History of Passive Treatment

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During the past 30 years, the possibility that mine water might be treated passively has developed from an experimental concept to full-scale field implementation at thousands of sites throughout the world. Passive treatment of mine water can be traced to two independent research projects, which showed that natural wetlands were ameliorating CMD without incurring any obvious ecological damage. Researchers at Wright State University studied a site in the Powelson Wildlife Area in Ohio where Sphagnum recurvem had volunteered and was growing in pH 2.5 water. As the water flowed through the boggy area, iron, magnesium, calcium, sulphate, and manganese all decreased, while the pH increased to 4.6. A natural outcrop of limestone located at the downstream end provided sufficient neutralization to raise the effluent pH to between 6 and 7 (Huntsman et al., 1978). Meanwhile, independently, a similar study was being conducted by a group at West Virginia University, working at a natural Sphagnum-dominated wetland, Tub Run Bog, in northern West Virginia. They were looking at the ecological damage to the wetland as a result of drainage water from an adjoining abandoned mine (Figure 1). They found no adverse ecological effects and that, in fact, within 20 to 50 m of the influent, the pH of the water rose from between 3.05 and 3.55 to between 5.45 and 6.05. Sulphate concentrations decreased to 15 mg/L or less, and iron decreased to less than 2 mg/L (Wieder and Lang, 1982). These field observations prompted the U.S. Bureau of Mines to examine whether wetlands might be constructed for the intentional treatment of coal mine drainage. It was thought that small seeps present at many abandoned mine sites could be passively treated in this manner. Research efforts were initiated by the U.S. Bureau of Mines, in cooperation with Wright State University (Kleinmann et al., 1983; Kleinmann, 1985) (Figure 2).

Figure 1. Wetland in West Virginia receiving and treating ARD from an adjoining surface mine.

Independently, West Virginia University, and subsequently, Pennsylvania State University conducted research as well (e.g., Gerber et al., 1985; Stone and Pesavento, 1985). Initially, all of these experimental wetlands were constructed to mimic the Sphagnum wetlands. However, Sphagnum moss was not readily available, proved difficult to transplant, and tended to accumulate metals to levels that were toxic to the Sphagnum after several months of exposure to mine drainage (Huntsman et al., 1985; Spratt and Wieder, 1988). Indeed, unless the CMD was relatively mild, the Sphagnum crusted over with metal deposits. Instead of abandoning the concept, researchers experimented with different kinds of constructed wetlands, using other plants that had shown the tendency to volunteer in CMD.

Eventually a wetland design evolved that proved tolerant to years of exposure to CMD and was effective at lowering concentrations of dissolved metals. Most of these treatment systems consisted of a series of small wetlands (< 1 ha) that were initially vegetated with Typha latifolia (Girts et al., 1987; Stark et al., 1990) (Figure 3). Although they were not as acid tolerant or as effective in removing metals as the Sphagnum systems, the Typha systems proved to be very hardy. These systems proved to be very cost effective in treating circumneutral and net alkaline mine water, where the primary objective was to precipitate the iron in the wetland, instead of downstream. Such wetlands are typically divided into cells to reduce short circuiting and thereby enhance retention time.

Figure 2. An early attempt to construct an aerobic wetland using Sphagnum moss to treat acidic CMD at Friendship Hill, PA (photograph by Bob Kleinmann)
Figure 3. A typical example of an aerobic wetland constructed to treat circumneutral or alkaline CMD (photograph by Bob Kleinmann)

Some of these wetlands were constructed with a compost and limestone substrate to provide a favourable environment for the Typha to root. Others were constructed without an exogenous organic substrate; emergent plants were rooted in whatever soil or spoil substrate was available on the site when the treatment system was constructed. Researchers soon realized that the Typha were generally collecting only a small component of accumulated metals internally (Sencindiver and Bhumbla, 1988), and that its principal functions were dispersing the flow of the water and filtering out the suspended floc of the precipitated metals. Subsequently, some systems were constructed that did not rely at all on the early wetland model. Ponds, ditches, and rock-filled basins were constructed without emergent plants and, in some cases, without soil or organic substrate.

In the late 1980s, two new approaches were developed that extended the treatment capabilities of wetlands to more acidic mine water. In the first case, U. S. Bureau of Mines researchers, assessing the performance of a wetland that had been constructed with a compost and limestone substrate in an attempt to treat very acidic CMD, found that in isolated locations, the mine water was being neutralized and iron was being precipitated as a sulphide. Apparently, water was flowing down through the compost/limestone substrate and then back up again, gaining alkalinity in the process; some of the alkalinity was generated by limestone dissolution while much of it was generated by bacterial sulphate reduction (Hedin et al., 1988) (Figure 4). An approach was developed to optimize this effect and was evaluated in the field; it was determined that these anaerobic or compost wetlands added alkalinity, but were not very efficient for iron removal. Removing high concentrations of iron required sequential placement of aerobic and anaerobic systems (McIntyre and Edenborn, 1990; Nawrot, 1990). Thus, anaerobic wetland systems are seldom constructed nowadays to treat CMD; however, they have proven very useful for treatment of metal mine drainage, since they provide a mechanism to remove metals such as cadmium, copper, lead, etc. Researchers from Colorado experimented with bioreactors that contained compost or other organic sources (Howard et al., 1989; Wildeman and Laudon, 1989; Wildeman et al., 1990, 1994). This was soon followed by larger scale field tests at metal mines in the western U.S. (Cohen and Staub, 1992; Gusek, 1998; Gusek et al., 2000), Canada (Kalin et al., 1996), and Europe (Brown et al., 2002; PIRAMID Consortium, 2003; Younger 2000; Younger et al., 2002).

Figure 4. An illustration of the wetland in which bacterial sulphide generation was first observed and a photograph
of the same site; the black precipitate seen in the photograph is iron sulphide and the white precipitate is aluminum,
precipitating out due to the elevated pH.

Illustrationofwetlandinwhichbacterialsulphidegeneration.jpg Photoofwetlandinwhichbacterialsulphidegeneration.jpg

The other new approach involved contacting the acidic water from an underground coal mine with limestone in an anoxic environment before it was exposed to the atmosphere and became oxygenated. Although limestone had previously been used many times to treat mine water (as mentioned in Section, limestone in passive systems typically became coated or “armoured” by iron hydroxide, which greatly decreases limestone dissolution. Turner and McCoy (1990) reasoned that if the CMD could be intercepted before it contacted the atmosphere, and was directed into a limestone-filled French drain, the dissolved iron would not oxidize to ferric hydroxide to armour the limestone, and the water would be neutralized. The water could then be discharged into an aeration pond and a wetland. A great number of anoxic limestone drains (ALDs) were subsequently constructed, and soon, sizing guidelines were developed (Hedin et al., 1994b; Watzlaf et al., 2004). Figure 5 illustrates their construction. However, these systems also have their limitations. They work well for mildly acidic water (pH > 4.5) that is anoxic, but highly acidic water tends to contain dissolved aluminum, which precipitates in the ALD and reduces permeability, often to the point of failure. In addition, if the pH of the water is below about 3.5, the dissolved iron is often already oxidized (ferric), so that armouring can occur even if no oxygen is present.

Figure 5. Three photographs of an anoxic limestone drain (ALD) being constructed. First a ditch is constructed, then
limestone is added and covered with clay or PVC, and then an effluent pipe is emplaced to allow the water to flow out
of the ALD without allowing air to flow back into it. Finally, the ALD is connected to the underground mine discharge
in such a way that no air is introduced into the water.


To compensate for dissolved oxygen and dissolved ferric iron, the concepts of the ALD and compost wetland were combined (Kepler and McCleary, 1994; Kepler, 1995). Compost was placed upgradient of the limestone. The bacterial activity in the compost consumed the dissolved oxygen and reduced the ferric iron to ferrous iron, allowing the ALD component to work as intended, even for very acidic water. These systems were initially referred to as sequential alkalinity-producing systems (SAPS); others have preferred to use the term reducing and alkalinity producing systems (RAPS) (GARD Guide to more accurately describe the process, and to include systems that did not put more than one unit in sequence. These systems are also called vertical flow ponds, vertical flow wetlands or vertical flow systems. A photograph of a RAPS is shown as Figure 6. Aluminum is still retained in these systems, and can cause permeability problems, so Kepler and McCleary (1997) suggested a simple gravity-powered flushing mechanism to extend their effective life span.

Figure 6. A photograph and a cross-section diagram of a RAPS installation

In general, it is difficult to argue with the long-term success of many of these passive treatment systems. However, failures can be very damaging to the perceived effectiveness of the technology. In general, it is fair to state that systems that were not effective or failed were undersized, improperly designed, or both. The key is to understand the limitations of each unit’s operation, to have reasonable expectations, and to use conservative sizing criteria to attain specific water quality goals. Even undersized passive systems can be useful, discharging water with significantly lower concentrations of metal contaminants than were present 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 abandoned sites have on receiving streams and lakes.

In practice, RAPS, ALDs, settling ponds, and aerobic wetlands are used as unit operations in a total CMD passive remediation system. Researchers are continuing to develop additional passive treatment technologies, such as steel slag leach beds (Simmons et al., 2002), which are adding to the passive treatment arsenal. Research is also being conducted on semi-passive approaches that have the potential to significantly reduce the land requirements of passive treatment systems.