d^k United States Environmental Protection L^l M % Agency TD 756 .5 .052 2002 Copy 2 FT MEADE GenCol1 Anaerobic Compost Constructed Wetlands System (CWS) Technology Innovative Technology Evaluation Report 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. Notice The information in this document has been funded by the U. S. Environmental Protection Agency (EPA) under Contract No. 68-C5-0037 to Tetra 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. / ^ 75 ^ HS- + 2HCO; + H + 4 2 3 M +2 + H 2 S 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 8 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 CWS 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 the perforated PVC piping from the compost. The compost was held in place 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. • 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 detennined 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 should be 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. 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: 9 Figure 2. Schematic cross-section of an anaerobic CWS upflow cell. 10 Proximity to a populated area—odors generally are associated with CWS treatment. The Clear Creek/Central City Superfund Site • Land availability near the source of the contaminated Michael Holmes - Remedial Pro J ect Manager water to avoid extended transport. The CWS U.S. Environmental Protection Agency typically requires more land than a conventional Region 8 treatment system. Consequently, locations with 999 18th Street, Suite 300 steep slopes and drainages would make construction o enver Colorado 80202 more difficult and costly. Telephone: (303)312-6607 • 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 11 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 or multicomponent 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 heavy metals. 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 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 percent used 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 not be 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 Sys terns If on-site facilities are not already available, a small storage building equipped with electricity may be 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 OOOCO'tCMOtJDCO-^-CNO (cud6) 8 ;dj/v\o|j Z.6 _AO N Z.6-d3S Z.6— un P ^6~Jdy Z6-q3J 96-390 96—P0 96-6nv 96 -unp 96-Jdv 96-qsj 56-390 96—PO 56-6nv 56—unp 56-Jdy S6-q9J t6-390 t6-P0 ■^6-& n V ^6-unp t6-Jdv ^6-qsj 56-39Q c o _o *4— c s o Q I 4- Q. 3 I 0) S o O 14 Figure 3. Flow rates measured for effluent cells. 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 upflow 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 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 up flow CWS can provide short-term protection of the 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 pg/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 Table 1. Evaluation of CWS Treatment Versus RI/FS Criteria Criterion Discussbn 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.theCWS may provide short- and tong-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 (50gpm) 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 CWSs to treat AMD. 17 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 must be 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 Sup^rfund 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 and regulate 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 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 aimed at 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 pg/L in the receiving stream (Clear Creek). In order to met this standard, the discharge from Burleigh Tunnel must contain less than 13,650 pg/L zinc under low-flow conditions and 65,700 pg/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 OSH A 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 (/) Y - £ § $ co <-> ^ p O. o •£ 0) *—> i w O $ §u _ co o c 0) E ■E CO CL 0) Q o T3 CO ka o o O "9) n g> X > tS g c o to C o E u Q w H 00 00 £ U to c o ■«—' fi c flj o c o U c . csi T— 'O’ 5 1 o o o CM CM O O in t- o s in co o o o - CM in CM 03 CO O O O CM CM O O CM in 03 co CM o CO CO o o o Co' ■O’ o CO co cm' (X. O- o co o_ co’ CM o o o o o o o N-" in o o CM CO o co CO o o o o' CM CO CO in O’ CM in O’ CM 3 c 'E .3 < § E 3 I ■O co o 03 Q. CL o o c o T3 CO 03 E 3 C/3 03 c 03 co 03 0) c co 03 c co 2 ’(/) E 03 0) T5 03 J> o c 0 3 5= c T3 c c3 T3 C 2 <55 c 0 E 0 ■© 0 3 C 0 3. o ^ o O) 03 w 0 > < C 0 E 0 LU in 00 i in cb in 00 i in cb § 1 • o II oo CD o o o i o o o o o 05 c 3 I a. CP 3. co CM f- O o co oo O O O o CD o (0 co c 0 O) >. X o co ■o is -S' C/5 0 o c <0 c -0 > ^ D TSS TDS 11 1 s . 2 3 i CD w Q5 CL E CD 2 0 C CD c E a c o O E 3 — E — X S ^ TO 2* CD CD TJ ■a c c CD o Tn o CD o CO z c/> 0 21 Section 3 Treatment Effectivemess The following sections discuss the treatment effectiveness of the CWS demonstration in Silver Plume, Colorado. The discussion includes abackground 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 stopes, 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 stopes. 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 9,152 feet, and the climate is typical of mountainous 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 expected by CWS treatment. The treatability study involved the construction, operation, and sampling of two bioreactors. Each bioreactor was 22 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 the bottom 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 compost and water. The lower 6 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 grading to install the two CWS treatment cells. 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 (AV S), 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 influent values and the up flow 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 comer 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 biobeds (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, "lhe 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 • 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 and effluent 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 to potential technology users but are not critical to evaluate the technology. The noncritical parameters of the C W S 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 of zinc 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 time allowable for the demonstration. This time frame is allowed because the CWS is a biological technology and performance depended, in part, on primary substances and nutrients 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 1 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 Sample Type Method Number Method Title Source Metals Aqueous and Substrate 6010A,6020, 7470 ICP, 1 CP/MS, or AA SW-846 1 Sulfate Aqueous 300.0 Ion chromatography MCAWW2 Fluoride Aqueous 9056 Ion chromatography SW-846 Nitrate/Nitrite Aqueous 353.2 and 354.1 Various MCAWW2 Chloride Aqueous 300.0 Ion chromatography MCAWW2 Total and Aqueous 365.3 Various MCAWW Orthophosphate pH Aqueous 9040 Electrometric MCAWW TSS Aqueous 160.2 Gravimetric MCAWW TDS Aqueous 160.1 Gravimetric MCAWW TOC Aqueous 9060 Various SW-846 Ammonia Aqueous 350.1 Various MCAWW2 Alkalinity Aqueous 310.1 Various MCAWW2 Sulfide Aqueous 376.2 Various MCAWW2 Aquatic Toxicity Aqueous EPA SOPs 3 EPA 5 Acid Volatile Sulfide Substrate EPA Method Acid volatile sulfide EPA 1991 (A VS) Sulfate reducing bacteria Substrate None Anaerobic deep tube CSM 3 count Toxicity leaching Substrate 1311 ICP, ICP-MS or AA SW-846 procedure Reactive sulfide Substrate EPA 4 Titration SW-846 Orthophosphate Substrate 365.3 Various MCAWW Sulfate Substrate 300.0 Various MCAWW Physical parameters Substrate Various3 Various3 ASTM Residence time Aqueous ND ND ND pH Aqueous SOP 3 12 Tetra Tech 6 Temperature Aqueous SOP 3 11 Tetra Tech 6 Dissolved oxygen Aqueous SOP 3 62 Tetra Tech6 Conductivity Aqueous SOP 3 99 Tetra Tech6 Notes: 1 Test Methods for Evaluating Solid Wastes, Volumes IA- 1C: Laboratory Manual, Physical/Chemical Methods; and 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. 6 These are field measurements made byTetra 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 C WS demonstration conducted from January 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, influent and effluent data for much of 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 lOOmg/L. The mine drainage contains high levels of zinc that typically range from 45 to 65 mg/L. However, in May and June 1995, a great 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, 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 a relatively 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 in pH suggest these mine drainage constituents originate from mineral dissolution. Calcite (CaC03) 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 CI/VS The downflow cell was operated for approximately 2'/ 2 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 (~|/6iu) oujz Z.6-dsS Z.6-unp Z.6— JD kN 96-3 s 0 96-das 96-unp 96-JDkN g6-3 s a 56-d3S 56—l n P 56— un P 56-Jdv 56-qsj 56-uop ^6-aon >6-PO t6-das t6—inp >6-unp *6—'dv ^6-Jdia| JZ c o J= c c E -C C/5 c o •*—< co V- c 6-JDn C o c o e X3 o 2 ■*—- c O c o o E 3 E •o CU u 00 £ U ici v l. 3 “I 30 Downflow Effluent Cadmium 0.06 0/ 6lu ) P Da_ > Q6- ds S 96-inp S6— un P 56-Jdv 96-qsj 96-uDp t6— AO N *6—PO t6-das >6—l n P 176-unp t6-Jdv _C c o 2 x> o C/5 C o c 6—AON > 6-100 >6-d3S >6-inn >6-Nnr >6—ddV >6~dVW x i— z o (H/OIAl) 3S3NV0NVW 32 Figure 7. CWS manganese removed by month. 500,000 a i6-inr 96-ddV S6~inr S6-Nnr 96-AVkN 96-ddV 96-dVW -- 96-NVr >6-030 >6—AON >6-100 >6-onv -- >6-inn >6-Nnn -- >6-AVAI >6—ddV >6-dVkN LlJ 96-833 H < o LU _J CL < (/) CM 2 ID o 2 £ o o o _l 3 C/5 CO £ u £ o c= c * o T3 g "C o O C3 -C W) c o -o u — I o ,C3 "5 C/5 3C 4/ U 3 d£ 33 Table 4. Average Downflow CWS Substrate Results Cadmium (mg/kg) Lead (mg/kg) Nickel (mg/kg) Zinc (mg/kg) Acid Volatile Sulfides (mg/kg) Sulfate- Reducing Bacteria (count) Ortho¬ phosphate (mg/kg) 0-6 months 2.7 18 3.1 1,100 180 8.5 x 10 4 34 6-12 months 8.0 31 6.1 3,400 120 1.1 x 10 5 12 12-18 months 23 74 7.0 5,200 460 3.3 x 10 5 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 A VS 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 96—i n r o cr> o oo o o l£> O ID O O ro O CM q}uoiu/6>| 6uipD0"i oujz 96—Xidiai 96— JD kM 96-uDp 96-aon 96-das S6-l n P 96-Xd^ 96— JD kM 96-uDp t6— AO N >6-d9S 1 ^ 6 —i n r t6-^DlA| >6--JD^ (L> O Q 35 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). The adsorption period; 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 compost substrate by the bacteria. During this period, the concentration of zinc in the upflow effluent increased from3.2 mg/L(October 12,1994)to 18 mg/L (March 15, 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 that the bacteria became 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 of 97.6 and 66.8, respectively. In May and June 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 ofboth 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 of4,800 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 and April 1995, leading to higher loading. The maximum 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 upflow 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 Upflow CWS Substrate Results Acid Volatile Sulfate- Ortho- Cadmium (mg/kg) Lead (mg/kg) Nickel (mg/kg) Zinc (mg/kg) Sulfides (mg/kg) Reducing Bacteria (count) phosphate (mg/kg) Year 1 0.17 9.9 1.9 40 210 7.2 x 10 6 55 Year 2 0.18 13 2.0 71 460 3.2 x 10 6 54 Year 3 5.0 40.0 4.1 1,500 1,300 2.2 x 10 5 6.3 Year 4 9.6 NR 6.2 4,800 1,000 6.2 x 10 4 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 25,000,000 Z.6-d3S z.6-mr 96-030 96-100 96-ddV 96-inr Q6-Nnr 96-AVW 96—ddV 96—dVW 96-033 96-NVP V6-03CI t6-AON * 6-100 *6-0DV *6-inn *6-Nnr *6—AVIA *6—ddV V6-0VIA o O o o o o o o o o o o o' o' o' o' o o o o o o o o o' in o' in' CM T— < o < CO AlVdO d3d SINROO NOliVindOd to z ■'*- o z <: o o o _l < o o o _j z LxJ _l —1 CL CL < < CO CO UJ LJ 1- I— < ■< cr QC h“ 1 — CO CO CD CD 3 3 CO CO <1 o I I 6- A °N t6-d9S t76-nr >6-^ d 1A| 176-JD^ 39 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 C WS 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 samples 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 C WS 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 Table 6. Clear Creek Upstream Cadmium (mg/L) Lead (mg/L) Nickel (mg/L) Zinc (mg/L) pH Conductivity (MS) Temperature ( # C) Average 0.0022 0.0034 0.0047 0.126 7.8 155.7 5.4 Maximum 0.0094 0.013 0.015 0.56 8.1 167.5 9.7 Minimum 0.0 0.0 0.0 0.11 7.6 144.0 0.9 Notes: °C Degrees Celsius mg/L Milligrams per liter pS MicroSiemens ND Not Detected pH Standard units Average sArithmetic Mean Table 7. Clear Creek Downstream Cadmium (mg/L) Lead (mg/L) Nickel (mg/L) Zinc (mg/L) pH Conductivity (MS) Temperature (°C) Average 0.00075 0.0013 0.0068 0.512 7.6 132.8 4.3 Maximum 0.0017 0.0024 0.026 0.56 8.1 173.3 9.7 Minimum ND ND ND 0.14 6.5 80.0 .. Notes: °C Degrees Celsius mg/L Milligrams per liter pS MicroSiemens ND Not Detected pH Standard units Average EArithmetic 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 Table 8. CWS Demonstration Toxicity (LC J0 ) Results Indicator Species Date Collected Influent Upflow Effluent Downflow Effluent Clear Creek Upstream Clear Creek Downstream Fathead Minnows 08/24/94 1.1 No toxicity NA 2 No toxicity No toxicity (Pimephalus promelas) 09/19/94 0.73 No toxicity No toxicity 02/22/95 1.6 No toxicity No toxicity 06/12/95 1.0 2.3 2.6 No toxicity No toxicity 09/05/95 0.62 1.2 30 12/10/96 0.62 1.6 NA 06/24/97 0.69 24 NA No toxicity No toxicity 10/29/97 1.4 14 NA 10/29/97 1 11 Water Fleas 08/24/94 0.46 No toxicity NA No toxicity No toxicity (Ceriodaphnia dubia) 09/19/94 0.31 No toxicity No toxicity 02/22/95 1 1.0 8.4 No toxicity 02/22/95 No toxicity 06/12/95 0.10 0.43 0.51 No toxicity No toxicity 12/10/96 0.09 0.22 NA 06/24/97 0.43 0.41 NA No toxicity No toxicity 09/05/95 0.10 <0.19 0.31 10/29/97 0.15 0.13 NA 10/29/97 1 0.19 NA 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 (ZnS04-7 H20) 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 A VS, 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 Visible observations of the up flow 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 C WS 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 objective was achieved in part. The demonstration zinc results indicate zinc removal is greater with an up flow 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. Unfortunately, it was not possible during this 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 detemiine 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 results 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 to be sulfate reduction. Section 3.4.3 discusses the upflow cell results for sulfate-reducing bacteria removal of metals. Data indicated an initial high 43 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 should be 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 to precipitation of metal oxides, hydroxides, and carbonates, settling of fme 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 fme 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, and 67 hours at a flow rate of 5 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 could not be readily measured as background levels of dissolved salts was somewhat high during the treatability study and the organic dye likely absorbed to 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. Dewatering 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 fme mesh geonet should alleviate the problem. 45 Section 4 Data Quality Review r> 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. Two equipment blanks (WEV090794EB and EBO12197) 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 of0.034 mg/L, slightly above the reporting limit of0.020 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 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 = J(A-B)| 0.5 (A + B) X 100 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. This process involves establishing data quality 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. 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 analyzed for 51 rounds of sampling activities. All laboratory 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)/T X 100 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 Toxicity 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 most prevalent toxic component present (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 by monthly 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 - in pH, 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 statistically different than the control when tested with Steel’s Many-One-Rank test at an = 0.05 (EPA 1993a). WED at 100 percent concentration did not exhibit sufficient mortality for the survival ratio to be statistically different than the control. The acute tests with P. promelas do not show any statistical difference from the control for WED or for WEDII; therefore, no toxicity for this species is evident. In general, C. dubia is more sensitive to environmental toxicants, so the absence of toxicity for P. promelas supports the presumption that WEDII is slightly toxic. Using the C. dubia results alone, it appears that there is a slight difference in the acute toxicity of the duplicate samples (WED and WEDII). Also, the arrival, initial, and final chemistry data show a difference in the characteristics in the ambient water between the two samples. Therefore, the duplicate analysis indicates that there is sufficient variability in the effluent stream to reflect a difference in the toxicity results of duplicate samples. However, this difference between duplicates is sufficiently small that the results of the acute toxicity tests, with LC50 as the endpoint, are not sensitive enough to calculate a coefficient of variation for effluent mine drainage samples. Representativeness For this project, representativeness for acute toxicity tests involved sample size, sampling times relative to seasonal temperature variation, and sampling locations. Most importantly, the changes due to seasonal environmental conditions needed to be documented to enable evaluation of zinc concentration reduction by biological conversion and uptake during cold stress conditions against warm temperature conditions. The QA goal was to obtain samples that represented biological water quality, measured by acute toxicity, in the treated and untreated mine drainage under typical seasonal environmental conditions. The primary seasonal environmental parameter of concern was temperature due to the regional extremes present at the demonstration location. Prior to the demonstration, it was known that three or four seasonal cycles would be required to conduct a statistical analysis of seasonal variation. The project budget and time schedule did not permit this type of data collection; consequently, the QA goal for representativeness was limited to successfully collecting data that would enable a limited evaluation of seasonal rise and fall of acute toxicity values in response to seasonal temperature stress. Since acute toxicity and zinc concentration data were obtained under environmental conditions representative of seasonal fluctuations in temperature in mine drainage influent and effluent, the QA objective for representativeness was met. Completeness Completeness is an assessment of the amount of valid data obtained from a measurement system compared to the amount of data expected to achieve a predefined quantity of information or level of confidence. The percent completeness is calculated by dividing the number of samples with acceptable data by the total number of samples planned to be collected and multiplying the result by 100. Greater than 90 percent completeness was achieved for all demonstration samples, and 100 percent of the critical samples for acute toxicity achieved acceptable results. Comparability The acute toxicity tests were conducted in accordance with the EPA guidance document “Methods for Measuring the Acute Toxicity of Effluents and Receiving Waters to Freshwater and Marine Organisms” (EPA 1991). All quality assurance guidance procedures have been adhered to, and the quantitative results for all QA criteria for reference toxicity fall within the specified limits. Therefore, the demonstration data are considered comparable to other acute toxicity data generated using these standard methods and adhering to the QA guidelines. 4.3 Noncritical Parameters Data Quality Review Data quality review for the first noncrtical objective of substrate utilization, and the third noncritical objective of effluent impact to Clear Creek were included in the review for the number one critical obj ecti ve data. Analytical results for these two noncritical parameters were within the quality assurance objectives stated in the Demonstration Plan (PRC 1995). Data quality results for noncritical obj ecti ve number two, the metal removal by sulfate-reducing bacteria were within the parameters cited in the Demonstration Plan. As stated in the plan, the evaluation of sulfate-reduction was expected to be more qualitative in nature. Results for the bacteria counts and acid-volatile sulfides are considered acceptable quality. Specific data quality assurance objectives for the fourth, and final noncritical ojbective, compiled capital and 50 operating costs, were not stated in the Demonstration Plan. However, cost tracking and compilation was performed using a best professional judgment approach. These data are considered accurate and usable within accepted professional standards. 51 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 Superfund and RCRA sites (Evans 1990). Costs are presented in year 1995 dollars and are considered estimates, with an accuracy of plus 50 percent and minus 30 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 zincat50 ppm was selected for this economic analysis. The following presents additional assumptions and conditions as they apply to each case. For each case, this analysis assumes that an up flow 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 [yd 3 ]), 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. 52 The site is located within 100 miles of a moderate¬ sized city. • 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 and handling; (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. 5.2.1 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 estimated mobilization 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 and regulatory 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 Table 9. CWS Costs for Different Treatment Flow Rates* System Life 15 Years Cost Categories 25 gpm_50 gpm_100 gpm Fixed Costs Site Preparation Administrative Mobilization $15,000 $10,000 5,000 $15,000 $10,000 5,000 $15,000 $10,000 5,000 Permitting and Regulatory Requirements $5,000 $5,000 $5,000 Capital Equipment System Design Excavation and Site Preparation Wetland Cell Construction Piping and Valves Storage Building $215,300 $50,000 9,800 120,000 25,500 10,000 $345,000 $50,000 19.500 240,000 25.500 10,000 $604,500 $50,000 39,000 480,000 25,500 10,000 Startup $1,500 $1,500 $1,500 Demobilization Excavation and Backfilling Substrate Disposal $52,250 $10,000 42,250 $104,500 $20,000 84,500 $209,000 $40,000 169,000 Total Fixed Costs $316,000 $492,000 $844,000 Variable Costs Labor Operations Staff $153,000 $153,000 $153,000 $153,000 $153,000 $153,000 Consumables and Supplies Personal Protective Equipment $39,000 $39,000 $39,000 $39,000 $39,000 $39,000 Utilities NA NA NA Residual and Waste Shipping and Handling Substrate Disposal $120,000 40,000 (3) $240,000 80,000 (3) $480,000 160,000 (3) Analytical Services $360,000 $360,000 $360,000 Maintenance and Modifications Annual Maintenance Substrate Removal and Replacement $247,550 $5,000 80,850 (3) $490,100 $5,000 161,700 (3) $975,200 $5,000 323,400 (3) Total Variable Costs $919,550 $1,282,100 $2,007,200 Total Costs $1,235,500 $1,774,100 $2,851,200 Total Cost Per Gallon Treated $0.0063 $0.0045 $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) N umber of removals 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 verify 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 yd 3 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. Oversight of 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, Q A/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 for this analysis are assumed to be $5,000 for each case. No modification costs are assumed to be 55 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, assuming native material from the 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 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 IC 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 Section 7 References Camp, Dresser, and McKee (CDM). 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. Report 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 Appendix A Analytical Results Summary Tables 59 Table A-l. Influent Results INFLUENT WI030994 WI032394 WI040694 WI042094 WI050594 W1051994 ANALYTICAL 03/09/94 03/23/94 04/06/94 04/20/94 05/05/94 05/19/94 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 ND ND ND ND ND 0.045 ARSENIC 6020 ND 0.0041 0.0068 0.020 0.060 0.052 CADMIUM 6020 0.10 0.099 0.10 0.10 0.098 0.081 CALCIUM 6010 84.8 88.0 91.7 96.9 89.9 83.2 IRON 6010 0.31 0.33 0.33 0.34 0.32 0.21 LEAD 6020 0.014 0.015 0.014 0.016 0.016 0.014 MAGNESIUM 6010 41.8 43.1 44.2 46.5 47.1 49.1 MANGANESE 6010 2.3 2.4 2.5 2.6 2.3 1.8 NICKEL 6010 0.045 0.039 0.042 0.047 0.043 0.035 POTASSIUM 6010 2.6 2.9 3.0 3.1 3.6 3.2 SILVER 6020 0.0011 0.00012 0.000066 0.000070 0.000098 0.00019 SODIUM 6010 10.3 9.3 10.9 9.1 14.0 10.5 ZINC 6010 55.0 56.1 60.1 64.0 56.1 44.8 ANIONS: SULFATE 300.0 386 374 387 384 317 314 SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 1.0 1.2 1.1 1.1 0.98 1.0 CHLORIDE 300.0 19.9 21.8 22.3 21.9 19.0 15.0 PHOSPHORUS, TOTAL 365.3 ND ND ND ND ND ND ORTHOPHOSPHATE 365.3 ND 0.30 ND ND ND 0.40 NITRATE PLUS NITRITE ASN 353.2 ND ND 0.060 0.11 ND ND NITRITE AS N 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND ND 0.060 0.11 ND ND AMMONIA 350.1 ND ND ND ND ND ND TOTAL SOLIDS: TSS 160.2 16.8 8.8 20.4 15.2 7.4 8.4 TDS 160.1 732 655 640 663 641 622 TOC 9060 1.1 NA NA ND NA NA ALKALINITY, TOTAL: AS CaC03 310.1 100 107 105 107 104 107 ALKALINITY, BICARB AS CAC03 310.1 100 107 105 107 104 107 DISSOLVED OXYGEN (mg/L) — 8.1 8.3 * oo vb NA NA PH — 7.4 7.5 7.5 7.4 7.5 CONDUCTIVITY (pS) -- 730 745 745 699 698 TEMPERATURE (degrees C) - 6.9 7.3 7.3 8.9 9.4 — = Not applicable NA = Not analyzed pS = MicroSiemens ND = Not detected mgL = Milligrams per liter 60 Table A-l (continued). Influent Results INFLUENT WI060194 WI062994 WI071394 WJ072894 WI081594 WI082494 ANALYTICAL 06/01/94 06/29/94 07/13/94 07/28/94 08/15/94 08/24/94 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 ND 0.068 ND ND ND ND ARSENIC 6020 ND ND ND ND ND ND CADMIUM 6020 0.092 0.089 0.086 0.098 0.10 0.0952 CALCIUM 6010 89.6 86.1 94.5 91.2 92.5 94.6 IRON 6010 0.25 0.23 0.23 0.30 0.24 0.25 LEAD 6020 0.020 0.017 0.013 0.017 0.016 0.014 MAGNESIUM 6010 50.6 45.4 48.3 46.4 47.7 48.1 MANGANESE 6010 1.9 2.1 2.2 2.2 2.3 2.4 NICKEL 6010 0.033 0.045 0.044 0.043 0.042 0.046 POTASSIUM 6010 3.6 3.0 3.1 2.9 2.9 3.2 SILVER 6020 0.00019 ND 0.00013 0.00015 0.00017 ND SODIUM 6010 13.2 12.8 13.00 12.0 14.4 15.3 ZINC 6010 49.1 54.2 56.8 59.1 54.7 57.5 ANIONS: SULFATE 300.0 357 378 377 397 374 403 SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 1.0 1.0 0.90 1.1 1.1 1.1 CHLORIDE 300.0 16.9 17.9 17.5 18.7 18.6 19.6 PHOSPHORUS, TOTAL 365.3 ND ND ND ND ND ND ORTHOPHOSPHATE 365.3 ND 0.44 ND 0.077 ND ND NITRATE PLUSNITRITE ASN 353.2 ND ND ND 2.0 1.7 1.9 NITRITE ASN 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND ND ND 2.0 1.7 1.9 AMMONIA 350.1 ND ND ND ND ND ND TOTAL SOLIDS: TSS 160.2 4.4 11.2 9.2 9.6 2.4 18.4 TDS 160.1 657 680 685 707 759 703 TOC 9060 NA NA NA NA NA NA ALKALINITY, TOTAL: AS CaC03 ALKALINITY, BICARB 310.1 109 107 109 103 105 102 ASCAC03 310.1 109 107 109 103 105 102 DISSOLVED OXYGEN (mg/L) - 8.7 NA 8.2 NA NA 7.6 pH — 7.6 7.57 7.5 NA 7.5 7.4 CONDUCTIVITY (pS) -- 775 980 950 927 948 920 TEMPERATURE (degrees C) - 9.4 9.5 9.4 9.5 9.4 9.4 ** = Degrees Farenheit NA = Not analyzed — = Not applicable ND = Not detected pS = micro Siemens mg/L= Milligrams per liter 61 Table A-l (continued). Influent Results I NFLUEN1 ANALYTE ANALYTICAL METHOD WI090794 09/07/94 mg/L WI091994 09/19/94 mg/L WI100494 10/04/94 mg/L WI101994 10/19/1994 mg/L Wll 10294 11/02/94 mg/L W1112094 11/20/94 mg/L AQUEOUS ALUMINUM 6010 ND ND ND ND 0.030 ND ARSENIC 6020 ND ND ND ND ND ND CADMIUM 6020 0.098 0.085 0.089 .10 0.10 0.091 CALCIUM 6010 90.2 89.7 92.6 92.4 89.2 93.5 IRON 6010 0.29 0.29 0.31 0.25 0.28 0.32 LEAD 6020 0.017 0.015 0.014 0.014 0.014 0.016 MAGNESIUM 6010 46.5 46.6 47.3 46.7 46.2 47.3 MANGANESE 6010 2.3 2.3 2.3 2.4 2.2 2.3 NICKEL 6010 0.047 0.042 0.052 0.046 0.051 0.050 POTASSIUM 6010 3.9 3.1 3.0 3.0 2.9 3.1 SILVER 6020 0.00040* 0.00041 0.00050 ND ND 0.00030 SODIUM 6010 12.1 12.5 11.6 13 14.8 14.4 ZINC 6010 56.4 57.6 59.7 57.6 56.5 58.2 ANIONS: SULFATE 300.0 416 404 400 409 410 407 SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 1.0 1.0 1.0 ND 1.0 1.1 CHLORIDE 300.0 20.2 19.6 19.8 19.5 20.1 21.3 PHOSPHORUS, TOTAL 365.3 ND ND ND ND ND ND ORTHOPHOSPHATE 365.3 ND ND ND ND 0.13 ND NITRATE PLUSNITRITE ASN 353.2 ND ND ND ND ND ND NITRITE ASN 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND ND ND ND ND ND AMMONIA 350.1 ND ND ND ND ND ND TOTAL SOLIDS: TSS 160.2 17.6 8.4 18.8 18.8 8.0 18.0 TDS 160.1 711 723 695 695 709 711 TOC 9060 NA NA NA NA NA NA ALKALINITY, TOTAL: AS CaC03 310.1 102 101 112 102 82.4 101 ALKALINITY, BICARB AS CAC03 310.1 102 101 112 102 82.4 101 DISSOLVED OXYGEN (mg/L) - 9.5 7.8 NA NA NA NA pH - 7.41 7.4 7.4 7.1 6.9 6.9 CONDUCTIVITY (pS) -- 922 930 935 750 900 NA TEMPERATURE (degrees C) -- 9.3 9.3 9.1 8.5 8.7 8.1 — = Not applicable NA = Not detected pS = MicroSicmcns ND = Not detected mgT, = Milligrams per liter 62 Table A-l (continued). Influent Results INFLUENT ANALYTE ANALYTICAL METHOD WI113094 11/30/94 mg/L WI121494 12/14/94 mg/L WI010495 01/04/95 mg/L WI011895 01/18/95 mg/L WI020195 02/01/95 mg/L WI021595 02/15/95 mg/L AQUEOUS ALUMINUM 6010 ND 0.036 0.032 0.038 0.047 0.043 ARSENIC 6020 ND ND ND ND ND ND CADMIUM 6020 0.086 0.092 0.82 0.076 0.089 0.084 CALCIUM 6010 95.4 98.1 87.7 90.8 90.1 100.0 IRON 6010 0.34 0.37 0.31 ND 0.34 0.39 LEAD 6020 0.014 0.018 0.016 0.015 0.016 0.015 MAGNESIUM 6010 47.7 48.9 46.5 45.4 44.1 49.4 MANGANESE 6010 2.5 2.5 2.3 2.4 2.4 2.7 NICKEL 6010 0.044 0.050 0.048 0.046 0.052 0.048 POTASSIUM 6010 2.8 3.3 2.9 3.0 2.8 3.5 SILVER 6020 0.00036 ND 0.00037 0.00021 ND ND SODIUM 6010 14.2 19.5 15.0 15.9 14.1 20.4 ZINC 6010 62.8 63.0 55.5 57.1 56.6 58.9 ANIONS: SULFATE 300.0 411 413 395 386 402 390 SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 1.1 1.0 1.1 1.1 1.1 1.1 CHLORIDE 300.0 21.4 21.2 21.6 21.7 22.5 22.8 PHOSPHORUS, TOTAL 365.3 ND ND ND ND ND ND ORTHOPHOSPHATE 365.3 0.13 0.36 ND ND ND 0.10 NITRATE PLUS NITRITE AS N 353.2 ND ND ND ND 1.7 ND NITRITE AS N 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND ND ND ND 1.7 ND AMMONIA 350.1 ND ND ND ND ND ND TOTAL SOLIDS: TSS 160.2 16.4 10.4 5.2 12.0 12.8 12.8 TDS 160.1 711 687 689 693 694 656 TOC 9060 NA NA NA NA NA NA ALKALINITY, TOTAL: AS CaC03 310.1 99.6 103 104 106 106 106 ALKALINITY, BICARB AS CAC03 310.1 99.6 103 104 106 106 106 DISSOLVED OXYGEN (mg/L) — NA 8.0 8.5 7.3 7.6 NA pH -- 6.9 7.54 7.5 7.5 7.9 7.0 CONDUCTIVITY (pS) -- 605 600 610 600 610 NA TEMPERATURE (degrees C) -- 7.9 8.0 6.5 9.0 7.9 8.1 * = Dissolved metals NA = Not analyzed — = Not applicable ND = Not detected pS= Microsiemens mg/L = Milligrams per liter 63 Table A-l (continued). Influent Results INFLUENT WI022795 W1031595 WI032995 W1041295 WI042695 WI051095 ANALYTICAL 02/27/95 03/15/95 03/29/95 04/12/95 04/26/95 05/10/95 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 0.024 0.049 ND ND 0.060 0.15 ARSENIC f> CADMIUM 6020 ND ND ND ND ND ND 6020 0.071 0.076 0.074 0.057 0.095 0.095 CALCIUM 6010 92.6 91.4 85.2 90.9 88.2 92.0 IRON 6010 0.33 0.36 0.33 0.32 0.41 0.48 LEAD 6020 0.014 0.016 0.014 0.015 0.022 0.026 MAGNESIUM 6010 45.1 44.4 41.9 42.9 41.2 41.9 MANGANESE 6010 2.5 2.5 2.3 2.4 2.6 3.0 NICKEL 6010 0.068 0.045 0.045 0.048 0.071 0.054 POTASSIUM 6010 2.9 2.9 2.8 3.0 2.9 3.1 SILVER 6020 ND ND ND ND ND ND SODIUM 6010 16.2 15.8 16.4 16.1 14.2 14.8 ZINC 6010 58.6 57.0 53.1 55.0 55.7 61.4 ANIONS: SULFATE 300.0 384.0 384.0 368.0 376.0 370.0 374 SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 1.1 1.1 1.0 1.0 1.1 1.1 CHLORIDE 300.0 22.6 22.4 23.1 22.4 23.8 20.5 PHOSPHORUS, TOTAL 365.3 ND ND ND ND ND ND ORTHOPHOSPHATE 365.3 ND ND ND 0.11 ND ND NITRATE PLUS NITRITE ASN 353.2 ND ND ND ND 0.14 ND NITRITE AS N 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND ND ND ND 0.14 ND AMMONIA 350.1 ND ND ND ND ND ND TOTAL SOLIDS: TSS 160.2 11.2 9.2 12.8 14.4 7.2 2.8 TDS 160.1 692 672 655 656 575 689 TOC 9060 NA NA NA NA NA ND ALKALINITY, TOTAL: AS CaC03 310.1 107 104 107 107 104 103 ALKALINITY, BICARB AS CAC03 310.1 107 104 107 107 104 103 DISSOLVED OXYGEN (mg/L) — 7.8 NA 7.5 8.6 7.5 pH -- 7.4 7.5 7.7 7.5 NA CONDUCTIVITY (pS) - 630 620 600 620 600 TEMPERATURE (degrees C) - 8.6 9.3 8.1 8.4 9.0 * = Dissolved metals NA = Not analyzed -- = Not applicable ND = Not detected pS = Microsiemens mg/L = Milligrams per liter 64 Table A-l (continued). Influent Results INFLUENT VVI061295 WI062895 WI071095 WI072695 WI080895 WI082395 ANALYTICAL 6/12/1995 6/28/1995 7/10/1995 7/26/1995 8/8/1995 8/23/1995 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 0.065 ND ND ND ND 0.079 ARSENIC 6020 ND ND ND ND ND ND CADMIUM 6020 0.25 0.26 0.25 0.24 0.26 0.240 CALCIUM 6010 94.4 111 119 129 123 125 IRON 6010 0.12 0.11 0.10 ND 0.15 0.19 LEAD 6020 0.058 0.051 0.050 0.038 0.043 0.039 MAGNESIUM 6010 58.3 61.4 64.0 64.2 61.7 61.3 MANGANESE 6010 3.9 4.4 5.0 5.5 5.2 5.2 NICKEL 6010 0.061 0.073 0.081 0.084 0.093 0.086 POTASSIUM 6010 4.1 ND 3.6 3.7 3.5 3.2 SILVER 6020 ND ND ND ND ND ND SODIUM 6010 9.9 14.2 14.8 13.2 14.1 15.2 ZINC 6010 75.5 86.8 99.8 105 109 108 ANIONS; SULFATE 300.0 499 502 582 596 638 630 SULFIDE TOTAL 376.2 FLUORIDE 340.2 0.8 0.89 0.96 0.88 0.87 0.95 CHLORIDE 300.0 6.9 8.8 10.2 11.7 13.1 PHOSPHORUS, TOTAL 365.3 ND ND ND ND ND 0.093 ORTHOPHOSPHATE 365.3 ND ND ND ND 0.095 ND NITRATE PLUS NITRITE ASN 353.2 0.13 0.10 ND 0.63 ND ND NITRITE ASN 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 0.13 ND ND 0.63 ND ND AMMONIA 350.1 ND ND ND ND ND ND TOTAL SOLIDS; TSS 160.2 20.4 20.4 24.8 22.4 18.8 32.0 TDS 160.1 838 967 1010 999 10.0 1050 TOC 9060 ALKALINITY, TOTAL; AS CaC03 310.1 120 125 118 107 107 107 ALKALINITY, BICARB AS CAC03 310.1 120 125 118 107 107 107 DISSOLVED OXYGEN (mg/L) — NA 7.1 NA NA NA NA pH - 7.4 7.2 7.4 NA NA NA CONDUCTIVITY (pS) - NA 700 NA NA 750 NA TEMPERATURE (degrees Q - 10.2 10.3 10.3 NA 10.4 NA * = Dissolved metals NA = Not analyzed — = Not applicable ND = Not detected pS = Microsiemens mgL = Milligrams per liter 65 Table A-l (continued). Influent Results INFLUENT ANALYTE ANALYTICAL WI090595 WI110995 CDPHE CDPHE CDPHE CDPHE METHOD 9/5/1995 11/9/1995 1/29/1996 2/29/1996 4/25/1996 5/31/1996 mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 ND ND NA NA NA NA ARSENIC 6020 ND ND NA NA NA NA CADMIUM CALCIUM 6020 0.24 0.20 0.160 0.200 0.12 0.14 6010 123 113 NA NA NA NA IRON 6010 0.28 0.18 0.24 0.26 0.18 0.17 LEAD 6020 0.038 0.027 NA NA NA NA MAGNESIUM 6010 60.2 56.2 NA NA NA NA MANGANESE 6010 5.2 5.2 3.60 3.50 2.4 2.7 NICKEL 6010 0.087 0.082 NA NA NA NA POTASSIUM 6010 ND 3.2 NA NA NA NA SILVER 6020 ND ND NA NA NA NA SODIUM 6010 12.4 15.6 NA NA NA NA ZINC 6010 107 105 73 69 46 56 ANIONS: SULFATE 300.0 652 591 490 450 NA NA SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 0.88 0.97 NA NA NA NA CHLORIDE 300.0 NA 17.7 NA NA NA NA PHOSPHORUS, TOTAL 365.3 0.067 0.060 NA NA NA NA ORTHOPHOSPHATE 365.3 ND 0.20 NA NA NA NA NITRATE PLUS NITRITE AS N 353.2 ND ND NA NA NA NA NITRITE AS N 354.1 ND ND NA NA NA NA NITRATE ASN 353.2/354.1 ND ND NA NA NA NA AMMONIA 350.1 ND ND NA NA NA NA TOTAL SOLIDS: TSS 160.2 18.4 14.4 NA NA NA NA TDS 160.1 1050 956 NA NA NA NA TOC 9060 NA NA NA ALKALINITY, TOTAL: AS CaC03 310.1 107 95.7 NA NA NA NA ALKALINITY, BICARB NA NA NA NA ASCAC03 310.1 107 95.7 NA NA NA NA DISSOLVED OXYGEN (mg/L) — NA NA NA NA pH - NA NA NA NA CONDUCTIVITY (pS) - NA NA NA NA TEMPERATURE (degrees C) - NA NA NA NA * = Dissolved metals NA = Not analyzed — = Not applicable ND = Not detected pS = Micro siemens mg/L = Milligrams per liter 66 Table A-l (continued). Influent Results INFLUENT CDPHE CDPHE CDPHE WI120996 WI012197 WI022097 ANALYTICAL 6/14/1996 7/19/1996 8/31/1996 12/9/1996 1/21/1997 2/20/1997 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 NA NA NA ND ND ND ARSENIC 6020 NA NA NA NA NA NA CADMIUM 6020 0.16 0.19 0.20 0.15 0.12 0.11 CALCIUM 6010 NA NA NA 104 100.0 105 IRON 6010 0.18 0.20 0.24 0.30 0.30 0.33 LEAD 6020 NA NA NA NA NA NA MAGNESIUM 6010 NA NA NA 52.8 51.2 52 MANGANESE 6010 2.9 3.5 4.1 3.7 3.5 3.7 NICKEL 6010 NA NA NA 0.07 0.06 0.06 POTASSIUM 6010 NA NA NA 3.1 J 3.0 J 3.0 J SILVER 6020 NA NA NA NA NA NA SODIUM 6010 NA NA NA 17.4 16.4 17.0 ZINC 6010 60 71 84 78 74 78 ANIONS: SULFATE 300.0 430 490 520 488 491 471 SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 NA NA NA NA NA NA CHLORIDE 300.0 NA NA NA 17.8 18.2 18.3 PHOSPHORUS, TOTAL 365.3 NA NA NA NA NA NA ORTHOPHOSPHATE 365.3 NA NA NA 0.31 0.17 0.22 NITRATE PLUSNITRITE ASN 353.2 NA NA NA ND ND ND NITRITE ASN 354.1 NA NA NA ND ND ND NITRATE ASN 353.2/354.1 NA NA NA ND ND ND AMMONIA 350.1 NA NA NA ND ND ND TOTAL SOLIDS TSS 160.2 NA NA NA NA 8.4 3.2 TDS 160.1 NA NA NA 849 796 809 TOC 9060 NA NA NA 0.8J 1.1 1.8 ALKALINITY, TOTAL: ASCaC03 310.1 NA NA NA 97.6 94.9 101 ALKALINITY, BICARB NA NA NA ASCAC03 310.1 NA NA NA 97.6 94.9 101 DISSOLVED OXYGEN (mg/L) — NA NA NA 7.4 8.8 8.6 pH — NA NA NA 7.2 5.1 7.5 CONDUCTIVITY (pS) -- NA NA NA NA NA NA TEMPERATURE (degrees Q — NA NA NA 10.0 8.2 3.2 * = Dissolved metals NA = Not analyzed — = Not applicable ND = Not detected pS = Microsiemens mg / L= Milligrams per liter 67 Table A-l (continued). Influent Results INFLUENT WI032097 WI042297 WI052897 WI062397 WI082897 WI093097 ANALYTICAL 3/20/1997 4/22/1997 5/28/1997 6/23/1997 8/28/1997 9/30/1997 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 ND 0.17 ND ND ND ND ARSENIC 6020 NA NA NA NA NA NA CADMIUM 6020 0.14 0.07 0.11 0.19 0.22 0.200 CALCIUM 6010 97.5 67.2 86.4 95.6 121 119 IRON 6010 0.34 0.34 0.24 0.26 0.3 0.33 LEAD 6020 NA NA NA NA NA NA MAGNESIUM 6010 48.8 37.3 53.8 52.3 61.9 58.4 MANGANESE 6010 3.6 2.0 2.7 3.3 4.9 4.9 NICKEL 6010 0.07 0.034 J 0.042 0.030 J 0.090 0.098 POTASSIUM 6010 ND 2.7 J 3.3 J 3.5 J 4.8 J 3.4 J SILVER 6020 NA NA NA NA NA NA SODIUM 6010 15.6 ND 14.9 ND ND 18.3 ZINC 6010 75 42 56 72 104 104 ANIONS: SULFATE 300.0 476 279 358 428 541 568 SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 NA NA NA NA NA NA CHLORIDE 300.0 18.7 9.3 7.2 9.2 13.8 16 PHOSPHORUS, TOTAL 365.3 NA NA NA NA NA NA ORTHOPHOSPHATE 365.3 0.15 ND ND 0.10 ND ND NITRATE PLUS NITRITE ASN 353.2 ND 0.14 ND 0.14 ND 0.19 NITRITE AS N 354.1 0.0021 J 0.0046 J 0.0024 J 0.0028 J 0.0037 J ND NITRATE ASN 353.2/354.1 ND ND ND ND ND ND AMMONIA 350.1 ND ND ND ND ND ND TOTAL SOLIDS: TSS 160.2 7.6 1.6 J 12.4 14.4 16.4 TDS 160.1 751 507 653 765 927 940 TOC 9060 0.20 J 1.30 1.4 0.98 J 0.80 J 0.58 J ALKALINITY, TOTAL: AS CaC03 310.1 96.3 99.7 107 121 102 ALKALINITY, BICARB AS CAC03 310.1 96.3 99.7 107 121 102 DISSOLVED OXYGEN (mg/L) — 7.8 7.3 7.3 8 8.7 NA pH - 6.9 7.4 7.4 7.5 6.9 6.9 CONDUCTIVITY (pS) -- NA NA NA NA NA NA TEMPERATURE (degrees C) - 8.6 9.7 10.5 9.7 9.6 9.4 * = Dissolved metals NA = Not analyzed — = Not applicable ND = Not detected pS = Micro siemens mg/L = Milligrams per liter 68 Table A-l (continued). Influent Results INFLUENT ANALYTE ANALYTICAL METHOD WI102997 10/29/1997 mg/L WT112597 11/25/1997 mg/L AQUEOUS ALUMINUM 6010 ND ND ARSENIC 6020 NA NA CADMIUM 6020 0.19 0.22 CALCIUM 6010 113 103 IRON 6010 0.37 0.39 LEAD 6020 NA NA MAGNESIUM 6010 58.8 50.4 MANGANESE 6010 4.9 4.2 NICKEL 6010 0.079 0.065 POTASSIUM 6010 3.4 J ND SILVER 6020 NA NA SODIUM 6010 18.3 16.5 ZINC 6010 95 86 ANIONS: SULFATE 300.0 571 548 SULFIDE TOTAL 376.2 NA NA FLUORIDE 340.2 NA NA CHLORIDE 300.0 17.5 17.8 PHOSPHORUS, TOTAL 365.3 NA NA ORTHOPHOSPHATE 365.3 ND 0.15 NITRATE PLUSNITRITE ASN 353.2 0.11 ND NITRITE AS N 354.1 0.002J 0.0025J NITRATE ASN 353.2/354.1 NA NA AMMONIA 350.1 ND ND TOTAL SOLIDS: TSS 160.2 10.4 14.8 TDS 160.1 940 869.0 TOC 9060 0.71J 1.8 ALKALINITY, TOTAL: AS CaC03 310.1 84 102 ALKALINITY, BICARB ASCAC03 310.1 84 102 DISSOLVED OXYGEN (mg/L) — 10.3 7.5 pH - 7.2 7.2 CONDUCTIVITY (pS) - NA NA TEMPERATURE (degrees C) - 9.2 8.9 * = Dissolved metals NA = Not analyzed — = Not applicable ND = Not detected pS = Microsiemens mg/L= Milligrams per liter 69 Table A-2. Downflow Effluent Results DOWNFLOW EFFLUENT WED030994 WED032394 WED 040694 WED042094 WED050594 ANALYTICAL 03/09/94 03/23/94 04/06/94 04/20/94 05/05/94 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 0.021 0.021 0.027 0.029 0.033 ARSENIC f ^ 6020 ND 0.00056 0.029 0.016 0.076 CADMIUM 6020 0.00034 0.00025 0.00028 0.00053 0.00072 CALCIUM 6010 105.0 107.0 110.0 113.0 113.0 IRON 6010 1.5 1.2 1.1 1.0 1.1 LEAD 6020 0.0015 0.0012 0.00065 0.0015 0.0017 MAGNESIUM 6010 56.7 56.9 58.6 58.3 58.9 MANGANESE 6010 1.6 1.5 1.5 1.4 1.4 NICKEL 6010 0.0073 0.0081 0.0086 0.010 0.0090 POTASSIUM 6010 55.8 56.6 54.0 50.6 48.3 SILVER 6020 0.0015 0.00012 0.000060 0.000089 0.0051 SODIUM 6010 19.0 17.1 18.1 15.3 18.6 ZINC 6010 14.2 14.9 15.6 15.3 13.1 ANIONS: SULFATE 300.0 350 357 338 337 280 SULFIDE TOTAL 376.2 4.1 5.2 5.7 2.1 0.74 FLUORIDE 340.2 0.82 0.93 0.88 0.90 0.87 CHLORIDE 300.0 15.6 28.4 27.2 28 22 PHOSPHORUS, TOTAL 365.3 9.9 10.6 11.0 10.8 10.4 ORTHOPHOSPHATE 365.3 10.6 12.4 10.7 11.1 11.1 NITRATE PLUS NITRITE AS N 353.2 0.24 ND ND ND ND NITRITE AS N 354.1 ND ND ND ND ND NITRATE AS N 353.2/354.1 0.24 ND ND ND ND AMMONIA 350.1 5.4 6.2 5.9 5.8 4.6 TOTAL SOLIDS: TSS 160.2 51.0 27.0 47.0 39.2 3.8 TDS 160.1 864 781 766 783 753 TOC 9060 60.4 20.6 29 28.2 20.8 ALKALINITY, TOTAL: AS CaC03 310.1 193 209 200 213 193 ALKALINITY, BICARB AS CAC03 310.1 193 209 200 213 193 ORP (mV) — -77.0 -180 -184 PH — 7.3 7.2 7.6 CONDUCTIVITY (pS) — 845 889 803 TEMPERATURE (degrees C) - 4.1 5.2 8.8 -- = Not applicable NA = Not analyzed pS = MicroSiemens ND = Not detected mgL = M illigrams per liter mV = M illivolts 70 Table A-2 (continued). Downflow Effluent Results DOWNFLOW EFFLUENT ANALYTE ANALYTICAL METHOD WED051994 05/19/94 mg/L VVED060194 06/01/94 mg/L YVED062994 06/29/94 mg/L WED071394 07/13/94 mg/L WED072894 07/28/94 mg/L WED081594 08/15/94 mg/L AQUEOUS ALUMINUM 6010 0.024 0.030 0.017 0.012 0.017 0.016 ARSENIC 6020 0.066 0.0013 0.0011 0.0010 0.0012 0.0011 CADMIUM 6020 0.0011 0.00073 ND ND ND 0.00033 CALCIUM 6010 107.0 112.0 106.0 118.0 116.0 114.0 IRON 6010 1.0 1.1 1.0 1.1 1.1 1.3 LEAD 6020 0.0013 0.0011 ND ND ND ND MAGNESIUM 6010 57.1 60.8 55.2 57.9 55.9 56.6 MANGANESE 6010 1.3 1.4 1.5 1.8 1.8 2.1 NICKEL 6010 0.0088 0.015 0.014 0.0089 0.013 0.013 POTASSIUM 6010 39.5 29.2 19.8 20.8 17.8 23.0 SILVER 6020 0.000063 ND 0.00010 0.00025 ND 0.00014 SODIUM 6010 15.4 15.2 13.8 14.7 14.5 15.5 ZINC 6010 9.9 10.3 12.6 15.3 16.5 14.5 ANIONS: SULFATE 300.0 270 319 338 337 354 311 SULFIDE TOTAL 376.2 3.2 2.4 2.1 1.3 6.9 1.5 FLUORIDE 340.2 0.91 0.95 0.80 0.90 1.1 1.0 CHLORIDE 300.0 17.4 18.4 19.6 17.8 19.8 19.2 PHOSPHORUS, TOTAL 365.3 11.4 10.1 8.9 9.5 7.8 8.7 ORTHOPHOSPHATE 365.3 10.6 9.2 8.6 8.6 7.5 6.7 NITRATE PLUS NITRITE AS N 353.2 ND ND ND ND 2.3 1.7 NITRITE AS N 354.1 ND ND ND ND ND ND NITRATE AS N 353.2/354.1 ND ND ND ND 2.3 1.7 AMMONIA 350.1 4.4 3.2 2.3 3.1 2.9 3.2 TOTAL SOLIDS: TSS 160.2 ND 3.6 33.6 43 45.6 43.2 TDS 160.1 739 741 709 722 747 759 TOC 9060 26.3 35.6 17.8 15.9 15.4 15.6 ALKALINITY, TOTAL: AS CaC03 310.1 196 208 188 190 188 194 ALKALINITY, BICARB AS CAC03 310.1 196 208 188 190 188 194 ORP (mV) — -271 -253 -250 NA NA pH — 7.28 7.10 7 NA 7.06 CONDUCTIVITY (pS) - 812 1040 1010 996 1006 | TEM PERATURE (degrees C) - 12.2 12.3 11.6 11.8 12.1 - = Not applicable NA = Not analyzed pS = MicroSiemens ND = Not detected mgL = Milligrams per liter mV = M illivolts 71 Table A-2 (continued). Downflow Effluent Results DOWN FLOW EF FLUENT ANALYTE ANALYTICAL METHOD WED082494 08/24/94 mg/L WED090794 09/07/94 mg/L WED091994 09/19/94 mg/L WED 100494 10/04/94 mg/L WED101994 10/19/1994 mg/L WED 110294 11/02/94 mg/L AQUEOUS ALUMINUM 6010 0.015 0.053 0.022 0.037 0.018 0.023 ARSENIC 6020 0.0011 ND 0.0011 0.0018 ND ND CADMIUM 6020 0.00030 ND ND 0.00038 0.00048 0.00041 CALCIUM 6010 117.0 113.0 124.0 115.0 112.0 112.0 IRON 6010 1.7 1.8 2.0 1.8 1.7 1.8 LEAD 6020 ND 0.0016 0.0023 0.0032 ND ND MAGNESIUM 6010 57.5 55.8 63.9 57.6 57.7 58.0 MANGANESE 6010 2.2 2.0 2.2 1.9 1.8 1.6 NICKEL 6010 0.014 0.013 0.020 0.019 0.020 0.020 POTASSIUM 6010 21.7 25.0 24.9 21.6 19.5 16.8 SILVER 6020 ND 0.00032* 0.00034 0.0012 ND ND SODIUM 6010 15.6 14.5 16.4 14.4 14.5 15.5 ZINC 6010 15.3 15.2 17.5 15.5 14.2 12.1 ANIONS: SULFATE 300.0 345 349 349 333 353 365 SULFIDE TOTAL 376.2 4.5 0.12 5.3 10.7 4.8 7.4 FLUORIDE 340.2 1.0 0.94 0.96 0.88 0.85 0.87 CHLORIDE 300.0 21.3 22.3 21.0 21.0 20.3 20.8 PHOSPHORUS, TOTAL 365.3 10.4 1.6 9.1 8.8 9.0 8.2 ORTHOPHOSPHATE 365.3 7.9 8.6 13.8 8.5 8.4 8.8 NITRATE PLUS NITRITE AS N 353.2 1.8 ND ND ND ND ND NITRITE AS N 354.1 ND ND ND ND ND ND NITRATE AS N 353.2/354.1 1.8 ND ND ND ND ND AMMONIA 350.1 3.2 2.6 2.5 2.9 2.2 1.5 TOTAL SOLIDS: TSS 160.2 48.8 49.6 47.2 52.0 45.6 40.0 TDS 160.1 713 741 738 716 698 734 TOC 9060 13.8 12.3 10.3 9.7 8.1 5.0 ALKALINITY, TOTAL: AS CaC03 310.1 191 194 184 200 174 152 ALKALINITY, BICARB AS CAC03 310.1 191 194 184 200 174 152 ORP (mV) — -125 -163 -216 -220 -331 -149 pH — 6.88 6.91 6.9 6.9 6.66 6.92 CONDUCTIVITY (pS) — 973 997 1010 960 750 890 TEMPERATURE (degrees C) - 13.4 12.4 10.7 9.0 6.8 4.9 -- = Not applicable NA = Not analyzed pS = M icroSiemens ND = Not detected mg/L = Milligrams per liter mV = M illivolts 72 Table A-2 (continued). Downflow Effluent Results DOWNFLOW EFFLUENT ANALYTE ANALYTICAL METHOD MED 112094 11/20/94 mg/L MED 113094 11/30/94 mg/L MED121494 12/14/94 mg/L MED010495 01/04/95 mg/L MED011895 01/18/95 mg/L MED02019* 02/01/95 mg/L AQUEOUS ALUMINUM 6010 0.018 0.023 0.013 0.013 0.014 0.022 ARSENIC 6020 ND ND ND 0.0039 0.0035 ND CADMIUM 6020 0.00030 0.00030 0.00088 ND ND ND CALCIUM 6010 120.0 118.0 120.0 117.0 119.0 115.0 IRON 6010 1.8 2.4 2.0 2.7 3.0 2.6 LEAD 6020 0.0054 0.0018 0.011 ND 0.0012 ND MAGNESIUM 6010 60.6 58.0 56.6 57.1 54.5 50.7 MANGANESE 6010 1.6 1.7 1.5 1.9 1.9 1.8 NICKEL 6010 0.019 0.019 0.017 0.013 0.014 0.018 POTASSIUM 6010 16.0 13.1 11.5 9.7 9.9 8.3 SILVER 6020 ND 0.00022 ND ND ND ND SODIUM 6010 14.6 14.5 15.0 14.3 14.9 15.0 ZINC 6010 10.9 11.7 8.8 8.3 9.7 10.5 ANIONS: SULFATE 300.0 357 391 391 386 386 380 SULFIDE TOTAL 376.2 0.11 5.8 3.1 3.3 1.6 2.3 FLUORIDE 340.2 0.90 1.1 0.99 1.1 1.0 1.0 CHLORIDE 300.0 21.0 22.0 21.2 22.1 22.1 21.9 PHOSPHORUS, TOTAL 365.3 6.5 7.2 7.3 6.6 6.4 6.3 ORTHOPHOSPHATE 365.3 3.1 5.0 6.2 5.5 4.9 6.0 NITRATE PLUS NITRITE AS N 353.2 ND ND ND ND ND ND NITRITE AS N 354.1 ND ND ND ND ND ND NITRATE AS N 353.2/354.1 ND ND ND ND ND ND AMMONIA 350.1 2.2 2.0 0.41 1.6 1.5 1.3 TOTAL SOLIDS: TSS 160.2 41.0 40.5 28.5 34.0 37.0 33.0 TDS 160.1 750 767 744 729 718 721 TOC 9060 6.9 20.4 5.7 4.8 5.6 4.8 ALKALINITY, TOTAL: AS CaC03 310.1 187 143 152 146 141 129 ALKALINITY, BICARB ASCAC03 310.1 187 143 152 146 141 129 ORP (mV) — -170 -220 -195 -20.0 -6.5 -7.3 pH -- 7.6 7.12 7.46 7.26 7.6 7.6 CONDUCTIVITY (pS) - NA 600 600 590 590 670 TEMPERATURE (degrees C) — 3.7 3.0 2.9 3.3 3.0 4.0 -- = Not applicable NA = Not analyzed (j.S = M icroSiemens ND = Not detected mgE = Milligrams per liter mV = Millivolts 73 Table A-2 (continued). Downflow Effluent Results DOWNFLOW EFFLUENT ANALYTE ANALYTICAL METHOD WED021595 02/15/95 mg/L WED02279* 02/27/95 mg/L WED03159S 03/15/95 mg/L WED032995 03/29/95 mg/L WED041295 04/12/95 mg/L WED04269f 04/26/95 mg/L AQUEOUS ALUMINUM 6010 0.018 0.011 0.011 ND 0.014 ND* ARSENIC 6020 0.0011 0.0019 ND ND 0.0021 ND CADMIUM 6020 0.00033 ND ND ND ND ND CALCIUM 6010 116.0 121.0 126.0 103.0 113.0 109.0 IRON 6010 2.4 2.1 2.2 1.8 1.8 1.7 LEAD 6020 0.0010 ND ND ND ND ND MAGNESIUM 6010 51.2 52.5 54.3 46.0 48.1 46.6 MANGANESE 6010 1.9 1.9 2.1 1.8 1.9 1.9 NICKEL 6010 0.016 0.015 0.018 0.019 0.014 0.014 POTASSIUM 6010 8.4 8.5 9.0 6.9 6.7 6.9 SILVER 6020 ND ND ND ND ND ND SODIUM 6010 15.9 15.1 16.5 14.7 14.1 14.1 ZINC 6010 10.7 11.7 13.0 12.2 12.6 11.9 ANIONS: SULFATE 300.0 359 346 370 341 338 341 SULFIDE TOTAL 376.2 1.9 1.9 3.1 3.1 0.099 1.6 FLUORIDE 340.2 1.1 1.0 1.1 1.1 1.0 1.1 CHLORIDE 300.0 22.1 22.7 24.4 22.5 21.8 23.8 PHOSPHORUS, TOTAL 365.3 17.5 5.9 5.7 5.2 4.7 4.7 ORTHOPHOSPHATE 365.3 5.1 5.8 5.7 3.8 5.4 2.4 NITRATE PLUS NITRITE AS N 353.2 ND ND ND ND ND ND NITRITE AS N 354.1 ND ND ND ND ND ND NITRATE AS N 353.2/354.1 ND ND ND ND ND ND AMMONIA 350.1 1.2 1.2 1.5 1.3 1.1 1.1 TOTAL SOLIDS: TSS 160.2 32.4 34.0 33.0 31.0 35.0 31.2 TDS 160.1 679 723 707 662 655 651 TOC 9060 4.3 5.5 5.4 5.8 6.9 6.8 ALKALINITY, TOTAL: AS CaC03 310.1 140 152 152 141 143 141 ALKALINITY, BICARB AS CAC03 310.1 140 152 152 141 143 141 ORP (mV) — 59.0 -82.0 -65.0 -81.1 35.0 NA pH - 8.8 7.1 7.1 7.3 7.2 NA CONDUCTIVITY (pS) - NA 620 680 580 580 600 TEM PERATURE (degrees C) - 2.8 5.6 6.8 5.6 4.8 7.0 -- = Not applicable NA = Not analyzed pS = M icroSiemens ND = Not detected mg/L = Milligrams per liter mV = M illivolts 74 Tabic A-2 (continued). Downflow Effluent Results DOWNFLOW EFFLUENT ANALYTE ANALYTICAL METHOD WED051095 05/10/95 mg/L WED061295 6/12/1995 mg/L WED062895 6/28/1995 mg/L YVED071095 7/10/1995 mg/L VYED072695 7/26/1995 mg/L VYED08089f 8/8/1995 mg/L AQUEOUS ALUMINUM 6010 ND* ND ND ND ND 0.015 ARSENIC 6020 ND ND ND ND ND ND CADMIUM 6020 ND ND ND ND ND ND CALCIUM 6010 121.0 125 142 144 157 148 IRON 6010 2.1 4.2 3.9 3.9 2.9 2.8 LEAD 6020 ND ND ND ND ND ND MAGNESIUM 6010 47.8 52.7 61.9 68.7 71.7 68.6 MANGANESE 6010 2.4 3.9 4.4 4.1 4.1 3.8 NICKEL 6010 0.016 0.017 0.020 0.021 0.020 0.022 POTASSIUM 6010 6.5 6.8 7.1 8.2 7.6 6.8 SILVER 6020 ND ND ND ND ND ND SODIUM 6010 14.1 8.7 10.6 12.8 12.6 12.5 ZINC 6010 13.3 26.5 31.2 30.8 29.7 33.1 ANIONS: SULFATE 300.0 348.0 425 453 525 537 535 SULFIDE TOTAL 376.2 0.38 0.054 6.9 5.7 0.83 10.0 FLUORIDE 340.2 1.1 0.87 0.80 0.96 0.86 0.91 CHLORIDE 300.0 22.6 7.0 7.2 8.6 10.1 11.1 PHOSPHORUS, TOTAL 365.3 4.3 3.7 4.7 3.5 2.6 2.5 ORTHOPHOSPHATE 365.3 4.1 2.2 1.5 3.7 2.0 1.6 NITRATE PLUS NITRITE AS N 353.2 ND ND ND ND ND ND NITRITE AS N 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND ND ND ND ND ND AMMONIA 350.1 0.96 0.90 0.94 1.0 0.50 0.64 TOTAL SOLIDS: TSS 160.2 29.2 43.0 53.6 48.0 28.0 38.8 TDS 160.1 707 763 918 946 959 1090 TOC 9060 4.4 6.6 11.4 5.4 7.2 4.7 ALKALINITY, TOTAL: AS CaC03 310.1 137 129 195 146 141 ALKALINITY, BICARB 129 195 AS CAC03 310.1 137 146 141 ORP (mV) -80 -68 -52 14 pH 6.8 6.6 6.7 7.1 CONDUCTIVITY (pS) NA 720 NA 850 TEMPERATURE (degrees C) 11.7 12.3 13.8 14.1 * - Aluminum was re-analyzed 6/2/95 due to blank contamination -- = Not applicable mV = Millivolts pS = M icroSiemens NA = Not analyzed mgd = M illigrams per liter ND = Not detected 75 Tabic A-2 (continued). Downflow Effluent Results DOWNFLOW EFFLUENT WED082395 VVED090595 WED110995 CDPHE CDPHE CDPHE ANALYTICAL 8/23/1995 9/5/1995 11/9/1995 1/29/1996 2/29/1996 4/25/1996 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 ND 0.016 ND NA NA NA ARSENIC 6020 ND ND ND NA NA NA CADMIUM 6020 ND ND 0.00030 0.00012 0.00072 0.15 CALCIUM 6010 155 147 149 NA NA NA IRON 6010 2.7 2.2 2.4 NA 0.28 1.7 LEAD 6020 ND ND 0.0016 NA NA NA MAGNESIUM 6010 70.2 66.3 66.2 NA NA NA MANGANESE 6010 3.9 3.7 4.0 3.2 3.0 2.2 NICKEL 6010 0.026 0.028 0.04 NA NA NA POTASSIUM 6010 6.2 6.2 5.3 NA NA NA SILVER 6020 ND ND ND NA NA NA SODIUM 6010 13.7 12.5 14.5 NA NA NA ZINC 6010 34.1 29.1 34.5 28 26 15 ANIONS: SULFATE 300.0 539 529 535 440 430 318 SULFIDE TOTAL 376.2 11.4 5.6 3.8 NA NA NA FLUORIDE 340.2 0.85 0.82 0.81 NA NA NA CHLORIDE 300.0 12.2 14 17.3 NA NA NA PHOSPHORUS, TOTAL 365.3 3.0 2.8 2.5 NA NA NA ORTHOPHOSPHATE 365.3 3.0 1.3 1.1 NA NA NA NITRATE PLUS NITRITE AS N 353.2 ND ND ND NA NA NA NITRITE AS N 354.1 0.0070 ND ND NA NA NA NITRATE AS N 353.2/354.1 ND ND ND NA NA NA AMMONIA 350.1 0.78 0.64 0.39 1.0 1.1 1.1 TOTAL SOLIDS: TSS 160.2 50.0 45.6 12.8 NA NA NA TDS 160.1 996 941 957 NA NA NA TOC 9060 4.2 4.9 4.2 NA NA NA ALKALINITY, TOTAL: AS CaC03 310.1 143 179 152 NA NA NA ALKALINITY, BICARB AS CAC03 310.1 143 179 152 NA NA NA ORP (mV) -60 NA NA NA pH 6.7 NA NA NA CONDUCTIVITY (pS) 750 NA NA NA TEMPERATURE (degrees C) 4.7 NA NA NA — = Not applicable NA = Not analyzed pS = M icroSiemens ND = Not detected mg/L = Milligrams per liter mV = M illivolts 76 Table A-2 (continued). Downflow Effluent Results DOWNFLOW EFFLUENT ANALYTE ANALYTICAL METHOD CDPHE 5/31/1996 mg/L CDPHE 6/14/1996 mg/L CDPHE 7/19/1996 mg/L CDPHE 8/31/1996 mg/L WED012197 1/21/1997 mg/1 WED022097 2/20/1997 mg/1 AQUEOUS ALUMINUM 6010 NA NA NA NA 0.098 ND ARSENIC 6020 NA NA NA NA NA NA CADMIUM 6020 0.00016 ND 0.00021 0.00013 0.016 0.034 CALCIUM 6010 NA NA NA NA 115 113 IRON 6010 0.87 0.92 1.10 1.60 0.53 0.72 LEAD 6020 NA NA NA NA 57.3 56.9 MAGNESIUM 6010 NA NA NA NA 3.3 5.0 MANGANESE 6010 1.8 2.00 2.10 2.20 NA NA NICKEL 6010 NA NA NA NA 0.05 0.035 POTASSRJM 6010 NA NA NA NA 0.39 3.80 SILVER 6020 NA NA NA NA NA NA SODIUM 6010 NA NA NA NA 16.6 16 ZINC 6010 11 9.7 8.7 5.8 55 59.7 ANIONS: SULFATE 300.0 230 82 340 350 421 322 SULFIDE TOTAL 376.2 NA NA NA NA 0.13 ND FLUORIDE 340.2 NA NA NA NA NA NA CHLORIDE 300.0 NA NA NA NA 18.6 18.6 PHOSPHORUS, TOTAL 365.3 NA NA NA NA NA NA ORTHOPHOSPHATE 365.3 NA NA NA NA 1.1 0.54 NITRATE PLUS NITRITE AS N 353.2 NA NA NA NA ND 0.2 NITRITE AS N 354.1 NA NA NA NA 0.0025 0.0055 NITRATE AS N 353.2/354.1 NA NA NA NA NA NA AMMONIA 350.1 0.67 1.2 0.90 ND 0.24 0.20 TOTAL SOLIDS: TSS 160.2 NA NA NA NA 7.2 6.0 TDS 160.1 NA NA NA NA 787 752 TOC 9060 NA NA NA NA 7.2 25.8 ALKALINITY, TOTAL: AS CaC03 310.1 NA NA NA NA 158 259 ALKALINITY, BICARB AS CAC03 310.1 NA NA NA NA 158 259 ORP(mV) NA NA NA NA 110 92.0 pH NA NA NA NA 5.3 7.0 CONDUCTIVITY (uS) NA NA NA NA NA NA TEMPERATURE (degrees C) NA NA NA NA 1.8 1.8 -- = Not applicable NA = Not analyzed |iS = M icroSiemens ND = Not detected mg/L = Milligrams per liter mV= Millivolts 77 Table A-3. Upflow Effluent Results UPFLOW EFFLUENT WEU030994 WEU032394 WEU040694 WEU042094 WEU050594 YVEU051994 ANALYTICAL 03/09/94 03/23/94 04/06/94 04/20/94 05/05/94 05/19/94 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L ALUMINUM 6010 0.077 0.20 0.078 0.39 0.062 0.028 ARSENIC 6020 0.0062 0.0071 0.036 0.028 0.085 0.067 CADMIUM 6020 0.00042 0.00049 0.00034 0.00036 0.00024 0.00020 CALCIUM 6010 75.3 96.2 112.0 115.0 123.0 115.0 IRON 6010 0.48 0.61 0.48 0.99 0.27 0.25 LEAD 6020 0.0042 0.0030 0.0038 0.020 0.0022 0.0015 MAGNESIUM 6010 72.7 71.4 69.3 63.1 66.0 60.1 MANGANESE 6010 0.051 0.072 0.065 0.16 0.17 0.25 NICKEL 6010 0.0054 0.0071 0.0095 0.0086 0.0086 0.0086 POTASSIUM 6010 223.0 188.0 150.0 108.0 91.2 49.4 SILVER 6020 0.0014 0.00015 0.000084 0.00048 0.000071 0.000072 SODIUM 6010 33.9 31.2 27.3 21.8 22 16.8 ZINC 6010 0.22 0.22 0.13 0.43 0.14 0.32 ANIONS SULFATE 300.0 354 388 364 343 292 265 SULFIDE TOTAL 376.2 0.38 7.9 9.4 1.9 0.47 2.4 FLUORIDE 340.2 0.30 0.57 0.62 0.72 0.71 0.88 CHLORIDE 300.0 83.2 76.0 59.7 50.0 35.5 21.8 PHOSPHORUS TOTAL 365.3 24.3 23.2 20.5 20.8 18.3 17.6 ORTHOPHOSPHATE 365.3 26.8 26.7 20.9 20.6 18.6 15.9 NITRATE PLUSNITRITE ASN 353.2 ND ND 0.060 ND ND ND NITRITE ASN 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND ND 0.060 ND ND ND AMMONIA 350.1 23.8 19.6 15.0 12.9 10.5 6.8 TOTAL SOLIDS TSS 160.2 6 12.0 6.0 25.2 ND ND TDS 160.1 1390 1200 1110 1010 934 804 TOC 9060 264 51.3 60.0 49.3 35.6 23.8 ALKALINITY, TOTAL: AS CaC03 310.1 367 347 310 308 265 230 ALKALINITY, BICARB ASCAC03 310.1 367 347 310 308 265 230 ORP (mV) - -377 -280 -269 -271 PH - 8 7.85 7.20 7.84 CONDUCTIVITY (pS) - 1410 1222 954 893 TEMPERATURE (degrees Q - 5 6.0 7.8 8.8 AQUEOUS — = Not applicable NA = Not analyzed p/s = MicroSemens ND = Not detected mgT = Milligrams per liter mV= Millivolts 78 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT ANALYTE ANALYTICAL METHOD WEU060194 06/01/94 mg/L WEU062994 06/29/94 mg/L WEU071394 07/13/95 mg/L WEU072894 07/28/95 mg/L WEU081594 08/15/94 mg/L WEU082494 08/24/94 mg/L AQUEOUS ALUMINUM 6010 0.045 0.021 ND 0.38 0.015 0.023 ARSENIC 6020 ND ND ND ND ND ND CADMIUM 6020 ND ND ND ND ND ND CALCIUM 6010 117 120 132 132 134 132 IRON 6010 0.26 0.47 0.79 1.4 2.7 3.3 LEAD 6020 0.0030 0.0017 ND ND ND ND MAGNESIUM 6010 61.5 61.7 61.4 58.6 58.3 57.1 MANGANESE 6010 0.33 0.79 1.3 1.7 2.1 2.3 NICKEL 6010 0.014 0.011 0.0052 0.0075 0.0089 0.0077 POTASSIUM 6010 37.3 24.2 17.3 13.7 12.8 11.3 SILVER 6020 ND 0.00014 0.00015 ND 0.00021 ND SODIUM 6010 15.7 15.6 15 14.2 14.4 14.4 ZINC 6010 0.20 0.35 0.18 0.29 0.38 0.58 ANIONS SULFATE 300.0 330 355 372 356 369 392 SULFIDE TOTAL 376.2 5 3.2 0.59 1.5 0.69 1.0 FLUORIDE 340.2 0.81 0.90 0.80 1.0 0.96 1.1 CHLORIDE 300.0 22.2 20.9 18.9 20.2 19.9 20.5 PHOSPHORUS TOTAL 365.3 27.3 12.8 13.3 10.8 10.5 9.8 ORTHOPHOSPHATE 365.3 14.9 21.3 19.5 10.5 7.8 9.2 NITRATE PLUSNITRITE ASN 353.2 ND ND ND 1.9 1.7 1.8 NITRITE ASN 354.1 ND ND ND ND 0.077 ND NITRATE ASN 353.2/354.1 ND ND ND 1.9 1.7 1.8 AMMONIA 350.1 5.6 3.0 3.0 2.6 1.6 1.3 TOTAL SOLIDS TSS 160.2 ND 2.4 2.0 18.8 7.6 27.2 TDS 160.1 808 759 766 816 802 767 TOC 9060 28.0 11.4 9.0 9.6 8.8 6.0 ALKALINITY, TOTAL: AS CaC03 310.1 244 220 211 206 194 183 ALKALINITY, BICARB ASCAC03 310.1 244 220 211 206 194 183 ORP (mV) — -275 -280 NA NA -344 PH - 7.7 7.6 NA 7.6 7.46 CONDUCTIVITY (pS) - 1115 1090 1049 1069 1037 TEMPERATURE (degrees Q - 9.7 9.4 9.7 9.4 10.0 - = Not applicable NA = Not analyzed p/s = MicroSemens ND = Not detected mg/L = Milligrams per liter mV= Millivolts 79 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT WEU090794 k'EU090794E WEU091994 WEU100494 WB5100494 WEU101994 ANALYTICAL 09/07/94 09/07/94 09/19/94 10/04/94 10/04/94 10/19/94 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 0.052 ND 0.023 0.017 6.1 0.015 ARSENIC 6020 ND ND ND 0.0011 0.0021 0.0011 CADMIUM 6020 ND ND ND ND 0.024 ND CALCIUM 6010 126 0.15 132 127 344 128 IRON 6010 4.3 ND 5.1 5.7 92.7 5.6 LEAD 6020 0.0011 ND 0.0015 ND 0.020 ND MAGNESIUM 6010 53.3 ND 56.6 54.5 139.0 54.1 MANGANESE 6010 2.4 ND 2.6 2.4 28.6 2.7 NICKEL 6010 0.0083 ND 0.015 0.015 0.20 0.019 POTASSIUM 6010 10.2 ND 9.0 11.9 7.4 7.7 SILVER 6020 0.00011* ND* 0.00046 0.00052 0.00099 ND SODIUM 6010 13.6 ND 14.2 13.8 46.8 14.6 ZINC 6010 0.82 0.019 1.4 2.4 9.4 3.1 ANIONS SULFATE 300.0 395 ND 391 369 1760 392 SULFIDE TOTAL 376.2 0.12 ND 0.23 5.0 NS 1.3 FLUORIDE 340.2 1.0 ND 1.1 0.99 1.0 0.95 CHLORIDE 300.0 21.1 ND 20.4 21.4 6.0 20.2 PHOSPHORUS TOTAL 365.3 1.6 ND 7.9 8.0 NS 6.8 ORTHOPHOSPHATE 365.3 8.8 ND 9.8 7.1 ND 6.8 NITRATE PLUSNITRITE ASN 353.2 ND ND ND ND NS ND NITRITE ASN 354.1 ND ND ND ND ND 0.018 NITRATE ASN 353.2/354.1 ND ND ND ND NS ND AMMONIA 350.1 1.0 ND 0.87 1.0 NS 0.51 TOTAL SOLIDS TSS 160.2 27.6 ND 28.8 37.6 49.6 40.8 TDS 160.1 787 ND 790 750 2520 734 TOC 9060 6.4 ND 5.8 7.4 NS 5.3 ALKALINITY, TOTAL: AS CaC03 ALKALINITY, BICARB 310.1 175 ND 164 182 ND 150 ASCAC03 310.1 175 ND 164 182 ND 150 ORP (mV) - -315 -267 -260 NA -344 pH -- 7.39 7.3 7.3 5.2 6.95 CONDUCTIVITY (pS) -- 1007 990 960 NA 760 TEMPERATURE (degrees C) -- 9.3 9.2 8.7 15.0 7.7 — = Not applicable NA = Not analyzed |Vs = MicroSiemens ND = Not detected mgT. = Milligrams per liter mV= Millivolts 80 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT ANALYTE ANALYTICAL METHOD VEU101994I 10/19/94 mg/L WEU110294 11/02/94 mg/L WEU112094 11/20/94 mg/L WEU113094 11/30/94 mg/1 WEU121494 12/14/94 mg/L WEU010495 01/04/95 mg/L AQUEOUS ALUMINUM 6010 0.025 0.025 0.012 0.013 0.015 0.020 ARSENIC 6020 ND 0.0011 0.0011 0.0014 0.0010 0.0035 CADMIUM 6020 ND ND ND ND ND ND CALCIUM 6010 130 122 127 123 127 116 IRON 6010 5.7 7.0 6.0 7.5 6.8 6.3 LEAD 6020 ND ND ND ND ND ND MAGNESIUM 6010 54.8 52.5 53.8 51.4 52.4 53.1 MANGANESE 6010 2.7 2.7 2.8 2.9 2.9 2.7 NICKEL 6010 0.018 0.018 0.016 0.018 0.016 0.012 POTASSIUM 6010 7.6 11.6 9.6 7.6 7.6 15.3 SILVER 6020 ND ND 0.00082 0.00028 ND 0.00033 SODIUM 6010 15 14.2 14.7 14.5 15.6 14.5 ZINC 6010 3.2 6.8 6.5 7.9 9.0 11.7 ANIONS; SULFATE 300.0 380 371 360 379 375 341 SULFIDE TOTAL 376.2 1.8 3.8 3.8 4.6 3.2 3.3 FLUORIDE 340.2 0.97 1.1 1.0 1.2 1.1 1.1 CHLORIDE 300.0 20.0 23.2 23.0 22.2 22.4 25.6 PHOSPHORUS, TOTAL 365.3 6.9 6.2 5.5 6.9 5.3 4.8 ORTHOPHOSPHATE 365.3 6.2 5.9 2.7 2.7 4.7 3.0 NITRATE PLUSNITRITE ASN 353.2 ND ND ND ND ND ND NITRITE ASN 354.1 0.017 0.016 ND ND ND ND NITRATE ASN 353.2/354.1 ND ND ND ND ND ND AMMONIA 350.1 0.52 0.38 0.74 0.55 1.5 0.68 TOTAL SOLIDS: TSS 160.2 36.8 52.0 49.0 47.0 44.0 51.0 TDS 160.1 742 727 745 729 729 707 TOC 9060 5.6 9.4 7.3 19.1 6.4 12.5 ALKALINITY, TOTAL: ASCaC03 310.1 148 141 185 142 157 171 ALKALINITY, BICARB ASCAC03 310.1 148 141 185 142 157 171 ORP (mV) - -344 -164 -160 -216 -196 -80 pH - 6.95 7.01 7.2 6.8 7.33 7.0 CONDUCTIVITY (pS) -- 760 935 NA 640 670 670 TEMPERATURE (degrees C) - 7.7 8.5 8.1 7.1 7.7 7.0 — = Not applicable NA - Not applicable ps = MicroSemens ND - Not detected mg/L = Milligrams per liter mV= Millivolts 81 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT WEU011895 WEU020195 WEU021595 WEU022795 WEU031595 WEU032995 ANALYTICAL 01/18/95 02/01/95 02/15/95 02/27/95 03/15/95 03/29/95 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 0.026 0.017 0.012 0.014 0.012 0.015 ARSENIC 6020 0.0043 0.0015 0.0020 0.0021 0.0012 0.0012 CADMIUM 6020 ND ND ND ND ND ND CALCIUM 6010 116 119 119 116 116 105 IRON 6010 5.4 4.9 4.3 4.0 4.0 3.5 LEAD 6020 0.0034 ND ND ND ND ND MAGNESIUM 6010 49.5 49.0 49.1 48.2 48.2 44.6 MANGANESE 6010 2.6 2.5 2.5 2.4 2.4 2.2 NICKEL 6010 0.012 0.016 0.016 0.016 0.019 0.019 POTASSIUM 6010 10.5 9.1 9.1 8.9 7.5 5.9 SILVER 6020 ND ND Nr ND ND ND SODIUM 6010 15 16.7 16 15.2 16.0 15.7 ZINC 6010 12.5 16.9 12.9 17.8 18.0 17.5 ANIONS: SULFATE 300.0 347 330 308 340 335 317 SULFIDE TOTAL 376.2 3.0 6.0 3.3 4.3 2.7 4.3 FLUORIDE 340.2 1.1 1.0 1.1 1.0 1.1 1.1 CHLORIDE 300.0 23.0 23.4 23.4 23.6 24.3 23.0 PHOSPHORUS, TOTAL 365.3 4.7 5.5 13.3 3.4 4.1 3.4 ORTHOPHOSPHATE 365.3 3.0 5.0 3.7 3.0 2.6 1.8 NITRATE PLUSNITRITE ASN 353.2 ND 1.4 ND ND ND ND NITRITE ASN 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND 1.4 ND ND ND ND AMMONIA 350.1 0.63 0.52 0.51 0.34 0.38 0.31 TOTAL SOLIDS: TSS 160.2 51.0 54.0 45.2 47.0 39.0 41.0 TDS 160.1 693 692 682 700 671 667 TOC 9060 7.8 6.8 6.2 5.8 4.3 7.0 ALKALINITY, TOTAL: AS CaC03 310.1 168 161 191 150 151 154 ALKALINITY. BICARB AS CAC03 310.1 168 161 191 150 151 154 ORP (mV) - 5 -11.7 -44.0 -65 -63 -81.1 pH - 7.1 7.4 7.2 6.9 6.9 7.3 CONDUCTIVITY (pS) -- 650 610 NA 680 650 580 TEMPERATURE (degrees C) -- 8.3 6.1 7.6 8.4 8.8 5.6 — = Not applicable NA = Not analyzed pS = MicroSiemens ND = Not detected mgl. = Milligrams per liter mV = Millivolts 82 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT ANALYTE ANALYTICAL METHOD WEU041295 04/12/95 mg/L WEU042695 04/26/95 mg/L WEU051095 05/10/95 mg/L WEU061295 06/12/95 mg/L WEU062895 6/28/1995 mg/L WEU071095 7/10/1995 mg/L AQ UEOUS ALUMINUM 6010 0.013 ND* ND* 0.028 ND ND ARSENIC 6020 0.0028 ND ND ND ND ND CADMIUM 6020 ND 0.00078 0.0094 0.0084 0.0045 ND CALCIUM 6010 114 106 110 103 121 130 IRON 6010 3.5 2.2 2.2 4.6 3.7 3.8 LEAD 6020 ND ND 0.0019 0.0018 ND ND MAGNESIUM 6010 46.5 45.3 44.5 45.2 60.2 68.2 MANGANESE 6010 2.5 2.0 2.5 3 4.0 4.1 NICKEL 6010 0.013 0.015 0.022 0.019 0.026 0.026 POTASSIUM 6010 7.1 11.7 18.1 7.5 6.0 5.4 SILVER 6020 ND ND ND ND ND ND SODIUM 6010 15.3 14.2 13.3 8.9 11.2 13.2 ZINC 6010 15.9 18.5 26.7 33.5 47.1 50.8 ANIONS: SULFATE 300.0 326 326 355 326 494 514 SULFIDE TOTAL 376.2 0.39 2.9 1.3 0.065 1.5 1.5 FLUORIDE 340.2 1.0 1.1 1.2 0.90 0.90 0.96 CHLORIDE 300.0 22.5 26.0 25.9 7.6 7.0 8.3 PHOSPHORUS, TOTAL 365.3 3.0 3.2 2.0 2.3 1.2 1.5 ORTHOPHOSPHATE 365.3 2.7 1.6 2.4 1.5 0.34 0.48 NITRATE PLUS NITRITE ASN 353.2 ND ND ND ND ND ND NITRITE ASN 354.1 ND ND ND ND ND ND NITRATE ASN 353.2/354.1 ND ND ND ND ND ND AMMONIA 350.1 0.33 0.31 0.42 0.36 0.20 0.20 TOTAL SOLIDS: TSS 160.2 41.9 31.2 29.0 47.3 18.8 25.6 TDS 160.1 657 607 724 668 885 944 TOC 9060 8.0 8.3 9.9 9.1 4.9 4.8 ALKALINITY, TOTAL: ASCaC03 310.1 152 147 138 181 136 147 ALKALINITY, BICARB ASCAC03 310.1 152 147 138 181 136 147 ORP (mV) — -7.0 NA -57 pH — 7.1 NA 6.7 6.9 6.9 CONDUCTIVITY (pS) - 620 620 NA TEMPERATURE (degrees C) - 8.5 8.0 10.1 * - Aluminum was re-analyzed 6/2/95 due to blank contamination — = Not applicable Not analyzed jjS = MicroSiemens ND = Not detected mg/L = Milligrams per liter mV = Millivolts 83 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT WEU072695 VVEU080895 WEU082395 WEU090595 WEU110995 CDPHE ANALYTICAL 7/26/1995 8/8/1995 8/23/1995 9/5/1995 11/9/1995 1/29/1996 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 ND ND ND ND ND NA ARSENIC 6020 ND ND ND ND ND NA CADMIUM 6020 0.0060 0.0046 0.0093 0.010 0.04400 0.037 CALCIUM 6010 144 135 141 137 133 NA IRON 6010 2.5 2.5 2.1 1.8 0.93 1.6 LEAD 6020 ND ND ND ND 0.0022 NA MAGNESIUM 6010 68.6 64.4 66.1 64.3 62.1 NA MANGANESE 6010 4.1 3.8 3.8 3.6 4.4 3.3 NICKEL 6010 0.028 0.032 0.036 0.04 0.059 NA POTASSIUM 6010 5.7 4.9 4.5 ND 4.3 NA SILVER 6020 ND ND ND ND ND NA SODIUM 6010 12.2 13.3 14.0 12.2 15.6 NA ZINC 6010 53.2 56.6 59.8 59.9 73.6 47 ANIONS: SULFATE 300.0 549 584 561 569 559 460 SULFIDE TOTAL 376.2 4.3 3.5 5.2 2.8 0.84 NA FLUORIDE 340.2 0.89 0.88 0.90 0.86 0.96 NA CHLORIDE 300.0 10.0 11.2 12.5 13.7 17.1 NA PHOSPHORUS, TOTAL 365.3 1.1 1.1 1.3 1.5 0.69 NA ORTHOPHOSPHATE 365.3 1.0 0.9 1.2 0.43 0.80 NA NITRATE PLUSNITRITE ASN 353.2 ND ND ND ND ND NA NITRITE ASN 354.1 ND ND 0.0080 ND ND NA NITRATE ASN 353.2/354.1 ND ND ND ND ND NA AMMONIA 350.1 0.11 ND 0.21 ND ND 0.2 TOTAL SOLIDS: TSS 160.2 16.8 17.6 30.0 26 5.2 NA TDS 160.1 961 999 1010 978 932 NA TOC 9060 5.7 3.4 3.2 3.7 2.1 NA ALKALINITY, TOTAL: AS CaC03 310.1 135 138 149 160 115 NA ALKALINITY, BICARB AS CAC03 310.1 135 138 149 160 115 NA ORP (mV) — PH — 6.8 7.1 7.1 6.9 7.0 NA CONDUCTIVITY (pS) -- NA TEMPERATURE (degrees C) — NA — = Not applicable NA = Not analyzed pS = MicroSemens ND = Not detected mg/L = Milligrams per liter mV = Millivolts 84 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT CDPHE CDPHE CDPHE CDPHE CDPHE CDPHE ANALYTICAL 2/29/1996 4/25/1996 5/31/1996 6/14/1996 7/19/1996 8/31/1996 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 NA NA NA NA NA NA ARSENIC 6020 NA NA NA NA NA NA CADMIUM 6020 0.035 0.030 0.140 0.031 0.051 0.053 CALCIUM 6010 NA NA NA NA NA NA IRON 6010 1.3 0.81 0.17 1.1 0.87 0.90 LEAD 6020 NA NA NA NA NA NA MAGNESIUM 6010 NA NA NA NA NA NA MANGANESE 6010 3.1 2.3 2.7 2.2 2.6 2.5 NICKEL 6010 NA NA NA NA NA NA POTASSIUM 6010 NA NA NA NA NA NA SILVER 6020 NA NA NA NA NA NA SODIUM 6010 NA NA NA NA NA NA ZINC 6010 42 31 56 30 41 43.0 ANIONS: SULFATE 300.0 430 329 420 310 410 45 SULFIDE TOTAL 376.2 NA NA NA NA NA NA FLUORIDE 340.2 NA NA NA NA NA NA CHLORIDE 300.0 NA NA NA NA NA NA PHOSPHORUS, TOTAL 365.3 NA NA NA NA NA NA ORTHOPHOSPHATE 365.3 NA NA NA NA NA NA NITRATE PLUSNITRITE ASN 353.2 NA NA NA NA NA NA NITRITE ASN 354.1 NA NA NA NA NA NA NITRATE ASN 353.2/354.1 NA NA NA NA NA NA AMMONIA 350.1 0.4 0.3 ND 0.2 0.2 ND TOTAL SOLIDS: TSS 160.2 NA NA NA NA NA NA TDS 160.1 NA NA NA NA NA NA TOC 9060 NA NA NA NA NA NA ALKALINITY, TOTAL: ASCaC03 310.1 NA NA NA NA NA NA ALKALINITY, BICARB ASCAC03 310.1 NA NA NA NA NA NA ORP (mV) — NA NA NA NA NA NA PH - NA NA NA NA NA NA CONDUCTIVITY (pS) — NA NA NA NA NA NA TEMPERATURE (degrees C) -- NA NA NA NA NA NA — = Not applicable NA = Not analyzed pS = MicroSiemens ND = Not detected mgT, = Milligrams per liter mV = Millivolts 85 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT WEU120996 YVEU012197 WEU022097 VVEU031997 WEU042297 ANALYTICAL 12/09/96 01/21/97 02/20/96 03/19/97 04/22/97 04/22/97 ANALYTE METHOD mg/L mg/L mg/L mg/L mg/L mg/L AQUEOUS ALUMINUM 6010 ND ND ND ND ND ND (D) ARSENIC 6020 NA NA NA NA NA NA CADMIUM , 6020 0.088 0.032 0.057 0.034 0.015 0.015 (D) CALCIUM 6010 115 116 119 109 95.9 97.3 (D) IRON 6010 0.99 1.2 0.8 1.1 0.98 0.99 (D) LEAD 6020 NA NA NA NA NA NA MAGNESIUM 6010 53.4 52.8 55.4 47.8 43.2 43.9 (D) MANGANESE 6010 2.9 2.7 3 2.8 2 2(D) NICKEL 6010 0.035 J 0.032 J 0.033 J 0.041 0.021 J 0.021 J (D) POTASSIUM 6010 3.6 J 3.5 J 4.4 ND 4.6 J 4.6 J SILVER 6020 NA NA NA NA NA NA SODIUM 6010 15.3 15.5 15.7 14.4 7.3 7(D) ZINC 6010 46 41.3 48.6 38 22.7 22.9 (D) ANIONS: SULFATE 300.0 434 400 413 392 252 248 (D) SULFIDE TOTAL 376.2 0.8 1.2 0.71 0.037 J 3.5 3.6(D) FLUORIDE 340.2 NA NA NA NA NA NA CHLORIDE 300.0 17.4 18.1 19.4 18.5 12.4 12.2 (D) PHOSPHORUS, TOTAL 365.3 NA NA NA NA NA NA ORTHOPHOSPHATE 365.3 0.84 1.7 1.1 2 1.7 1.8(D) NITRATE PLUSNITRITE ASN 353.2 ND ND ND ND 0.020 J ND (D) NITRITE ASN 354.1 ND 0.0057 J 0.0055 J 0.0058 J 0.0038 J .0036 J(D) NITRATE ASN 353.2/354.1 NA NA NA NA NA NA AMMONIA 350.1 0.070 J 0.17 0.11 0.19 0.22 0.21 (D) TOTAL SOLIDS TSS 160.2 NA 17.2 7.2 16.4 21.6 23.2 (D) TDS 160.1 806 773 800 712 575 574 (D) TOC 9060 5.6 5.2 5.3 5 7.6 7.7 (D) ALKALINITY, TOTAL: ASCaC03 310.1 158 178 160 153 180 179 (D) ALKALINITY, BICARB ASCAC03 310.1 158 178 160 153 180 179 (D) ORP (mV) - 94 108 80 82 72 72 pH - 6.5 5.4 6.7 6.3 6.2 6.2 CONDUCTIVITY (pS) - NA NA NA NA NA NA TEMPERATURE (degrees C) - 5.1 3.2 5.0 6.0 7.0 7.0 — = Not applicable NA = Not analyzed jjS = MicroSiemens ND = Not detected mg/1 = Milligrams per liter mV = Millivolts 86 Table A-3 (continued). Upflow Effluent Results UPFLOW EFFLUENT ANALYTE ANALYTICAL METHOD WEU052897 05/28/97 mg/L WEU062397 06/23/97 mg/L WEU082897 8/28/1997 mg/L WEU093097 9/30/1997 mg/L WEU102997 10/29/1997 mg/L WEU112597 11/25/1997 mg/L AQUEOUS ALUMINUM 6010 ND ND 0.10 0.078 ND ND ARSENIC 6020 NA NA NA NA NA NA CADMIUM 6020 0.2 ND 0.0063 0.0040 0.010 0.016 CALCIUM 6010 99.6 113 153 152 144 138 IRON 6010 3.3 1.2 4.0 2.9 1.2 1 LEAD 6020 NA NA NA NA NA NA MAGNESIUM 6010 48.9 53.7 64.6 64.8 65.6 56.6 MANGANESE 6010 2.1 2.2 3.0 2.7 3.6 3.0 NICKEL 6010 0.022 J ND 0.023 ND ND ND POTASSIUM 6010 4.0 J 4.9 J 5.2 4.5 3.9 ND SILVER 6020 NA NA NA NA NA NA SODIUM 6010 14.8 0.94 J 6.9 13.5 14.7 15.0 ZINC 6010 60.1 25.4 21.2 14.8 26.4 24.6 ANIONS: SULFATE 300.0 250 275 308 311 484 460 SULFIDE TOTAL 376.2 17 6.1 2.4 4.1 2.2 2.4 FLUORIDE 340.2 NA NA NA NA NA NA CHLORIDE 300.0 6.9 9.2 12.9 15.5 16.7 17.8 PHOSPHORUS, TOTAL 365.3 NA NA NA NA NA NA ORTHOPHOSPHATE 365.3 1.1 1.9 0.70 2.7 1.1 1.9 NITRATE PLUSNITRITE ASN 353.2 ND ND 0.034 ND ND NITRITE ASN 354.1 0.0051 J 0.0047J 0.0040 ND 0.0020J 0.0032J NITRATE ASN 353.2/354.1 NA NA NA NA NA NA AMMONIA 350.1 ND 0.4 1.7 1.3 0.64 1.0 TOTAL SOLIDS: TSS 160.2 102 38 31.6 13.6 27.2 TDS 160.1 566 683 808 865 892 887 TOC 9060 5.3 16 29.7 18.8 6.7 6.6 ALKALINITY, TOTAL: AS CaC03 310.1 190 228 NA 317 166 199 ALKALINITY, BICARB ASCAC03 310.1 190 228 NA 317 166 199 ORP (mV) — -58 47 30 -37 NA 49 PH — 6.4 6.8 5.7 6.2 6.7 6.7 CONDUCTIVITY (pS) - NA NA NA NA NA NA TEMPERATURE (degrees C) - 9.2 12.7 12.0 10.3 5.7 5.5 — = Not applicable NA = Not analyzed pS = MicroSiemens ND = Not detected mg/L = Milligrams per liter mV = Millivolts 87 Table A-4. Substrate Results - Downflow Cell SUBSTRATE - DOWNFLOW CE1 LL SD2032394 SD2062994 SD5062994 SD5082594 ANALYTICAL 03/23/94 06/29/94 06/29/94 08/25/94 ANALYTE METHOD mg/kg mg/kg mg/kg mg/kg SEDIMENT ALUMINUM 6010 1410.0 65.6 423.0 2580.0 ARSENIC 6020 2.9 0.14 ND 0.59 CADMIUM 6020 2.2 0.56 4.8 5.1 CALCIUM 6010 7040.0 406 2330.0 7650.0 IRON 6010 2250.0 88.7 653.0 3650.0 LEAD 6020 7.4 3.1 53.4 16.2 MAGNESIUM 6010 2140.0 145 571.0 2120.0 MANGANESE 6010 99.2 4.1 36.0 140.0 NICKEL 6010 3.9 ND 1.9 4.9 POTASSIUM 6010 890.0 149.0 184.0 1360.0 SILVER 6020 0.061 0.024 0.79 0.16 SODIUM 6010 ND 76.3 ND ND ZINC 6010 1560.0 59.7 1000.0 2650.0 ANIONS: SULFATE 300.0 214 56.5 143.0 214 SULFIDE, REACTIVE EPA/OSW 0.40 19.1 18.6 3.2 SULFIDE, ACID VOLATILE EPA (Draft) NA 226 178.0 ND FLUORIDE 340.2 NA NA NA NA CHLORIDE 300.0 NA NA NA NA PHOSPHORUS, TOTAL 365.3 NA NA NA NA ORTHOPHOSPHATE 365.3 25.8 63.4 30.5 18.8 NITRATE PLUS NITRITE AS N 353.2 NA NA NA NA NITRITE AS N 354.1 NA NA NA NA NITRATE AS N 353.2/354.1 NA NA NA NA AMMONIA 350.1 NA NA NA NA WATER (%) ILMOl.l 82 62 70 75 NA = Not analyzed ND = Not detected 88 Table A-4 (continued). Substrate Results - Downflow Cell SUBSTRATE - DOWN! FLOW CE1 LL SD2100494 SD5100494 SD2110294 SD2010495 ANALYTICAL 10/04/94 10/04/94 11/02/94 01/04/95 ANALYTE METHOD mg/kg mg/kg mg/kg mg/kg SEDIMENT ALUMINUM 6010 2640.0 3200.0 3200.0 2430.0 ARSENIC 6020 1.5 0.97 1.3 1.5 CADMIUM 6020 4.6 10.5 4.3 4.3 CALCIUM 6010 8460.0 4890.0 11700.0 8770.0 IRON 6010 3410.0 4640.0 4860.0 3460.0 LEAD 6020 46.4 30.8 11.3 18.2 MAGNESIUM 6010 2180.0 1800.0 2910.0 2190.0 MANGANESE 6010 160.0 151.0 232.0 144.0 NICKEL 6010 3.7 6.4 7.0 4.9 POTASSIUM 6010 930.0 1410.0 1140.0 729.0 SILVER 6020 0.17 0.29 0.069 0.28 SODIUM 6010 ND 108.0 92.8 ND ZINC 6010 1510.0 2850.0 3170.0 3250.0 ANIONS: SULFATE 300.0 86.8 187.0 159.0 184.0 SULFIDE, REACTIVE EPA/OSW 103.0 79.3 1.1 15.3 SULFIDE, ACID VOLATILE EPA (Draft) 190.0 70.6 171.0 117.0 FLUORIDE 340.2 NA NA NA NA CHLORIDE 300.0 NA NA NA NA PHOSPHORUS, TOTAL 365.3 NA NA NA NA ORTHOPHOSPHATE 365.3 39.0 3.3 12.6 6.4 NITRATE PLUS NITRITE AS N 353.2 NA NA NA NA NITRITE AS N 354.1 NA NA NA NA NITRATE AS N 353.2/354.1 NA NA NA NA AMMONIA 350.1 NA NA NA NA WATER (%) ILM01.0 62 70 NA 63 NA = Not analyzed ND = Not detected 89 Table A-4 (continued). Substrate Results - Downflow Cell SUBSTRATE - DOWNFLOW CELL ANALYTE ANALYTICAL METHOD SD2061295 06/12/95 mg/kg SD2082395 34934 mg/kg SD093097 09/30/97 mg/kg SEDIMENT ALUMINUM 6010 2050 1660 2200 ARSENIC 6020 0.59 0.75 NA CADMIUM 6020 11.3 31.4 219 CALCIUM 6010 7860 4720 7680 IRON 6010 3200 2490 4400 LEAD 6020 21.4 177 NA MAGNESIUM 6010 1860 1360 2070 MANGANESE 6010 149 108 1950 NICKEL 6010 7.0 6.2 22.5 POTASSIUM 6010 646 463 666 SILVER 6020 0.11 ND NA SODIUM 6010 119 ND 1930 ZINC 6010 4990 4680 37500 ANIONS: SULFATE 300.0 93.0 154 154 SULFIDE, REACTIVE EPA/OSW 5.3 2.5 NA SULFIDE, ACID VOLATILE EPA (Draft) 528 687 187 FLUORIDE 340.2 MNA NA NA CHLORIDE 300.0 NA NA NA PHOSPHORUS, TOTAL 365.3 NA NA NA ORTHOPHOSPHATE 365.3 5.0 1.8 NA NITRATE PLUS NITRITE AS N 353.2 NA NA NA NITRITE AS N 354.1 NA NA NA NITRATE AS N 353.2/354.1 NA NA NA AMMONIA 350.1 NA NA NA WATER (%) ILM01.0 60 64 NA = Not analyzed ND = Not detected 90 Appendix B Case Study 91 BUREAU OF MINES INFORMATION CIRCULAR/1994 PB94173341 II III II Mill llll II llllll Passive Treatment of Coal Mine Drainage By Robert S. Hedln, Robert W. Narin, and Robert L. P. Kleinmann UNITED STATES DEPARTMENT OF THE INTERIOR REPHOOUCIO «Y: U.S. Otpartmcnt of Commorco National Tocfcncal Information Sorvtot (/Ma* fit 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 our national 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. X 93 information Circular 9389 Passive Treatment of Coal Mine Drainage By Robert S. Hedin, Robert W. Nairn, and Robert L. P. Kleinmann UNITED STATES DEPARTMENT OF THE INTERIOR Bruce Babbitt, Secretary BUREAU OF MINES Library of Congress Cataloging in Publication Data: Hedin, Robert S., 1956- Passive treatment of coal mine drainage / by Robert S. Iledin, Robert W. Naim, and Robert L.P. Kieinmann. p. cm. — (Information circular; 9389) Includes bibliographical references (p. 34). Supt. of Docs, no.: I 28.27:9389. 1. Mine drainage. 2. Coal mines and mining—Waste disposal. 3. Mine water— Purification. 4. Water—Purification—Biological treatment. I. Naim, Robert W. II. Kieinmann, Robert L P. III. Title. IV. Series: Information circular (United States. Bureau of Mines); 9389. TN295.U4 [TN321] 622 s-dc20 [622*_5J 93-23717 CIP 7 /,' 95 REPORT DOCUMENTATION PAGE Form Approved 0MB No. 0704-0188 Public reporting burden for this collection of information is estimated to average 1 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 other aspect of this collection of information, including suggestions for reducing this burden to Washington Headquartt Operations and Reports, 1213 Jefferson Davis Highway, Suite 1204, Arlington, VA 22202-4302. and to the Office of Mana Project (0704-0188), Washington, DC 20503. ^ III ill!!llllllllllllllllllllll!!!llll 2. REPORT DATE 3. REPORT TYPE ANO DATE* COVB*ED PB94-173341 October 4, 1993 Information Circular 9389 4. TITL£ AND SUBTITLE Passive Treatment of Coal Mine Drainage 5. FUNblNQ NUMBER* 6. AUTHOmSI Hedin, R. S., Nairn, R. W., and Kleinmann, R. L. P. (. PERFORMING ORGANIZATION NAME(S) AND AODRt&SlESI U.S. Bureau of Mines Pittsburgh Research Center P.O. Box 18070 Pittsburgh, PA 15236 b. PERFORMING ORGANIZATION REPORT NUMBER IC 9389 9. SPONSOWNG/MON1 TORINO AGENCY INAM&S) AND AODflE3*(Cil U.S. Bureau of Mines Research 810 7th Street, NW Washington, DC 20241 ^0. SPONSORING/MONITOAlNQ AGENCY REPORT NUMBER SUPH-EMEnTARV NOTES None 12A. DJSTRlAUTION/AVAl LABILITY STATEMENT 12B. tXSlRJBUTION CODE Passive methods of treating mine water utilize chemical and biological processes that decrease metal concentrations and neutralize acidity. 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 g • m' J • d' 1 and 0.5-1.0 g*m- J »d'’, respectively. Wetlands with a composted organic substrate remove acidity from mine water at rates of 3-9 g*m* 2, d*’. A model for the design and sizing of passive treatment systems is presented in this report. 14. SUBJECT TERMS Acid mine drainage, constructed wetlands, anoxic limestone drains, iron, sulfate, manganese, water treatment IS. NUMBER OF PACES 38 16. PRICE COPE T7. SECURJT Y CLASSIFICATION OF REPORT lb. SECURITY CLASSIFICATION OF THIS PAGE 19. SECURITY CLASSIFICATION OF ABSTRACT 20. LIMITATION OF ABSTRACT *7ir 96 GENERAL INSTRUCTIONS TOR COMPLETING SF 298 The Report Documentation page (RDP) is used in announcing and cataloging reports. It is important that this information be consistent with the rest of the report, particularly the cover and title page. Instructions for filling in each block of the form follow. It is important to stay within the lines to meet optical scanning requirements. Block 1. Agency Use Only (Leave Blank). Block 2. Report Date. Full publication date including day, month, and year, if available (e.g. 1 Jan 88). Must cite at least the year. Block 3. Type of Report and Dates Covered. State whether report is interim, final, etc. If applicable, enter inclusive report dates (e.g. lOJun 87 - 30 Jun 88). Block 4. Title and Subtitle. 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If blank, the abstract is assumed to be unlimited. Standard Form 298 Back (Rev. 97 CONTENTS Page Abstract. Introduction. Treatment of mine water . 2 Background of passive treatment. 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. Metal removal in aerobic environments. 7 Iron oxidation and hydrolysis . 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 Fe 2 * from acidic and alkaline mine waters in a laboratory experiment . 8 3. Concentrations of Fe 3+ and field pH for water samples collected from Emlenton wetland. 8 ii ILLUSTRATIONS—Continued Page 4. Concentrations of Fe lc * 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 Fe 2 * and Mn 2 *. 10 7. Concentrations of Ca, Fe, 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 and A, 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 anoxic limestone drain, January 23, 1992 ... 11 6. Solubility products of some metal sulfides. 12 7. Sinks for HjS in constructed wetlands and their net effect on 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 . 18 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 S0 4 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 IN THIS REPORT cm centimeter L^min* 1 liter per minute °C degree Celsius m meter ft foot m 2 square meter g gram fim micrometer g»cm' 3 gram per cubic centimeter meq milliequivalent g*d‘ gram per day mg milligram g.m* 2 gram per square meter mg»L _1 milligram per liter g^m'^d' 1 gram per square meter per day mg*L _1 »h' 1 milligram per liter per hour g»m' 2 »yr -1 gram per square meter per year mL milliliter gpm gallon per minute min minute ha hectare nmol nanomole h kg kg*d _1 kg»m' 3 hour kilogram (concentration) kilogram per day kilogram per cubic meter nmol»cm‘ 3 «d' 1 yd 2 yr nanomole per cubic centimeter per day square yard year L liter PASSIVE TREATMENT OF COAL MINE DRAINAGE By Robert S. Hedin , 1 Robert W. Nairn , 2 and Robert L P. Kleinmann 3 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' J «d' 1 , 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 treatment systems is presented in this report. 1 Rcscarch biologist. 2 Re search biologist (now with The Ohio State University, Columbus, OH). 3 Research supervisor. Pittsburgh Research Center, U.S. Bureau of Mines, Pittsburgh, PA. 2 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 (l). 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 effluent limitation* for coal mine drainage Pollutant or Maximum tor any Average of daily values pollutant 1 day, for 30 consecutive property mg-L' 1 days mg-l' 1 Fe total . 6.0 3.0 Mn total. 4.0 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 all 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), 4 Italic 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 detrital material black. Dissolved A1 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 their 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 in 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 3 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 (4-5). 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 (6-8), 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 (9-10). 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 plarits, 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 (15). 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 (12). 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-contaminated 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 Kleinhenz, 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. 4 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. 5 Alkalinity was meas¬ ured in the field using a pH meter and an Orion Total Alkalinity Test Kit. At sites where particulates were vis¬ ible in water samples, an extra sample was collected that was filtered through a 0.22 -pun membrane filter before acidification. All samples were immediately placed on ice in an 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 Ciller (19). 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-/um 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 unpre¬ 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 lo specific products does not imply endorsement by the U.S. 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% H 2 0 2 (hydrogen peroxide), and then titrating the solution with 0.1 N NaOH (sodium hydroxide) to pH 8.3 (27). Acidity and alkalinity are reported as mg»L _1 CaC0 3 equivalents. ANALYTICAL QUALITY CONTROL For each set of samples for a particular site, 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-typc 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 5 Coal mining can promote pyrite oxidation and result in drainage containing high concentrations of Fe, Mn, and Al, as well as S0 4 , Ca, Mg, and Na. The solubilities of Fe, Mn, and Al are generally very low (<1 mg»L' 1 ) 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, Mn, 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 is 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\ Fe r + 1/40 2 + 3/2H 2 0 - FeOOH + 2H* (A) Fe 3 * + 2H 2 0 - FeOOH + 3H* (B) Al 3 * + 3H 2 0 - Al(OH) 3 + 3H* (C) Mn 2 * + 1/40 2 ♦ 3/2H 2 0 - 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 milliequivalents of acidic metals. For most coal mine drainages, the calculation is as follows: Acid^ = 50(2Fe 2+ /56 ♦ 3Fe 3 756 (1) ♦ 3A1/27 + 2Mn/55 + 1000(10"P H )) where all metal concentrations are in milligram per liter and 50 is the equivalent weight of CaC0 3 , and thus trans¬ forms milliequivalent per liter of acidity into milligram per liter CaC0 3 equivalent. For water samples with pH <4.5 (no alkalinity present), equation 1 cakulates a mine water acidity that corresponds closely with measurements of acidity made using the standard H 2 0 2 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 add 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 <4.5 was 693 mg'L' 1 , while the predicted acidities for these samples averaged 655 mg^L* 1 , a difference of only 6% (figure 1). 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 aridity 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 total aridity. Mine drainage at the Friendship Hill wetland had extremely low pH (2.7), but the acidity of the 6 Figure 1.—Comparison of calculated and measured acidities for water samples 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 circumneutral, ferrous iron accounted for 98% of the acidity, while the hydrogen ion accounted for < 1% of mine water acidity. 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 ligands, phosphate, and ammonia (23). The principal source of alkalinity in mine water is dissolved carbonate, which can exist in a bicarbonate (HC0 3 ') or carbonate form (C0 3 2 ). Both can neutralize proton acidity. H + + HC0 3 _ - H 2 0 + C0 2 (E) 2H* + C0 3 ' -* H 2 0 + C0 2 (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 levels 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 Composition 1 Acidity pH Fe 2 * Fe 3 * Al Mn Calculated 2 Measured 3 Diff. 4 3.9 98 1 0 0 181 184 -2% 3.9 0 0 106 0 598 578 + 3% 3.6 0 0 0 97 192 186 + 3% 3.8 13 0 47 42 370 335 + 9% Measured values are the average of three tests. Metal concentrations are mg-L' 1 . Acidities are mg'l 1 CaC0 3 equivalent. 2 From reaction 1. 3 0ata determined by the hot H 2 0 2 acidity method (27). 4 (1.00 - meas/cal) x 100. Table 3.—Acidic components of mine drainage influent at three passive treatment systems Parameter Friendship Hill Somerset Cedar Grove Concen¬ tration, mg-L' 1 Acid equivalent, 1 mg-L' 1 % of total acidity Concen¬ tration, mg-L' 1 Acid equivalent, 1 mg-L' 1 % of total acidity Concen¬ tration, mg-L' 1 Acid equivalent, 1 mg-L' 1 % of total acidity Fe 2+ . 7 13 1 193 345 69 95 170 98 Fe 3+ . 153 434 49 9 24 5 <1 <1 <1 Al 3+ . 58 317 36 3 17 3 <1 <1 <1 Mn 2+ . 9 16 1 59 107 21 2 4 2 PH. 2.6 112 13 3.7 10 2 6.3 <1 <1 'CaC0 3 equivalents calculated from the stoichiometry of reactions A-D. 7 Table 4.—Chemical compositions of mine drainages that contain high concentrations of alkalinity Location pH Alkalinity, mg-L 1 Al, mg-L 1 Fe 2 *, mg-L 4 Fe 3 *, mg-L 1 Mn, mg-L' 1 S0 4 , mg-L' 1 Net alkalinity, 1 mg-L' 1 Ohio: Coshocton. Pennsylvania: 152 <1 119 <1 2 1,325 -50 Cross Creek. 300 <1 96 <1 2 1,260 140 Donegal. 214 <1 39 <1 8 830 130 Fallston. 6.2 120 <1 30 <1 3 390 66 Keystone. 6.5 106 <1 37 <1 1 331 72 Latrobe . 204 <1 102 <1 6 1,200 15 New Bethlehem. 6.1 163 <1 51 <1 28 493 51 Possum Hollow. _ 6.4 263 <1 32 <1 1 620 209 Sligo. 93 <1 43 <1 26 1,720 -31 Somerset. 275 <1 2 <1 6 750 265 St. Petersburg. 255 <1 29 <1 9 250 203 Uniontown. 6.3 220 <1 70 <1 3 950 95 'Alkalinity minus acidity. Alkalinity and acidity arc 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 Fe 2 *, the net reaction for the oxidation, hydrolysis and neutralization reactions is complex because it differs between metals and also between abiotic and biotic processes. METAL REMOVAL IN AEROBIC ENVIRONMENTS Iron Oxidation and Hydrolysis The most common contaminant of coal mine drainage is ferrous iron. In oxidizing environments common to most 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. Fe 2 * + Y*0 2 + 2HC0 3 ' - FeOOH + Vi H 2 0 + 2C0 2 (G) Reaction G indicates that net alkaline waters contain at least 1.8 mg»L _1 alkalinity for each 1.0 mg»L"‘ 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 Fe 2 *, Fe 3 *, 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 metal 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 Fe 2 * + ‘/«0 2 + H* - Fe 3 * + 1/2H 2 0 (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- cumncutral waters, oxidation of ferrous iron occurred at a 8 Figure 2.—Removal of Fe 2 ’ from acidic and alkaline mine water* In laboratory experiment Raw mine drainage was col¬ lected from A, acidic Latrobe site; 8, alkaline Cedar Grove site. Splits of each sample were filter-sterilized (0J22-nm Alter). The Latrobe sample* were shaken throughout experiment; air was bubbled through Cedar Grove samples during experiment rate of 18 mg»L' 1 «h' 1 , while the rate for the raw acidic samples was only 1.4 mg’L'^h' 1 . To evaluate the signi¬ ficance of bacterial processes in iron oxidation, splits of both samples were filter-sterilized ( 0.22-fim membrane filter) before the experiment was begun. Removal of bac¬ teria had no 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 abiotically, 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 mg’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 Fe 3 * 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 Fe 3 * at a site where pH varied by almost 3 units. Ferric iron was not generally indicated unless the pH was <4, and the highest 40 - 1 - 1 —— - 1 - r— - 35 • 30 • L 25 a • | 20 *• *if. i * 4-* 15 ro Fe o * Y* 5 <4 • • • • V - 1_1-1 ! ..1- FIELD, pH Figurs 3.— Concentrations of F# 3 * and field pH for water samples collected from Emlenton wetland. 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. As water flowed through the constructed wetland, iron concentrations decreased from 95 to 15 mg*L _1 , and pH decreased from 5.5 to 3.2. Figure 42? shows Fe concen¬ trations and pH within a wetland that received mine water with a net alkalinity. Despite the removal of 60 mg’L' 1 Fe 2 * 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- hydroxides. The specific mechanism(s) by which Mn 2 ' precipitates from aerobic mine water in the absence of chemical additions is uncertain. Mn 2 * 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 MnO z . In alka¬ line environments, Mn 2 * can precipitate as a carbonate, which may also be oxidized by oxygen to Mn0 2 (25). Mn 2 * + HC0 3 - MnC0 3 + H + (I) MnC0 3 + V0 2 -* Mn0 2 + C0 2 (J) 9 Regardless of the mechanism by which Mn 2 * is oxidized to Mn 4 *, the removal of one mole of Mn 2 * from solution results in the release of two moles of H* or an equivalent decrease in alkalinity (HC0 3 *). The kinetics of Mn 2 * oxidation reactions are strongly affected by pH. Abiotic oxidation reactions are very slow at pH <8 (24). Microorganisms can catalyze Mn 2 * 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 Mn 2 * under acidic conditions. If pH falls below 6, Mn 2 * oxidation virtually ceases, the proton-producing hydrolysis reaction ceases, and pH stabilizes. The oxidation and precipitation of Mn 2 * from solution is accelerated by the presence of Mn0 2 and FeOOH (24, 30). Both solids reportedly act as adsorption surfaces for Mn 2 * and catalyze the oxidation mechanism. While addi¬ tions of FeOOH to Mn-containing water might accelerate Mn oxidation, the direct precipitation of FeOOH from mine water containing Fe 2 * does not generally stimulate SAMPLING STATION Figure 4.—Concentration* of Fe' 0 * end field pH at two con¬ structed wetlands. A, Emlenton wetland; B, Cedar Grove wetland. Mn-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, Fe decreased from 150 to <1 mg*L~ l , while Mn decreased from 42 to 11 mg»L* 1 . 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. Mn0 2 + 2Fe 2 * + 2H z O - 2FeOOH + Mn 2 * + 2H* (K) or MnOOH + Fe 2 * ^ FeOOH + Mn 2 * (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-oxides were Figure 5.—Mean concentrations of Fe, Mn, and Mg at the Morrison Wetland. Mine water flows linearly from station 1 to station 5. Verticte bars are one standard error of the mean. 10 0 20 40 60 80 100 TIME, h Figure 6.—Changes in concentrations of Fe 2 * and Mn 2 *. A, absence; B, presence of MnOOH. Mine water was collected from influent pipe of Biair wetland. MnOOH was collected from inside of final effluent pipe. collected from a wetland that removed Fe and Mn in a sequential manner. The wetland influent was alkaline (pH 6.2, 162 mg'L' 1 alkalinity) and contaminated with 50 mg'L* 1 Fe and 32 mg^L* 1 Mn. Two flasks of mine water received MnO z additions, while the controls did not receive Mn0 2 . 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 Mn0 2 , concentrations of Fe decreased to <3 mg’L' 1 in only 22 h. No change in concentrations of Mn occurred in the control flasks. Concentrations of Mn in the Mn0 2 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 solubilization of Mn 2 * suggests that the Mn0 2 oxidized Fe 2 * 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 2 * oxidation reactions, are not stable in the presence of Fe 2 *. Concentrations of ferrous iron must decrease to very low levels before Mn 2 * oxidation processes can result in a stable solid precipitate. In the absence of Fe 2 *, 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 0 2 is absent, Fe 2 * and Mn 2 * 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 circumneutral 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. CaC0 3 + H + - Ca 2+ -*• HC0 3 (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 Fe 3 * in most anaerobic environments limits the formation of FeOOH coatings that armor carbonate surfaces and inhibit further carbonate dissolution in aero¬ bic environments (31). Second, the solubilities of carbon¬ ate compounds are directly affected by the partial pressure of dissolved C0 2 (23-24, 32). Anaerobic mine water en¬ vironments commonly contain high C0 2 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 CO z within the treatment system, allowing the development of high C0 2 partial pressures (36). 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 clay. 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 CaC0 3 dissolution. Table 5.—Chemistry of mine water flowing through the Howe Bridge anoxic limestone drain, January 23, 1992 Parameter In Well 1 Well 2 Well 3 Well 4 Eff PH. 5.9 6.1 6.4 6.5 6.5 6.3 Alkalinity . .. 39 75 141 179 183 176 Ca. 140 150 183 201 206 198 Fe 2+ . 249 237 246 246 245 244 Fe 3+ . <1 <1 <1 <1 <1 <1 Mn. 34 33 34 34 34 34 Al. <1 <1 <1 <1 <1 <1 Mg. 90 87 91 91 90 90 Na. 11 11 11 11 11 11 S0 4 . 1175 1175 1200 1150 1200 1200 C0 2 . 6.3 4.0 4.7 4.3 4.7 NA NA Not available. NOTE.—Water flows linearly from the influent (In) through wells 1, 2, 3, and 4 and out the effluent (Eff). C0 2 values are the partial pressure percentages (atmosphere) of gas samples collected from the headspace within the sampling wells. No gas sample could be collected for the effluent because it is an open pipe. Figure 7.—Concentrations of Ca, and alkalinity for water as It flows through the Howe Bridge ALO. Water flows linearly from influent to effluent Dissolution of CaC0 3 within the ALD was greater than would be expected from an open system in equilibrium with atmospheric concentrations of C0 2 (0.035%). An equilibrated open system would only produce alkalinity in the range of 50 to 60 mg»L' 1 , and increase Ca concen¬ trations by 4 to 8 mg»L _1 . Observations of elevated C0 2 gas concentrations within the ALD, and the higher sol¬ ubility of CaC0 3 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 CaC0 3 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 C0 2 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, 2CH 2 0 + S0 4 2 " - H 2 S + 2HC0 3 (N) where CH z O 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¬ fate, suitable concentrations of low-molecular weight car¬ bon compounds, pH >4, and the absence of oxidizing agents such as O^, Fe 3 * and Mn 4 \ 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 (lactate, 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. M 2 * + H 2 S + 2HC0 3 ~ MS + 2H 2 0 + 2C0 2 (O) For Fe, the formation of pyrite is also possible Fe 2 * + H 2 S + S° - FeS 2 + 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 to form only when the con¬ centrations of all other metals in the table are very low («1 mg*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 6.—Solubility products of some metal sulfides Metal sulfide CdS. CuS. FeS. MnS. NiS . PbS. ZnS. ^ee reference 38. Solubility product 1 1.4 x 10' 23 4.0 x 10 38 1.0 x 10‘ 19 5.6 x 10 w 3.0 x 10* 1.0 x 10 29 4.5 x 10 24 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 H 2 S dissociation process (HjS -» 2H* + S 2 ‘) are neutralized by an equal release of bicarbonate during sulfate reduction. An organic substrate in which 100% of the H 2 S 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.—Sinks for HjS In constructed wetlands and their net effect on mine water acidity and alkalinity Reaction Acidity 1 Effect Alkalinity 2 H 2 S ♦ 2 HCO 3 - - H 2 S(g) ♦ 2HC0 3 ‘ 0 ♦ 100 H 2 S ♦ 2HC0 3 ‘ ♦ Fe 2 * - FeS ♦ 2H 2 0 ♦ 2C0 2 -100 0 H 2 S ♦ 2HCCV ♦ 2Fe 3 * - S° ♦ 2 Fe 2 * ♦ 2H 2 0 ♦ 2C0 2 -100 0 H 2 S * 2HC0 3 ' * 2 Fe(OHJ 3 - S° * 2 Fe 2 * ♦ 2H 2 0 ♦ 40H' ♦ 2HC0 3 ' ♦200 ♦300 H 2 $ ♦ 2HC0 3 - + V*Oj - S° ♦ H 2 0 ♦ 2HCO,~ 0 ♦ 100 H 2 S ♦ 2HC0 3 " ♦ FeS ♦ 'A0 2 - FeS 2 ♦ H 2 0 ♦ 2HC0 3 ' 0 ♦100 H 2 S ♦ 2HC0 3 " ♦ 20 2 - S0 4 2 * ♦ 2H 2 0 ♦ 2C0 2 0 0 1 Effect based on change in mineral acidity. 2 Effect based on summed change in bicarbonate and hydroxyl alkalinity. 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 50 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 chemistry at the Latrobe wetland Parameter Pore water 1 Surface water 2 Mean Std dev Mean Std dev Al. 1 5 35 5 Ca. 467 188 308 29 Fe 2+ . 215 183 73 39 Fe 3+ . 2 9 24 16 h 2 s. 37 75 <1 0 Mg. 175 48 166 9 Mn. 24 10 42 2 Na. 11 10 5 1 S0 4 . 1,674 532 1,967 115 Acidity 3 . 493 340 503 86 Alkalinity . 885 296 0 0 Net Alkalinity 4 ... 392 NAp -503 NAp pH. 6.8 .8 3.1 .1 NAp Not applicable. Std dev Standard deviation. 1 A total of 52 water samples were collected on July 25 and August 11, 1988, by the dialysis tube method. Metals were ana¬ lyzed for every sample. Field 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, Fe 2+ , Fe 3+ , Al, Mn, and HjS for pore water samples and measured by the Hj 0 2 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 (std 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 cf 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 Fe 2 * likely resulted from the dissolution of ferric oxyhydroxides at the redox boundary. FeOOH can be reduced by direct heterotrophic bacterial activity (< 2 ), CH 2 0 + 4FeOOH + H 2 0 -♦ 4Fe 2+ + 80H“ + C0 2 (Q) and also by H 2 S that results from sulfate reduction. H 2 S + 2FeOOH -* 2Fe 2+ + 40H~ + 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 C0 2 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. 4Fe 2+ + 80H’ + 0 2 - 4FeOOH + 2H 2 0 (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 in 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 Fe 2 * 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 6 The difference between surface and pore water concentrations of sulfate averaged 293 mg*L _1 , which is equivalent to 305 mg^L' 1 CaCOj alkalinity (reaction N); the difference in calcium concentrations averaged 159 mg*L -1 , which is equivalent to 398 mg»L _1 CaC0 3 alkalinity (reaction M). TOT 14 -Oi Figure 8.—Influent and effluent concentrations at the Latrobe wetland during the summer of 1988. A, Fe; B, acidity. concentrations of dissolved Al. Instead, concentrations of A1 in mine waters are primarily influenced by the solubility of Al(OH)j (23, 43). At pH levels between 5 and 8, Al(OH) 3 is highly insoluble and concentrations of dissolved A3 are usually <1 mg'L' 1 . At pH values <4, Al(OH) 3 is highly soluble and concentrations >2 mg*!/ 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 clays 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 Al(OH) 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 that 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 (CE%) has been a common measure of performance (11-12). Using iron concentration as an example, the calculation is 15 CE% « x 100 (2) ^ e in where the subscripts "in" and "efP 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' x 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 Lamin' 1 ) or for dif¬ ferences in system size (which vary from <0.1 to >10 ha) ( 12 ). 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 ) in = 1.44 x flow (L • min -1 ) x Fe (mg • L-') in ) (3) where g* d' 1 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 _1 ) in = 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, Fe fe * - (Fc g • - (Fe g • 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 • d -1 ) rem = 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~ J ). 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 Wetland Fe Concentration Fe Loading Fe removal System size, Flow rate In Eff In Eff performance m 2 L*min' 1 mg'L' 1 mg'L' 1 Kg*d' 1 Kg-d 1 CE Rate % g*m' 2 *d' 1 A . . , 400 10 100 40 1.4 0.6 60 2.2 B . . 400 40 100 40 5.8 2.3 60 8.6 Cl .. 500 30 40 <1 1.7 <0.1 99 3.5 C 2 . . 500 80 35 10 4.0 1.2 71 5.8 C3 . . 500 150 30 16 6.5 3.5 47 6.0 n .. 750 50 100 25 7.2 1.8 75 7.2 In Influent. Eft Effluent. CE Concentration efficiency. 16 Lastly, consider a comparison of wetland systems of 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^k-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' 2 , while concentrations in r ainfall 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, MgS0 4 , MgCOj, and CaMgtCOj)* do not form at the concentrations and pH conditions found in the systems (25). 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*m* 2 *yr -1 . The uptake of dis¬ solved Mg by plants in constructed wetlands can only account for 5 to 10 g Mg»m' a# yr _1 . This estimate as¬ sumes that the net primary productivity of the constructed wetlands is 2,000 g»m' 2 *yr' 1 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 Mg by 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 day substrate that has a density of 13 g* cm' 3 , 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. Table 10.—Influent and effluent concentrations of Ca, Mg, Na, and sulfate at eight constructed wetlands _Ca_ _Mg_ _Na_ _SOj_ In, Eff, Change, In, Eff, Change, In, Eff, Change, In, Eff, Change, mg'L' 1 mg-L* 1 % mg*L' 1 mg*L 4 % mg-L' 1 mg-l' 1 % mg*L* 1 mg*L‘ x % Donegal. 244 241 -1 81 79 -2 6 6 0 729 729 0 Emlenton_ 429 433 *-1 308 306 -1 11 10 -2 2,810 2,770 -1 FH. 122 189 +55 51 51 0 5 7 +2 1,125 842 -25 Gourley. 117 120 +3 114 117 +3 3 4 +6 1,000 1,030 +3 Latrobe . 244 256 +14 127 125 -2 6 11 +8 1,525 1,225 -20 PineyA. 416 426 +2 251 262 +4 15 16 +4 2,190 2,120 -3 PineyB. 355 354 0 217 216 0 27 27 -2 2,050 2,100 +2 Somerset .... 307 469 +53 _ 312 312 _0_6_ 7 +15 2,740 2,300 -16 Eff Effluent. In Influent. FH Friendship Hill National Historical Site. 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: DF = Mg,, /Mg. . (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 &Fc da =■ Fe m - (Fe,„/DF) (8) where AFe DA 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, AFe DA = (Fein xDF ) - Fe eff ( 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 • day' 1 )^ = (AFe DA xFlow x 1.44 )/SA. (10) LOADING LIMITATIONS 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*m' a «d‘ l . The data from Cl are not sufficient to estimate whether the wetland could have removed 10 or 100 g»m' 2 »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~ l at the system influent station (the ALD dis¬ charge) to <1 mg«L _1 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' a »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 concentration* of Fe, Mn, and Mg at the Morrison passive treatment system Cumulative Row, Concentration, Removal rate 1 , Station area, m 2 L*m _1 mg*L' 1 g*m - 2 *d 1 Fe Mn Mg Fe Mn Influent . 0 6.6 151 42 102 NA NA Ditch Effluent .... 43 NA 56 37 91 19.2 0.17 Pond Effluent .... 461 NA 5 24 72 2.3 0.14 Final Effluent .... 1,076 NA <1 71 71 1.3 0.13 NA Not available. 1 Removal 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' 1 ) 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 occassions 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»min _1 (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 arc referred to as REM-Lower. Mine water at the Howe Bridge system is characterized at two locations. The "upper" analysis describes mine water discharging from an ALD that flows into aerobic settling ponds. The Tower* 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 ^.- Construction characteristics of the constructed wetlands Site Constructed year Design Substrate Emergent vegetation m z Water depth, cm Row rate , 1 L*min ' 1 Est. ret time , 2 days Donegal. 1987 Pond, 8 Cells LS, SMC Typha 8,100 15 501 1.7 Cedar . 1989 5 Cells Clay, LS .. do. 1,360 15 156 0.9 Keystone . 1989 Ditch Topsoil None 4,200 100 8,606 .3 Blair. 1989 Ditch Manure, straw Mixed 1,080 5 11 3.4 Shade . 1989 ALD, 2 Cells LS None 880 10 10 6.4 Piney. 1987 1 Cell HB Mixed 2,500 50 468 1.9 Morrison . 1990 ALD, 3 Cells Clay, manure Typha 1,075 30 7 33.9 Emlenton. 1987 9 Cells LS, manure .. do. 643 50 55 4.1 Somerset. 1984 2 Cells HB, LS, SMC .. do. 1,005 15 47 2.2 Howe. 1991 ALD, 3 Cells Clay, LS, SMC None 3,000 50 130 8.0 Latrobe . 1987 3 Cells HB, LS, SMC Typha 2,800 15 86 3.4 REM . 1992 2 ALDs, 9 Cells SMC .. do. 4,849 30 206 4.9 FH. 1988 6 Cells LS. SMC .. do. 667 15 15 4.6 Est. Estimated. FH Friendship Hill National Historical Site. HB Haybales. LS Limestone, ret. 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 Tabto 13.—Average chemical characteristic* of Influent water at the constructed wetlands (sites are arranged according to the net acidity) Site Number of samples pH Aik Fe Composition, mg«L _1 Mn A1 Mg SO, Net Acidity , 1,2 mg-L ' 1 Donegal. 7.1 202 5 8 <1 81 738 -182 Cedar . 6.3 336 92 2 <1 54 1,251 -140 Keystone . 6.3 142 37 <1 <1 14 330 -73 Blair. 6.2 166 52 30 <1 77 645 -51 Shade . 6.0 31 <2 22 <1 125 966 -17 Piney. 5.8 60 1 15 <1 225 1.845 -6 Morrison . 6.3 271 150 42 <1 102 1,087 75 REM - L. 20 6.1 128 190 50 <1 118 1,275 258 Howe - Lower. 13 5.6 22 185 34 <1 91 1,128 312 Emlenton. 4.7 15 89 77 8 249 2,317 320 Somerset. .... 43 4.4 0 162 50 3 193 1,691 373 Howe - Upper. . . . . 13 6.2 160 272 39 <1 105 1,315 375 REM-Lower . 9 3.5 0 246 92 2 171 1,875 496 Latrobe . 43 3.5 0 125 32 43 125 1,655 617 REM - R. . . . . 18 5.5 57 473 130 3 232 2,495 867 FH. .... 73 2.6 0 153 9 58 85 1,733 929 Aik Alkalinity. FH Friendship Hill National Historical Site. ’CaC 03 equivalent. 2 Negative 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 geochemical reactions within the mine spoil (Donegal and Blair) or the flooded deep mine (Cedar, Keystone, and Pincy). 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 Fe 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 low pH were contaminated with Fe, Mn, and AJ. EFFECTS OF TREATMENT SYSTEMS ON CONTAMINANT CONCENTRATIONS the systems decreased Fe concentrations by more than 50 mg^L' 1 . 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 ). Both the Shade and Blair wetland effluents flowed into settling ponds which discharged water in compliance with regu¬ latory criteria. On occassions, the discharges of the Morrison and Pincy treatment systems met compliance criteria. Every wetland system decreased concentrations of acidity. The Morrison system, which received mine water that contained 75 mg«L‘ l acidity, always discharged net alkaline water. None of the constructed wetlands that received highly acidic water (net acidity >100 mg*L' 1 ) 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 . 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 mg^L" 1 . Nine of 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 conjuction 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. Few 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, their presence did not markedly affect the average contaminant removal rates for the constructed wetland study areas. Table 14.—Mean water quality (or sampling stations at the constructed wetlands Site Sampling station n 1 pH Fe Mn Acidity Mg Donegal. 6 6.4 34 9 NAp 83 Wetland influent 29 7.1 5 8 NAp 81 Effluent 28 7.4 <1 2 NAp 80 Cedar . 26 6.3 92 2 NAp 54 Effluent 27 6.4 41 2 NAp 53 Keystone. 28 6.3 37 1 NAp 14 Effluent 28 6.4 32 1 NAp 14 Blair. 12 6.2 52 30 NAp 77 Effluent 8 7.0 <1 5 NAp 59 Shade . 20 6.0 2 23 NAp 128 LC effluent 20 6.8 <1 10 NAp 122 Piney. 21 6.4 32 25 NAp 201 Wetland influent 39 5.8 1 15 NAp 225 Wetland effluent 39 6.1 <1 11 NAp 225 Morrison . 24 6.3 151 42 75 102 Ditch 24 6.4 56 37 64 91 Effluent 24 6.6 <1 11 -1 71 REM-L. 20 6.1 190 50 258 118 Left effluent 20 3.8 84 48 225 112 Emlenton. 46 4.7 89 77 320 249 Effluent 40 3.2 15 73 271 234 Somerset. 43 4.4 162 50 373 193 Effluent 40 5.5 18 33 69 139 Howe. 13 6.0 265 37 373 101 Upper effluent 13 5.6 185 34 312 91 Lower effluent 13 6.2 68 33 112 91 REM-lower . 9 3.5 246 92 496 171 Effluent 9 2.9 115 88 436 166 Latrobe . 43 3.5 125 32 617 125 Cell 3 effluent 43 3.7 56 29 343 122 REM-R. 18 5.5 473 130 867 232 Right effluent 18 3.3 338 113 712 201 FH. 73 2.6 153 10 929 85 Effluent 73 2.9 137 10 674 85 FH Friendship Hill National Historical Site. LC Limestone cell. NAp Not applicable, dumber of samples. ^e flow-weighted average of two discharges. 21 Table 15.—Dilution factors for the constructed wetlands Site Average Sd Minimum Maximum Donegal. 0.99 0.05 0.76 1.04 Cedar . 0.99 0.03 0.92 1.05 Keystone. 0.99 0.04 0.91 1.15 Blair. 0.83 0.10 0.70 1.01 Shade . 0.96 0.08 0.76 1.09 Piney. 1.00 0.06 0.92 1.31 Morrison Ditch ., 0.87 0.18 0.40 1.05 Morrison Wetland 0.69 0.25 0.27 1.12 REM-L. 0.95 0.09 0.70 1.13 Howe Lower .... 1.00 0.10 0.80 1.25 Emlenton. 0.94 0.09 0.66 1.04 Somerset. 0.73 0.30 0.30 1.76 Howe Upper .... 0.89 0.08 0.73 0.99 REM-Lower .... 0.93 0.09 0.72 1.01 Latrobe . 0.95 0.08 0.75 1.14 REM-R. 0.86 0.16 0.36 1.00 FH. 1.00 0.12 0.58 1.34 FH Friendship Hill National Historical Site. REMOVAL OF METALS FROM ALKALINE 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 >50 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' 2 *d‘ l 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 lowered Fe concentrations in a very large deep mine discharge by 5 mg L 1 , removed Fe at a rate of 21 g*m -2 * d' 1 . The shallow-water Morrison ditch, which decreased concentrations of Fe in a low-flow seep by al¬ most 100 mg*L~ l , had an average Fe removal rate of 19 g«m' 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 28 g»m' 2 *d‘ l . Table 16.—Fe and Mn removal rate* at constructed wetland Site Fe removal rate Mn removal rate Mean Std dev n sig? 1 Mean Std dev n sig? Donegal. NAp NAp NAp NAp 0.50 0.25 9 yes Cedar . 6.3 2.2 7 yes 0.17 0.41 7 no Keystone. 20.7 5.1 15 yes NAp NAp NAp NAp Blair. NAp NAp NAp NAp 0.43 0.37 6 yes Shade . NAp NAp NAp NAp 0.72 0.64 17 yes Piney. NAp NAp NAp NAp 1.07 1.34 33 yes Morrison Dit. 19.2 10.6 24 yes 0.17 0.41 24 yes Morrison Wet. NAp NAp NAp NAp 0.20 0.18 24 yes REM-L. 28.3 5.7 20 yes •0.05 0.13 20 no Howe-Lower... 8.1 1.9 13 yes 0.06 0.16 13 no Emlenton. 9.1 3.3 39 yes •0.09 0.19 39 no Somerset. 5.0 4.9 34 yes -0.01 0.54 34 no Howe-Upper. 42.7 82 13 yes -0.43 0.49 13 no REM-Lower . 12.0 3.4 9 yes -0.05 0.14 9 no Latrobe . 2.1 1.0 21 yes 0.03 0.09 21 no REM-R. 20.1 4.0 18 yes 0.10 0.33 18 no PH •«•»•••••«•••••**••• 0.5 0.5 73 yes 0.00 0.02 73 no NAp Not applicable. FH Friendship Hill National Historical Site. n Sample size. sig? Significant at 0.05 level. Std dev Standard deviation. ‘Yes, rate is significantly greater than zero (t-test); no, rate is not significantly greater than zero (t-test). 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' 2 »d _1 , while the Howe Bridge Upper site removed Fe at a rate of 43 g»m‘ 2 »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 upwells 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 0.3-m-high pipe, flows 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. INFLUENT ALKALINITY 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). > o 2 UJ oc -50 0 50 100 150 200 250 300 350 400 INFLUENT ALKALINITY Figure 9.—Relationship between mean Fe removal rates and A, mean influent pH and B, mean Influent alkalinity concen¬ trations. Vertical bars are one standard error above and below the mean. "H-L* Is the Howe-Lower site. Figure 10.—Relationship between mean Mn removal rates and A mean influent pH and B, mean influent alkalinity concen¬ trations. Vertical bars are one standard error above and below mean. Fe values next to the bars are effluent Fe 2 * values. 23 Alkaline sites that contained high concentrations of Fe throughout the treatment system (Howc-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^L' 1 Fc, 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 uncertainities 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 Piney 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 to 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. Mn 2 * can 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'L' 1 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 sites and very minor removal of Mn occurred at the fifth site (Morrison ditch). Although the processes that remove Mn and Fc 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, bioaccumulation of metals in plant biomass is an insignificant component of Fe and Mn removal in con¬ structed wetlands (49). The ability of emergent plants to 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 oxygenation 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 Fc 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-removal 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 4 A, 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 sulfate 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 COj and promote carbonate disso¬ lution. The presence of organic matter also promotes the activity of sulfate-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-prccipitatc 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:S0 4 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:S0 4 ratio of 9.7:1, then changes in sulfate concentrations attributable to the co- predpitation process amount to only 5 to 30 mg*L‘‘ at the add mine water sites. Dilution-adjusted changes in sulfate concentrations at the Somerset, Latrobe, Friendship Hill (FH), and Howe-Lower wetlands were commonly 200 to 500 mg*L'\ Rates of addity removal, sulfate removal and caldum 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 addity 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.—Fe and S0 4 content of ferric oxyhydroxlde deposits; sites are arranged by pH Site pH Composition, ppm dry weight _Fe_ S0 4 Fe:SQ 4 Emlenton. 3X) 471,779 64,213 7 X” Latrobe . 3.5 288,939 27,991 10.3 Somerset. 3.5 461,583 48,263 9.6 Cedar . 6.4 362,300 8,946 40.5 Keystone. 6.6 398,337 6,888 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 S0 4 ' 2 »ra' 2 *d' 1 , is equivalent to a sulfate reduction rate of ~180 nmol*cm -3 * 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 18.—Average rates of acidity removal, sulfate removal, and calcium addition at sites receiving acidic mine water Site n Acidity removal rate Sulfate removal rate Calcium addition rate _ mean Std dev sig? 1 mean Std dev sig? mean Std dev sig? Emlenton. 25 3.1 2.4 yes 1.5 5.7 no 0.8 1.21 yes Somerset. 34 9.9 8.6 yes 5.1 5.7 yes 1.7 1.20 yes Howe Lower. 13 15.4 4.1 yes 8.9 7.2 yes 3.9 1.40 yes REM-Lower . 9 7.1 7.2 yes 2.9 2.4 yes 2.6 1.03 yes Latrobe . 21 6.9 4.4 yes 5.9 6.4 yes 0.9 0.07 yes FH. 72 7.0 3.8 yes 3.4 2.6 yes 1.2 0.80 yes FH Friendship Hill National Historical Site, n Sample size. Std dev Standard deviation. ’Yes, rate is significantly greater than zero (t-test); no, rate is not significantly greater than zero (t-test). 25 The highest rates of acidity removal, sulfate reduction, and limestone dissolution all occurred at the Howe-Lower site. This system differs from the others by its subsurface flow 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 surflcially 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. Figure 11 .—Measured rates of alkalinity generation and acidh removal at the Friendship Hill wetland. Units are g«m~ 2 *d CaC0 3 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, plantlcss 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, Sphagnum peat, haybales, and manure. The ALD is a buried bed of limestone that is intended to add alkalinity to the mine water (75, 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, not 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^rn'^d' 1 ) is based on data from sue sites, only one of which lowers Fe concentrations to compliance. Table 19.—Recommended sizing for passive treatment systems AML criteria, Compliance criteria, Alkaline Acid Alkaline Acid Fe. 20 NAp 10 NAp Mn. 1.0 NAp 0.5 NAp Acidity .._NAp_7_NAp_^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 2 »d*\ is in agreement with the findings of Stark (77) 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*nr 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 warmer 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 1 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 (H 2 0 2 method) measurements. If an anoxic limestone drain is being considered, the acidified sample should be analyzed for Fe 3 * 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 Figure 12.—Flow chert showing chemical determinations nec¬ essary for the 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, Acidity] x 1.44 (11) If the concentrations are milligrams per liter and flow 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 Water : 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, 18). In these systems, alkalinity either enters the treatment system with diluting water or 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 0 2 concentrations. Standard features that can aerate the drainage, such as waterfalls or steps, should be followed by quiescent areas. Aeration only provides enough dissolved 0 2 to oxidize about 50 mg»L _1 Fe 2 *. Mine drainage with higher concentrations of Fe 2 * will require a series of aeration structures and wetland basins. The wetland cells allow time 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 (m 2 ) = 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 50 cm. 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 (18). 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 (56). 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 intercell connections may also serve as aeration devices. If 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 particulates, prevent flow channelization and provide wildlife benefits that are valued by regulatory and environmental groups. PASSIVE TREATMENT OF NET ACiO 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 (Fe 3 *), aluminum (Al) and dissolved oxygen (DO). When acidic water containing any Fe 3 * or Al contacts limestone, metal hydroxide particulates (FeOOH or Al(OH) 3 ) 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 Fe 3 *, 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, Fe 3 * 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' 1 , the risk that the ALD will fail prematurely also increases. Recently, two ALD’s constructed to treat mine water that contained 20 rag^L' 1 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 > 5 characteristically have con¬ centrations of DO, Fe 3 *, and Al that are all <1 mg»L‘ l . 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 Fe 3 * and Al. Mine waters with pH <5 can contain dissolved Al; mine waters with pH <3.5 can contain Fe 3 *. In northern Appalachia, most mine drainages that have pH <3 contain high concentrations of Fe 3 * 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 CaC0 3 content because of its higher reactivity compared with a limestone with a high MgCOj or CaMg(C0 3 ) 2 content. The lime¬ stones used in most successful ALD’s have 80% to 95% CaC0 3 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 C0 2 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 (35). 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 29 Figure 13.—Longitudinal-section and cross-section of the Morrison ALO. Wells are for sampling purposes and have no Importance to drain’s functioning. 0.3 to 3 m of on-sitc clay to restore the original surface topography (34, 36). At sites where linear ALD’s are not possible, anoxic 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 insure 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% CaCOj 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 unchannelized flow for a prolonged period, whether 100% of the CaC0 3 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 field equipment (0 to 1 mg’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 30 circumneutral levels and increase concentrations of cal¬ cium and alkalinity. Dissolved Fe 2 * 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 mg*L _1 are readily treated with an ALD and aerobic wetland system. If the mine water is contaminated with only Fe 2 * and Mn, and the acidity exceeds 300 mg»L~ l , it is unlikely that an ALD constructed using current practices will dis-charge net alkaline water. When this partially neutralized water is treated aerobically, the Fe will precipitate rapidly, but the absence of sufficient bufferring can result in a discharge with low pH. Building a second ALD, to re¬ charge the mine water with additional alkalinity after it flows out of the aerobic system, is currently not feasible because of the high DO content of water flowing out of aerobic systems. If the treatment goal is to neutralize all of the acidity passively, then a compost wetland should be built so that additional alkalinity can be generated. Such a treatment system thus contains all three passive tech¬ nologies. The mine water flows through an ALD, into a settling pond and an aerobic system, and then into a com¬ post wetland. If the mine water is contaminated with ferric iron (Fe 3 *) or Al, higher concentrations of acidity can be treated with an ALD than when the water is contaminated with only Fe 2 * and Mn. This enhanced performance re¬ sults from a decrease in mineral acidity because of the hydrolysis and precipitation of Fe 3 * and Al within the ALD. These metal-removing reactions decrease the min¬ eral acidity of the water. ALD’s constructed to treat mine water contaminated with Fe 3 * and Al and having acidities greater than 1,000 mg'L" 1 have discharged net alkaline water. The long-term prognosis for these metal-retaining systems has been questioned (34). However, even if cal¬ culations of system longevity (as described above) are inaccurate for waters contaminated with Fe 3 * and Al, their treatment with an ALD may turn out to be cost-effective when compared with chemical alternatives (35). When a mine water is contaminated with Fe 2 * and Mn and has an acidity betweem 150 and 300 mg^L' 1 , the ability of an ALD to discharge net alkaline water will depend on the concentration of alkalinity produced by the limestone system. The amount of alkalinity generated by a properly constructed and sized ALD is dependent on chemical characteristics of the acid mine water. An ex¬ perimental method has been developed that results in an accurate assessment of the amount of alkalinity that will be generated when a particular mine water contacts a particular limestone (58). The method involves the anoxic incubation of the mine water in a container filled with limestone gravel. In experiments at two sites, the con¬ centration of alkalini ty that developed in these containers after 48 h correlated well with the concentrations of alkalinity measured in the ALD effluents at both sites. TREATING MINE WATER WITH COMPOST WETLAND When mine water contains DO, Fe 3 * or Al, or contains concentrations of acidity >300 mg»L _1 , construction of a compost wetland is recommended. Compost wetlands generate alkalinity through a combination of bacterial ac¬ tivity and limestone dissolution. The desired sulfate- reducing 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 CaC0 3 content (about 10% dry weight), but mixing in more limestone may increase the alkalinity generated by CaC0 3 dissolution. Compost sub¬ strates that do not have a high CaC0 3 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 m 2 to a depth of 45 cm thick. This is equivalent to one ton per 3.5 yd 2 . 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*m" 2 *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 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 (m 2 ) = Acidity Loading (g • d _1 /7). (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 (18). 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. CHAPTER 5. 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 (13). Similarly, the Simco wet¬ land in Ohio, which has discharged compliance water for several years (77), receives water containing -160 mg»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 32 contains an organic substrate in 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 undersiring. 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 m 2 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 m 2 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 in 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 (75, 18). 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-free waters, 0.5 g»m~ 2 «d‘ l , is consistent with rates reported by the TVA for aerobic wetlands in southern Appalachia (18) and by the Pennsylvania Department of Environmental Resources (DER) for constructed wetlands in Pennsylvania (60). The average Fe-removal rate .re¬ ported in this study for alkaline waters, 20 g«m" 2 «d _l , is only slightly greater than has been reported in other studies. The rates of Fe removal for aerobic wetlands in southern Appalachia ranged from 6 to 20g«m' 2 *d' 1 (18). Some of the lower rates reported by TVA investi¬ gators, however, are from wetland systems that discharge water with <1 mg»L' 1 Fe and thus are loading limited with respect to Fe. Such sites were intentionally avoided in this study. Stark (77), in their studies of a constructed wetland in Ohio, reported Fe removal rates over a range of loading conditions. When the wetland system dis¬ charged >15 mg*L* 1 Fe, and thus was overloaded with Fe, the removal rate averaged 21 g^m’^d" 1 . When the wetland effluent contained <15 mg»L _1 Fe, the removal rate averaged only 11 g»m' 2 *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 siring 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 in 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 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 are 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~* (65). If a wetland containing a 40 cm depth of compost generates CaC0 3 - derived alkalinity at a mean rate of 3 g»m' 2 »d~ l (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»m' 2 «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 (66). 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-buffcring 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 ii 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 arc 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 34 <3 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. 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Proceedings of the 1994 International Land Reclamation and Mine Drainage Conference, Pittsburgh, PA, U.S. Bureau of Mines SP GA-94, pp. 195-204. INT.BU.OF MINES,PGH.,PA 29916 ■0 vvEPA United States Environmental Protection Agency Please make all necessary changes on the below label, detach or copy, and return to the address in the upper left-hand corner. If you do not wish to receive these reports CHECK HERE □ ; detach, or copy this cover, and return to the address in the upper left-hand corner. PRESORTED STANDARD POSTAGE & FEES PAID EPA PERMIT No. G-35 National Risk Management Research Laboratory Cincinnati, OH 45268 Official Business Penalty for Private Use $300 LIBRARY OF CONGRESS E PA/540/R-02/506 December 2002