidized
or highly reduced environments has no effect on
concentrations of Al unless the pH also changes. In those
cases where the pH of mine water decreases (due to iron
oxidation and hydrolysis), concentrations of Al can in-
crease because of the dissolution of alumino-silicate days
by the acidic water. When acidic mine water passes
through anaerobic environments, the increased pH that
can result from carbonate dissolution or microbial activity
causes the precipitation of A1(QH)3.
CHAPTER 3. REMOVAL OF CONTAMINANTS BY PASSIVE TREATMENT SYSTEMS
Chapter 2 described chemical and biological processes
that decrease concentrations of mine water contaminants
in aquatic environments. The successful utilization of
these processes in a mine water treatment system depends,
however, on their kinetics. Chemical treatment systems
function by creating chemical environments where metal
removal processes are very rapid. The rates of chemical
and biological processes Chat underlie passive systems are
often slower than their chemical system counterparts and
thus require that mine water be retained longer before it
can be discharged. Retention time is gained by building
large systems such as wetlands. Because the land area
available for wetlands on minesites is often limited, the
sizing of passive treatment systems is a crucial aspect of
their design. Unfortunately, in the past, most passive
treatment systems have been sized based on guidelines
that ignored water chemistry or on available space, rather
than on comparisons of contaminant production by the
mine water discharge and expected contaminant removal
by the treatment system. Given the absence of quantita-
tive sizing standards, wetlands have been constructed that
are both vastly undersized and oversized.
In this chapter, rates of contaminated removal are
described for 13 passive treatment systems in western
Pennsylvania. The systems were selected to represent the
wide diversity of mine water chemical compositions that
exist in the eastern United States. The rates that are
reported from these sites are the basis of treatment system
sizing criteria suggested in chapter 4.
The analytical approach used to quantify the perform-
ance of passive treatment systems in this chapter differs
from the approach used by other researchers in several
respects. First, contaminant removal is evaluated from a
rate perspective, not a concentration perspective. Second,
changes in contaminant concentrations are partitioned into
two components: because of dilution from inputs of fresh-
water, and because of chemical and biological processes in
the wetland. In the evaluations of wetland performance,
only the chemical and biological components are consid-
ered. Third, treatment systems, or portions of systems,
were included in the case studies only if contaminant
concentrations were high enough to ensure that contam-
inant removal rates were not limited by the absence of the
contaminant. These unique aspects of the research are
discussed in further detail below.
EVALUATION OF TREATMENT SYSTEM
PERFORMANCE
To make reliable evaluations of wetland performance,
a measure should be used that allows comparison of con-
taminant removal between systems that vary in size and
the chemical composition and flow rate of mine water they
receive. In the past, concentration efficiency (C£%) has
been a common measure of performance (11-12), Using
iron concentration as an example, the calculation is
-------
15
CE% -
Fe,... - Fc.
"in
'eft
Fe
X 100
(2)
in
where the subscripts "in" and "eff represent wetland in-
fluent and effluent sampling stations and Fe concentra-
tions are in milligram per liter.
Except in carefully controlled environments, CE% is a
very poor measure of wetland performance. The efficiency
calculation results in the same measure of performance
for a system that lowers Fe concentrations from 300 to
100 mg-L'1 as one that lowers concentrations from 3 to
1 mg'L'1. Neither the flow rate of the drainage nor the
size of the treatment system are incorporated into the cal-
culation. As a result, the performances of systems have
been compared without accounting for differences in flow
rate (which vary from <10 to >1000 L-mur1) or for dif-
ferences in system size (which vary from <0.1 to >10 ha)
(72).
A more appropriate method for measuring the per-
formance of treatment systems calculates contaminant
removal from a loading perspective. The daily load of
contaminant received by a wetland is calculated from the
product of concentration and flow rate data. For Fe, the
calculation is
Fe (g • d'1)^ * 1.44 X flow (L • min"1)
X Fe (mg
(3)
where g*d*' is gram per day and 1.44 is the unit conver-
sion factor needed to convert minutes to days and milli-
grams to grams.
The contaminant load is apportioned to the down flow
treatment system by dividing by a measure of the system's
size. In this study, treatment systems are sized based on
their surface area (SA) measured in square meter,
Fe (g - m-2 - d-X - Fe (g - d'^/SA. (4)
The daily mass of Fe removed by the wetland between two
sampling stations, Fe(g* d'1)^, is calculated by comparing
contaminant loadings at the two points,
(g
g'Ota-CFeg-d'V (5)
An area-adjusted daily Fe removal rate is then calculated
by dividing the load removed by the surface area of the
treatment system lying between the sampling points,
Fe(g
• m
-2
Fe(g- d-^/SA. (6)
To illustrate the use of contaminant loading and con-
taminant removal calculations, consider the hypothetical
water quality data presented in table 9.
In systems A and B, changes in Fe concentrations are
the same (60 mg-L'1), but because system B receives four
times more flow and thus higher Fe loading, it actually
removes four times more Fe from the water. The concen-
tration efficiencies of the two wetlands are equivalent, but
the masses of Fe removed are quite different.
Data are shown for system C for three sampling dates
on which flow rates and influent iron concentrations vary.
On the first date (Cl), the wetland removes all of the Fe
that it receives. On the next two dates (C2 and C3), Fe
loadings are higher and the wetland effluent contains Fe.
From an efficiency standpoint, performance is best on the
first date and is worst on the third date. From an Fe-
removal perspective, the system is removing the least
amount of Fe on the first date. On the second and third
dates, the wetland removes similar amounts of iron (2,880
and 3,024 g*d"*). Variation in effluent chemistry results,
not from changes in wetland's Fe-removal performance,
but from variation in influent Fe loading.
Table 9.-HypothetIcal wetland data and performance evaluations
A
R
C1 .
C2.
C3 ,
n .
In
Eff
CE
Wetland
System size,
m2
400
400
500
, 500
, 500
750
Influent.
Effluent.
Concentration efficiency.
Flow rate
Lfnln4
10
40
30
80
150
50
Fe Concentration
In
mg-L'1
100
100
40
35
30
100
Eff
40
40
10
16
25
Fe Loading
In
Kg'd'1
1.4
5.8
17
4.0
6.5
7.2
Kg-tfl
0.6
2.3
1.2
3.5
1.8
Fe removal
performance
CE
60
60
99
71
47
75
Rate
g-m^-d'1
2.2
8.6
3.5
5,8
6.0
7.2
-------
16
Lastly, consider a comparison of wetland systems oE dif-
ferent sizes. System D removes more iron than any wet-
land considered (5,400 g'd'1), but it is also larger. One
would expect that, all other factors being equal, the largest
wetland would remove the most Fe. When wetland area
is incorporated into the measure by calculating area-
adjusted Fe removal rates (gram per square meter per
day), System B emerges as the most efficient wetland
considered.
DILUTION ADJUSTMENTS
Contaminant concentrations decrease as water flows
through treatment systems because chemical and biolog-
ical processes remove contaminants from solution and
because the concentrations are diluted by inputs of fresh-
water. To recognize and quantify the removal of contam-
inants by biological and chemical processes in passive
treatment systems, it is necessary to remove the effects
of dilution. Ideally, studies of treatment systems include
the development of detailed hydrologic and chemical
budgets so that dilution effects are readily apparent. In
practice, the hydrologic information needed to develop
these budgets is rarely available, except when systems
are built for research purposes. Treatment systems con-
structed by mining companies and reclamation groups are
rarely designed to facilitate flow measurements at all water
sampling locations, so estimating dilution from hydrologic
information? ^highly inaccurate or impossible.
An alternative method for distinguishing the effects
of dilution from those of chemical and biological processes
is through the use of a conservative ion (44-45). By de-
finition, the concentration of a conservative ion changes
between two sampling points only because of dilution or
evaporation. Changes in concentrations of contaminant
ions that proportionately exceed those of conservative ions
can then be attributed to biological and chemical wetland
processes.
In this study, Mg was used as a conservative ion. Mag-
nesium was considered a good indicator of dilution in
these systems for both theoretical and empirical reasons.
In northern Appalachia, concentrations of Mg in coal mine
drainage are often >50 mg-L'1, while concentrations in
rainfall are <1 mg-L'1 and in surface runoff are usually
<5 mg-L'1, Magnesium is unlikely to precipitate in pas-
sive treatment systems because the potential solid pre-
cipitates, MgSO4, MgCO3, and CaMg^COj)^ do not form
at the concentrations and pH conditions found in the
systems (23). While biological and soil processes exist that
may remove Mg in wetlands, their significance is negligi-
ble relative to the high Mg loadings that most mine water
treatment systems in northern Appalachia receive. The
average Mg loading for wetland systems included in this
study was -7,000 g Mg-nr^yr1. The uptake of dis-
solved Mg by plants in constructed wetlands can only
account for 5 to 10 g Mg-nr^yr1. This estimate as-
sumes that the net primary productivity of the constructed
wetlands is 2,000 g«nr2«yrl dry weight (46) and that the
Mg content of this biomass is 0.25% to 0.50% (47). The
estimate ignores mineralization processes that would
decrease the net retention of Mg to lower values. Most
constructed wetlands have a clay base that can adsorb Mg
by cation exchange processes, but the total removal of
Mgby this process is limited to about 100 g*m~2. This
estimate assumes that the mine water is in contact with a
5-cm-deep clay substrate that has a density of 1.5 g'cnr3,
a cation exchange capacity of 25 meq per 100 g, and 50%
of the available sites are occupied by Mg (48). These con-
servative calculations indicate that less than 2% of the
annual Mg loading at the study sites is likely affected by
biological and soil processes within the systems.
Empirical data also indicate that Mg is conservative in
the wetlands monitored in this study. Table 10 shows
influent and effluent concentrations of major noncontam-
inant ions at eight constructed wetlands. No precipita-
tion had occurred in the study area for 2 weeks previous
to collection of the samples, so dilution from rainfall,
surface water, or shallow ground water seeps was minimal.
Magnesium was the most conservative ion measured.
Concentrations of Mg changed by <5% with flow through
every wetland, while concentrations of all other ions mon-
itored changed by at least 15% at at least one site.
Tabl* 10.—Influwrt and •fflumt conowitratlofia of Ca, Mg, Na, and tulfato at eight constructed w*ttand*
Donegal .....
Emlenton ....
FH
Qourtey
Plney B
Somerset ....
h'.i
mg-L*1
244
429
122
117
244
416
355
307
Ca
Eff,
mg-L'1
241
433
189
120
256
426
354
469
Change,
%
.1
+1
+55
+3
+14
+2
0
+53
In,
mg-i:1
81
308
51
114
127
251
217
312
Mg
EH,
mg-L'1
79
306
51
117
125
262
216
312
Chang*,
%
-2
-1
0
+3
-2
+4
0
0
ta'4
mg'L-1
6
11
5
3
6
15
27
6
Na
Eff,
mg-l'1
6
10
7
4
11
16
27
7
Chang*,
%
0
-2
4-2
+6
+8
+4
-2
+15
In,
mg-L*1
729
2,810
1,125
1,000
1,523
2,190
2,050
2,740
SO,
Eff
mg'L'1
729
2,770
842
1,030
1,226
2,120
2,100
2,300
Chang*,
%
0
-1
-25
+3
-20
-3
+2
-16
Eff Effluent.
In Influant.
FH Friendship Hill National Historic*! SH*.
-------
17
Changes in concentrations of Mg were used to adjust
for dilution effects by the following method. For each set
of water samples from a constructed wetland, a dilution
factor (DF) was calculated from changes in concentrations
of Mg between the influent and effluent station:
(7)
Contaminant concentrations were adjusted to account for
dilution using the DF. When only an influent flow rate
was available, the chemical composition of the effluent
water sample was adjusted. For Fe, the adjustment cal-
culation was
(8)
where AFe^ is expressed in milligram per liter. When
only an effluent flow rate was available, the chemical com-
position of the influent water sample was adjusted,
AFeDA.(FeinxDF)-Feetf.
(9)
Because most of the DF values were <1.00, the adjust-
ment procedures generally resulted in smaller estimates of
changes in contaminant concentrations than would have
been calculated without the dilution adjustment.
Rates of contaminant removal, expressed as gram per
square meter per day, were then calculated from the
dilution-adjusted change in concentrations, the flow rate
measurement liter per minute, and the SA of the system,
in square meter
Fe(g • m
"2
X 1.44 )/SA.
LOADING LIMITATIONS
(10)
A primary purpose of this chapter is to define the
contaminant removal capabilities of passive treatment
systems. Accurate assessments of these capabilities re-
quire that the treatment systems studied contain excessive
concentrations of the contaminants. A system that is com-
pletely effective (lowers a contaminant to <2 mg-L'1)
may provide an indication that contaminant removal occurs
(if dilution is not the cause of concentration changes), but
cannot provide an estimate of the capabilities of the re-
moval processes, as the rate of contaminant removal may
be limited simply by the contaminant loading rate. For
example, in table 9, the removal rate of Fe for wetland Cl
is 3.5 g>m~2>d~1. This rate is not an accurate estimate
of the capability of the wetland to remove Fe because
the loading rate on this day was also only 3.5 g-nr'-d-1.
The data from Cl are not sufficient to estimate whether
the wetland could have removed 10 or 100 g»nra-d'1 of
Fe. Only when the wetland is overloaded with Fe (days
C2 and C3), can the Fe removal capabilities of the wetland
be assessed.
The Morrison passive treatment system demonstrates
the necessity of recognizing both dilution and loading-
limiting situations in the evaluation of the kinetics of metal
removal processes. The Morrison system consists of an
anoxic limestone drain followed by a ditch, a settling pond,
and two wetland cells. Figure 5, previously presented in
chapter 2, shows average concentrations of Fe, Mn, and
Mg at the sampling stations. Iron loading and removal
rates for the sampling stations are shown in table 11. The
treatment system decreased concentrations of Fe from
151 mg'L"1 at the system influent station (the ALD dis-
charge) to <1 mg'ir1 at the final wetland effluent sta-
tion. Most of the change in Fe chemistry occurred in the
ditch, a portion of the system that only accounted for 4%
of the total treatment system SA. Calculations of the rate
of Fe removal based on the entire treatment system re-
sulted in a value of 13 g*m~2>d~1. Because this removal
rate is equivalent to the load, it does not represent a
reliable approximation of the system's Fe-removal capa-
blity. Only when an Fe removal rate is calculated for the
ditch, an area where Fe loading exceeded Fe removal,
does an accurate assessment of the Fe removal capabilities
result.
Table 11 .-Average concentrations of Fe, Mn, and Mg at the Morrison passive treatment system
Station
Ditch Effluent ....
Pond Effluent
Final Effluent
Cumulative
area, mz
0
43
461
1,076
Flow,
L-rn'1
6.6
NA
NA
NA
Concentration,
nmV1
Fe
151
56
5
<1
Mn
42
37
24
71
Mg
102
91
72
71
Removal rate1,
gtn-Sr1
Fe
NA
19.2
2-3
1.3
Mn
NA
0.17
0.14
0.13
NA Not available.
*F«movaJ rate based on cumulative area.
-------
18
Concentrations of Mn at the Morrison effluent station
were generally above discharge limits. Manganese was
detectable in every effluent water sample (>.4 mg»L~l)
and >2 mg'L'1 in 75% of the samples. Thus, it was
reasonable to evaluate the kinetics of Mn removal based
on the SA of the entire treatment system. Concentrations
of Mg, however, decreased with flow through the treat-
ment system, suggesting an important dilution component.
Effluent water samples contained, on average, 31% lower
concentrations of Mg than did the influent samples. On
several occasions when the site was sampled in conjunc-
tion with a rainstorm, differences between effluent and in-
fluent concentrations of Mg were larger than 50%. Meas-
urements of metal removal by the Morrision treatment
system that did not attempt to account for dilution would
significantly overestimate the actual kinetics of metal
removal processes.
Dilution adjustments were possible for every set of
water samples collected from a treatment system because
concentrations of Mg were determined for every water
sample. Problems with loading limitations, however, could
not be corrected at every site. At two sites where com-
plete removal of Fe occurred, the Blair and Donegal wet-
lands, the designs of the systems were not conducive for
the establishment of intermediate sampling stations. For
these two systems, no Fe removal rates were calculated
because complete removal of Fe occurred over an unde-
termined area of treatment system.
STUDY SITES
The design characteristics of the 13 passive treatment
systems monitored during this study are shown in table 12.
At four of the sites, acidic mine water was pretreated with
anoxic limestone drains (ALD's) before it flowed into
constructed wetlands. The construction materials for the
wetlands ranged from mineral substances, such as clay and
limestone rocks, to organic substances such as spent mush-
room compost, manure, and hay bales. Cattails (Typha
latifolia and, less commonly, T. angustifolia) were the most
common emergent plants growing in the systems. Three
sites contained few emergent plants. Most of the wetland
systems consisted of several cells or ponds connected seri-
ally. Two systems, however, each consisted of a single
long ditch.
The mean influent flow rates of mine drainage at the
study sites ranged from 7 to 8,600 L-mur1 (table 12).
The highest flow rates occurred where drainage discharged
from abandoned and flooded underground mines. The
lowest flow rates occurred at surface mining sites. Esti-
mated average retention times ranged from 8 h to more
than 30 days.
The average chemistry of the influents to the 16 con-
structed wetlands are shown in table 13. Data from 15
sampling points are shown. At the REM site, two dis-
charges are treated by distinct ALD-wetland systems that
eventually merge into a single flow. The combined flows
are referred to as REM-Lower. Mine water at the Howe
Bridge system U characterized at two locations. The
"upper" analysis describes mine water discharging from an
ALD that flows into aerobic settling ponds. The "lower"
analysis describes the chemistry of water flowing out of the
last settling pond and into a large compost-limestone
wetland that is constructed so that mine water flows in a
subsurface manner.
Table 1&p-Construction characteristics of the constructed wetlands
Constructed
Sit* year
Cedar 1988
Keystone 1969
Blair 1969
Shade 1969
Ptney 1987
Morrison 1990
Emlenton 1987
Somerset 1984
Howe 1991
latrobe 1987
REM 1992
FH 1988
Design
Pond, 8 Cells
S Cells
Ditch
Ditch
ALD, 2 Oils
iCell
ALD,3Ce«s
9 Cells
2 Cells
ALD, 3 Cells
3 Cells
2ALDs,9Cells
e Cells
Substrate
LS, SMC
Clay, LS
Topsoll
Manure, straw
LS
HB
Clay, manure
LS, manure
HB. LS, SMC
day, LS, SMC
HB, LS, SMC
SMC
LS, SMC
Emergent
vegetation
Typha
..do.
None
Mixed
None
Mixed
Typfta
..do.
..do.
None
TfctfW
..do.
..do.
^
nr
8,100
1,360
4,200
1,080
880
2,600
1,075
643
1,006
3.000
2,800
4,849
667
Water
depth,
om
15
16
100
6
10
50
30
90
15
50
15
30
15
Flow
rate,1
Lfnln4
501
156
8,606
11
10
468
7
55
47
130
86
206
15
Est. ret
time*
days
1.7
o.g
.3
3.4
8.4
1.9
33.9
4.1
2.2
8.0
3.4
4.9
4.6
Est Estimated.
FH Friendship Hill National Historical Site.
HB HaybeJee.
LS Limestone.
rst Retention.
SA Surface area of wet area.
SMC Spent mushroom compost.
1 Average values.
2 Calculated from the water holding capacity and Influent flow rate.
-------
19
Tteto 13.-Aver«g« chemical characteristic* of influent water at the constructed wetland*
{•He* «ra arranged according to the net acidity)
Site
REM - R
FH
Number at
samples
29
26
28
12
20
39
34
20
13
40
43
13
g
43
18
73
pH
7.1
6,3
6.3
6.2
6.0
5.8
6,3
6.1
5.6
4.7
4.4
6.2
3.5
3.5
5.5
2.6
Composition, mg-L'1
m
202
336
142
166
31
60
271
128
22
15
0
160
0
0
57
0
Fe
5
92
37
52
<2
1
150
190
185
89
162
272
246
125
473
153
Mn
8
2
30
22
15
42
50
34
77
50
39
92
32
130
9
At
8
3
2
43
3
58
Mg
81
54
14
77
125
225
102
118
91
249
193
105
171
125
232
85
SO4
738
1,251
330
645
966
1,845
1,087
1,275
1,128
2,317
1,691
1,315
1,875
1,655
2,495
1,733
Net Acidity, w
mg-L*1
-182
-140
-73
-51
-17
-6
75
258
312
320
373
375
496
617
867
929
Alk Alkalinity,
FH Friendship Hill National Historical Site.
'03003 equivalent.
2Negative values Indicate alkaline conditions.
Ten of the influents to the constructed wetlands had pH
>5 and concentrations of alkalinity >25 mg-L"1. The
alkaline character of five of these discharges resulted from
pretreatment of the mine water with ALD's. The high
concentrations of alkalinity contained by five discharges
not pretreated with ALD's arose from natural geochcmical
reactions within the mine spoil (Donegal and Blair) or the
flooded deep mine (Cedar, Keystone, and Piney). For
mine waters that contained appreciable alkalinity, the
principal contaminants were Fe and Mn,
Concentrations of alkalinity for six of the influents
were high enough to result in a net alkaline conditions
(negative net acidity in table 13). A seventh alkaline
influent, Morrison, was only slightly net acidic. For these
seven influents, enough alkalinity existed in the mine
waters to offset the mineral acidity associated with Fc
oxidation and hydrolysis.
Nine of the influents were highly acidic. Five of the
acidic influents contained alkalinity, but mineral acidity
associated with dissolved Fe and Mn caused the solutions
to be highly net acidic. These inadequately buffered
waters were contaminated with Fe and Mn. Four of the
waters contained no appreciable alkalinity (pH <4.5) and
high concentrations of acidity. Mine waters with tow pH
were contaminated with Fe, Mn, and Al.
EFFECTS OF TREATMENT SYSTEMS
ON CONTAMINANT CONCENTRATIONS
The effects of the treatment systems on contaminant
concentrations are shown in table 14. Every system de-
creased concentrations of Fe. At four sites where the
original mine discharge contained elevated concentrations
of Fe, the final discharges contained <1 rag*!/1. Nine of
the systems decreased Fe concentrations by more than
50mg*L~l. The largest change in Fe occurred at the
Howe Bridge system where concentrations decreased by
197 mg-L"1. From a compliance perspective, the most
impressive decrease in Fe occurred at the Morrison system
where 151 mg-L"1 decreased to <1 mg-L'1.
Fourteen of the passive systems received mine water
contaminated with Mn. Eleven of these systems decreased
concentrations of Mn. Changes in Mn were smaller than
changes in Fe. The largest change in Mn concentration,
31 mg-L"1, occurred at the Morrison site. Only the
Donegal treatment system discharged water that con-
sistently met effluent criteria for Mn (<2 mg-L"1). Botb
the Shade and Blair wetland effluents flowed into settling
ponds which discharged water in compliance with regu-
latory criteria. On occasions, the discharges of the
Morrison and Piney treatment systems met compliance
criteria.
Every wetland system decreased concentrations of
acidity. The Morrison system, which received mine water
that contained 75 mg-L-1 acidity, always discharged net
alkaline water. None of the constructed wetlands that
received highly acidic water (net acidity >100 mg-L *)
regularly discharged water with a net alkalinity. During
low-flow periods, the Somerset, Latrobe, and FH systems
discharged net alkaline water. The largest change in
acidity occurred at the Somerset wetland where concen-
trations decreased by an average 304 mg-L*1.
DILUTION FACTORS
While contaminant concentrations decreased with flow
through every constructed wetland, concentrations of Mg
also decreased at many of the sites. Decreases in Mg
-------
20
indicated that part of the improvement in water quality
was because of dilution. Average dilution factors for the
treatment systems are shown in table 15. For 9 of the 17
systems, average dilution factors were 0.95 to 1.00 and
dilution adjustments were minor. At the remaining eight
systems, mean DF values were less than 0.95 and dilution
adjustments averaged more than 5%. Water quality data
from the Morrison and Somerset constructed wetlands
were adjusted, on average, by more than 25%.
Dilution factors varied widely between sampling days.
Dilution adjustments were higher for pairs of samples
collected in conduction with precipitation events or thaws.
Every system was adjusted by more than 5% on at least
one occassion (see minimum dilution factors in table 15).
Adjustments of more than 20% occurred on at least one
occasion at 13 of the 17 study sites.
Pew dilution adjustments were >1.00 (see maximum
dilution factors in table 15). Of the 390 dilution factors
that were calculated for the entire data set, 13 exceeded
1.05. These high dilution factors could have resulted from
evaporation or freezing out of uncontaminated water with-
in the treatment system, from temporal changes in water
chemistry, or from sampling errors. Most of the high
dilution factors were associated with rainstorm events, sug-
gesting temporal changes in water quality. When dilution
factors were >1.00, the calculated rates of contaminant
removal were greater than would have been estimated
without any dilution adjustment. Because of the limited
number of sample pairs with high dilution factors, then-
presence did not markedly affect the average contaminant
removal rates for the constructed wetland study areas.
Tabto 14.—Mean water quality lor sampling stations at the constructed wtttanda
Site
Donegal
Cedar
Keystone
Blair
Shade
Plney
Morrison
RHM-L
Emtonton
Somerset
Howe
REM-Lower
Lalrobe
REM-R
FH
Sampling
station
Pond Influent
Wetland Influent
Effluent
Influent
Effluent
Influent
Effluent
Influmt
Effluent
1C influent
LC effluent
Setp
Wetland influent
Wetland effluent
Influent
Ditch
Effluent
left influent
Lett effluent
Influent
Effluent
Influent
Effluent
Influents2
Upper effluent
Lower effluent
Influent
Effluent
Influent
Cell 3 effluent
Right influent
Right effluent
Influent
Effluent
n1
6
29
28
26
27
26
28
12
8
20
20
21
39
39
24
24
24
20
20
46
40
43
40
13
13
13
9
9
43
43
18
18
73
73
PH
6.4
7.1
7.4
6.3
6.4
6.3
6.4
6.2
7.0
6.0
6.8
6.4
9.8
6.1
6.3
6.4
6.6
6.1
3.8
4.7
3.2
4.4
5.5
6.0
5.6
6.2
3.5
2.9
3.5
3.7
5.5
3.3
2.6
2.9
Fe
34
5
<1
92
41
37
32
52
<1
2
<1
32
1
<1
151
56
<1
190
84
89
15
162
18
265
185
68
246
115
125
56
473
338
153
137
Mn
9
8
2
2
2
1
1
30
5
23
10
25
15
11
42
37
11
50
48
77
73
50
33
37
34
33
92
88
32
29
130
113
10
10
Acldfty
NAp
NAp
NAp
NAp
NAp
NAp
NAp
NAp
NAp
NAp
NAp
NAp
NAp
NAp
78
64
-1
256
225
320
271
373
68
373
312
112
496
436
617
343
867
712
929
674
MO
83
81
80
54
53
14
14
77
59
128
122
201
225
225
102
91
71
118
112
249
234
193
139
101
91
91
171
166
125
122
232
201
85
85
FH Friendship Hill National Historical Site.
LC Limestone cell,
NAp Not applicable.
'Number of samples.
*Tne flow-weighted average of two discharges.
-------
21
Tab!* 15.-OUution factors for th* constructed wetlands
Site
Donegal
Cedar ........
Keystone
Blair
Shads
Plney
Morrison Ditch . .
Morrison Wetland
REM-L
Howe tower
Emlerrton
Somerset ......
Howe Upper
REM-Lower
Latrobe
REM-R
FH
Average
0.99
0.99
0.99
0.83
0.96
1.00
0,87
0.69
0.95
1.00
0.94
0.73
0.89
0.93
0.95
0.86
1.00
sd
0.05
0,03
0-04
0.10
0.08
0.06
0.18
0.25
0.09
0.10
0.09
0.30
0.08
0.09
0.08
0.16
0.12
Minimum
0.76
0.92
0.91
0.70
0.76
0.92
0.40
0.27
0.70
0.80
0.66
0.30
0.73
0.72
0.75
0.36
0.58
Maximum
1.04
1.05
1.15
1.01
1.09
1.31
1.05
1.12
1.13
1.25
1.04
1.76
0.99
1.01
1.14
1.00
1.34
FH Friendship Hill National Historical Site.
REMOVAL OF METALS PROM AUCAUNE
MINE WATER
Rates of Fe and Mn removal for the study systems are
shown in table 16. Significant removal of Fe occurred at
every study site. Fe removal rates were directly correlated
with pH and the presence of bicarbonate alkalinity (fig-
ure 9). These two water quality parameters are closely
related because the buffering effect of bicarbonate alka-
linity causes mine waters with >SO mg*L alkalinity to
typically have a pH between 6.0 and 6.5. Within the group
of sites that received alkaline mine water, there was not a
significant relationship between the Fe removal rate and
the concentration of alkalinity.
Removal of Fe at the alkaline mine water sites ap-
peared to occur principly through the oxidation of ferrous
iron and the precipitation of ferric hydroxide (reaction A,
chapter 2). Mine water within the systems was turbid
with suspended ferric hydroxides. By the cessation of the
studies, each of the alkaline water sites had developed
thick accumulations of iron oxyhydroxides. Laboratory
experiments, discussed in chapter 2, demonstrated that
abiotic ferrous iron oxidation processes are rapid in aer-
ated alkaline mine waters. No evidence was found that
microbially-mediated anaerobic Fe removal processes,
which require the presence of an organic substrate, con-
tributed significantly to Fe removal at the alkaline sites.
Fe removal rates at the REM wetlands, which were con-
structed with fertile compost substrates, did not differ
from rates at sites constructed with mineral substrates
(Morrison, Howe-Upper, Keystone).
Rates of Fe removal averaged 23 g-m^-d"1 at the six
sites that contained alkaline, Fe-contaminated water. Four
of the alkaline systems displayed similar rates despite
widely varying flow conditions, water chemistry and sys-
tem designs. The Keystone system, a deep plantless ditch
that towered Fe concentrations in a very large deep mine
discharge by 5 mg ir1, removed Fe at a rate of
21 g-m*J«d'1. The shallow-water Morrison ditch, which
decreased concentrations of Fe in a low-flow seep by al-
most 100 mg'L-1, had an average Fe removal rate of
19 g-ra-3-d-1. The REM-L and REM-R wetlands, which
were constructed almost identically, but received water
with contaminant concentrations and flow rates that var-
ied by 200%, displayed Fe removal rates of 20 and
TabUk 16.—Fe and Mn removal rates at constructed wetland
Site
Fe removal rate
Mn removal rate
Cedar
Blair
Shade
REM-L
Howe-Upper ,
REM-R
FH
Mean
NAp
6.3
20.7
NAp
NAp
NAp
19.2
NAp
28.3
8.1
9.1
S.O
42.7
12.0
2.1
20.1
0.5
Std dev
NAp
2.2
5.1
NAp
NAp
NAp
10.6
NAp
5.7
1.9
3.3
4.9
8.2
3.4
1.0
4.0
0.5
n
NAp
7
15
NAp
NAp
NAp
24
NAp
20
13
39
34
13
9
21
18
73
slg?1
NAp
yee
yes
NAp
NAp
NAp
yee
NAp
yes
ye*
yea
yes
yes
yes
yes
yes
yes
Mean
0.50
0.17
NAp
0.43
0.72
1.07
0.17
0.20
-0.05
0.06
-0-09
-0.01
-0.43
-0.05
0.03
0.10
0.00
Std dev
0.25
0.41
NAp
0.37
0.64
1.34
0.41
0.18
0.13
0.16
0.19
0.54
0.49
0.14
0.09
0,33
0.02
n
9
7
NAp
6
17
33
24
24
20
13
39
34
13
9
21
18
73
slg?
yes
no
NAp
yes
yes
yes
yes
V»
no
no
no
no
no
no
no
no
no
NAp Not applicable.
FH Friendship Hill National Historical Site.
n Sample size.
slg? Significant at 0.05 level.
Std dev Standard deviation.
i, rate Is significantly greater than zero (West); no, rate is not significantly greater than zero (west).
-------
22
Two alkaline mine water sites varied considerably from
the other sites in their Fe removal capabilities. The Cedar
Grove wetland removed Fe at a rate of 6 g«m*J"d~\
while the Howe Bridge Upper site removed Fe at a rate
of 43 g"DV*»d'1. The Cedar Grove system consists of a
series of square cells that may have more short-circuiting
flow paths than the rectangular-shaped cells of the other
systems. The Cedar Grove system also contains less aera-
tion structures than the other systems. Mine water at the
site upweUs from a flooded underground mine into a pond
that dicharges into a three-cell wetland. Limited topo-
graphic relief prevented the inclusion of structures that
efficiently aerate the water (i.e., waterfalls, steps). The
Howe Bridge Upper system, in contrast, very effectively
aerates water. Drainage drops out of a 03-m-high pipe,
Hows down a cascading ditch and through a V-notch weir
before it enters a large settling pond. Because the rate of
abiotic ferrous iron oxidation is directly proportional to
the concentration of dissolved oxygen, insufficient oxygen
transfer may explain the low rate of Fe removal at the
Cedar Site, while exceptionally good oxygen transfer at the
Howe Bridge Upper site may explain its high rate of Fe
removal.
At sites where the buffering capacity of bicarbonate
alkalinity exceeded the mineral acidity associated with iron
hydrolysis, precipitation of Fe did not result in decreased
pH. This neutralization was evident at the Morrison,
Cedar, Keystone, Blair, Piney, and Donegal sites (ta-
ble 14). At the Howe Bridge and REM wetlands, the
mine water was insufficiently buffered and iron hydrolysis
eventually exhausted the alkalinity and pH fell to low
levels. The effluents of both REM systems had pH <3.5.
The Howe Bridge Upper system discharged marginally
alkaline water (<25 mg-L'1 alkalinity; pH 5.6). Spot
checks of the pH of surface water 20 m into the Howe
Bridge Lower wetland (which receives the Upper system
effluent) always indicated pH values <3.5.
Significant removal of Mn only occurred at five of the
constructed wetlands (table 13). Each of these sites re-
ceived alkaline mine water (figure 10). Each site also
either received water with low concentrations of Fe (Piney
and Shade) or developed low concentrations of Fe within
the treatment system (Blair, Donegal, and Morrison).
40
30
20
^-
TJ
M '°
'E
/ I .
L I
* H-L- j
. , • , *
,-.
N
1
en
^ °2 3 4 5 6 7 J
^ INFLUENT pH §
> rn °
UJ
°= 40
o>
u.
30
20
10
n
i . > i i i >
.* I
,
I
, I
I 5
S
or
c
s
I.O
1.2
.9
.6
.3
.0
-.3
'
™
. L_
-G2 3
1.2
.9
.3
.0
-.6
-f,
* *
' B I
' I1*
|FeJ80-j r^e
' *li
^iFe^O
, ,
i.ii
L
\
1 l
i t \l
ill £
. . .1 .
4567
INFLUENT pH
• i i i i
<)
<1 I {Fe300 lFe>50 •
Ire^oU
JF«>ISO
*
8
1
;
Fa>40
,
-50 0 50 100 150 200 250 300 350 400
INFLUENT ALKALINITY
Figure 9.—Relationship between mean Fe removal rate* and
A, mean Influent pH and B, mean Influent alkalinity concen-
tration*. Vertical bar* are one standard error above and below
the mean. "H-L" le the Howe-Lower srte.
-50 0 50 100 150 200 250 300 350 400
INFLUENT ALKALINITY
Figure 10.—Relationship between mean Mn removal ratea and
A, mean Influent pH and B, mean Influent alkaJtntty concen-
tration*. Vertical bars are one standard error above and below
mean. Fe value* next to the bar* are effluent Fe2* value*.
-------
23
Alkaline sites that contained high concentrations of Fe
throughout the treatment system (Howe-Upper, REM-L,
REM-R, and Cedar), did not remove significant amounts
of Mn. The Morrison ditch, which contained water with
an average 56 mg'lr* Fe, had a significant Mn removal
rate. This rate, however, was derived from an average
dilution-adjusted decrease in Mn concentrations of only
1.2 mg-L"1 or 3% of the influent concentrations. Because
of uncertainties with sampling, analysis, and dilution-
adjustment procedures that could reasonably bias Mn data
by 2-3%, the authors do not currently place much practical
confidence in this value.
The five sites that markedly decreased concentrations
of Mn had variable designs. The Donegal wetland has a
thick organic and limestone substrate and is densely veg-
etated with cattails. The Blair and Morrison wetlands
contain manure substrates and are densely vegetated with
emergent vegetation. The Finey wetland was not con-
structed with an organic substrate and includes deep open
water areas and shallow vegetated areas. The Shade treat-
ment system contains limestone rocks, no organic sub-
strate, and few emergent plants. Thus, chemical aspects
of the water, not particular design parameters, appear lo
principally control Mn removal in constructed wetlands.
The removal of Mn from aerobic mine waters appeared
to result from oxidation and hydrolysis processes. Black
Mn-rich sediments were visually abundant in the Shade,
Donegal, and Blair wetlands. As discussed in chapter 2,
the specific mechanism by which these oxidized Mn solids
form is unclear. The amorphous nature of the solids pre-
vented identification by standard X-ray diffraction meth-
ods. However, samples of Mn-rich solids collected from
the Shade and Blair wetlands were readily dissolved by
alkaline ferrous iron solutions, indicating the presence of
oxidized Mn compounds.
Mn2* caa reportedly be removed from water by its
sorption to charged FeOOH (ferric oxydroxide) particles
(23, 30). If this process is occurring at the study wetlands,
it is not a significant sink for Mn removal. The bottoms
of the Morrison ditch, Howe-Upper, Cedar, REM-L, and
REM-R wetlands were covered with precipitated FeOOH
and the mine water within these wetlands commonly con-
tained 5 to 10 mg'I"J of suspended FeOOH (difference
of the Fe content of unfiltered and filtered water samples).
After mine water concentrations were adjusted to reflect
dilution, no removal of Mn was indicated at four of the
sties and very minor removal of Mn occurred at the fifth
site (Morrison ditch).
Although the processes that remove Mn and Fe from
alkaline mine water appears to be mechanistically similar
(both involve oxidation and hydrolysis reactions), the ob-
served kinetics of the metal removal processes arc quite
different. In the alkaline mine waters studied, Mn removal
rates were 20 to 40 times slower than Fe removal.
The presence or absence of emergent plants in the wet-
lands did not have a significant effect on rates of either Fe
or Mn removal at the alkaline mine water sites. In gen-
eral, bioaccum ulation of metals in plant biomass is an
insignificant component of Fe and Mn removal in con-
structed wetlands (49). The ability of emergent plants lo
oxygenate sediments and the water column (50) has been
proposed as an important indirect plant function in wet-
lands constructed to treat polluted water (57). Either
oxygeuation of the water column is not a rate limiting
aspect of metal oxidation at the constructed wetlands that
received alkaline mine water, or physical oxygen transfer
processes are more rapid than plant-induced processes.
REMOVAL OF METALS AND ACIDITY
FROM ACID MINE DRAINAGE
Metal removal was slower at constructed wetlands that
received acidic mine water than at those that received
alkaline mine water. Removal of Mn did not occur at any
site that received highly acidic water (figure 10). Removal
of Fe occurred at every wetland that received acidic mine
water, but the Fe removal rates were less than one-half
those determined at alkaline wetlands (figure 9). Because
abiotic ferrous iron oxidation processes are extremely slow
at pH values <5, virtually all the Fe removal observed at
the acidic sites must arise from direct or indirect microbial
activity. Microbially-mediated Fe removal under acidic
conditions is, however slower than abiotic Fe-remova!
processes under alkaline conditions.
Wetlands that treat acidic mine water must both pre-
cipitate metal contaminants and neutralize acidity. At
most wetland sites, acidity neutralization was the slower
process. At the Emlenton and REM wetlands, Fe removal
processes were accompanied on every sampling occasion
by an increase in proton acidity which markedly decreased
pH (see figure 44, chapter 2). Mine water pH occasion-
ally decreased with flow through the Latrobe and Somerset
wetlands. Thus, for the wetlands included in this study,
the limiting aspect of acid mine water treatment was the
generation of alkalinity or the removal of acidity (which
were considered in this report to be equivalent, sec chap-
ter 2). The best measure of the effectiveness of the acid
water treatment systems was through the calculation of
acidity removal rates.
Acidity can be neutralized in wetlands through the
alkalinity-producing processes of carbonate dissolution and
bacterial suifate reduction. As was discussed in chapter 2,
the presence of an organic substrate where reduced Eh
conditions develop promotes both alkalinity-generating
processes. In highly reduced environments where dis-
solved oxygen and ferric iron are not present, carbonate
surfaces are not passivated by FeOOH armoring. Decom-
position of the organic substrate can result in elevated
-------
24
partial pressures of CO2 and promote carbonate disso-
lution. The presence of organic matter also promotes the
activity of sulfatc-reducing bacteria.
The rates of alkalinity generated from these two
processes in the constructed wetlands were determined
based on dilution-adjusted changes in the concentrations
of dissolved Ca and sulfate, the stoichiometry of the
alkalinity-generating reactions, and measured flow rates.
The calculations are based on the assumption that Ca con-
centrations only increase because of carbonate dissolution
and that sulfate concentrations only decrease because of
bacterial sulfate reduction. One possible error in this
approach is that sulfate can co-precipitate with ferric
hydroxides in low-pH aerobic environments (52), The Fe
and sulfate content of surface deposits collected from the
constructed wetlands indicate that sulfate is incorporated
into the precipitates collected from acidic environments
at an average Fe:SO4 ratio of 9.7 (table 17). If all of
the Fe removed from mine water is assumed to precipitate
as ferric hydroxide with a Fe:SO4 ratio of 9.7:1, then
changes in sulfate concentrations attributable to the co-
precipitation process amount to only 5 to 30 mg«L' at
the acid mine water sites. Dilution-adjusted changes in
sulfate concentrations at the Somerset, Latrobe, Friendship
Hill (FH), and Howe-Lower wetlands were commonly 200
to500mg-IA
Rates of acidity removal, sulfate removal and calcium
addition for six constructed wetlands that received acidic
mine water are shown in table 18. Significant removal of
acidity occurred at all sites. The lowest rates of acidity
removal occurred at the Emlenton wetland. This site con-
sists of cattails growing in a manure and limestone sub-
strate. No sulfate reduction was indicated (the rate was
not significantly >0). Dissolution of the limestone was
indicated, but the rate was the lowest observed.
Table 17.-f« and S04 content of ferric oxyhydroxld* deposits;
•Res are arranged by pH
Site
PH
Composition, ppm dry weight
Latrobe . , .
Somerset . .
Cedar
Keystone . .
3.0
.... 3.S
3.5
.... 6.4
e.e
Fe
471,779
266,039
461,583
362,300
398,337
S04
64,213
27,991
48,263
8,946
6,886
Fe:S04
7.4
10.3
9.6
40,5
57.8
1 Field pH measured where substrate sample collected.
The Latrobe, Somerset, FH, Howe-Lower, and REM
systems were each constructed with a spent mushroom
compost and limestone substrate. Spent mushroom com-
post is a good substrate for microbial growth and has a
high limestone content (10% dry weight). At these five
wetlands, sulfate reduction and limestone dissolution both
occurred at significant rates (table 18). The summed
amount of alkalinity generated by sulfate reduction and
limestone dissolution processes (Reactions M and N,
chapter 2) correlated strongly with the measured rate of
acidity removal at these four sites (r >0,90 at each site).
At the FH wetland, 94% of the measured acidity removal
could be explained by these two processes (figure 11).
On average, sulfate reduction and limestone dissolution
contributed equally to alkalinity generation at these five
sites (51% versus 49%, respectively). The average sulfate
removal rate calculated for the compost sites, 5.2 g
S04~2*m'2*d"S is equivalent to a sulfate reduction rate
of -180 nmol-cnr^d"1. This value is consistent with
measurements of sulfate reduction made at the constructed
wetlands using isotope methods (41) as well as measure-
ments of sulfate reduction made for coastal ecosystems
(55).
Table 1fl.-Av»rag« rates of acidity removal, sulfate removal, and calcium addition at attes receiving acidic mint water
Site
Acidity removal rate
Sulfate removal rate
Calcium addition rate
mean Std dev slg?1 mean Std dev a!0? mean Std dev slg?
Somerset
REM-Lower
Latrobe
FH
25
34
13
g
. , 21
72
3 1
99
154
7 1
6.9
7.0
24
8.6
41
72
4.4
3.8
yes
1 5
6.1
5.9
3.4
57
5.7
72
24
64
2.6
ye»
0.8
1,7
26
0,9
1.2
1.21
t,20
1 40
1 03
0.07
0.80
yes
FH Friendship Hill National Historical Site.
n Sample size.
Std dev Standard deviation,
'Yea, raw is significantly greater than zero (west); no, rate Is not significantly greater than zero (t-test}.
-------
25
The highest rates of acidity removal, sulfcte reduction,
and limestone dissolution a!! occurred at the Howe-tower
site. This system differs from the others by its subsurface
How system. Drainage pipes, buried in the limestone that
underlies the compost, cause the mine water to flow
directly through the substrate. At the Somerset, Latrobe,
REM, and FH systems, water flows surficially through the
wetlands. Mixing of the acidic surface water and alkaline
substrate waters presumably occurs by diffusion processes
at the surface-flow sites. By directly contacting contam-
inated water and alkaline substrate, the Howe-Lower site
is extracting alkalinity from the substrate at a significantly
higher rate than occurs in surface flow systems. How long
the Howe-Upper system can continue to generate alka-
linity at the present rates is unknown. Monitoring of
the system, currently in its third year of operation, is
continuing.
40
2 J 30
< 'E 25
20
15
10
5
-5 0 5 10 15 20 25
MEASURED AClDtTt REMOVAL,
30 35 40
Figure 11 .-Measured rate* of alkalinity generation and acidity
iov«l at the Friendship Hill wetland. Unit* are a»m~z»dT*
removal at the Friendship
CaCOj equivalent
CHAPTER 4. DESIGN AND SIZING OF PASSIVE TREATMENT SYSTEMS
Three principal types of passive technologies currently
exist for the treatment of coal mine drainage: aerobic
wetland systems, wetlands that contain an organic sub-
strate, and anoxic limestone drains. In aerobic wetland
systems, oxidation reactions occur and metals precipitate
primarily as oxides and hydroxides. Most aerobic wetlands
contain cattails growing in a clay or spoil substrate. How-
ever, plantless systems have also been constructed and at
least in the case of alkaline influent water, function sim-
ilarly to those containing plants (chapter 3).
Wetlands that contain an organic substrate are similar
to aerobic wetlands in form, but also contain a thick layer
of organic substrate. This substrate promotes chemical
and microbial processes that generate alkalinity and neu-
tralize acidic components of mine drainage. The term
"compost wetland" is often used in this report to describe
any constructed wetland that contains an organic substrate
in which biological alkalinity-generating processes occur.
Typical substrates used in these wetlands include spent
mushroom compost, Sphafgtum peat, haybales, and
manure.
The ALD is a buried bed of limestone that is intended
to add alkalinity to the mine water (7^ 33-34). The lime-
stone and mine water are kept anoxic so that dissolution
can occur without armoring of limestone by ferric oxy-
hydroxides. ALD's are only intended to generate alka-
linity, and must be followed by an aerobic system in which
metals are removed through oxidation and hydrolysis
reactions.
Each of the three passive technologies is most ap-
propriate for a particular type of mine water problem.
Often, they are most effectively used in combination with
each other. In this chapter, a model is presented that is
useful in deciding whether a mine water problem is suited
to passive treatment, and also, in designing effective pas-
sive treatment systems.
Two sets of sizing criteria are provided (table 19). The
"abandoned mined land (AML) criteria" are intended for
groups that are attempting to cost-effectively decrease
contaminant concentrations. In many AML situations, the
goal is to improve water quality, noi consistently achieve
a specific effluent concentration. The AML sizing criteria
are based on measurements of contaminant removal by
existing constructed wetlands (chapter 3). Most of the
removal rates were measured for treatment systems (or
parts of treatment systems) that did not consistently lower
concentrations of contaminants to compliance with OSM
effluent standards. In particular, the Fe sizing factor for
alkaline mine water (20 g"m"2-d-l) is based on data
from six sites, only one of which lowers Fe concentrations
to compliance.
Table 19.-~Heeommended sizing for paislve treatment system*
AML criteria,
Compliance criteria,
Fe
Mn
Acidity . .
Alkaline
20
1.0
NAp
Acid
NAp
NAp
7
Alkaline
10
0.5
NAp
Acid
NAp
NAp
3.5
NAp Not applicable.
It is possible that Fe removal rates are a function of Fe
concentration; i.e., as concentrations get lower, the size of
-------
26
system necessary to remove a unit of Fe contamination
(e.g., 1 g*d"') gets larger. To account for this possibility,
a more conservative sizing value for systems where the
effluent must meet regulatory guidelines was provided
(table 1). These are referred to as "compliance criteria."
The sizing value for Fe, 10 g-nr^d"1, is in agreement
with the findings of Stark (17) for a constructed compost
wetland in Ohio that receives marginally acidic water.
This rate is larger, by a factor of 2, than the Fe removal
rate reported by Brodie (18} for aerobic systems in
southern Appalachia that are regularly in compliance.
The Mn removal rate used for compliance,
0.5 g»m'2'd~l, is based on the performance of five
treatment systems, three of which consistently lower Mn
concentrations to compliance levels. A higher removal
value, 1 g»m~2»d~l, is suggested for AML sites. Because
the toxic effects of Mn at moderate concentrations
(<50 mg-L"1) are generally not significant, except in very
soft water (54), and the size of wetland necessary to treat
Mn-contaminated water is so large, AML sites with Fe
problems should receive a higher priority than those with
only Mn problems.
The acidity removal rate presented for compost wet-
lands is influenced by seasonal variations that cannot
currently be corrected with wetland design (55). This is
not a problem for mildly acidic water, where the wetland
can be sized in accordance with winter performance, nor
should it be a major problem in wanner climates. In
northern Appalachia, however, no compost wetland that
consistently transforms highly acidic water (>300 mg-L"1
acidity) into alkaline water is known. One of the study
sites, which receives water with an average of 600 mg-L~l
acidity and does not need to meet a Mn standard, has
discharged water that only required chemical treatment
during winter months. While considerable cost savings are
realized at the site because of the compost wetland, the
passive system must be supported by conventional treat-
ment during a portion of the year.
Because long-term metal-removal capabilities of passive
treatment systems are currently uncertain, current Federal
regulations require that the capability for chemical treat-
ment exist at all bonded sites. This provision is usually
met by placing a "polishing pond" after the passive treat-
ment system. The design and sizing model does not cur-
rently account for such a polishing pond.
All passive treatment systems constructed at active sites
need not be sized according to the compliance criteria pro-
vided in table 19. Sizing becomes a question of balancing
available space and system construction costs versus in-
fluent water quality and chemical treatment costs. Mine
water can be treated passively before the water enters a
chemical treatment system to reduce water treatment costs
or as a potential part-time alternative to full-time chemical
treatment. In those cases where both passive and chemical
treatment methodologies are utilized, many operators find
that they recoup the cost of the passive treatment system
in less than a year by using simpler, less expensive chem-
ical treatment systems and/or by decreasing the amount of
chemicals used,
A flow chart that summarizes the design and sizing
model is shown in figure 12. The model uses mine drain-
age chemistry to determine system design, and contam-
inant loadings combined with the expected removal rates
in table 19 to define system size. The following text de-
tails the use of this flow chart and also discusses aspects
of the model that are currently under investigation.
CHARACTERIZATION OF MINE
DRAINAGE DISCHARGES
To design and construct an effluent treatment system,
the mine water must be characterized. An accurate meas-
urement of the flow rate of the mine discharge or seep is
required. Water samples should be collected at the dis-
charge or seepage point for chemical analysis. Initial
water analyses should include pH, alkalinity, Fe, Mn, and
hot acidity (H2O3 method) measurements. If an anoxic
limestone drain is being considered, the acidified sample
should be analyzed for Fe3* and Al, and a field meas-
urement of dissolved oxygen should be made.
Both the flow rate and chemical composition of a
discharge can vary seasonally and in response to storm
Analyze row water chemistry
and determine flow rate
Net alkaline
water
Net acidic
water
DO, Fe3* Al
acceptable
Anoxic
limestone
drain
DO, Fe Al
unacceptable
pH>4
pH<4
-Influent Influent
acidity acidity
<300
>300
Figure 12,—Flow chart showing chemical determination* nec-
essary for th« design of passive treatment systems.
-------
27
events. If the passive treatment system is expected to
be operative during all weather conditions, then the dis-
charge flow rates and water quality should be measured
in different seasons and under representative weather
conditions.
CALCULATIONS OF CONTAMINANT LOADINGS
The size of the passive treatment system depends on
the loading rate of contaminants. Calculate contaminant
(Fe, Mn, acidity) loads by multiplying contaminant con-
centrations by the flow rate. If the concentrations are
milligrams per liter and flow rates are liters per minute,
the calculation is
[Fe,Mn,Acidity] g • d'1 = flow
X [Fe,Mn,Atidity] X 1.44 (11)
If the concentrations are milligrams per liter and How
rates are gallons per minute, the calculation is
[Fe,Mn, Acidity] g • d'1 « flow
X [Fe,Mn,Acidity] x 5.45 (12)
Calculate loadings for average data and for those days
when flows and contaminant concentrations are highest.
CLASSIFICATION OF DISCHARGES
The design of the passive treatment system depends
largely on whether the mine water is acidic or alkaline.
One can classify the water by comparing concentrations of
acidity and alkalinity.
Net Alkaline Water: alkalinity > acidity
Net^Acidic Watgr: acidity > alkalinity
The successful treatment of mine waters with net acidities
of 0 to 100 mg-L"1 using aerobic wetlands has been
documented in this report and elsewhere (14, IS). In
these systems, alkalinity either enters the treatment system
with diluting water cr alkalinity is generated within the
system by undetermined processes. Currently, there is no
method to predict which of these marginally acidic waters
can be treated successfully with an aerobic system only.
For waters with a net acidity >0, the incorporation of
alkalinity-generating features (either an ALD or a com-
post wetland) is appropriate.
PASSIVE TREATMENT OF NET ALKALINE WATER
Net alkaline water contains enough alkalinity to buffer
the acidity produced by metal hydrolysis reactions. The
metal contaminants (Fe and Mn) will precipitate given
enough time. The generation of additional alkalinity is
unnecessary so incorporation of limestone or an organic
substrate into the passive treatment system is also un-
necessary. The goal of the treatment system is to aer-
ate the water and promote metal oxidation processes. In
many existing treatment systems where the water is net
alkaline, the removal of Fe appears to be limited by
dissolved Q2 concentrations. Standard features that can
aerate the drainage, such as waterfalls or steps, should be
followed by quiescent areas. Aeration only provides
enough dissolved O2 to oxidize about 50 mg-L-1 Fe2*.
Mine drainage with higher concentrations of Fe3* will
require a series of aeration structures and wetland basins.
The wetland cells allow tune for Fe oxidation and hydrol-
ysis to occur and space in which the Fe floe can settle out
of suspension. The entire system can be sized based on
the Fe removal rates shown in table 19. For example, a
system being designed to improve water quality on an
AML site should be sized by the following calculation:
Minimum wetland size (m2)
= Fe loading (g - d"1)/20 (g • m"2 • d"1). (13)
If Mn removal is desired, size the system based on the Mn
removal rates in table 19. Removal of Fe and Mn occurs
sequentially in passive systems. If both Fe and Mn re-
moval are necessary, add the two wetland sizes together.
A typical aerobic wetland is constructed by planting
cattail rhizomes in soil or alkaline spoil obtained on-site.
Some systems have been planted by simply spreading
cattail seeds, with good plant growth attained after 2 years.
The depth of the water in a typical aerobic system is 10 to
50 cm. Ideally, a cell should not be of uniform depth,
but should include shallow and deep marsh areas and a
few deep (1 to 2 m) spots. Most readily available aquatic
vegetation cannot tolerate water depths greater than
50cm.
Often, several wetland cells are connected by flow
through a V-notch weir, lined railroad tie steps, or down
a ditch. Spillways should be designed to pass the maxi-
mum probable flow. Spillways should consist of wide cuts
in the dike with side slopes no steeper than 2H:1V, lined
with nonbiodegradable erosion control fabric, and coarse
rip rap if high flows are expected (IS). Proper spillway
design can preclude future maintenance costs because of
erosion and/or failed dikes. If pipes are used, small
diameter (<30 cm) pipes should be avoided because they
can plug with litter and FeOOH deposits. Pipes should be
made of polyvinyl chloride (PVC). More details on the
construction of aerobic wetland systems can be found in a
text by Hammer (5(5).
The geometry of the wetland site as well as flow con-
trol and water treatment considerations may dictate the
-------
28
use of multiple wetland cells. The interceil connections
may also serve as aeration devices. It there are elevation
differences between the cells, the interconnection should
dissipate kinetic energy and be designed to avoid erosion
and/or the mobilization of precipitates.
It is recommended that the freeboard of aerobic wet-
lands constructed for the removal of Fe be at least 1 m.
Observations of sludge accumulation in existing wetlands
suggest that a 1-m freeboard should be adequate to con-
tain 20 to 25 years of FeOOH accumulation.
The floor of the wetland cell may be sloped up to about
3% grade. If a level cell floor is used, then the water level
and flow are controlled by the downstream dam spillway
and/or adjustable riser pipes.
As discussed in chapter 3, some of the aerobic systems
that have been constructed to treat alkaline mine water
have little emergent plant growth. Metal removal rates in
these plantless, aerobic systems appears to be similar to
what is observed in aerobic systems containing plants.
However, plants may provide values that are not reflected
in measurements of contaminant removal rates. For ex-
ample, plants can facilitate the filtration of participates,
prevent flow channelization and provide v/ildlife benefits
that are valued by regulatory and environmental groups.
PASSIVE TREATMENT OF NET AGIO WATER
Treatment of acidic mine water requires the generation
of enough alkalinity to neutralize the excess acidity. Cur-
rently, there are two passive methods for generating alka-
linity; construction of a compost wetland or pretreatment
of acidic drainage by use of an ALD. In some cases, the
combination of an ALD and a compost wetland may be
necessary to treat the mine water.
ALD's produce alkalinity at a lower cost than do
compost wetlands. However, not all water is suitable for
pretreatment with ALD's. The primary chemical factors
believed to limit the utility of ALD's are the presence of
ferric iron (Fe3+), aluminum (Al) and dissolved oxygen
(DO). When acidic water containing any Fe3* or Al
contacts limestone, metal hydroxide particulates (FeOOH
or Al(OH)a) will form. No oxygen is necessary. Ferric hy-
droxide can armor the limestone, limiting its further dis-
solution. Whether aluminum hydroxides armor limestone
has not been determined. The buildup of both precipitates
within the ALD can eventually decrease the drain perme-
ability and cause plugging. The presence of dissolved
oxygen in mine water will promote the oxidation of ferrous
iron to ferric iron within the ALD, and thus potentially
cause armoring and plugging. While the short-term per-
formance of ALD's that receive water containing elevated
levels of Fe3*, Al, or DO can be spectacular (total
removal of the metals within the ALD) (34), the long-term
performance of these ALD's is questionable.
Mine water that contains very low concentrations of
DO, Fe3* and Al (all <1 mg«L-1) is ideally suited for
pretreatment with an ALD. As concentrations of these
parameters rise above 1 mg«L"S the risk that the ALD
will fail prematurely also increases. Recently, two ALD's
constructed to treat mine water that contained 20 tng*L'*
Al became plugged after 6-8 months of operation.
In some cases, the suitability of mine water for pre-
treatment with an ALD can be evaluated based on the
type of discharge and measurements of field pH. Mine
waters that seep from spoils and flooded underground
mines and have a field pH >S characteristically have con-
centrations of DO, Fe3*, and Al that are all <1 rag-L"1.
Such sites are generally good candidates for pretreatment
with an ALD. Mine waters that discharge from open drift
mines or have pH <5 must be analyzed for Fe3* and Al,
Mine waters with pH <5 can contain dissolved Al; mine
waters with pH <3.5 can contain Fe3*. In northern
Appalachia, most mine drainages that have pH <3 contain
high concentrations of Fe3* and Al*
PRETREATMENT OF ACIDIC WATER WITH ALD
In an ALD, alkalinity is produced when the acidic water
contacts the limestone in an anoxic, closed environment.
It is important to use limestone with a high CaCO3 content
because of its higher reactivity compared with a limestone
with a high MgCO3 or CaMg(CO3)2 content. The lime-
stones used in most successful ALD's have 80% to 95%
CaCO3 content. Most effective systems have used number
3 or 4 (baseball-size) limestone. Some systems con-
structed with limestone fines and small gravel have failed,
apparently because of plugging problems. The ALD must
be sealed so that inputs of atmospheric oxygen are min-
imized and the accumulation of CO2 within the ALD is
maximized. This is usually accomplished by burying the
ALD under several feet of clay. Plastic is commonly
placed between the limestone and clay as an additional gas
barrier. In some cases, the ALD has been completely
wrapped in plastic before burial (55). The ALD should be
designed so that the limestone is inundated with water at
all times. Clay dikes within the ALD or riser pipes at the
outflow of the ALD will help ensure inundation.
The dimensions of existing ALD's vary considerably.
Most older ALD's were constructed as long narrow drains,
approximately 0.6 to 1.0 m wide. A longitudinal section
and cross section of such an ALD is shown in figure 13.
The ALD shown was constructed in October 1990, and is
1 m wide, 46 m long and contains about 1 m depth of
number 4 limestone. The limestone was covered with two
layers of 5 mil plastic, which in turn was covered with
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29
Vegetated
crown
High quality
limestone
(>90%CaC03)
CROSS-SECTION VIEW
Figure 13.—Longitudinal-section and crow-section of ihe Morrison ALD. Wells are for sampling purposes and have no Importance
to drain's functioning.
0.3 to 3 m of on-site clay to restore the original surface
topography (34, 3(5).
At sites where linear ALD's are not possible, anoric
limestone beds have been constructed that are 10 to 20 m
wide. These bed systems have produced alkalinity concen-
trations similar to those produced by the more conven-
tional drain systems.
The mass of limestone required to neutralize a certain
discharge for a specified period can be readily calculated
from the mine water flow rate and assumptions about the
ALD's alkalinity-generating performance. Recent USBM
research indicates that approximately 14 h of contact time
between mine water and limestone in an ALD is necessary
to achieve a maximum concentration of alkalinity (57). To
achieve 14 h of contact time within an ALD, -3,000 kg of
limestone rock is required for each liter per minute of
mine water flow. An ALD that produces 275 mg-L"1 of
alkalinity (the maximum sustained concentration thus far
observed for an ALD), dissolves -1,600 kg of limestone a
decade per each liter per minute of mine water flow. To
construct an ALD that contains sufficient limestone to
unsure a 14-h retention time throughout a 30-yr period, the
limestone bed should contain -7,800 kg of limestone for
each liter per minute of flow. This is equivalent to 30 tons
of limestone for each gallon per minute of flow. The
calculation assumes that the ALD is constructed with 90%
CaCO3 limestone rock that has a porosity of 50%. The
calculation also assumes that the original mine water does
not contain ferric iron or aluminum. The presence of
these ions would result in potential problems with armor-
ing and plugging, as previously discussed.
Because the oldest ALD's are only 3 to 4 yr old, it is
difficult to assess how realistic these theoretical calcu-
lations are. Questions about the ability of ALD's to main-
tain unchannelizcd flow for a prolonged period, whether
100% of the CaCO3 content of the limestone can be ex-
pected to dissolve, whether the ALD's will collapse after
significant dissolution of the limestone, and whether inputs
of DO that are not generally detectable with standard Held
equipment (0 to Img'L'1) might eventually result in
armoring of the limestone with ferric hydroxides, have not
yet been addressed.
The anoxic limestone drain is one. component of a pas-
sive treatment system. When the ALD operates ideally, its
only effect on mine water chemistry is to raise pH to
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30
circumneutral levels and increase concentrations of cal-
cium and alkalinity. Dissolved Fe2* and Mn should be
unaffected by flow through the ALD. The ALD must be
followed by a settling basin or wetland system in which
metal oxidation, hydrolysis and precipitation can occur.
The type of post-ALD treatment system depends on the
acidity of the mine water and the amount of alkalinity
generated by the ALD. If the ALD generates enough
alkalinity to transform the acid mine drainage to a net
alkaline condition, then the ALD effluent can then be
treated with a settling basin and an aerobic wetland. If
possible, the water should be aerated as soon as it exits
the ALD and directed into a settling pond. An aerobic
wetland should follow the settling pond. The total post-
ALD system should be sized according to the criteria
provided earlier for net alkaline mine water. At this time,
it appears that mine waters with acidities <150 mg300 mg"L~l, construction of a
compost wetland is recommended. Compost wetlands
generate alkalinity through a combination of bacterial ac-
tivity and Umestone dissolution. The desired sulfate-
reducmg bacteria require a rich organic substrate in which
anoxic conditions will develop. Limestone dissolution also
occurs readily within this anoxic environment. A substance
commonly used in these wetlands is spent mushroom
compost, a substrate that is readily available in western
Pennsylvania. However, any well-composted equivalent
should serve as a good bacterial substrate. Spent mush-
room compost has a high CaC03 content (about 10% dry
weight), but mixing in more limestone may increase the
alkalinity generated by CaCO3 dissolution. Compost sub-
strates that do not have a high CaCO3 content should
be supplemented with limestone. The compost depth used
in most wetlands is 30 to 45 cm. Typically, a metric ton
of compost will cover about 3.5 ra2 to a depth of 45 cm
thick. This is equivalent to one ton per 3.5 yd2. Cattails
or other emergent vegetation are planted in the substrate
to stabilize it and to provide additional organic matter
to "fuel" the sulfate reduction process. As a practical tip,
cattail plant-rhizomes should be planted well into the
substrate prior to flooding the wetland cell.
Compost wetlands in which water flows on the surface
of the compost remove acidity (e.g., generate alkalinity)
at rates of approximately 2-12 g-nr'-d'1. This range in
performance is largely a result of seasonal variation: lower
rates of acidity removal occur in winter than in summer
(55). Research in progress indicates that supplementing
the compost with limestone and incorporating system
designs that cause most of the water to flow through the
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31
compost (as opposed to on the surface) may result in
higher rates of limestone dissolution and better winter
performance.
Compost wetlands should be sized based on the re-
moval rates in table 19. For an AML site, the calculation
is
Minimum Wetland Size (m2) =
-l
Acidity Loading (g • d"1/?)- (14)
In many wetland systems, the compost cells are pre-
ceded with a single aerobic pond in which Fe oxidation
and precipitation occur. This feature is useful where the
influent to the wetland is of circumneutral pH (either
naturally or because of pretreatment with an ALD), and
rapid, significant removal of Fe is expected as soon as the
mine water is aerated. Aerobic ponds are not useful when
the water entering the wetland system has a pH <4. At
such low pH, Fe oxidation and precipitation reactions are
quite slow and significant removal of Fe in the aerobic
pond would not be expected.
OPERATION AND MAINTENANCE
Operational problems with passive treatment systems
can be attributed to inadequate design, unrealistic ex-
pectations, pests, inadequate construction methods, or
natural problems. If properly designed and constructed, a
passive treatment system can be operated with a minimum
amount of attention and money.
Probably the most common maintenance problem is
dike and spillway stability. Reworking slopes, rebuilding
spillways, and increasing freeboard can all be avoided by
proper design and construction using existing guidelines
for such construction.
Pests can plague wetlands with operational problems.
Muskrats will burrow into dikes, causing leakage and
potentially catastrophic failure problems, and will uproot
significant amounts of cattails and other aquatic vegetation.
Muskrats can be discouraged by lining dike inslopes with
chainlink fence or riprap to prevent burrowing (13).
Beavers cause water level disruptions because of damming
and also seriously damage vegetation. They are very dif-
ficult to control once established. Small diameter pipes
traversing wide spillways ("three-log structure") and trap-
ping have had limited success in beaver control. Large
pipes with 90° elbows on the upstream end have been used
as discharge structures in beaver-prone areas (IS). Other-
wise, shallow ponds with dikes with shallow slopes toward
wide, riprapped spillways may be the best design for a
beaver-infested system.
Mosquitos can be a problem where mine water is alka-
line, In southern Appalachia, mosquitofish (Gambusia
affinis) have been introduced into alkaline-water wetlands.
Other insects, such as the armyworm, have devastated
monocultural wetlands with their appetite for cattails (59).
The use of a variety of plants in a system will minimize
such problems.
CHAPTERS. SUMMARY AND CONCLUSIONS
The treatment of contaminated coal mine drainage
requires the precipitation of metal contaminants and the
neutralization of acidity. In conventional treatment sys-
tems, distinctions between these two treatment objectives
are blurred by additions of highly basic chemicals that
simultaneously cause the rapid precipitation of metal con-
taminants and the neutralization of acidity. Passive treat-
ment differs from conventional treatment by its distinction
between these two treatment objectives. It is possible to
passively precipitate Fe contaminants from mine water, but
have little effect on the mine water acidity. Alternatively,
it is possible to passively add neutralizing capacity to acidic
mine water without decreasing metal concentrations.
Waters that contain high concentrations of bicarbonate
alkalinity are most amenable to treatment with constructed
wetlands. Bicarbonate acts as a buffer that neutralizes the
acidity produced when Fe and Mn precipitate and main-
tains a pH between 5.5 and 6.5. At this circumneutral pH,
Fe and Mn precipitation processes are more rapid than
under acidic pH conditions. Given the ability of bi-
carbonate alkalinity to positively impact both the metal
precipitation and neutralization aspects of mine water
treatment, it is not surprising that the most noteworthy
applications of passive treatment have been at sites where
the mine water was net alkaline. The most successful wet-
lands constructed in western Pennsylvania in the early
1980's treated mine waters that contained alkalinity. All
of the early successes of the TVA were, likewise, with
waters that were alkaline (23). Similarly, the Simco wet-
land in Ohio, which has discharged compliance water for
several years (17), receives water containing -160 rag • L'1
alkalinity. In this study, the two treatment systems that
met all effluent discharge requirements (Donegal and
Blair) both received alkaline, metal-contaminated water.
When mine water is acidic, enough alkalinity must be
generated by the passive treatment system to neutralize
the acidity. The most common method used to passively
generate alkalinity is the construction of a wetland that
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32
contains an organic substrate la which alkalinity-generating
microbial processes occur. If the substrate contains
limestone, as spent mushroom compost does, then alka-
linity will be generated by both calcite dissolution and
bacterial sulfate reduction reactions. These alkalinity
generating processes are slow relative to processes that
remove Fe. Thus, the performance of the constructed wet-
lands that receive acidic water is usually limited by the rate
at which alkalinity is generated within the substrate. While
wetlands can significantly improve water quality, and have
proven to be effective at moderately acidic sites, no wet-
land systems that consistently and completely transform
highly acidic water to compliance quality are known.
Inconsistent or partial treatment indicates undersizing.
The authors believe this is because of a lack of awareness
of how much larger wetlands constructed to treat acidic
water must be than ones constructed to treat alkaline
water. The Fe and acidity removal rates measured in this
study indicate that the treatment of 5,000 g'd'1 of Fe in
alkaline water requires - 250 ma of aerobic wetland. The
treatment of the same Fe load in acidic water (where
treatment requires both precipitation of the Fe and neu-
tralization of the associated acidity) requires -1,300 m2 of
compost wetland. Thus wetlands constructed to treat
acidic water need to be six times larger than ones con-
structed to treat similarly contaminated alkaline water.
The recent development of limestone pretreatment sys-
tems, e.g., the anoxic limestone drain, is a significant ad-
vancement in passive treatment technology. When suc-
cessful, ALD's can lower acidities or actually transform
acidic water into alkaline water, and markedly decrease the
sizing demands of the wetlands constructed to precipitate
the metal contaminants. Because limestone is inexpensive,
the cost of an ALD-aerobic wetland passive treatment
system is typically much less than the compost wetland
alternative. Thus, when the influent water is appropriate,
ALD's should be the preferred method for generating
alkalinity in passive treatment systems.
Anoxic limestone drains have also been used to increase
the performance of existing constructed wetlands. At
many poorly performing wetlands that receive acidic water,
the wetland was built too small to treat an acidic, metal-
contaminated influent, but is large enough for an alkaline,
metal-contaminated influent. One of the study sites, the
Morrison wetland, was undersized for the highly acidic
water that it received. As a result, the wetland effluent
required supplemental treatment with chemicals. Since
construction of an ALD, and its addition of 275 mg-L"1
of bicarbonate alkalinity to the water, the discharge of the
wetland has been alkaline, low hi dissolved metals, and
does not require any supplemental chemical treatment.
Similar enhancements in wetland performance through
the addition of ALD's have been reported elsewhere in
Appalachia (IS, IS).
KINETICS OF CONTAMINANT
REMOVAL PROCESSES
This report presents an intensive analysis of con-
taminant removal kinetics in passive treatment systems.
The rates presented are generally in agreement with those
reported by other investigators. For example, the average
Mn-removal rate measured in this study for alkaline,
Fe-rrcc waters, 0.5 g«m**»d*1, is consistent with rates
reported by the TVA for aerobic wetlands in southern
Appalachia (75) and by the Pennsylvania Department of
Environmental Resources (DER) for constructed wetlands
in Pennsylvania (<$0). The average Fe-15 mg«L-1 Fe, and thus was overloaded with
Fe, the removal rate averaged 21 g^nr^d'1. When the
wetland effluent contained <15 mg-L'1 Fe, the removal
rate averaged only 11 g^m^-d'1.
LONG-TERM PERFORMANCE
Passive treatment systems cannot be expected to per-
form indefinitely. In the long term, wetland systems will
fill up with metal precipitates or the conditions that
facilitate contaminant removal may be compromised.
None of the treatment systems considered in this study
demonstrated any downward trends in contaminant re-
moval performance. Therefore, estimates of the long-
term performance of passive systems must be made by
extrapolating available data. Like the design and sizing
of passive treatment systems, estimates of long-term per-
formance vary with the chemistry of the mine water. Sys-
tems receiving alkaline water precipitate Fe and Mn con-
taminants by oxidative processes. The rapid removal of
Fe that occurs in alkaline treatment systems means that
such systems will inevitably fill up. Stark (61} reports that
the Fe sludge in a constructed wetland hi Ohio is in-
creasing by 3 to 4 cm per year. Similar measurements at
Pennsylvania wetlands indicate an increase in sludge depth
of 2 to 3 cm per year (62). These measurements suggest
that dikes that provide 1 m of freeboard should provide
sufficient volume for 25 to 50 years of performance.
At some surface mines, water quality tends to improve
within a decade after regrading and reclamation arc
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33
completed (63-64), At these surface minesites, 25 to
50 years of passive treatment may be adequate to mitigate
the contaminant problem. At surface mine sites where
contaminant production is continual, or at systems con-
structed to treat drainage from underground mines or coal
refuse disposal areas, the system can either be built with
greater freeboard or rebuilt when it eventually fills up.
Site conditions will determine whether it is more econom-
ical to simply bury the wetland system in place and con-
struct a new one, or to excavate and haul away the ac-
cumulated solids for proper disposal. Disposal of these
excavated sludges is not difficult or unduly expensive
because the material is not considered a hazardous waste.
Wetlands that receive acidic water, and function
through the alkalinity-generating processes associated with
an organic substrate, may decline in performance as the
components of the organic substrate that generate alka-
linity arc exhausted. The compost wetlands described in
this report neutralize acidity through the dissolution of
limestone and the bacterial reduction of sulfate. Lime-
stone dissolution is limited by the amount of limestone
present in the substrate. The limestone content of spent
mushroom compost is -30 kg»m*3 ((55). If a wetland
containing a 40 cm depth of compost generates CaCO3-
derived alkalinity at a mean rate of 3 g»nr*»d"1 (the
average rate measured in this study), then the limestone
content of the compost will be exhausted in 11 years. The
same volume of compost contains -40 kg of organic car-
bon. If bacterial sulfate reduction mineralizes 100% of
this carbon to bicarbonate at a rate of 5 g"nr2«d~l, then
the carbon will be exhausted in 91 years. This estimate is
Increased by the carbon input of the net primary produc-
tion of the wetland system, but decreased by the fact that
some of the carbon is mineralized by reactions other than
sulfate reduction. Studies of a salt marsh on Cape Cod,
MA, indicated that 75% of the carbon was eventually min-
eralized by sulfate reduction processes (6tf). Another sig-
nificant factor that decreases the available carbon is that
a portion of the carbon pool is recalcitrant.
A realistic scenario for the long-term performance of
a compost wetland is that sulfate reduction is linked, in
a dependent manner, to limestone dissolution. Sulfate-
reducing bacteria are inactive at pH less than 5 (37).
Their activity in a wetland receiving lower pH water may
depend, in part, on the presence of pH-buffering supplied
by limestone dissolution. Thus, limestone dissolution may
create alkaline zones in which sulfate reduction can
proceed and produce further alkalinity. If this scenario is
accurate, then the long-term performance of a compost
wetland may be limited by the amount of limestone in the
substrate, or according to the above calculations, about
11 years of performance. Under these conditions h would
be advisable to increase the chemical buffering capability
of the wetland substrate by adding additional limestone
during wetland construction. In fact, this procedure is
commonly practiced at many constructed compost wetland
sites.
The performance of anoxic limestone drains has many
aspects that make long-term expectations uncertain. An-
oxic limestone drains function through the dissolution,
and thus removal, of limestone. Eventually, this chemical
reaction will exhaust the limestone. Long-term scenarios
about ALD performance fail to consider the hydrologic
implications of the gradual structural failure of the sys-
tems. In large ALD's, most of the limestone dissolution
occurs in the upgradient portion of the limestone bed. It
is unknown whether this preferential dissolution will
produce partial failure of the integrity of the system or
whether the permeability will be adversely affected.
Another aspect that affects long-term ALD performance
is the fact that ALD's retain ferric iron and aluminum (34-
35). This retention has raised concerns about the ar-
moring of limestone or the plugging of flow paths long
before the limestone is exhausted by dissolution reactions
(34), No methods are currently available to predict exactly
how the retention of these metals affects the performance
of ALD's.
CONTINUALLY EVOLVING PASSIVE
TECHNOLOGIES
This document reports the current state of passive mine
water treatment technologies. The design and sizing rec-
ommendations presented herein represent current meth-
odologies that will subsequently be replaced with more
efficient techniques. For example, important experiments
are underway in Pennsylvania, Virginia, and West Virginia
testing "hybrid" ALD-compost wetland systems. In these
experimental systems, organic substrates are used to re-
duce ferric iron to ferrous iron and strip dissolved oxygen
from the water so that the mine water is suitable for flow
through an anoxic limestone drain. If these systems prove
successful, it may be possible to treat highly acidic water
by cycling it between anoxic alkalinity-generating environ-
ments and aerobic, metal-removal environments. Experi-
mental systems using this design have recently been con-
structed in western Pennsylvania (67).
While the specific tools of passive treatment are likely
to evolve in the coming years, the fundamental mech-
anisms of passive treatment that have been identified in
this report will probably not change markedly. Research
has shown that the treatment of contaminated coal mine
drainage by constructed wetlands can be explained by well-
known chemical and biological processes. Passive treat-
ment, like active treatment with chemicals, requires that
the metal contaminants be precipitated and that the acidity
associated with these ions be neutralized. By recognizing
that these treatment goals need not be accomplished
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34
simultaneously, one can focus on optimization of the
individual objectives. As a result, the performance and
cost effectiveness of passive treatment systems is rapidly
improving. Today, most mine operators who install prop-
erly designed passive treatment systems rapidly recoup the
cost of their investment through decreased water treatment
costs. There is no reason to doubt that this technology
will continue to improve and that, over time, passive
treatment will be used in applications that are not possible
today.
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INT.BU.OF MINES,PGH.,PA 29916
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