Chapter 2

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2.0 The Acid Rock Drainage Process

2.1 Introduction
2.2 Acid Rock Drainage, Neutral Mine Drainage, and Saline Drainage
2.2.1 Definition of Acid Rock Drainage, Neutral Mine Drainage, and Saline Drainage
2.2.2 Natural vs Anthropogenic Drainage
2.3 History of Acid Rock Drainage
2.4 The Acid Generation Process
2.4.1 Characteristics of Acid Rock Drainage, Neutral Mine Drainage, and Saline Drainage
2.4.2 The Global and Geochemical Sulphur Cycles
2.4.3 Acid Rock Drainage Sources
2.4.4 The Sulphide Oxidation Process
2.4.5 Reaction Products from Sulphide Oxidation
2.4.6 Neutralization Reactions
2.4.7 The Acid Rock Drainage Process and Migration
2.5 Concluding Statement
2.6 References
List of Tables
List of Figures

2.0 The Acid Rock Drainage Process

2.1 Introduction

Chapter 2 presents an introduction to acid rock drainage (ARD), the history of ARD, and an overview of ARD processes and definitions. This chapter also provides a description of the sulphide oxidation process, including the biological, chemical, and physical factors that govern sulphide oxidation, control migration of ARD, and that modify the compositional characteristics of mine discharges along flow paths. A brief evaluation of receptors and potential impacts resulting from ARD is presented in closing.

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2.2 Acid Rock Drainage, Neutral Mine Drainage, and Saline Drainage

2.2.1 Definition of Acid Rock Drainage, Neutral Mine Drainage, and Saline Drainage

Throughout this document, the terms acid rock drainage (ARD), neutral mine drainage (NMD), and saline drainage (SD) are used. All three types of drainage can be produced by oxidation of sulphide minerals. In the context of this GARD Guide, ARD, NMD, and SD do not represent consecutive stages in the evolution of mine waters, but instead reflect different end points in terms of water quality that may have different effects on the environment and may necessitate different forms of management.

Figure 1-1 in Chapter 1 provides an overview of the compositional characteristics of each of these drainage types. Although formal guidelines for quantitative definitions of ARD, NMD, and SD are lacking, for the purpose of this GARD Guide, the following approximate thresholds between the three types of mine drainage are applied:

  • Above pH 6: NMD and SD; and below pH 6: ARD
  • Sulphate concentration of 1,000 milligrams per liter (mg/L): threshold between NMD vs. SD (in accordance with U.S. Geological Survey [USGS] cut-off between fresh water and slightly saline water of 1,000 mg/L TDS [USGS, 2004])

A more detailed description of the characteristics of ARD, NMD, and SD and the factors that control their development is provided in Section 2.4.

A term frequently used as being synonymous with ARD is acid mine drainage (AMD). “AMD” continues to be commonly used, especially where ARD is derived from coal mines. However, use of "acid mine drainage" is discouraged as a general term because it implies that acid drainage originates from mining activities alone. Acid drainage can result from a variety of natural sources and anthropogenic activities; therefore, acid mine drainage is too narrow a descriptor for many situations. For example, highway construction and associated excavations have led to a generation of acid rock drainage in such locations as British Columbia, Pennsylvania, and Northumberland because of exposure of sulphide-bearing mineralized zones. Waste from sand and oil shale extraction in Australia has also been noted to produce ARD on occasion.

Another term that broadly refers to waters that have been impacted by mining or mineral processing facilities is mining influenced water (MIW). This term includes ARD, NMD, SD, and metallurgical process waters of potential concern. In Australia, the term acid and metalliferous drainage (AMD) is used as a synonym for ARD. A key characteristic of most of these waters is that they contain elevated metals that have leached from surrounding solids (e.g., waste rock, tailings, mine surfaces, or mineral surfaces in their pathways). This fact is commonly acknowledged by the phrase “metal leaching” (ML), frequently resulting in acronyms such as ARD/ML.

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2.2.2 Natural vs Anthropogenic Drainage

The primary process responsible for generation of ARD, NMD, or SD of concern is weathering of sulphide minerals, in particular pyrite. In some cases, the generation of ARD, NMD, or SD may also be due to oxidation of elemental sulphur. Weathering, or oxidation, of pyrite occurs naturally when exposed to atmospheric conditions, either through geologic processes or anthropogenic activities that involve removal of material (e.g., mining, highway construction). Historical records clearly indicate that many mineralized areas contained natural waters with low pH and elevated concentrations of metals and sulphate before the onset of mining, and geographic names such as Rio Tinto (Red River), Rio Agrio (Sour River), Sulphur Creek, Rötlbach (Red Brook) and Copper Creek are testimony to the presence of low pH and elevated metal concentrations.

Mining and other forms of earth moving, however, greatly accelerate the weathering of reactive sulphides because they create conditions that tend to facilitate movement of air and water, expose large volumes of material, increase the surface area of the reactive component, and create the opportunity for colonization by microorganisms that catalyze the oxidation processes in the presence of acidity. As a consequence, the potential environmental consequences of human activities can be significantly more noticeable than those resulting from natural processes.

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2.3 History of Acid Rock Drainage

Like other human endeavours, mining and the use of mineral resources has resulted in environmental consequences. While organized mining may have originated around 6,000 BC (principally associated with extraction of alluvial gold and flint stone), there is substantial evidence that by the third millennium BC (early Bronze Age) there was contamination resulting from copper mining and smelting in the Iberian Pyrite Belt (Figure 2-1). On a regional scale, the effects of mining and smelting included heavy metal pollution in river water and sediments, increased erosion, and deforestation. Atmospheric pollution present as metal-rich layers in ice cores from Greenland signifies more global effects of these historical mining and smelting activities (Nocete et al., 2005). Similar observations have been made regarding Bronze Age copper mining in Ireland, Great Britain, and Austria.

Figure 2-1: Roman Portal with Acid Rock Drainage – Spain


As mining progressed throughout the Iron Age, the Roman Empire, and medieval times, the environmental effects of mining continued without controls. Specific references to reactive sulphides and their degradation to acid and salts date from as early as the Roman era, and by the time Georgius Agricola published his seminal and oft-quoted work on mining and metallurgy in the mid-16th century (Agricola, 1556), ARD and its effects on human health and the environment were known. The Industrial Revolution was made possible through, and required extraction of, vast amounts of mineral resources; in particular coal, and the associated water and air pollution became more widespread. Little was done at the time to mitigate these impacts because of a lack of knowledge about the fundamentals of ARD generation, and because not much was known about the consequences of environmental effects in general.

Modern mining has yet again increased the degree of resource recovery (and thereby the potential for ARD generation). However, this resource recovery increase has been accompanied by an increased awareness and understanding of potential environmental consequences and an increased scientific and technical understanding of ARD management tools. As a consequence, organized mining is currently conducted in accordance with a (generally extensive) set of guidelines and regulations aimed at operating in an environmentally responsible manner.

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2.4 The Acid Generation Process

This section presents a summary of the acid generation process, including the sulphur cycle and its weathering products and the factors that control generation and migration of ARD, NMD, and SD. Potential impacts from ARD, NMD, and SD on specific receptors are also briefly discussed. More detail on the genesis of coal mine drainage (CMD) is presented here: Introduction to Coal Mine Drainage.

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2.4.1 Characteristics of Acid Rock Drainage, Neutral Mine Drainage, and Saline Drainage

The generation, release, mobility and attenuation of ARD, NMD, and SD are complex processes governed by a combination of physical, chemical and biological factors. Whether ARD, NMD or SD enters the environment depends largely on the characteristics of the sources, pathways, and receptors involved. A generalized conceptual model of sources, pathways, and receiving environments is shown in Figure 2-2. These sources, pathways and receiving environments vary by commodity, climate, mine facility, and mine phase. The sources include the mine and process wastes and mine and process facilities that contain reactive sulphide and potentially neutralizing minerals involved in mitigation of acidity. The characteristics and relative abundance of these sulphides and neutralizing minerals, which play a critical role in determining the nature of the discharge being generated, may vary as a function of commodity and ore-deposit type, type of mining, and waste-disposal strategy. The pathways and transport mechanisms are related to climate and seasonal effects and the hydraulic characteristics of the mine or process waste/facility that represents the source. Climate and seasonal effects may determine whether a mine discharge is continuous or intermittent, dilute or highly concentrated, which has an effect on the nature of the drainage. The hydraulic characteristics of a mine or process waste/facility may determine the contact time between solid and solution (e.g., rapid preferential flow vs. gradual matrix flow) or the proportion of mine waste being flushed. The receptors (i.e., the receiving environment) may also alter the nature of the mine drainage. Examples of receiving environments include groundwater, surface water, or wetlands. All of these receiving environments can alter the original characteristics of the mine discharge through a combination of physical mixing, chemical, and biological reaction.

Figure 2-2: Generalized Conceptual Model of Sources, Pathways and Receiving Environment at a Mine or Processing Site

The influence of commodity, climate, mine or process facility, and mine phase on the nature of the mine drainage (ARD, NMD or SD) can be illustrated using Ficklin diagrams or analogue versions. Ficklin diagrams are plots that can be used to interpret variations in mine drainage water chemistry between different deposits (Plumlee et al., 1999). These diagrams were developed in support of the use of geoenvironmental models, which are constructs that interpret the environmental characteristics of an ore deposit in a geologic context. Geoenvironmental models provide a very useful way to interpret and summarize the environmental signatures of mining and mineral deposits in a systematic geologic context. Geo-environmental models can also be used to anticipate potential environmental problems at future mines, operating mines, and orphan sites.

The traditional Ficklin plot is a scattergram in which the sum of the base metals zinc (Zn), copper (Cu), lead (Pb), cadmium (Cd), cobalt (Co), and nickel (Ni) is plotted against pH. These parameters were selected rather than more common metals such as iron (Fe), aluminum (Al), and manganese (Mn) because they have proven the most diagnostic in differentiating between different geologic controls. However, similar plots using parameters other than Zn, Cu, Pb, Cd, Co, and Ni can also be used to demonstrate the effect of commodity, climate, mine facility, and mine phase.

Figure 2-3 shows a Ficklin plot that represents a compilation of data provided in Plumlee et al., (1999) for a wide variety of ore deposit types. Individual data points are not presented, but instead the shaded outline presents the range of major and trace metal concentrations and pH for all deposit types in this publication. Figure 2-4 is a Ficklin analogue, which now shows the range of sulphate concentrations observed in mine waters, also based on data from Plumlee et al. (1999). Superimposed on both plots are the approximate outlines of the ARD, NMD, and SD fields. These outlines should not be construed as representing strict classifications because there are no formal guidelines for quantitative definitions of ARD, NMD, and SD.

Figure 2-3: Ficklin Diagram Showing ARD, NMD, and SD as a Function of
Dissolved Base Metal Concentrations (adapted from Plumlee et al., 1999)


Figure 2-4: Diagram Showing ARD, NMD, and SD as a Function of Sulphate Concentrations

The data compilations presented in Figure 2-3 and Figure 2-4 also include mine water qualities not resulting from sulphide oxidation. However, the fields for ARD, NMD, and SD are drawn so that this nomenclature covers the entire range of water qualities observed, including water types that may deviate from the proper definitions. For example, acidic water with low metal and sulphate levels is captured in the ARD field even though acidic water may not originate from sulphide oxidation but may, for instance, reflect weathering of soils rich in hydrous iron oxides such as laterites.

Typical ore-deposit types most commonly associated with ARD include volcanogenic massive sulphide (VMS) deposits, high sulphidation epithermal deposits, porphyry copper deposits and skarn deposits. Coal deposits also frequently generate ARD. Typical deposit types associated with SD include Mississippi-Valley Type (MVT) deposits, low-sulphide gold-quartz vein deposits, and “clean” skarns. NMD can be generated by a wide variety of ore deposits, depending on the type of alteration and sulphide content, including most types listed for ARD and SD.

Ficklin diagrams can also be used to illustrate a number of principles that govern mine water quality (Figure 2-5). In this figure, a number of trend lines demonstrate the generic effect of increasing pyrite content, increasing base-metal sulphide content, and increasing carbonate content on mine water quality. As portrayed in the diagrams, an increase in pyrite content tends to result in more acidic waters. An increase in base-metal sulphide content tends to result in an increase in trace metal concentrations, and an increase in carbonate content tends to lead to more alkaline waters. However, these trends must be interpreted with caution. For example, some deposits can be carbonate rich but can still generate acidic waters if the acid-buffering carbonates are physically separate from the sulphides, if a reaction barrier of iron (hydr)oxides coats the carbonates and prevents their dissolution, or if the carbonates are associated with metals that release acid when precipitated as hydroxides. Therefore, site-specific evaluation of geochemical and geological characteristics of the ore and mine wastes is required.

Figure 2-5: Ficklin Diagram Showing Selected Principles that Govern
Mine Water Quality (adapted from Plumlee et al., 1999)


Generic effects of climate are superimposed on Figure 2-5. In very general terms, mine waters from acid generating deposits in arid climates tend to be more acidic and metalliferous due to enhanced evaporation and a greater solid to water ratio during water/rock interaction. Conversely, the greater dilution and reduced solid to water ratio in wetter climates generally leads to mine waters from acid generating deposits with a less acidic and concentrated character. The less common evapoconcentration of mine waters with an alkaline nature (not shown on the figure) tends to result in waters that are more basic, while dilution in wetter climates tends to reduce the alkalinity. Cryoconcentration (i.e., concentration due to freezing) in arctic environments may lead to mine waters with elevated concentrations of trace metals and sulphate. In addition, cryoconcentration tends to increase either the acidic or alkaline nature of the mine effluent. Although climate is a key control on mine water quality, according to Plumlee (1999), the relative shifts in pH and metal content for a given deposit type in different climatic settings are generally of lesser importance than the changes due to the differences in geologic characteristics. Also, the effect of climate on environmental impacts downstream from a mineral deposit should not be ignored. Such effects can be quite significant. For example, downstream dilution is much enhanced in wetter climates relative to dry climates, while seasonal occurrences such as the spring freshet and intense rainfall events can produce short-term high loads of contaminants with potentially dramatic effects on downstream mine water quality.

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2.4.2 The Global and Geochemical Sulphur Cycles

Sulphur plays an important role in the formation of ARD, NMD, and SD. Sulphur is a very versatile element that can occur in many different chemical forms and oxidation states. The chemical forms of most importance with respect to mine discharges are elemental sulphur, sulphate (in mineral form as well as aqueous), and sulphide (in mineral form and, to a lesser degree, aqueous and gaseous). The corresponding oxidation states of sulphur in minerals are So, S6+ and S2-, respectively. Sulphur speciation, and its associated potential environmental impacts, is therefore strongly related to the reduction-oxidation (redox) properties of the aqueous systems with which it interacts. The oxidation and reduction processes that involve sulphur species tend to be slow unless mediated by microorganisms. However, oxidation of certain sulphide minerals present in mine and process wastes (e.g., pyrrhotite) can be very rapid even in the absence of biotic mediation. In extreme cases, oxidation can result in self-heating and combustion, which necessitates the use of special precautions when handling these materials.

The global cycle of sulphur is characterized by a rather rapid recycling of aqueous forms in water and is also characterized by gases and aerosols in the atmosphere. Sulphur present in reduced form in sulphide minerals is relatively immobile. However, after sulphur is exposed through mining or other earthmoving activities, following oxidation, it can have a significant effect on the receiving environment through formation of ARD, NMD, or SD. A large labile reservoir of sulphate species exists incorporated in sediments and in dissolved form in the world’s oceans.

A simplified global sulphur cycle is provided in Figure 2-6 (Stumm and Morgan, 1996), but the quantities presented are approximate at best. In this diagram, the weathering of sulphur-bearing minerals (gypsum and sulphides), volcanic emissions, and the recirculation of sea salt and biogenic gases represent the most important natural segments of the cycle. The global sulphur cycle is strongly affected by anthropogenic inputs. Although the exact quantities and ecological consequences of human contributions are only partially understood, it is generally accepted that human contributions are significant, particularly on the continental land masses where human activity is concentrated and the impacts of such activity are felt. Major factors in the sulphur cycle are the combustion of coal and petroleum. Other factors include industrial processes such as smelting and refining of sulphide ores, which release sulphur oxides into the atmosphere. These emissions are a substantial contributor to the formation of acid rain, which has had many undesirable effects in Europe, the United States, and Canada. More recent undesirable effects have been seen in nations industrializing at a very rapid rate such as China and India. Human activities also add to the natural flux of dissolved sulphur in river water due to increased erosion and industrial activity.

Figure 2-6: The Global Sulphur Cycle (Stumm and Morgan, 1996). Global Fluxes in
Millions Tons of Sulphur per Year and Inventories in Millions Tons of Sulphur


The biogeochemical cycle of sulphur is relatively simple in concept, as illustrated in Figure 2-7, with all components of the cycle being heavily influenced by microorganisms. Most of the sulphur in the earth’s sediments and crust is present in the form of primary elemental sulphur and sulphide minerals, which can be oxidized into sulphate through both biotic and abiotic processes. This is the process responsible for formation of ARD, NMD, or SD. Sulphate in soils can be taken up by plants and assimilated into proteins. When plants die and decay, microorganisms mineralize the sulphur in the proteins into hydrogen sulphide or sulphate. The hydrogen sulphide can then be combined with metals to form metal sulphides, or the hydrogen sulphide can be oxidized to elemental sulphur or sulphur dioxide, depending on redox conditions and involvement of biota. In cases where hydrogen sulphide combines with metals, authigenic or secondary sulphide minerals are formed. In the atmosphere, sulphur dioxide may be oxidized and combine with water to form sulphuric acid, which may report to the terrestrial and aqueous environment as acid rain. Direct transformation between sulphate and hydrogen sulphide can be accomplished through a variety of processes that are generally biologically mediated.

Figure 2-7: The Biogeochemical Sulphur Cycle

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2.4.3 Acid Rock Drainage Sources

In the context of this GARD Guide, which focuses on drainages produced by sulphide mineral oxidation, the potential sources for generation of ARD, NMD, and SD are reactive sulphide minerals and their oxidation products. Although other naturally occurring minerals and byproducts from pyrometallurgical and hydrometallurgical processes may generate acidic solutions (e.g., elemental sulfur, jarosites, and other hydroxyl-sulphates such as alunites), these are not discussed in this GARD Guide. The most common sulphide mineral is pyrite [FeS2]. Pyrite is the mineral of most relevance from an acid-generation perspective, because its concentration, grain size, and distribution may be the most important factors affecting the production of acidic mine waters (Nordstrom and Alpers, 1999). Other sulphides commonly found in ore deposits are listed in Table 2-1 (Plumlee, 1999). These sulphides may produce ARD, NMD, or SD. Secondary minerals resulting from sulphide oxidation include a complex array of soluble sulphates, hydrous sulphates, hydroxysulphates, metal oxides and hydroxides, clays, carbonates and supergene, and diagenetic sulphides. Some of these secondary minerals may have deleterious effects on water quality because of the release of additional acidity during their formation (e.g., metal (hydr)oxides) or release of stored acidity, sulphate or metals (or both sulphate and metals) during their dissolution (e.g., iron and aluminum hydroxysulphates).

Table 2-1: Common Sulphides Known or Inferred to Generate Acid when Oxidized (Plumlee, 1999)



Common sulphides known (inferred) to generate acid with oxygen as the oxidant:

  Pyrite, marcasite


















Common sulphides that may generate acid with ferric iron as the oxidant:

  All of the above plus:

















The type and distribution of sulphide minerals can vary widely according to the type of ore deposit, nature of the mine waste, and mine stage. Certain types of ore deposits can be devoid of sulphides (e.g., oxide-facies banded iron formation [BIF] deposits) while others contain very substantial amounts of sulphide (e.g., VMS deposits). Similarly, within an individual ore deposit, the distribution of sulphide minerals can range from disseminated evenly throughout the deposit to the sulphides being confined to specific zones. The distribution depends on the nature of the original ore-forming processes or subsequent alteration. The techniques used to extract and process ores can also significantly affect the type and distribution of sulphide minerals, which in turn has ramifications regarding the nature of corresponding mine discharges. The large volume of waste rock resulting from an open pit mine may have a lower overall sulphide content than the smaller volume originating from an underground mine. This lower overall sulphide content would be due to a greater dilution with rock not directly associated with the mineralization.

A wide variety of mineral processing methods are being used. Some of these methods result in very different types of source materials, especially regarding tailings. For example, where the economic resource is associated with one or more sulphide minerals, its recovery may result in a tailings stream that is low in that particular sulphur mineral association. However, less commonly, where this is not the case, the tailings may represent a material that is concentrated in sulphur relative to the original ore. The type and distribution of sulphide minerals can vary over the life of the mine. This may be due to a change in ore type or processing methodology (or both ore type and processing methodology) over time (e.g., from milling to heap leach as the ore grade decreases or the nature of the ore changes from an oxide-rich to a more sulphidic variant). In addition, because of the ongoing weathering of sulphides, they may become depleted during the life of the mine if sufficiently reactive or, more likely, at some point after mine closure. As illustrated earlier with the various Ficklin diagrams in Section 2.1, all of the above implies that mine drainage generated through sulphide oxidation can show considerable compositional range over the life of the mine as a function of the type of ore deposit and mine or process waste. These site-specific considerations need to be acknowledged before mine development, during operation, and after closure so that they can be accounted for in the identification and implementation of appropriate mine waste and water management practices.

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2.4.4 The Sulphide Oxidation Process

The process of sulphide oxidation and formation of ARD, NMD, and SD has been discussed in many references (Stumm and Morgan, 1996; Nordstrom and Alpers, 1999), and only a summary is presented here. Sulphide minerals in ore deposits are formed under reducing conditions in an absence of oxygen. When exposed to atmospheric oxygen or oxygenated waters due to mining, mineral processing, excavation, or other earthmoving processes, sulphide minerals can become unstable and oxidize.

Figure 2-8 presents a model describing the oxidation of pyrite (Stumm and Morgan, 1981). The reactions shown are schematic and may not represent the exact mechanisms, but the illustration represents a useful visual aid for discussing pyrite oxidation.

Figure 2-8: Model for the Oxidation of Pyrite (Stumm and Morgan, 1981).
(The numbers in brackets refer to the reactions presented in Section 2.4.4)


The chemical reaction representing pyrite oxidation requires three basic ingredients: pyrite, oxygen, and water. The overall pyrite oxidation reaction generally is written as:

FeS2 + 7/2O2 + H2O = Fe2+ + 2SO42- + 2H+ [1]

This reaction can occur both abiotically or biotically (i.e., mediated through microorganisms). In addition to direct oxidation, pyrite can also be dissolved and then oxidized (reaction [1a] on Figure 2-8).

Under most circumstances, atmospheric oxygen acts as the oxidant. Oxygen dissolved in water can also result in pyrite oxidation but due to its limited solubility in water, this process is much less prominent. Aqueous ferric iron can oxidize pyrite as well according to the following reaction:

FeS2 + 14Fe3+ + 8H2O = 15Fe2+ + 2SO42- + 16H+ [2]

This reaction is considerably faster (2 to 3 orders of magnitude) than the reaction with oxygen and generates substantially more acidity per mole of pyrite oxidized but it is limited to conditions in which significant amounts of dissolved ferric iron occur (i.e., acidic conditions). Therefore, pyrite oxidation is generally initiated through reaction [1] at circumneutral or higher pH, followed by reaction [2] when conditions have become sufficiently acidic (approximately pH 4.5 and lower). A third reaction is required to generate and replenish ferric iron, through oxidation of ferrous iron by oxygen as follows:

Fe2+ + ¼O2 + H+ = Fe3+ +½H2O [3]

A common misunderstanding is that ferric iron can oxidize pyrite indefinitely in the absence of oxygen. As indicated by reaction [3], oxygen is required to generate ferric iron from ferrous iron. Also, the bacteria that may catalyze this reaction (primarily members of the Acidithiobacillus genus) are obligate aerobes (i.e., they require oxygen for aerobic cellular respiration). Therefore, some nominal amount of oxygen is needed for this process to be effective even when catalyzed by bacteria, although the oxygen requirement is less than for abiotic oxidation.

A process of environmental importance related to pyrite oxidation pertains to the fate of ferrous iron generated through reaction [1]. Ferrous iron can be removed from solution under slightly acidic to alkaline conditions through oxidation and subsequent hydrolysis and the formation of a relatively insoluble iron (hydr)oxide. Assuming the nominal composition of ferrihydrite [Fe(OH)3] for the latter phase, this reaction can be summarized as:

Fe2+ + ¼O2 + 2½H2O = Fe(OH)3+ 2H+ [4]

When reactions [1] and [4] are combined, as is generally the case when conditions are not acidic (i.e., pH > 4.5), it can be seen that oxidation of pyrite generates double the amount of acidity relative to reaction [1] as follows:

FeS2 + 15/4O2 + 7/2H2O = Fe(OH)3 + 2SO42- + 4H+ [5]

A variety of microorganisms are abundant in mine waters and when conditions become highly acidic they may be the only form of life. Included in the bacterial fauna are iron and sulphur-oxidizing bacteria (e.g., A. ferrooxidans and A. thiooxidans) These microbes play an important role in sulphide oxidation and in the formation of ARD, NMD, or SD. Due to microbial mediation, many important geochemical reactions take place against thermodynamic expectations (Mills, 1999) because bacteria can couple a thermodynamically unfavourable reaction with a reaction that yields net energy. Rates of reactions, such as iron oxidation, which in turn affects the rate of pyrite oxidation, may be increased by many orders of magnitude relative to the corresponding abiotic rates (Nordstrom and Alpers, 1999; Nordstrom, 2003; Gould and Kapoor, 2003). For example, the oxidation rate of ferrous iron to ferric iron (reaction [3]) can be increased by 5 to 6 orders of magnitude in the presence of iron-oxidizing bacteria. Although the exact reaction mechanism of pyrite oxidation on a molecular level is still under investigation (a recent discussion of the state of knowledge and research is given in Wolkersdorfer [2008]), the rate-limiting step is the production of ferric iron from ferrous iron through microbial catalysis. Figure 2-9 (Robertson and Broughton, 1992) provides a schematic illustration of the normalized relative oxidation rates with and without bacterial mediation as a function of pH.

Figure 2-9: Schematic Illustration of Normalized Sulphide Oxidation Rates with and without
Bacterial Mediation (after Robertson and Broughton, 1992)


Although pyrite is by far the dominant sulphide responsible for the generation of acidity, different ore deposits contain different types of sulphide minerals, not all of which generate acidity when being oxidized. As a general rule, iron sulphides (pyrite, marcasite, pyrrhotite), sulphides with molar metal/sulphur ratios < 1, and sulphosalts (e.g., enargite) generate acid when they react with oxygen and water. Sulphides with metal/sulphur ratios = 1 (e.g., sphalerite, galena, chalcopyrite) tend not to produce acidity when oxygen is the oxidant. However, when aqueous ferric iron is the oxidant, all sulphides are capable of generating acidity. Therefore, the amount of iron sulphide present in an ore deposit or mine waste plays a crucial role in determining the characteristics of the mine drainage. While pyrite is the most common source mineral for iron, sulphides such as chalcopyrite and ferriferous sphalerite can also act as iron donors. As a result, mine waters originating from such materials tend to be significantly more acidic than discharges from sulphide assemblages that primarily include sphalerite and galena. Oxidation of the sphalerite and galena still occurs, resulting in release of sulphate and trace metals such as zinc and lead, respectively. Should these metals remain in solution, NMD will be generated. Table 2-1 (Plumlee, 1999) provides an overview of sulphide minerals and their known or inferred potential to generate acid when oxidized by either oxygen or ferric iron.

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2.4.5 Reaction Products from Sulphide Oxidation

The potential reaction products from sulphide oxidation include acidity, sulphur species, total dissolved solids, and metals. The degree to which these reaction products are being generated and persevere in the receiving environment determines whether ARD, NMD, or SD results.

The production and persistence of acidity largely depends on the nature of the sulphide mineral being oxidized, the reaction mechanism (i.e., oxygen vs. ferric iron as the oxidant), and the presence of acid-consuming minerals. In most ore deposits and mine wastes, sulphide minerals occur in a mineral assemblage that also includes acid-consuming minerals such as carbonates and aluminosilicates. More detail on acid neutralization mechanisms is provided in Section 2.6.6.

The sulphur species generated from sulphide oxidation is sulphate. Under the acidic conditions commonly encountered at some mining sites, dissolved sulphate concentrations can be up to approximately 10,000 mg/L (Figure 2-4). However, in extreme cases, concentrations over 100,000 mg/L have been observed. An example of an extreme case is at the Iron Mountain VMS deposit in California, where pH values may be as low as -2 to -3 (Nordstrom and Alpers, 1999). As conditions become more alkaline, sulphate concentrations are usually governed by the solubility product of gypsum [CaSO4•2H2O], which tends to limit sulphate levels to a few thousand milligrams per litre. Other dissolved sulphur species known to occur in discharges associated with mining activity and during mineral processing are bisulphide (HS-), sulphide (S2-) and thiosalts (sulphur oxyanions, including polysulphides [Sn2-], sulphoxy anions such as thiosulphate [S2O32-], polythionates [SnO62-], and sulphite [SO32-]). The thiosalts occupy a metastable position between sulphate and sulphide, and both sulphide and thiosalts will naturally oxidize to sulphate under atmospheric conditions, generating acidity in the process. There is little evidence that dissolved sulphides and thiosalts originate from oxidation of sulphide minerals in ore deposits or mine wastes. Instead, dissolved sulphide tends to be a by-product of active water treatment at mining sites (for instance, metal precipitation using Na2S or NaSH). Sulphides can also occur because of interaction between sulphate and organic matter in reducing environments, either in natural settings, active sulphate reducing bacteria (SRB) treatment systems, or in passive treatment systems such as constructed wetlands. Thiosalts usually result from partial oxidation of sulphide minerals during ore processing. The presence of sulphides or thiosalts in process water can be a significant concern in acidification of tailings supernatant and pore water, and can also potentially be of concern in downstream receiving waters.

Total dissolved solids (TDS) concentrations in mine and process discharges are usually directly related to the amount of sulphate, chloride, or bicarbonate present in solution. Although other constituents may also increasingly contribute to TDS when pH decreases and mineral dissolution becomes more effective, sulphate, chloride, or bicarbonate tend to represent the dominant anion contribution to TDS. In the case of alkaline mine discharges (e.g., from kimberlites), alkalinity in the form of carbonate and bicarbonate ions may represent the most important proportion of TDS.

Major and trace metals in ARD, NMD, and SD are sourced from the oxidizing sulphides and dissolving acid-consuming minerals. In the case of ARD, Fe and Al are usually the principal major dissolved metals, with concentrations that can range from 1,000s to 10,000s mg/L . Trace metals such as Cu, Pb, Zn, Cd, Mn, Co, and Ni can also achieve elevated concentrations in ARD, reaching levels from 100s to 1,000s of mg/L (Figure 2-3). In mine discharges with a more circumneutral character, trace metal concentrations tend to be lower due to formation of secondary mineral phases and increased sorption of trace metals onto a variety of sorbents such as metal (oxy)hydroxides, clay minerals, and reactive particulate carbon (Smith, 1999). However, certain parameters remain in solution as the pH increases, in particular the metalloids As, Se, and Sb as well as other trace metals (e.g., Cd, Cr, Mn, Mo, and Zn). The resulting mine or process discharge is NMD, and treatment for these parameters can be challenging. As conditions become even more alkaline, some of these species will precipitate as carbonates or hydroxides (e.g., Zn and Mn) but others may remain in solution (e.g., Cr, As, Se, and Sb) while others (e.g., Al) may become remobilized, such as in alkaline drainages from kimberlite deposits. The mobility (and toxicity) of several environmentally significant trace metals is governed by their oxidation state, for instance for As, Se, uranium (U), and chromium (Cr). Mine waters tend to have an oxidized character, which favours the less mobile arsenic species [As(V)], but enhances the mobility of chromium, selenium, and uranium in the form of [Cr(VI)], [Se(VI)] and [U(VI)], respectively.

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2.4.6 Neutralization Reactions

Neutralization reactions play a key role in determining the compositional characteristics of drainage originating from sulphide oxidation. Generic reactions for consumption of acid because of dissolution of carbonate and silicate minerals (using plagioclase as an example) can be written as:

MeCO3 + H+ = Me2+ + HCO3- [6] (where Me represents a divalent cation, such as calcium or magnesium, but not iron or manganese because these release acidity after subsequent hydrolysis/precipitation)


CaAl2Si2O8 + 8H+ = Ca2+ + 2Al3+ + 2H4SiO4 [7]

As for sulphide minerals, the reactivity, and accordingly the effectiveness with which these minerals are able to buffer any acid being generated, can vary widely. Most carbonate minerals are capable of dissolving rapidly, making them effective acid consumers. Although generally more common, aluminosilicate minerals tend to be less reactive, and their buffering may only succeed in stabilizing the pH when low values have been achieved. In some cases, when sulphide oxidation rates and flushing rates are very low, certain silicate minerals, in particular calcium-magnesium (Ca-Mg) silicates, have been known to buffer mine effluents at neutral pH (Jambor, 2003).

Hydrolysis of dissolved Fe, Mn, or Al following dissolution of acid-consuming minerals and subsequent precipitation of a secondary mineral according to reaction [4] may generate acidity. As a consequence, the net effect of their dissolution in terms of acid consumption may be significantly less than expected from reactions such as [6] and [7], or they may even generate net acidity at formation of the secondary mineral phase. Examples of such incongruent dissolution reactions that are the equivalents of congruent reactions [6] and [7] are as follows:

FeCO3 + ¼O2 + 2½H2O = Fe(OH)3 + H+ + HCO3- [8]


CaAl2Si2O8 + 2H+ + H2O = Ca2+ + Al2Si2O5(OH)4 (kaolinite) [9]

Table 2-2 provides an overview of the ranges of neutralization potential and buffering pH for a number of common minerals. As is immediately obvious, carbonate minerals generate significantly more neutralization potential than silicate minerals, while they also tend to buffer at higher pH values. Effective neutralization, in practice is therefore generally directly related to the abundance of non-Fe/Mn carbonate minerals.

Table 2-2: Typical NP Values and pH Buffering Ranges for Some Common Minerals (Jambor, 2003; Blowes et al., 2003; BCAMDTF, 1989)



Buffer pH

Potential Range

(kg CaCO3/tonne))





   calcite, aragonite


5.5 – 6.9




5.1 – 6.0




5.1 – 6.0








3.7 – 4.3




3.0 – 3.7




2.8 – 3.0




1.7 – 2.0




0.5 – 1.5


Feldspar Group
















Pyroxene Group




Amphibole Group




Mica Group












Chlorite Group




Clay Group




Garnet Group




Apatite Group
























Me = monovalent, divalent or trivalent cation

The combination of acid generation and acid neutralization reactions typically leads to development of ARD, as illustrated in Figure 2-10 (Broughton and Robertson, 1992). Over time, pH decreases along a series of plateaus governed by the buffering of a range of mineral assemblages. Stage I is characterized by a circumneutral pH range and consumption of acid by carbonate minerals such as calcite according to generic reaction [6]. Reactions [1], [3], and [4] describe the sulphide oxidation process, which results in sulphate release. As this alkalinity is consumed, the pH declines in stages depending on the nature of the neutralizing minerals (Stage II). Generally, at this stage sulphate, acidity, and trace metal levels increase although for some metals (e.g., Cu, Pb), concentrations may be limited by mineral solubility controls. Buffering may be provided by metal hydroxides. At pH values of approximately 4.5 and below, microbially mediated oxidation predominates (reactions [2] and [3]), resulting in a rapid acceleration of acid generation. In Stage 3, final buffering is usually limited to dissolution of silicate minerals (reaction [7]), and solubility controls on trace metal concentrations are largely absent.

Figure 2-10: Stages in the Formation of ARD (after Broughton and Robertson, 1992).
Note: The numbers in brackets refer to the reactions presented in Chapter 2.6.4

The lag time to acid generation, as shown in Figure 2-10, is an important consideration in ARD prevention. It is far more effective (and less costly in the long term) to control ARD generation during Stage I than during the subsequent stages. The lag time also has important ramifications for interpretation of test results. It is critical to recognize the stage of oxidation when predicting ARD potential: the first stage may last for a very long time (up to years), even for materials that will eventually be highly acid generating. The early results of geochemical testing, therefore, may not be representative of long-term behaviour and discharge quality.

2.4.7 The Acid Rock Drainage Process and Migration

This section describes the factors that control ARD formation, migration of ARD, and potential impacts of ARD to receptors. The following three aspects of ARD formation and migration are identified and discussed:

  • Factors that govern the rate of sulphide oxidation
  • Factors that modify the composition of resulting drainage in the mine or process waste
  • Factors that modify the composition of drainage after exiting the mine or process waste facility

These three aspects are schematically illustrated in Figure 2-11. The factors that govern the rate of sulphide oxidation are described individually in Section and represent the source term in Figure 2-2. The factors that modify the composition of the resulting drainage both within the mine or process waste and after exiting the mine or process facility are illustrated in Section These factors are represented by the pathways in Figure 2-2. Finally, the receiving environment is discussed in Section

Figure 2-11: Schematic Illustration of Factors that Affect Sulphide Oxidation and Modify Mine Drainage during Transport

Top of this page Factors directly involved in rate of sulphide oxidation

A large number of factors control the rate of sulphide oxidation. For the purpose of this discussion, these factors are classified as chemical factors, physical factors, and biological factors. However, in reality these aspects form part of a very complicated and interwoven environmental system, and as such don’t operate independently but tend to be highly interrelated. Because of this interdependence, the chemical and physical factors are described and summarized simultaneously. While the description of chemical and physical factors also includes some references with regard to biological aspects, more detail on biological factors is provided under separate headings.

Top of this page Chemical and physical factors

The principal chemical and physical factors governing the rate of sulphide oxidation include the following:

  • Sulphide mineral:
    • Type of mineral
    • Surface area
    • Encapsulation
    • Crystallinity
    • Morphology
    • Assemblage
  • Ambient environment:
    • pH
    • Oxidation-reduction (redox) potential
    • Temperature
    • Source of water
  • Oxidant:
    • Type of oxidant
    • Oxygen
    • Ferric iron
    • Availability of oxidants

The type of sulphide mineral has a significant bearing on oxidation rates. As illustrated in Nordstrom and Alpers (1999) and Plumlee (1999), measured oxidation rates for sulphide minerals vary widely. Although laboratory studies usually agree with the general ranking of sulphides in terms of their reactivity, individual studies can vary considerably in detail when considering observed reaction rates. This disagreement may be due to differences in other aspects related to the sulphide mineral, including surface area, crystallinity, morphology, mineral composition, and sulphide assemblage, all of which can greatly influence resistance to oxidation. The description of these factors in the following paragraphs equally applies to non-sulphides, including acid-consuming minerals such as carbonates and silicates.

Grain size is a fundamental control on the reactivity of waste rock (Smith and Beckie, 2003). Because of the greater surface area available for interaction with oxidizing agents, fine-grained sulphides tend to oxidize more rapidly than their coarser equivalents. A larger surface area can also result from rapid growth of sulphide minerals under highly supersaturated conditions or bacterial activity, which may lead to formation of, for instance, framboidal pyrite (i.e., “raspberry-textured”: agglomerations of many microscopic spherules). Framboidal pyrite is most commonly found in deposits that formed in sedimentary environments such as coal, or colloform (intergrown radiating fibres) textures. In contrast, slower growth in a less supersaturated environment results in euhedral crystals with a smaller surface area, such as the typical cubic form of pyrite. Rapid crystal growth also promotes formation of amorphous rather than crystalline mineral phases. The reactivity of amorphous phases tends to be elevated relative to their crystalline counterparts because of their higher surface energy. The “reactive surface area” (i.e., the portion of the surface area available for chemical reaction) can be considerably less than the surface area measured using standard techniques because of intergranular contacts or inclusion within other minerals. An additional complication in mine and process facilities is that not all exposed surface area is in the flow path of water, thereby further lowering the reactive surface area. The generalized effect of particle size on system reactivity is often counterbalanced by the effect that smaller particle size has on decreasing permeability of the system to oxygen and water.

The composition of the sulphides can also affect oxidation rates. The presence of impurities and defects causes strain in the crystal structure, which in turn diminishes the resistance of the sulphide mineral to oxidation. Sulphides forming part of an assemblage consisting of different sulphide mineral phases can weather preferentially because of galvanic reactions caused by differences in standard electrode potentials of sulphides. This galvanic effect is similar in concept to the galvanic protection applied in many industries that rely on longevity of metals, where sacrificial metals are used to prevent oxidation of the metal of interest (e.g., zinc blocs on steel surfaces). Minerals with lower standard electrode potentials include pyrrhotite, sphalerite, and galena, and these minerals will oxidize preferentially when in electrochemical contact with minerals such as argentite and pyrite, which have higher potentials.

Temperature and the pH of the ambient environment are important controls on the rate of sulphide oxidation. Temperature effects can be described in terms of the Arrhenius equation, which relates chemical reactivity to temperature and activation energy. As a general rule, reaction rates approximately double for every 10°C increase in temperature. However, this generalization needs to be applied with caution, because reaction rates depend on the actual reaction paths (simple vs. complex, homogeneous vs. heterogeneous), reaction order (1st, 2nd, 3rd), and the mechanisms that govern movement of reactants and reaction products (e.g., diffusion). Figure 2-12 (Robertson and Broughton, 1992) presents a schematic illustration of the effect of temperature on abiotic sulphide oxidation rates (normalized) and provides a comparison against the temperature effect on bacterially mediated oxidation.

Figure 2-12: Schematic Illustration of the Effect of Temperature on Normalized
Sulphide Oxidation Rates (after Robertson and Broughton, 1992)


The hydrogen activity (and therefore the pH) features prominently in sulphide oxidation reactions. According to Le Chatelier’s principle, for a system at equilibrium, when a condition is changed, the equilibrium will shift in the direction that will reduce, in part, the applied change. For the overall pyrite oxidation reaction [5], this would indicate that pyrite oxidation will decelerate as conditions become more acidic, yet there is little evidence for this occurring in mining materials. One reason for this apparent discrepancy is that the overall reaction as represented by equation [5] does not adequately describe the actual reaction steps and mechanism, which implies that the pH-dependency of the reaction may not be as straightforward as suggested. In addition, at many mine waste facilities, oxidation products are periodically removed by flushing, which allows the oxidation reaction to proceed. The most important effect of pH, however, is that acidic conditions promote the activity of bacteria that can catalyze the various reactions associated with pyrite oxidation. For instance, A. ferrooxidans cannot use ferrous iron for metabolic purposes above a pH of approximately 3.5. However, it should be noted that the microenvironment at the reaction site may be very different from what is measured in the bulk aqueous environment because the pH at the pyrite surface, where the bacteria may be active, could be much lower. It is likely that one of the critical catalytic effects that bacteria may have relates to the ability of bacteria to actively maintain a microenviroment at the reaction site that suits their specific metabolic requirements. Measurements of pH, temperature, or redox state in the bulk liquid may therefore not be a reliable indicator of environmental conditions at the reaction site, and care must be taken in reaching conclusions based on these bulk liquid measurements.

The effect of the type of oxidant on the rate of sulphide oxidation is described in Section 2.4.4. In summary, ferric iron accelerates the oxidation of pyrite by approximately 2 to 3 orders of magnitude relative to oxygen. The availability of the oxidant can also have a pronounced effect on pyrite oxidation rates, and, in the case of oxygen, forms the basis for a number of mitigation alternatives.

Ritchie (2003) has demonstrated that the transport of oxygen is the limiting factor in sulphide oxidation rates within mine waste facilities. The principal mechanisms contributing to airflow and oxygen transport include the following:

  • Diffusion
  • Convection due to a thermal gradient
  • Advection due to a wind gradient
  • Barometric pumping

Diffusion is usually limited to a near-surface zone of a few meters depth, while convection and barometric pumping have the capability to move air and oxygen to much greater depths. Diffusion and thermally induced convection are largely governed by the reactivity of the mine or process waste and its air permeability, whereas barometric pumping and advection represent external factors whose effect is primarily controlled by only porosity.

Oxygen diffusion is caused by a concentration gradient resulting from oxygen depletion due to sulphide oxidation. Diffusion is described by Fick’s Law, which includes a diffusion coefficient that is specific to the properties of the diffusion medium. The degree of saturation within the pore space has a dramatic effect on the value for the oxygen diffusion coefficient, which decreases over approximately 5 orders of magnitude from air to fully saturated conditions. This reduction in diffusivity explains the benefits of maintaining near fully saturated conditions to control sulphide oxidation. An example of a reduction in diffusivity is the use of an overlying water cover or through maintaining saturation within the pore space. Although oxygenated rainwater or pore water represent a potential source of oxygen for sulphide oxidation, this supply is very limited and it cannot be replenished with sufficient efficiency so that it results in the generation of ARD, NMD, or SD. There is increasing evidence that the oxygen present in the water in reactions [1] and [2] partakes in the sulphide oxidation reaction, but this participation is likely to be of negligible importance relative to the control of atmospheric or even dissolved oxygen on the sulphide reaction rate. The degree of oxygen depletion through sulphide oxidation depends on the reactivity of the sulphide minerals. Therefore, oxygen diffusion into waste rock piles tends to be less efficient compared to process tailings because the former are typically less reactive. Conversely, the oxidation front within tailings frequently stalls at a relatively shallow depth (10s of centimetres to several meters), in particular in tailings with significant sulphide content. This is because oxygen is consumed by the reactive sulphides before it can penetrate to greater depths. At some point in time, a steady state may be reached, governed by the rate of oxygen consumption and by the rate of oxygen transport and replenishment. Waste products from coal mining also have a strong oxygen consuming capacity related to the abundant presence of organic material that further depletes oxygen available for sulphide oxidation.

In addition to diffusion of oxygen through the pore space, a second diffusion step is occasionally invoked according to the shrinking core model from Davis and Ritchie (1986). This step involves the diffusion of the oxygen into the unoxidized core of a sulphide particle through an oxidized shell. The shrinking core model has been used successfully to simulate the observed advancement of an oxidation front in mine processing tailings. Figure 2-13 (Wunderly et al., 1996) illustrates this two-stage process for pyrite oxidation in a tailings impoundment, where oxygen diffuses through the bulk pore space, partitions into a water film, and then diffuses into pyrite grains through the oxidized coating formed around the unoxidized core as governed by the oxygen concentration gradient between the particle surface and the unreacted core.

Figure 2-13: Two-Stage Process for Pyrite Oxidation in a Tailings Impoundment (Wunderly et al., 1996)

Pyrite oxidation is a strongly exothermic reaction, and the release of heat drives temperature up within waste rock piles, typically up to approximately 70°C. In colder climates or in the winter, this increased temperature can lead to formation of “fumaroles,” which may appear as small-scale equivalents of volcanic vents on waste rock facilities and provide evidence that convection indeed can be an important gas transport mechanism. Although the reason for this temperature ceiling is not fully understood, a reasonable hypothesis seems to be that at higher temperatures some of the microbes responsible for the pyrite oxidation perish, after which temperatures decline until the microbial community can be re-established and the process is repeated. The increase in temperature can modify the mechanism for oxygen transfer because of the creation of thermally and density-driven convective air flow. The resulting air movement draws atmospheric oxygen into the waste rock pile much more efficiently than diffusion, and convection is considered a significant oxygen-supply mechanism, at least for the outer portions of waste rock facilities. When these strongly exothermic reactions occur in coal waste deposits, spontaneous combustion may occur.

Advective air gradients and gas transport can be generated because of wind. Because the time scale for wind velocities is much reduced relative to convective air movement due to temperature gradients, this process is probably of less significance.

Barometric pumping results from changes in atmospheric pressure. The process involved is the compression of the gas phase within a waste rock pile because of an increase in atmospheric pressure, which then allows ingress of atmospheric oxygen. Oxygen egress will occur when the atmospheric pressure declines. The net effect of barometric pumping on oxygen supply and ARD generation is only partially understood and further investigation is required, but it is generally not considered a dominant mechanism.

Based on the considerations presented above regarding availability of oxygen in mine and process waste facilities, Ritchie (1994) coined the concepts of “global oxidation rate” and “intrinsic oxidation rate (IOR),” which describe the overall flux rate of ARD from a waste rock pile and the observed oxygen consumption rate of the waste, respectively. The intrinsic oxidation rate is a measure of the oxidation rate at the mineral surface controlled largely by the reactivity of the sulphide minerals. Ritchie found global oxidation rates to be insensitive to changes in the intrinsic oxidation rate, which suggested that oxygen diffusion is the dominant rate-limiting process for sulphide oxidation, particularly in a newly built pile. With time, convective gas transport will gain access further into the waste rock facility as it ages, but even so, the global oxidation rate will remain approximately constant.

The availability and effectiveness of ferric iron as an oxidant is controlled by the amount of ferric iron present, which in turn is related to the bacterial activity, the pH of the solution, and the residence time of the ferric iron. Acidic conditions and the presence of an active microbial community promote oxidation of ferrous iron to ferric iron. Physical removal mechanisms, such as flushing of contact water, also have an effect on the availability of ferric iron as an oxidant.

Top of this page Biological factors

Certain bacteria may accelerate the rate at which some of the reactions involved with sulphide oxidation proceed. Bacteria of the Acidithiobacillus species (formerly referred to as Thiobacillus) are of particular importance with regard to sulphide oxidation. This is because A. ferrooxidans is capable of catalyzing both the oxidation of sulphur and ferrous iron (reactions [1], [2], and [3]), while A. thiooxidans can oxidize sulphur only (e.g., reaction [1]). Other members of the Acidithiobacillus species are also capable of catalyzing pyrite oxidation, as are certain members of the genera Sulfolobus and Leptospirillium (Gould and Kapoor, 2003; Mills, 1999). In situations where bacterial acceleration of sulphide oxidation is significant (principally at low pH – see Figure 2-9), the bacterial population density and rate of population growth determine the bacterial activity and the associated rate of acid generation. Population density and growth for bacteria such as Acidithiobacillus are functions of the following:

  • Carbon availability (in the form of carbon dioxide)
  • An electron donor (ferrous iron or sulphur)
  • Nutrient availability (i.e., nitrogen, phosphorus for production of biomass)
  • Oxygen (promotes growth of aerobic bacteria and is an electron acceptor; kills strictly anaerobic bacteria)
  • Temperature (most bacteria demonstrate optimal growth below approximately 70°C)

A. ferrooxidans can be characterized as an aerobic autotrophic bacterium (i.e., it requires oxygen and must reduce atmospheric carbon dioxide (CO2) to an organic carbon form to generate biomass). A. ferrooxidans has a temperature optimum near 35°C and a maximum temperature for growth of 40°C (Gould and Kapoor, 2003). A. ferrooxidans is an obligate acidophile: it requires acidic conditions (pH range of 1.0 to 3.5 with an optimum pH near 2.0) to survive. A. thiooxidans demonstrates similar characteristics, but is tolerant of a wider range of acidic conditions (pH between 0.5 and 4.0) while its optimum growth temperature is between 25 and 30°C. Figure 2-12 (Robertson and Broughton, 1992) presents a schematic illustration of the temperature effect on bacterially mediated sulphide oxidation rates (normalized) and provides a comparison against the effect of temperature on abiotic oxidation. The principal nutritional requirements for Acidithiobacilli (nitrogen, carbon dioxide) are ubiquitous. Sulphur and iron are readily available in mining environments, while only small amounts of phosphorous are required. As a result, sulphur and iron are virtually omnipresent at mining sites, and have been identified in mine effluents of different compositions from quite different mines and in different climatological environments. This implies that microbial mediation of sulphide oxidation is the norm rather than the exception. It must be emphasized that microorganisms are highly efficient at manipulating their immediate environment, either on their own or in a symbiotic relationship with other microorganisms, and that actual reaction site environmental conditions may, therefore, be much more conducive to elevated oxidation rates than would be predicted from measurements in the bulk liquid phase.

Top of this page Factors that modify drainage resulting from sulphide oxidation

This section presents the factors that modify the composition of drainage resulting from sulphide oxidation both within the mine waste and after exiting the mine facility. These factors are described simultaneously because the processes that affect the drainage composition during transport within and outside of the mine or process facility are very similar in concept. Transport of dissolved and particulate constituents takes place along pathways, which are physical or biological conduits that allow movement of these constituents.

As for the sulphide oxidation process, the factors that affect mine drainage composition are classified as chemical factors, physical factors, and biological factors. These aspects form part of a very complicated and interwoven environmental system, and as such don’t operate independently but instead tend to be highly interrelated. For the purpose of this discussion, the drainage generated from sulphide oxidation at the grain surface can be ARD, NMD, or SD, depending on the type of sulphide mineral and the reactions taking place in the microenvironment immediately at the grain/water interface.

The transport of ARD, NMD, or SD through and away from the mine or processing facility can take many forms. The pathways of main interest include runoff and overland flow with eventual discharge into surface water and transport via surface water, infiltration through the mine or processing waste facility and into the soil/vadose zone followed by transport in groundwater, uptake by biota, and physical movement of mine waters as part of mine water management (Figure 2-2). This section focuses on the infiltration, surface water, and groundwater pathways.

Top of this page Chemical factors

The principal chemical factors that can modify the composition of drainage during transport in the mine waste facility and beyond include the following:

  • pH
  • Redox conditions
  • Chemical composition of drainage
  • Secondary mineral formation
  • Sorption
  • Neutralization reactions
  • Photochemistry

The most important chemical control on metal mobility is pH, but the other factors listed above also influence drainage composition. Redox conditions determine the speciation of redox-sensitive species, such as many base and trace metals and sulphur and nitrogen species. Redox conditions can have a pronounced effect on the mobility and toxicity of redox sensitive species. The mobility of constituents of interest may be affected by other species present. For example, complexation with anionic and organic ligands may increase the mobility of base and certain trace metals. A good example of this is the formation of Al-F complexes, which greatly enhances the mobility of aluminum, resulting in much higher dissolved aluminum concentrations than would occur in the absence of fluoride. Similarly, many dissolved organic compounds can promote metal mobility.

Attenuation mechanisms are processes that reduce the mobility and concentration of dissolved constituents in water. The most important attenuation mechanisms include formation of secondary minerals and sorption reactions. A typical attenuation sequence for ARD in surface water is described in Plumlee (1999). As an acidic mine discharge enters a stream, it is progressively diluted, which causes an increase in pH. This leads to precipitation of the typical orange ferric (hydr)oxide colloids and coatings commonly observed in acidic mine effluents. As the pH of the water continues to rise, aluminum and manganese precipitates form, while sorption onto the suspended Fe, Mn, and Al particulates becomes more effective. Other sorbents of potential interest are clay minerals and particulate organic matter. If the pH continues to increase, formation of secondary carbonate minerals may occur, for example, the copper carbonates malachite and azurite frequently encountered. Sulphate removal in such a sequence may occur through formation of Fe/Al hydroxysulphate minerals at low pH or, more commonly, in the form of gypsum if sufficient calcium is present. Most of these processes are reversible. Should conditions change, remobilization of these attenuated trace metals into the water column may occur.

In addition to pH changes because of the mixing of different water types, chemical interaction between water and neutralizing minerals in the solid matrix (waste rock, process tailings, stream sediment, aquifer solids) may also result in a pH increase. The neutralization mechanisms potentially occurring can be described by such reactions as presented in equations [8] and [9]. Photochemical reactions are of importance with respect to reduction of iron, which may lead to iron rerelease as well as the concomitant release of sorbed metals. Where ARD is contacted with organically enriched water, sulphate reduction processes may occur spontaneously, resulting in depletion of sulphate and an increase in alkalinity and precipitation of trace metals in the form of sparingly soluble sulphides.

ARD, NMD, and SD do not represent consecutive stages in the evolution of a mine water, but instead reflect endpoints in terms of water quality that may have different effects on the environment and may necessitate different forms of management. However, in concept, the above progression presented for the hypothetical surface water describes the potential evolution of ARD to NMD or ARD to SD. In reality, in almost all mine water occurrences this sequence is not brought to completion. If the receiving environment (including the mine or process waste facility) demonstrates a lack of neutralization potential or dilution (or both neutralization potential and dilution), the composition of the original ARD may not change appreciably with distance from the original reactive sulphide grain, and the ensuing acidic contamination can have a large spatial extent. NMD will be the outcome if some of the trace metals are removed from solution, while, according to its definition, SD requires almost complete metal removal. NMD frequently occurs in mine waste environments and surface water and groundwater systems with sufficient readily available neutralization potential to counter the acid being generated, such as many skarns, marine sedimentary strata, and carbonate-hosted replacement deposits. Saline drainage is most commonly associated with deposits that contain little or no sulphide minerals, other than pyrite and sufficient buffering capacity. Examples of such deposits include certain coal deposits as well as “clean” skarns.

Top of this page Physical factors

The principal physical factors that can modify the composition of drainage during transport in the mine waste facility and beyond include the following:

  • Climate conditions
  • Precipitation events
  • Water movement
  • Temperature

While chemical factors tend to be the most important controls on sulphide oxidation rates and the nature of the resulting discharge (ARD, NMD, or SD), physical factors tend to govern transport of the reaction products and the type and effectiveness of reactions that occur along the flow path.

Climate types (e.g., wet, arid, arctic) and climatic variations (e.g., storm events, seasonal) can have a significant effect on the hydrologic regime within a mine waste and in the receiving environment. Climate types and variations also affect transport of sulphide oxidation products and any changes in drainage composition while being transported. The more water available for transport, the more likely it is that a mine discharge will be generated, and the more likely that this discharge may travel off site and into the receiving environment. Climatic variations may result in transport that is in essence continuous vs. episodic. Therefore, the water balance associated with each climate type and climate variation will be very different, and each will pose its own challenges for water management. The effect of climate type and climate variation on chemical modification of mine drainage is also an important consideration. For example, in drier climates, high evaporation rates tend to increase the acid-buffering capacities of waters draining most rock types. Therefore, a smaller volume of alkaline water in a dry climate may mitigate ARD as effectively as a larger volume of less alkaline water in a wetter climate. Mixing, degree of dilution, and contact time between a mine discharge and the solid matrix (waste rock, process tailings, stream sediment, aquifer solids) are all affected by climate and seasonal events.

Water movement is also affected by the physical characteristics of the mine waste and receiving environment. If water is essentially stagnant, transport of sulphide oxidation products is limited to diffusion, which is a very slow process resulting in a reaction front that may not appreciably move over time. If the water is in motion, these reaction products can move with the velocity of the water in the absence of chemical attenuation mechanisms.

The most important factor involving water movement is the hydraulic conductivity. In principle, the greater the hydraulic conductivity of the source material, the greater the potential for effective transport of the sulphide oxidation products. Similarly, the greater the hydraulic conductivity of the receiving environment (e.g., a fractured vs. a non-fractured medium), the greater the potential for ingress of the mine discharge. Changes in drainage composition along a flow path also depend on the movement of water. Contact time is an important factor, and rapid flow through preferential coarser-grained channels in a waste rock pile will result in a different discharge quality (i.e., likely to be more dilute and less buffered) than gradual flow through the finer-grained matrix. Seepage from process tailings may display geochemical near-equilibrium with the tailings solids because of the extensive contact between tailings pore water and tailings solids before exiting the facility.

The effect of temperature is manifold. Higher temperatures promote evaporation, thereby reducing the amount of transport in liquid form, while also increasing the rates of chemical reactions. Lower temperatures may also prevent water movement (in the case of frozen conditions), while reducing reaction rates or even arresting sulphide oxidation altogether when the subzero environment prevents transport of reactants and reaction products. One of the most dramatic effects of temperature on transport of sulphide oxidation products is during the spring freshet in cold/polar, humid temperate/marine and humid cold/continental climates where snowfall occurs. During the spring melt, very large volumes of water can occur that contain large amounts of oxidation products accumulated during the summer and fall because of the reduced flushing rates. In addition to temperature fluctuation, as governed by external climate and seasonal conditions, sulphide oxidation can result in generation of heat, which may have an effect on transport and modification of mine discharges internal to the mine waste environment.

Top of this page Biological factors

The principal biological factors that can modify the composition of drainage during transport in the mine waste facility and beyond include the following:

  • Microbial ecology
  • Microbial growth kinetics

As explained in Section 2.6.4, biological factors greatly affect the rate of sulphide and iron oxidation because of the catalytic functioning of many different types of microbes. During transport of oxidation products, the importance of these bacteria tends to lessen while that of other types of microorganisms and biota tends to increase.

Bacterially mediated activity of potential importance along flow paths includes reduction of both iron and sulphate. Reduction of iron may lead to release of iron from previously insoluble minerals, such as iron (hydr)oxides, and increased mobility. Conversely, reduction of iron coupled with reduction of sulphate may also lead to formation of insoluble iron sulphide and other metal sulphides if other metal cations are available. Both reaction mechanisms result in release of alkalinity, providing neutralization capacity, and both operate with the greatest efficiency in anaerobic water-saturated environments such as submerged sediments and soils. In particular, the sulphate reduction process forms the basis for use of constructed wetlands and other sulphate-reducing bioreactors for passive and active treatment of ARD. In all of these cases, the type of bacteria (i.e., microbial ecology) and the growth kinetics determine the effectiveness of the various biologically mediated reactions. These biological factors are intrinsically linked to the chemical factors as the nature and type of microorganisms present are directly related to factors such as pH or redox.

Other biological processes that may affect drainage composition along a transport pathway include sorption and ion exchange onto particulate organic matter, direct uptake by plants through roots and leaves, formation of metal (hydr)oxides (aerobic conditions – results in a reduction in pH), formation of metal carbonates (anaerobic conditions – results in a reduction in pH), and filtration by organic substrate.

Top of this page Receiving environment

The final component of the ARD migration process is the receiving environment or receptors (Figure 2-2). The principal receiving environments associated with mine discharges water are groundwater resources, air, soil and sediments. More detail on receiving environment and ARD is presented in Chapter 8. Potential impacts associated with mine discharges are related primarily to sulphate/TDS (ARD, NMD, SD), metals (ARD, NMD) and acidity (ARD). In terms of chemical parameters, therefore, these impacts are not unique to the mining industry, and the toxicological consequences of any of these impacts on the ecological and human receptors identified in Figure 2-2 are usually understood to be the same as if they originated from a non-mining source.

The nature and extent of any impacts is related to the location of a receiving environment relative to the mine release, the degree of sensitivity to the mine release, and the nature of the mine release. The environment has a certain inherent capacity to absorb and sustain a level of ARD release without undergoing significant damage. For example, through natural neutralization, and dilution, it may be possible to contain an ARD plume of a certain magnitude without widespread environmental impact. However, as the reservoir of alkaline material is consumed close to the point of discharge, the ARD plume will migrate further away, resulting in an ever-increasing zone of impact. For a receiving environment with significant acid neutralizing capacity, the extent of the impact will be less than for an environment with little or no buffering ability, and it will take longer for an equivalent impact to manifest itself at any given location. The impact of an SD discharge will be less pronounced than that of an ARD discharge, all else being equal, because of the generally more benign chemical composition of SD. The identification and implementation of mitigation measures are aimed at preventing impact altogether, or reducing impacts to a level that can be sustained by the receiving environment.

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2.5 Concluding Statement

During the past two decades, much has been learned about ARD, NMD, and SD and how the reactions responsible for their development might be prevented. The complexity of these reactions and the unique issues associated with each mine and process site do not allow for application of a simple single strategy or a “one-size-fits-all” solution. Instead, each site needs to be examined on a case-by-case basis, leading to implementation of a site-specific ARD management plan from the initial decision to proceed with the mining operation through to post-mining beneficial land use. This plan must be a key focus of the operational, technical and managerial resources at the mine throughout its life cycle. The development of an ARD management plan is described in more detail in Chapter 9 of this GARD Guide.

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2.6 References

Agricola, 1556. De Re Metallica. Translated by H.C. Hoover and L.H. Hoover, 1950. Dover Publications, New York, NY.

Berner, R.A., 1984. Sedimentary pyrite formation: An update. Geochimica et Cosmochimica Acta, 48(4):605-615.

Blowes, D.W., Ptacek, C.J., and J. Jurjovec, 2003. Mill Tailings: Hydrogeology and Geochemistry. In: J.L. Jambor, D.W. Blowes and A.I.M. Ritchie (Eds.), Environmental Aspects of Mine Wastes, Short Course Series Vol. 31, Mineralogical Association of Canada.

British Columbia Acid Mine Drainage Task Force (BCAMDTF), 1989. Draft Acid Rock Drainage Technical Guide – Volume 1. Prepared by Steffen Robertson and Kirsten (SRK), Vancouver, BC.

Broughton, L.M., and A.M. Robertson, 1992. Acid Rock Drainage from Mines - Where We Are Now. IMM Minerals, Metals and Environment Conference, February 4-6, Manchester, UK.

Carrucio, F., Hossner, L.R., and G. Geidel, 1988. Pyritic materials: acid drainage, soil acidity, and liming. In: L.R. Hossner (Ed.), Reclamation of Surface Mined Lands, CRC Press, Boca Raton, FL, 1:159-190.

Casagrande, D.J., Finkelman, R.B., and F.T. Carrucio, 1989. The non-participation of organic sulfur in acid mine drainage generation. Environmental Geochemistry and Health, 7:187-192.

Davis, G.B., and A.I.M. Ritchie, 1986. A model of oxidation in pyritic mine wastes, 1. Equations and approximate solution. Applied Mathematical Modelling, 10:314-322.

Goldhaber, M.B., and I.R. Kaplan, 1982. Controls and consequences of sulfate reduction in recent marine sediments. Acid Sulfate Weathering, Soil Science Society of America Special Publication, 10:19-36.

Gould, W.D., and A. Kapoor, 2003. The Microbiology of Acid Mine Drainage. In: J.L. Jambor, D.W. Blowes and A.I.M. Ritchie (Eds.), Environmental Aspects of Mine Wastes, Short Course Series Vol. 31, Mineralogical Association of Canada, 203-226.

Jambor, J.L., 2003. Mine-Waste Mineralogy and Mineralogical Perspectives of Acid-Base Accounting. In: J.L. Jambor, D.W. Blowes and A.I.M. Ritchie (Eds.), Environmental Aspects of Mine Wastes, Short Course Series Vol. 31, Mineralogical Association of Canada, 117-146.

Kleinmann, R.L.P., and G.R. Watzlaf, 1988. Should the effluent limits for manganese be modified? In: Proceedings 1988 Mine Drainage and Surface Mine Reclamation Conference, U.S. Bureau of Mines IC 9184, Pittsburgh, PA, 2:305-310.

Mills, A.L., 1999. The Role of Bacteria in Environmental Geochemistry. In: G.S. Plumlee and M.J. Logsdon (Eds.), The Environmental Geochemistry of Mineral Deposits, Part A: Processes, Techniques and Health Issues, Reviews in Economic Geology Vol. 6A, Society of Economic Geologists, Inc., 125-132.

Nocete, F., Álex, E., Nieto, J.M., Sáez, R., and M.R. Bayona, 2005. An archaeological approach to regional environmental pollution in the south-western Iberian Peninsula related to Third millennium BC mining and metallurgy. Journal of Archaeological Science, 32:1566-1576.

Nordstrom, D.K., 2003. Effects of Microbiological and Geochemical Interactions in Mine Drainage. In: J.L. Jambor, D.W. Blowes and A.I.M. Ritchie (Eds.), Environmental Aspects of Mine Wastes, Short Course Series Vol. 31, Mineralogical Association of Canada, 227-238.

Nordstrom, D.K., and C.N. Alpers, 1999. Geochemistry of Acid Mine Waters. In: G.S. Plumlee and M.J. Logsdon (Eds.), The Environmental Geochemistry of Mineral Deposits, Part A: Processes, Techniques and Health Issues, Reviews in Economic Geology Vol. 6A, Society of Economic Geologists, Inc., 133-160.

Plumlee, G.S., 1999. The Environmental Geology of Mineral Deposits. In: G.S. Plumlee and M.J. Logsdon (Eds.), The Environmental Geochemistry of Mineral Deposits, Part A: Processes, Techniques and Health Issues, Reviews in Economic Geology Vol. 6A, Society of Economic Geologists, Inc., 71-116.

Plumlee, G.S., Smith, K.S., Montour, M.R., Ficklin, W.H., and E.L. Mosier, 1999. Geologic Controls on the Composition of Natural Waters and Mine Waters Draining Diverse Mineral-Deposit Types. In: L.H. Filipek and G.S. Plumlee (Eds.), The Environmental Geochemistry of Mineral Deposits, Part B: Case Studies and Research Topics, Reviews in Economic Geology Vol. 6B, Society of Economic Geologists, Inc., 373-432.

Raiswell, R., and R.A. Berner, 1986. Pyrite and organic matter in Phanerozoic normal marine shales. Geochimica et Cosmochimica Acta, 50:1967-1976.

Ritchie, A.I.M., 1994. The waste rock environment. In: J.L. Jambor and D.W. Blowes (Eds.), Environmental Geochemistry of Sulfide Mine-Wastes, Short Course Handbook Vol. 22, Mineralogical Association of Canada.

Ritchie, A.I.M., 2003. Oxidation and Gas Transport in Piles of Sulfidic Material. In: J.L. Jambor, D.W. Blowes and A.I.M. Ritchie (Eds.), Environmental Aspects of Mine Wastes, Short Course Series Vol. 31, Mineralogical Association of Canada, 73-94.

Robertson, A.M., and L.M. Broughton, 1992. Reliability of Acid Rock Drainage Testing. Workshop on U.S. EPA Specifications for Tests to Predict Acid Generation from Non-Coal Mining Wastes, July 30-31, Las Vegas, NV.

Rose, A.W. and C.A. Cravotta III, 1998. Geochemistry of coal mine drainage. In: K.B.C. Brady, M.W. Smith and J. Schueck (Eds.), Coal Mine Drainage Prediction and Prevention in Pennsylvania, Pennsylvania Department of Environmental Protection, Harrisburg, PA, 1:1-22.

Smith, K.S., 1999. Metal Sorption on Mineral Surfaces: An Overview with Examples Relating to Mineral Deposits. In: G.S. Plumlee and M.J. Logsdon (Eds.), The Environmental Geochemistry of Mineral Deposits, Part A: Processes, Techniques and Health Issues, Reviews in Economic Geology Vol. 6A, Society of Economic Geologists, Inc., 161-182.

Smith, L., and R. Beckie, 2003. Hydrologic and Geochemical Transport Processes in Mine Waste Rocks. In: J.L. Jambor, D.W. Blowes and A.I.M. Ritchie (Eds.), Environmental Aspects of Mine Wastes, Short Course Series Vol. 31, Mineralogical Association of Canada, 51-72.

Stumm, W., and J.J. Morgan, 1981. Aquatic Chemistry, First Edition. John Wiley & Sons, New York, NY.

Stumm, W., and J.J. Morgan, 1996. Aquatic Chemistry, Third Edition. John Wiley & Sons, New York, NY.

United States Geological Survey (USGS), 2004. Water Science Glossary of Terms.

Wolkersdorfer, C., 2008. Water Management at Abandoned Flooded Underground Mines – Fundamentals, Tracer Tests, Modelling, Water Treatment. Springer, Heidelberg, Germany.

Wunderly, M.D., Blowes, D.W., Frind, E.O., and C.J. Ptacek, 1996. Sulfide Mineral Oxidation and Subsequent Reactive Transport of Oxidation Products in Mine Tailings Impoundments: A Numerical Model. Water Resources Research, 32(10):3173-3187.

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List of Tables

Table 2-1: Common Sulphides Known or Inferred to GenerateAcid when Oxidized (Plumlee, 1999)
Table 2-2: Typical NP Values and pH Buffering Ranges for Some Common Minerals (Jambor, 2003; Blowes et al., 2003; BCAMDTF, 1989)

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List of Figures

Figure 2-1: Roman Portal with Acid Rock Drainage – Spain
Figure 2-2: Generalized Conceptual Model of Sources, Pathways and Receiving Environment at a Mine or Processing Site
Figure 2-3: Ficklin Diagram Showing ARD, NMD, and SD as a Function of Dissolved Base Metal Concentrations (adapted from Plumlee et al., 1999)
Figure 2-4: Diagram Showing ARD, NMD, and SD as a Function of Sulphate Concentrations
Figure 2-5: Ficklin Diagram Showing Selected Principles that Govern Mine Water Quality (adapted from Plumlee et al., 1999)
Figure 2-6: The Global Sulphur Cycle (Stumm and Morgan, 1996). Global Fluxes in Millions Tons of Sulphur per Year and Inventories in Millions Tons of Sulphur
Figure 2-7: The Biogeochemical Sulphur Cycle
Figure 2-8: Model for the Oxidation of Pyrite (Stumm and Morgan, 1981). (The numbers in brackets refer to the reactions presented in Section 2.6.4)
Figure 2-9: Schematic Illustration of Normalized Sulphide Oxidation Rates with and without Bacterial Mediation (after Robertson and Broughton, 1992)
Figure 2-10: Stages in the Formation of ARD (after Broughton and Robertson, 1992).
Figure 2-11: Schematic Illustration of Factors that Affect Sulphide Oxidation and Modify Mine Drainage during Transport
Figure 2-12: Schematic Illustration of the Effect of Temperature on Normalized Sulphide Oxidation Rates (after Robertson and Broughton, 1992)
Figure 2-13: Two-Stage Process for Pyrite Oxidation in a Tailings Impoundment (Wunderly et al., 1996)

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