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date: 20 July 2018

Mining, Ecological Engineering, and Metals Extraction for the 21st Century

Summary and Keywords

The first treatise on mining and extractive metallurgy, published by Georgius Agricola in 1556, was also the first to highlight the destructive environmental side effects of mining and metals extraction, namely dead fish and poisoned water. These effects, unfortunately, are still with us. Since 1556, mining methods, knowledge of metal extraction, and chemical and microbial processes leading to the environmental deterioration have grown tremendously. Man’s insatiable appetite for metals and energy has resulted in mines vastly larger than those envisioned in 1556, compounding the deterioration. The annual amount of mined ore and waste rock is estimated to be 20 billion tons, covering 1,000 km2. The industry also annually consumes 80 km3 of freshwater, which becomes contaminated.

Since metals are essential in modern society, cost-effective, sustainable remediation measures need to be developed. Engineered covers and dams enclose wastes and slow the weathering process, but, with time, become permeable. Neutralization of acid mine drainage produces metal-laden sludges that, in time, release the metals again. These measures are stopgaps at best, and are not sustainable. Focus should be on inhibiting or reducing the weathering rate, recycling, and curtailing water usage. The extraction of only the principal economic mineral or metal generally drives the economics, with scant attention being paid to other potential commodities contained in the deposit. Technology exists for recovering more valuable products and enhancing the project economics, resulting in a reduction of wastes and water consumption of up to 80% compared to “conventional processing.”

Implementation of such improvements requires a drastic change, a paradigm shift, in the way that the industry approaches metals extraction. Combining new extraction approaches, more efficient water usage, and ecological engineering methods to deal with wastes will increase the sustainability of the industry and reduce the pressure on water and land resources.

From an ecological perspective, waste rock and tailings need to be thought of as primitive ecosystems. These habitats are populated by heat-, acid- and saline-loving microbes (extremophiles). Ecological engineering utilizes geomicrobiological, physical, and chemical processes to change the mineral surface to encourage biofilm growth (the microbial growth form) within wastes by enhancing the growth of oxygen-consuming microbes. This reduces oxygen available for oxidation, leading to improved drainage quality. At the water–sediment interface, microbes assist in the neutralization of acid water (Acid Reduction Using Microbiology). To remove metals from the waste water column, indigenous biota are promoted (Biological Polishing) with inorganic particulate matter as flocculation agents. This ecological approach generates organic matter, which upon death settles with the adsorbed metals to the sediment. Once the metals reach the deeper, reducing zones of the sediments, microbial biomineralization processes convert the metals to relatively stable secondary minerals, forming biogenic ores for future generations.

The mining industry has developed and thrived in an age when resources, space, and water appeared limitless. With the widely accepted rise of the Anthropocene global land and water shortages, the mining industry must become more sustainable. Not only is a paradigm shift in thinking needed, but also the will to implement such a shift is required for the future of the industry.

Keywords: extreme environments, mine drainage, geomicrobiology, ecological principles and stressors, mining wastes, mining water use, ecological remediation processes, chloride metal extraction, ecological engineering, restoration, reclamation

The Origin of Mining Wastes

Mining and the extraction of metals have been a large part of human activity since prehistoric times, and the modern world has an almost unquenchable thirst for more and more metals, not only the traditional metals, such as iron, copper, and aluminum, but also the rare and so-called rare-earth metals, which are important for mobile phones and the like. However, recovering metals has not come without a cost. The environmental consequences of even ancient activity are still evident today, for example, from evidence in Spain of mining by the Iberians in 3000 bc (Davis et al., 2000), in the old mines of the Incas in South America (Strosnider et al., 2011), and more recently, in the vast “red mud” ponds from bauxite mining found scattered worldwide (Ritter, 2014), as well as in many coal, precious-metal, and base-metal mining areas. There is, perhaps, no greater and more poignant reminder of this latter environmental legacy than the fact that October 2016 was the 50th anniversary of the terrible Aberfan disaster in Wales, where an unstable coal mining tip engulfed a school, killing 116 children and 28 adults.

The practice of mining and metals extraction is that the mined broken rock is separated into rocks that do not contain a sufficient content of metal to be economically extracted, referred to as waste rock, and rocks that contain ore. These latter rocks are ground, the ore is extracted, and the remaining ground rock disposed of as tailings. The wastes generate effluent known as mine or rock drainage, which is either alkaline or acidic. Since ancient times, mining methods and practices have, of course, changed dramatically, currently being much larger scale and much more mechanized, and with commensurately greater environmental issues. Ore was once high-graded (collected in nearly pure mineral form) or mined in underground tunnels, collected in glory holes, and hauled by rail to the mill. A glory hole with its haulage tunnel is relatively small in comparison to an open pit, where the overburden (soil and rock without sufficient economic metal content) is removed to gain access to the ore-bearing strata. Today, most mines are either open pit, or have a network of underground tunnels, or both.

The first step before mining can begin is exploration, which is carried out today mainly by air, with sophisticated instrumentation, and covering large areas of the globe. Ground truthing, when the geology is promising, is carried out with borehole drilling and investigative trenching. Generally, these activities have low environmental impact. However, important environmental parameters could be collected in this early phase and later could be used to assess environmental consequences should a mine be developed. For example, water quality of boreholes and drainage characteristics of trenches could be used to plan for the future mine’s waste rock pile and tailings basin siting. This potential is rarely utilized.

The second step is mining. It is the costliest of all activities, and one with a long-lasting environmental impact. It disturbs the hydrological conditions of the underground, contaminates the groundwater emerging from underground workings, and destroys surface landscapes with waste rock and tailings deposits.

The third step is mineral processing. This consists of first crushing the rocks to an even size and then grinding them in a ball mill. Once the desired sandlike size of the rocks is reached, physical separation methods, such as flotation, gravity separation, tabling, dense media, etc., are applied to the sand-water slurry. These methods result in a mineral concentrate that is further processed in the fourth step. The ground rock remaining after extraction of the target economic mineral is regarded as being of little or no economic value and is discharged, generally as a 30% sand-water slurry, as tailings. The vast quantities of tailings and waste rock removed from the ground expose a very large surface area for weathering (oxidation) and hence represent long-term environmental liabilities, as evidenced by legacy sites worldwide.

The fourth step is collectively referred to as “extractive metallurgical processes.” These processes differ from element to element, but take two forms, pyrometallurgy (smelting) and hydrometallurgy (leaching). The former has serious occupational health and safety issues in terms of gaseous emissions (air pollution) and slag (material left over after smelting). Slag is accumulated on land, but generally has not been considered a serious long-term environmental issue. Slags from smelting are extensively reused as building materials (Piatak, Parsons, & Seal, 2015). Most of the hydrometallurgical processes use chemical agents in relatively small quantities. The chemicals do not have nearly the long-term environmental consequences that tailings and waste rock piles have. In hydrometallurgical processes, only accidents during extraction are of concern. This is particularly true for the extraction of gold, which uses highly toxic cyanide. Although spills of tailings or process liquor are generally disastrous, the toxicity of cyanide is short-lived.

Dimensions of Global Mining Waste Generation and Water Use

Mining and mineral processing are vital activities in an industrialized world, but their activities are mostly conducted in locations relatively remote from urban society, thereby attracting little attention except when, on relatively rare occasions, a major incident, such as a tailings dam failure or a rock failure, attracts wide media attention. The industry has developed procedures to minimize the risk of such incidents. A less spectacular but increasingly important aspect of mining is the need for land to store wastes, and the need for water to transport and process ores, especially when these needs compete with a limited area of arable land and/or water supplies for irrigation. This competition is becoming more intense as the world’s population increases, requiring more mineral resources and water, and requiring with more agricultural production from a finite arable land area.

This section is a first attempt to put together a set of global statistics to quantify, at least approximately, the dimensions of this competition and to identify some of the ways that mining, and agriculture might cooperate to the mutual benefit of society in general. The quantification of these dimensions is complex, but approximations are adequate to define the global challenge and to put mining activities into a global perspective. Worldwide estimates are offered, based on the scant data available in the literature. Where no literature has been found, estimates are made using the experience gained by the author over a lifetime in key positions in mining companies worldwide. What’s important here is not an exact figure, but an educated estimate of the order of magnitude of tonnages of waste rock and tailings produced with the concurrent use of freshwater. The global land use is presented as waste generation in units that are relatively easily to comprehend.

There are many organizations worldwide that collect statistics on mining wastes, including the UN Statistics Division and the U.S. Geological Survey, which provide global production figures for minerals and metals. Additionally, mining companies often provide information on solid waste production and water use in annual environmental or sustainable development reports. The data are framed within the global mining context covering a 25-year period, as this is a typical mine lifespan. Emphasis is placed on water resources supply and use by mining operations that appear to create competition, as they represent a close link to land use for agriculture. Abandoned or orphaned mine sites and their wastes, mankind’s shared global historic inheritance, are not considered in these estimates.

Population growth, increased prosperity, resource demand, and resource competition are forcing the mining industry to rethink the future of mineral extraction. These factors are inextricably linked, with consequences for people and the future. Which should be emphasized—irrigation for agriculture, or mine water and waste (Bebbington & Williams, 2008)?

The population of the globe is predicted to increase from the current 7.4 billion to 9.2 billion by 2040—an increase of about 20%. This will increase the demand for food, water, and raw materials of all kinds, including minerals and metals. Global freshwater consumption per capita increases in proportion to the increase in per capita income (UN Water). The global average per capita income has been increasing exponentially since the start of the Industrial Revolution and is expected to increase by 33% in the next 25 years (The Maddison-Project, 2013 version). This will result in a 60% increase in global water consumption. This estimate discounts major natural or human-generated disasters.

Water is vital in many mining operations, as it is used for dust control, drilling, transportation of solids, furnace cooling, and quenching slag and off gases (H2S emissions from the tower of the refinery) as well as in refining operations (Mudd, 2008). Efforts are underway to reduce freshwater use in mining, but it is not yet common practice (Bruce & Seaman, 2014).

With 80 km3/y (Table 1), the industry’s global water consumption is relatively small compared to other industrial sectors (Table 2). However, this consumption leads inevitably to contamination of groundwater and surface water. Increased surface area of waste rock and tailings exposed to weathering releases not only soluble elements, but also large quantities of suspended solids. Furthermore, during mine development, dewatering is often needed to access the ore to be mined, which in turn may deplete freshwater aquifers. The current global use of water of all major industries is estimated to be around 4,500 km3 per year (Table 2), or about 10% of the net precipitation (rain and snow) falling on land.

Table 1. First Estimate of Freshwater Consumption by the Global Mineral Industry (Gross Site Input, Including Recycle and Excluding Dewatering)

Mineral

Ore (106)

Process Water (m3/t)

Water Use (km3/y = 109t/y)

Coal

7,000

2

14

Ferro-alloys

1,000

1

1

Iron ore

2,600

5

13

Gold

1,700

5

8.5

Copper

1,100

5

5.5

Oil sands

1,000

10

10

Other

5,700

5

28.5

Total

20,100

80.2

Table 2. First Estimate of Global Freshwater Consumption

Industry

Product

Water Use

Per Person

km3/y (109t/y)

(m3/day)

Irrigation

Food

1,500

0.6

Forest

Paper

10

0.004

Power

Heat sink

70

0.03

Desalination

Freshwater

(−28)

(−0.01)

Municipal

Services

500

0.2

Mining

Metals

80

0.03

Manufacturing

Various

2,340

0.9

Global Use

4,500

1.76

United States

2,000

5.5

Currently, the water supplied by atmospheric precipitation is supplemented by draw-down of aquifers. Depletion of aquifers will increase competition for surface water supplies and lead to increased reliance on desalination plants in arid locations. These estimates highlight the fact that mining and milling might well be on a collision course with civilization’s water requirements, given the well-documented global water scarcity. Concomitantly, the dollar value of water will increase and the degradation of resources will gain importance. In Table 3, the global mined tonnages of minerals and their associated wastes are estimated. The global area covered annually with mine wastes can be estimated to be on the order of 1,000 km2, assuming an average loading of 20 tons per square meter.

Table 3. Major Land Disturbances due to Mining

Material

Ore (106 t/y)

Waste Rock/Tailings (106 t/y)

Coal

Ferro-alloys(1)

7,000

7,000

Ferro-alloys

1,000

1,000

Iron ore

2,600

1,000

Gold

1,700

3,000

Base metals

1,200

2,100

Oil sands

1,000

1,000

Industrial-agricultural minerals

5,700

5,700

Total World

20,100

20,800

Note: * Ferro-alloy ores include nickel laterites and beach sands.

Global satellite imagery might provide a more substantive and definitive estimate. The estimated annual total land area committed to mining is small compared to the total land area of the Earth, which is one hundred and forty-nine million square kilometers. However, it is large enough to be a serious local issue and becomes even larger if extrapolated over a century or more. A time trend of the available arable land area with the growing world population is presented in Table 4.

Table 4. Competition for arable land

Year

World Population (billions)

Arable Land (106 × km2)

hectares – billions (million km2)

Arable Land per Capita (hectares per capita)

1950

2.5

13

5200

1975

4.1

14

3400

2000

6.1

15

2500

2025

8.0

15

1900

2050

9.2

15

1600

Source: Fast facts: The state of the world’s land and water resources. United Nations Food and Agriculture Organization.

This emphasizes that arable land is being lost at an unsustainable pace. For example, between 1950 (0.52 ha/capita) and 2050, it is estimated that 0.36 ha per person will be lost. It can be expected that conflicts between agriculture and mines and their wastes will increase (Hilson, 2002). Already there are localized conflicts, as some groups like Mining Watch document the conflicts between Canadian mining companies and local landscapes. Mine wastes not only consume land but also create dust storms and silt streams, and contaminate surface water and/or groundwater. Failure of tailings dams is the cause of many disasters, as the long-term stability of dams is an acknowledged engineering challenge.

The rising global population is placing increased pressure on the finite area of arable land for food production and will increase the demand for irrigation (Table 4). The area of arable land is also shrinking as an increasing population requires more land for infrastructure. Some forest lands could be converted to arable land, but this would bring about a loss of carbon dioxide sinks, a loss of water-holding capacity, loss of wildlife habitat, and increased erosion, producing desertification and loss of livelihood for aboriginal peoples. This alternative is not generally regarded as desirable or viable. Some grassland may be suitable for crops, but usually only with irrigation.

Many mines exist in the tropical or subtropical deserts that cover a total area of 15.3 million km2. The combined area of tropical and subtropical deserts is the same as the global area of arable land, and with an adequate water supply, could presumably be equally productive (“Desert Farming,” from Wikipedia). In these arid environments, water supply and effluent discharge are major issues that must be balanced against lucrative ore bodies that are mined in the same areas. During mine operation, the local community can benefit from sharing a supply of freshwater, as the mining industry is gradually adopting desalination to guarantee a supply of freshwater for ore processing.

The total cost for a large desalination plant (20,000 m3/day), with power costs at 10 cents per kilowatt-hour, can be approximated as 1$/m3 (Table 5). This cost applies to the plant capital and operating costs only, and does not include the cost of delivering seawater, returning brine, and delivering desalinated water to the point of use. It should be noted that pipeline capital and operating costs to deliver freshwater to a mine at a high elevation can triple the final delivered cost (Soruco & Philippe, 2012).

Table 5. Seawater Desalinization and Energy Trends

Year

Cost

Energy

(US$/m3)

(Kwh/m3)

1982

1.5

8.1

1992

1.1

5.3

2002

0.7

4.5

2012

0.6

3.8

Base cost

0.82

2.2−3.0

Note: *The cost of desalinating sea water has been falling and desalination plants are being constructed by mining operations to avoid competition with scarce local water supplies.

Source: Soruco and Philippe (2012).

Mine Waste Management: A Brief History

When the ore body is mined out, underground workings are often force flooded, or fill gradually with rain and groundwater, or remain open voids. Pits are transformed to pit lakes when located within the groundwater table, which must be lowered during operation to keep the pit dry.

Pit lakes are not like regular lakes, because they lack shorelines with semi-aquatic vegetation, a drainage basin, and a microbially active sediment. All these components are essential for a functioning limnological ecosystem. Mine management practices have evolved with the rise of environmental awareness. In Table 6 the changes are summarized, starting in the early 1980s in Canada, mainly with the uranium industry. These measures are not necessarily implemented worldwide. The emphasis is clearly on containment and consolidation of the wastes.

Table 6. Comparison of Past and Present Mining Waste Management Site Selection and Design for Waste Rock and Tailings

Present

Past

Site selection for waste rock and tailings with hydrological and economic considerations

Economic considerations only (e.g., proximity to mine)

Ore stockpiles placement and exposure

Not considered

Run-off drainage systems isolated from contaminated flows

Sometimes considered

Progressive reclamation of site during operations

Not considered

General mine-closure plan considered

Not considered

Strict design criteria for storage facilities for chemicals and fuel

Sometimes considered

Segregation and stockpiling of rock types according to acid-generating potential

Unsegregated waste rock piles

Improved dam design, including liners and leak detection systems

Dams constructed from coarse tailings, overburden, or waste rock

Thickened tailings; Underwater tailings management facilities

Above-ground tailings management facilities, no thickening

Tailings cleaning—sulfide separation

Not available

Tailings: high-density paste backfill

Not available

Highly acid-generating material used as backfill

Conventional backfilling, using only coarse fraction of the tailings

Source: Kalin (2004).

Although initially the practice of backfilling high sulfide wastes was considered an environmentally desirable option, experience in some Canadian mines (e.g., Cyprus and Cominco) showed that high-sulfide waste used for backfill may catch fire and prevent continuation of the mining operation. Thus, it is normal practice for safety reasons to severely limit the sulfur content of mine backfill. Generally, though, the volume of broken rock and tailings exceed the volume of the cavities created by mining. Hence, it is not possible for all generated wastes to be accommodated in the mine voids from which they were extracted. The surplus must be stockpiled unless a use can be found for it as aggregate (sand or gravel), if the sulfur content is negligible. The overburden and waste rock from open pit operations must be stockpiled outside the pit during active pit operation and can comprise up to ten times the ore volume. Returning this waste material to the pit when it is mined out is generally prohibitively expensive. The challenge of isolating or otherwise finding beneficial uses for waste rocks and tailings from open pit operations remains.

There are companies that segregate sulfides in tailings for subaqueous storage, pending the day when they can be economically processed. It is also common practice to use and isolate waste rock and tailings as backfill in underground mining operations. Tailings-paste fill operations use the nonsulfide fraction of the tailings backfill after thickening, allowing immediate recycling of water. Sand-fill operations have the option of thickening the slime portion at the concentrate to permit immediate water recycling, producing a thickened product as a valuable impermeable cover for old tailings and waste rock deposits.

The uranium industry was one of the first to include environmental issues in their close-out plans, as public awareness raised these issues in the late 1960s. From these efforts, several principles were developed to govern management practices, such as ALARA (As Low as Reasonably Achievable) for radiation safety at uranium operations. It was later followed by BATEA (Best Available Technology Economically Achievable) for all other mining operations. Comprehensive historical reviews of risks and environmental policy have been written by Faber and Wagenhals (1988) and Kamieniecki and Kraft (2013). These efforts are commendable and have brought about significant change in the mining industry.

INAP, the International Network for Acid Prevention, has created a guideline, the GARD Guide (Global Acid Rock Drainage: GARD), which is an internationally recognized guide to the prediction, prevention, and management of drainage produced from sulfide mineral oxidation (Kleinmann & Chatwin, 2011). In accordance with the guide, most current mine operation practices emphasize containment of the wastes, thereby reducing the volume of effluent, not its quality. These containment practices require significant financial commitment from the operating mining company. But, while the management practices outlined in the GARD Guide certainly reduce the immediate environmental impacts, they may, in many ways, delay the onset of longer-term mine drainage issues. These entrenched practices are a hindrance to novel approaches to mine waste management and the acid challenge. Remediation efforts and drainage treatment are viewed by some as “the price to be paid and accepted as part of mining and metals.” To some degree, physical and chemical aspects of natural weathering processes are abated by the present practices, but the fundamental contribution of the microbial populations is ignored. Herein lies the long-term challenge. Only when the microbial oxidation is controlled will long-term weathering processes subside. Hence, current mine environmental management practices are, in a true sense, not sustainable.

Weathering Processes: Waste Rock Drainage or Tailings Seepage Generation

Minerals are the source of most of the elements present in all living organisms and are essential for growth and reproduction. Weathering, or the release of elements from rocks, occurs due to physicochemical forces (heat, wind, freezing, snow, rain, and erosion) and biogeochemical factors (vegetation exudates and microbial activity). These are primarily oxidative processes, driven by oxygen contained in the moist air, by water, and by microbes and they lead to the gradual breakdown of rocks and their minerals and supply elements to water and soil to support all life on the planet. Not all rocks display the same weatherability. The mineral composition of a rock and its weatherability are determined by a rock’s history or genesis over geological timescales (also known as the rock cycle). The weatherability of rock determines the buffering capacity and elemental composition of the surrounding ground- and surface water, which, in turn, with climate, define the characteristics of ecozones within ecosystems (e.g., arctic or tropical) around the globe. Together with the growth and decay of vegetation, these processes govern the characteristics of surface- and groundwater and soil formation.

Exudates of higher plants’ roots can alter the pH in the root zone and/or house microbes and fungi that assist in dissolving minerals in the soil to increase nutrient availability for plant growth (van Schöll et al., 2008). Lichens, fungi, and microbes grow attached to rock surfaces, exuding organic acids to liberate elements from minerals (Barker & Banfield, 1998; Uroz et al., 2009). These organisms control the biogeochemical cycles of elements and determine the distribution of elements in water, air, and soil. An extensive discussion on weathering is given by Drever (2005).

Mine wastes represent a very large amount of exposed rock surface, much larger than the land area they occupy. Hence, weathering of the rock surfaces releases disproportionally greater amounts of elements to ground- and surface water. Concentrations of the elements weathered from the rocks drastically increase the toxicity of receiving waters, to the detriment of aquatic life. Further, the oxidation of sulfides in waste rocks, such as pyrite (FeS2) and pyrrhotite (FeSn), produces a highly exothermic reaction that generates heat (Blowes et al., 2003). This can lead to steaming/burning waste rock piles (Kuenzer & Stracher, 2012; Rosenblum et al., 2015). In the high arctic, the heat generated prevents tailings from freezing completely. Microbes living in these tailings can adapt to the lower temperatures, suggesting that metabolic activity does not cease (Elberling, 2004).

Oxidation is primarily a geochemical reaction between mineral, oxygen, and water, supported by microbes (Nordstrom, 2011). Of interest is a group of microbes, the Archaea (chemolithotrophs), which derive energy from breaking mineral bonds, such as those in sulfidic minerals, increasing oxidation rates a thousandfold (Dave & Tipre, 2012). Weathering processes in mine waste management areas resemble natural extreme environments, dominated by sulfate and iron generated by volcanoes and their hot springs (King, 2003). Life is hypothesized to have originated in these iron-rich environments (Deamer & Weber, 2010). Mine sites are similar, and consequently control of microbes should be the primary issue in mine waste management. Since chemolithotrophic microbes are ubiquitous, they colonize any surface on the planet, and when they find the proper conditions, they flourish as evidenced by the sulfate oxidizing microbes in mine wastes.

Waste management approaches are generally planned by the mining operations, based on Acid-Base Accounting (ABA) tests. Predicting seepage or drainage water characteristics is difficult. Many test procedures have been developed over time to improve accuracy. A detailed review of ABA techniques is presented by Dold (2017), with emphasis on mineralogy.

Generally, ABA test work is based on ground or segregated rock. Segregation is based on particle size and mineralogy. This ensures a relatively homogenous distribution of all rock types occurring in the mine wastes, with both neutralizing and acid-generating minerals. Grinding or segregating rocks creates a relatively homogeneous sample and also a close proximity of the mineral surfaces. In a waste deposit, there is no homogeneity or proximity of neutralizing and acid generating minerals. Drainage from a waste rock pile generally emerges in an oxygenated form, whereas tailings seepage emerges highly reduced, with low Eh and circumneutral pH. Weathering products are generated mainly in the vadose zone of the tailings and the flow path of the waste rock pile, as mineral surfaces weather. Seepage or drainage is generated only at locations in the wastes through which atmospheric precipitation passes and emerges at foot of the piles. In the stockpiles, both waste rock and tailings, water can form perched water tables and develop distinct flow paths within the stockpiles. Within the paths, the contaminated water encounters different minerals, causing different precipitation reactions. Furthermore, neutralizing and acid-generating rocks do not release their minerals at the same rate, which would be required if they were to interact. In addition, as the seepage/drainage passes through the stockpiles, internal chemical precipitation occurs, leading to secondary minerals, some highly water soluble, and further altering the chemical composition of the emerging drainage or seepage. ABA procedures hardly account for all these interactions, which take place within the stockpiles. Hence, ABA test procedures cannot always serve as reliable predictors of the seepage or drainage characteristics.

An example of drainage characteristics from a waste rock pile that are predicted by conventional tests to generate neutral drainage is presented in Table 7. Water samples collected from seepages from a waste rock pile between 1992 and 1994 were filtered (0.45 µm) and analyzed. Large variations in seepage or drainage composition were found between different areas of the waste rock pile, suggesting that the rock mineralization was different, the microclimate was different, or the precipitation pathways were different. In any case, both sides of the waste rock pile produced acidic drainage where none was predicted, one (SW) with more extreme variations in pH as the other side (Table 7).

Table 7. Heterotrophs in the Milling Circuit to Tailings and AMD Drainage Ditches

Sample

Heterotrophs (bacteria/mL)

pH

Thickener underflow

0

11.5

Thickener overflow

0

12.5

Barrel Tailings─surface

< 1

n.d.

Barrel Tailings─subsurface

6

n.d.

Barrel Water

3

11.5

Tailings Beach

1,404,000

8.1

Tailings Seepage Ditch 1

40,500

2.1

Tailings Seepage Ditch 2

10,800

4.8

Old Pond

9

2.2

Source: Kalin et al. (1993).

Potentially, more reliable results could be obtained from assessments of mineralization exposed during the exploration stage, from weathered trenches and drill holes that have weathered, as often years elapse between exploration, mine development, and actual waste generation. Later, as operations start, large, outdoor test piles may provide a more realistic assessment of the long-term seepage/drainage characteristics.

Early Experiments: Discovering Extreme Environments

A summer student project at the Institute of Environmental Studies at the University of Toronto addressed the question: what, if any, indigenous species invade reclaimed or barren uranium tailings areas after they have been abandoned for 10 to 15 years? The answer: diverse flora of indigenous terrestrial and aquatic biota. During the initial study, such a flora were found on tailings and alongside completely barren areas. This raised questions about the colonization or invasion processes: were the differences chemical, physical, or ecological? The questions trigged the investigation of 15 uranium mill tailings sites in Ontario, Canada. A 4-year- study was funded by three government agencies, motivated not by the colonization or invasion processes, but by the potential food-chain contamination by long-lived radionuclides contained in the tailings. Ecologists had raised environmental concerns for some time and identified metal contamination in vegetation, especially by metal tolerant plants growing on precious- and base metal tailings. Soon, uranium tailings sites in the Northwest Territories and the Province of Saskatchewan were added to the investigation (Kalin, 1984). On completion of the 4-year study, it was evident that indigenous plants posed no threat to the food chain, as the radionuclides remained mainly in the roots.

At the time, the generally accepted agricultural restoration technique applied lime, fertilizer, and commercial grass seed to slow and reduce water penetration into the tailings and prevent wind erosion (Peters, 1988), but this did not prevent acidic seepages from the tailings. The accepted treatment of contaminated water was to add neutralizing materials, such as lime, to both wastes and their acid streams, which continue in perpetuity, identified as the major long-term environmental liability for the mining industry.

The Canadian government funded a 5-year program, the National Uranium Tailings Program (NUTP), in 1981, addressing the long-term environmental impact of uranium tailings. It was followed in 1983 by the Reactive Acid Tailings Sulphide Program (RATS), which focused on modeling, prediction, and methodologies to reduce or remediate the long-term environmental effects of acid-generating materials (John & Joe, 1987). With these programs in mind, and with the encouragement of the uranium industry, Boojum Research Ltd. was founded in 1982 as an R&D company. It had the mandate to find long-term, sustainable, economic solutions to the closure of mining operations. Closing mines, after the ore body is exhausted, and only waste and contaminated water remain, is called decommissioning (Kalin, 1989). Boojum’s goal has been to search for less costly, low maintenance, and sustainable decommissioning technologies.

Boojum Research’s objective under its first RATS project was to find a decommissioning scenario for pyrrhotite-covered tailings area that produced acid run-off from the oxidized surface when rain fell. When unoxidized pyrrhotite is exposed to moist air and rain, it starts to burn. The hard crust of an oxidized pyrrhotite-covered site could not be traversed by heavy equipment used to establish the conventional neutralized and seeded vegetation cover, hence dedicated as an experimental area for ecological engineering trials, a clear challenge.

When the site was visited for the first time, we found encouragement in our approach. Cattails were found in some locations, and a moss cover was found on other areas. A similar pattern was noted at other tailings sites. Running over the site was a small, alkaline stream that contained a high content of particulates from shotcrete used underground. Moss and horsetails colonized the banks of the stream and the shores of a pond covered with blue-green algae.

Another slow-running creek on the site with a pH of 2.5 was chosen as an experimental area (Figure 1a). Loose straw, a form of organic carbon, was added to curtained sections of the creek as microbial nutrient (Figure 1b). Later that winter, when the ice was drilled through, a strong hydrogen sulfide odor emerged from the hole and in spring, the water was clear (Figure 1c). We jumped with joy with both discoveries as the rotten egg smell indicated the iron had precipitated and the clear water meant that most of the iron was reduced.

Stumm and Morgan (1996; Figure 8.14, p. 477) provided an explanation for what had happened. Sediments contain microbial populations, but they can only grow when appropriate redox conditions exist. Each group of microbes alters the surrounding conditions to provide the proper redox conditions for the next group. This microbial reaction chain starts with aerobic respiration that reduces oxygen, and proceeds to sulfate reduction and thence to methane fermentation. As the Eh in the creek section with the straw gradually dropped, the pH increased. The emerging smell noted in the winter indicated that not enough oxidized iron was available to form biogenic pyrite, which led to the emergence of H2S, the rotten egg smell of bog gas. If the pH had been high enough and the Eh low enough, biogenic pyrite would have formed and the mineral siderite, an iron carbonate, could have formed on the straw (Fernández-Remolar et al., 2003; Reiter & Thiel, 2011).

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Figure 1. (a) Pyrrhotite tailings crust. In the foreground is the creek in which the first straw addition was made. (b) Straw addition in ARUM creek before winter at the time of setup. (c) Clear section of creek with straw in spring.

What function did the ice cover and organic carbon play? The ice cover on the creek slowed the flow to a trickle, while reducing wind-driven mixing and oxygen diffusion. Heterotrophic microbes growing on the straw removed oxygen through respiration, lowering the redox potential of the water. The combination of low oxygen and organic carbon fostered the growth of anaerobic, iron- and sulfate-reducing microbes. These experiments provided key observations for progress:

  • No microbes were needed to seed the acidic water.

  • Iron precipitate covered the straw, reducing access to the organic carbon.

  • Ice cover reduced oxygen access, giving anaerobic microbes a chance to flourish.

Iron reduction by microbes raises pH, and this in turn leads to the precipitation of some metals. To reproduce these conditions, two things needed to be developed. First, a living, floating vegetation cover would replace the ice cover. This would provide a continuous supply of organics, through decomposing litter and root exudates, as straw nutrients would eventually be exhausted. It would also reduce wind mixing. Second, an iron-precipitation pond was needed upstream of the living cover to prevent intense iron encrustation of the root systems. In the creek, the straw became encrusted with iron, forming round secondary mineral balls (Figure 2). The microbial-based treatment system thus developed was named Acid Reduction Using Microbiology (or ARUM). By providing reducing conditions (low Eh), a balance was introduced into the system that is generally present at the water sediment interphase.

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Figure 2. View of a piece of straw under the dissecting microscope.

The official objective of the company that provided the tailings areas for experimentation was to define a decommissioning approach for the pyrrhotite tailings site. Hence, Boojum experimented with altering the surface of the tailings with the particulates in the shotcrete stream, making small dikes behind which an alkaline cover developed (Figure 3a). The pyrrhotite tailings were fertilized in different seasons with different fertilizers (Figure 3b), to stimulate colonization of indigenous flora (Figure 3c). Boojum’s recommendations at the end of the RATS program were to create permeable, waste rock dikes, behind which a mine slime cover would be developed, but through which the water would still pass. No equipment was needed to foster the colonization of native flora, comprised of moss, cattails, and horsetails, which stabilized the surface after several growing seasons (Figure 3d).

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Figure 3. (a) Using alkaline mine slimes for self-seeding or planting cattails. (b) Fertilizing pyrrhotite surface at various times during the seasons. Fall fertilization was successful. (c) Horsetails colonized the alkaline mine slimes without any fertilizer. In background are the permeable waste rock dikes to accumulate mine slimes, based on final decommissioning recommendations. (d) Indigenous self-seeded growth covering the site. Photo taken after 10 years of implemented decommissioning recommendations.

On termination of the RATS program, a new Canadian government program was created, the Mine Environment Neutral Drainage (MEND) program. One of the areas supported by the program was constructed wetlands or passive treatment systems. These were acclaimed as an economically sustainable approach for municipal, industrial, and stormwater treatment.

The use of constructed wetlands for mine drainage, however, ignores the fact that in municipal and stormwater treatment wetlands, organic contaminants are degraded into biomass, water, and gases. In contrast, with mine effluents, the inorganic contaminants are precipitated as suspended solids. This results in plugging; the flow through the wetland is short-circuited and treatment is impaired relatively quickly. Sometimes organics and neutralizing agents added during construction need to be replaced. Larger flows require sizeable land areas, so wetland systems are relegated to treating smaller flows. For these reasons, wetland systems for the treatment of acid mine drainage have remained somewhat of a disappointment. An exhaustive review of passive treatment systems for mine effluents by Skousen et al. (2017) highlighted the continuing challenges of these passive systems and stated that further improvements are necessary before they can be economically used.

A Decommissioning Methodology: Ecological Engineering Approach

In 1980, Cairns published pioneering work in which he described the recovery processes of damaged ecosystems. In the same year, Bradshaw and Chadwick (1980) summarized seeding and stabilizing techniques, both natural and with amendments. Although these authors had mainly addressed coal wastes, similarities to uranium and base metal tailings exist. Not only are they all disturbed ecosystems, but they are extreme ecosystems dominated by extremophiles, as defined by Rothschild and Mancinelli (2001). Amendments of lime, seeds, and fertilizer are superficial fixes, not addressing the underlying biogeochemistry or ecology. Ecosystem development is based on evolutionary processes that need to be brought to bear on remediation work in general, and mine wastes specifically.

Ecosystem development principles were defined by Odum in 1962 as “ecological engineering tools.” What must take place are subtle actions by man, not dramatic change introduced with neutralizing agents and fertilizer. Odum stated that “environmental manipulation by man using small amounts of energy to control systems” is all that should be required. This approach was later consolidated by Mitsch and Jorgensen (1989) into “ecotechnology,” with 13 guiding principles, of which three apply specially to mining waste sites:

  1. 1. Ecosystem structure and function are determined by forcing functions of the system. Alteration of these causes the most drastic changes in the ecosystem.

  2. 2. Ecosystems are self-designing systems. The more one works with the self-design of nature, the lower the cost of energy to maintain that system.

  3. 3. Elements are recycled in ecosystems. Matching humanity and ecosystems in recycling pathways will ultimately reduce the effect of pollution. (Mitch & Jorgenson, Chapter 3, pp. 21–25).

These three principles underpin Boojum’s approach to working with the extreme mine wastes and water. Only when the forcing functions (or missing resources) are recognized and dealt with, can the close-out of mine waste management areas be truly realized.

The divergence between the ecological engineering and passive treatment approaches (constructed wetlands) widened during the MEND program, which emphasized treatment of the contaminated water emerging from the wastes, as opposed to developing processes that contain wastes and reduce their release. Unfortunately, Boojum’s work under the MEND program fell under the same category as passive treatment systems, and was evaluated under the category of Microbial Reactor Systems (Kilborn, 1996).

Boojum, with the support of many mining companies continued to define forcing functions within mine waste and water management areas and delivered the necessary ingredients tested in large scale R&D demonstration sites. After more than 30 years of developing decommissioning approaches for mines located around the world—bc, Saskatchewan, Ontario, Quebec, Nova Scotia, Newfoundland (Canada), the Minas Gerais (a tropical and semi-arid region of Brazil), the Sahel of Burkina Faso, the humid subtropical region of Guiyang in China, the Rocky Mountains, West Virginia, Northern Queensland in Australia, and Germany. Although many of the decommissioning approached developed were not implemented and remained on the shelves, we established that the same ecological principles persist worldwide, and most forcing functions are similar. This has enabled Boojum to develop a standard methodology for curtailing mining challenges, wherever they may be. Some decommissioning scenarios developed have ended up on corporate office shelves and have not been implemented (mostly due to management changes, common in the industry). The reports submitted to Canadian companies and to the regulatory offices are accessible at the Boojum Research Limited—Virtual Library at Laurentian University at Laurentian University. Other non-Canadian reports are forthcoming.

It is a general belief in the industry that every mine site has very site-specific environmental issues. This is like the belief that biological systems do not work in the winter (they work more slowly), or that deserts and the altiplano are ecosystems that are essentially void of biota (each has highly specialized flora and fauna). These beliefs are not based on ecological evidence. Indeed, different mines have varied hydrology and mineralogy, are in different climates, and hence differ in external ecosystems. But, just as the weathering processes are fundamental, so are the ecological processes that govern the functions of colonization, invasion, and with that, a recovery of the extreme conditions in mine waste management areas. A systematic approach has been formulated to address mine waste challenges. Each project’s common features are divided into three phases (Table 8).

Table 8. Water Quality Parameters of Seepages on the Northwest and Southwest Side of a Waste Rock Pile in Saskatchewan

NW Toe Seepages

SW Toe Seepages

Min

Max

N

Min

Max

N

Temperature (°C)

0.8

23

54

0.7

21

66

pH

3

6

58

1.9

6.3

69

Cond (µs/cm)

550

4620

58

273

4550

69

Eh (mV)

388

539

37

205

746

42

Acidity (mg/L)

4.7

653

32

10

1723

35

TDS (mg/L)

362

2490

35

327

4290

42

TSS (mg/L)

< 1

110

5

< 1

2300

9

Total Hardness (mg/L)

284

903

14

292

968

15

Cl (mg/L)

< 1

7

38

2

5

47

F (mg/L)

0.23

1

0.18

0.47

3

HCO3 (mg/L)

< 1

7

40

< 1

6

48

NH3-N (mg/L)

0.3

16

30

2.5

21

37

NO3-N (mg/L)

0.87

31

38

1.2

70

45

SO4 (mg/L)

219

1700

45

108

3590

59

Total PO4 (mg/L)

0.24

545

30

0.03

138

37

TKN (mg/L)

0.53

1.8

2

3.5

21

5

Source: Boojum Research (1994) CA066.

Phase I (the feasibility study) consists of a site visit to identify colonizing biota on and in the effluent streams, on the tailings, and the waste rock piles. This is carried out also with a survey of the undisturbed surroundings, because they reflect the resources available for the waste site recovery or colonization with respect to biota. Furthermore, the mine history, mineral extraction processes, the physical layout, hydrology, and former drainage basins are identified. Meteorological data from nearby weather stations, spanning several years, are requested in order to determine the amounts of precipitation that will fall onto the wastes when the mine operation ceases. Most of this information is available from the company or is found in historic files. Water, tailings, and biological samples found on the tailings or in the seepages are analyzed for contaminant concentrations to determine their potential contaminant-removal capacity. Concurrent with data collection, the chemistry, geochemistry, and geomicrobiology of the elements of concern are summarized. Here, references like the Encyclopedia of Geobiology (Reiter & Thiel, 2011) are useful. The authors described the biomineralizing processes for each element. These are compared to conventional-chemical accepted treatment systems. A report is issued describing the ecological potential to remove the elements of concern and the field and laboratory experimentation needed to develop the decommissioning approach. Upon approval of the report, Phase II (field and laboratory testing) commences. This includes a combination of laboratory and field testing of materials that trigger precipitation of the contaminants with additions of target substances (mostly organics, fertilizers). Field pilots are carried out on a small waste stream to determine residence times and growth rates. For the entire site, the contaminant loadings are calculated based on the size of the drainage basin(s) containing the wastes and the atmospheric precipitation expected. The residence times for the water are estimated based on the field-pilot growth rates achieved, and areas of surfaces (suspended curtains or brush) that need to be provided for attachment of submerged vegetation and its growth. Field pilots are generally run for 2 to 3 years and are regularly monitored. Often, adjustments are needed or one-time addition of nutrients is needed to kick-start the system, requirements often also determined in concurrent laboratory work.

An example of a field pilot is given in Figure 4 in a schematic form, representing in sequence several of the Boojum processes. A report is issued to the company and the regulatory bodies, and, depending on their response, the project proceeds to Phase III, with a gradual scale-up based on the field-pilot results. A monitoring program is carried out for several years, until the ecosystem is homeostatic or has reached self-design.

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Figure 4. Schematic of a pilot system constructed by an operating mine to determine site-specific design criteria. Gravity feed from the supply tank flows into the head tank and metered flow from the head tank enters phosphate rock to remove iron. The water then flows into the biological polishing tank, with baffles or curtains for biological polishing. After this step, it feeds into the ARUM tank to increase pH and flows into a second biological polishing tank, envisaged at scale up to be the open pit.

Ecological Engineering Processes and Applications

The emerging conflicts between agriculture and mining over land and freshwater use cannot be averted without a paradigm shift in the understanding of metal extraction, water usage, and the control of waste generation. Rarely can a shift take place without demonstrating that alternatives exist. This was recognized by a group of scientists promoting bioleaching, more than three decades ago. The book Biogeotechnology of Metals (Karavaiko & Groudev, 1985), opened with a quote from Pasteur in 1871:

No thousand times no, there is no category of science which could be named applied science. There is science and application of science related to it as the fruit to the fruit-bearing tree.

What Pasteur called “the fruit” are here called the ecological engineering processes used in mine waste management decommissioning. These processes have been demonstrated empirically, and, when finances allowed, they have been bolstered with extensive literature reviews and standard classical hypothesis testing.

Constructing ARUM Sediments and Floating Living Covers

Numerous publications on the selection of plants and organic substrates for passive treatment systems proposed that the nutritional value of different organics was relevant. To test this hypothesis, organic materials that were most commonly used to construct wetlands were subjected to sequential nutritional analysis. The results obtained from analyzing sawdust, peat, straw, cattail litter, and alfalfa straw revealed a composition of 1% to 9% lipids, 22% to 40% sugars, starch, and amino acids, and 17% to 44% recalcitrant cutins, lignin, and silica (Smith & Kalin, 1991). All contained the essential organic molecules for microbial growth, which suggested that it really didn’t matter which organic material was used as food material for microbial sediment construction. Recalcitrance (or degradability) of the organic debris is essential. It must give the sediment structure, without compacting, like alfalfa straw. Organic material that degrades easily is sometimes needed to “jump start’ the degradation, for example, guinea pig pellets (food for guinea pigs), which are rich in nitrogen, which sink to the bottom of the pond where degradation takes place. To maintain long-term microbial growth, a continuous supply of organic carbon is needed. This is the role of the floating, living cover. Floating vegetation islands on lakes and ponds occur naturally, mostly starting from shores, and are formed during a process of terrestrialization (Figure 5a). For example, such terrestrialization occurs in northern Canada, where floating muskeg gradually covers the surface of lakes and narrows of lakes Floating living islands exist worldwide. This process can be initiated along the shorelines of a lake or an open pit by dumping truckloads of cut brush onto the ice in winter, for distribution to the shores upon ice breakup (Figure 5b). After several years, significant terrestrialization is evident (Figure 5c), as described, for example, in the outflow area of an acidic lake at one of Boojum’s R&D demonstration sites in northern Ontario, Canada.

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Figure 5. (a) Terrestrialization in northern Canada in narrows of lakes. (b) Promoting terrestrialization with cut brush in an acid lake of an R&D site in northern Ontario. (c) With time, shoreline vegetation is evident in the shallow outflow channel from the acid lake where cut brush was added. Below the vegetation, an active ARUM sediment has formed.

To select the plants for the living floating covers, the primary focus is tolerance to mine water conditions, either alkaline or acidic. The second focus is to select a plant with a well-developed, loosely bound root system. Both Typha latifolia (common cattail) and Phalaris arundinacea (reed canary grass) tolerate the adverse chemical environments of mine waste water, but Typha has a better-suited, more loosely bound root system.

An operating mine site often has ditches and ponds in which drainage collects to be treated in a neutralization plant. These ditches and resulting ponds, when mining and milling ceases, continue to collect contaminated drainage. To treat them “in situ,” they must have organic sediments that are microbially active. Ditches may have to be equipped with flow obstructing structures to reduce the velocity within the ditches. Since these ponds and ditches were excavated, they have no organic sediments. One active mine operation was all enthusiastic about ecological engineering and placed hay bales in all collection ditches. Those were soon encrusted with iron (Figure 6), a means of reducing iron in the treatment plant, but the alkalinity and pH increases were not generated. Hence a precipitation pond would have to be integrated into the ditch to remove some of the iron.

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Figure 6. Compact hay bales in ditch, a bad approach but a good example of oversimplifying matters.

Boojum constructed pilot systems under the MEND program at two mines in northern Ontario to define design criteria for the application of the ARUM process: one in Elliot Lake to address sediment construction, and one in Sudbury promoting alkalinity and pH increases.

The site at Elliot Lake involved a very slow-flowing acid pool at the foot of a tailings dam. Plastic drums, with open bottoms, were filled with loose straw, and inserted into the pond sediment (tailings). Standpipes were added to the drums to sample water for analysis (Figure 7a). Analyses included, Eh, pH, and elemental composition. Redox conditions change rapidly when groundwater or sediment porewater is exposed to air. Hence porewater from the standpipes was sampled and stored under nitrogen atmosphere. Unfiltered porewater samples were sent to a certified laboratory for ICP analysis (inductively coupled plasma spectroscopy).

Microbial metabolism removes oxygen from sediments when organics degrade. The redox level drops to a negative value as microbes catalyze a series of chemical reactions that generate alkalinity through sulfate and iron reduction. The results from the porewater analysis collected from the drums after 2 years in the field were used as input to an aqueous geochemical calculation program, PHREEQC. The program calculated the saturation indices of compounds in aqueous solutions under the Eh and pH conditions that were measured in situ in the standpipes of the drums. The program calculated indices for each compound that express the potential or the probability that it will precipitate as a solid and therefore will be removed from the sediment porewater. The drum porewater had positive saturation indices for iron and sulfur, suggesting that these elements were likely being precipitated. Once this was confirmed, hay bales were installed as the basis of a microbially active sediment in the shallow pool at the foot of the tailings dam. The honeycomb pattern provided alternating oxygenating and reducing sections in the pool (Figure 7b). After 3 years, the acid drainage at the bottom of the pool, starting with a pH of 2.3, had improved to a pH of 5.2 at an electrical conductivity of 11,000 µS/cm.

Although these results confirmed ecological engineering principles, the experiments were never understood by the funding agencies or the program operators, because nothing green or “wetland-like” could be seen. Funding was cut, and 4 years later, upon Boojum staff return to the site, the experiment was found to have been destroyed. Only a photographic record of the black, biogenic pyrite in the straw remnants demonstrates that the process was successful (Figure 7c).

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Figure 7. (a) Drums with open bottom and perforated standpipes inserted into tailings filled with straw to sample pore water to determine conditions that potentially lead to biomineralization of pyrite. (b) Scale up from the results from the drums with a honeycomb placement of hay bales in winter (because access in summer was not possible) into the acid pool to develop a sediment. (c) No green, no wetland, so why not dig it up? The black-stained wood close to the road is visual evidence of biogenic pyrite, which is also somewhat evident in the darker color below the exposed hay bales.

The pilot system in Sudbury, Ontario, was not eliminated from the MEND program. The system was installed at one of the seepage pumping stations below a tailings dam. The tailings dam encompassed 2,225 ha of tailings. The intention of the mining company was to install ARUM systems, if they worked, at seepage stations around the tailings dam. Water that was clean enough would be released to the environment; otherwise, a much-improved effluent would be pumped back to the large tailings pond, thereby no longer recycling the high concentrations of contaminants in the seepage.

This pilot system was used to develop design criteria for ARUM (Figure 8). Since tailing seepages emerge containing reduced iron (Fe2+), the residence time needed for oxidation, which is reported to be a fast first-order reaction, needed to be determined, to prevent entrustment of the straw or roots. Visually, iron oxidation is fast as the solid iron hydroxide precipitates and the pH drops, but, in fact, only a fraction of the iron reacts, the rest remains reduced and stays suspended. This is especially the case for coal waste seepage, and it is suspected that the iron is often complexed (combined with other compounds, e.g., humic acids). Hence it is wise to experiment in the laboratory and or in the field regarding the oxidation behavior of reduced iron in the drainage or seepage.

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Figure 8. (a) Schematic of ARUM pilot test system for tailings seepage treatment aimed at pH increases and alkalinity generation. (b) ARUM pilot test system after 3 years, with floating cattail covers on the ARUM cells. (c) Summary of floating island development steps. Upper right panel shows the natural occurrence of cattails floating. Upper left and lower left show first trials of cattail survival in an iron precipitation pond. Lower right shows seed germination in a clearwater pond. Upper right shows root formation, with stained bands reflecting different water levels in the cells. Stained brown oxidation and black reduction during flow testing in the system, for residence time estimates.

The pilot system schematic is shown in Figure 8a. The first two cells are connected through a permeable dike, containing curtains suspended to provide surface to precipitate iron. In the second cell, smaller iron particles are still contained in the nearly clear water. The next dike is connected by an impermeable dike to an intermediate cell, which was planned for potential addition of soluble organics, to kick-start the microbial system, but it was never needed.

Another design question was, how long do the microbes need to raise the pH from 3 to 4.5 in the ARUM cells? The acid seepage from the first two cells, the precipitation pond followed by the acidification enters through a permeable dike to two ARUM cells in which a sediment is created and a living floating cover installed. These cells are separated by an impermeable dike and flow passes through a pipe. With the pipe we could control the flow between both cells and determine the residence time needed to increase the Ph. The permeable dikes are not needed when upscaling the system, but for each site the residence times of the water has to be tested, both the precipitation of iron with the acidification and for the pH increase. Those differ as each drainage contains different species of iron, complexed and proportions of reduced and oxidized iron and metals. In Figure 8b, the floating covers are shown after three years. In Figure 8c (insert upper right-hand side) the root system is depicted, showing the seasonal bands of oxidation and reduction, (brown and black) stained stripes.

The two ARUM cells are also used to develop the floating vegetation covers. The first attempts at developing the floating vegetation covers involved transplanting cattails with the root mass along with peat and living moss onto PVC rafts to provide flotation (Figure 8c bottom left). This was tedious and not especially scalable to large installations, but the cattails grew. To remedy this, cattails seeds were germinated in relatively clean water on site, and after germination, were transferred directly into the contaminated acid or alkaline water (Figure 8c bottom right; Smith & Kalin, 2000).

Nearly two decades later, Pavlineri et al. (2017) described the use of floating, living covers for water treatment in other remediation situations. Such covers have reached commercial scale. The parameters derived from the pilot systems were summarized in a final report (Kalin, 1993). The MEND program was terminated by the government, and a review of the applicability of passive treatment systems was commissioned (Kilborn, 1996). This review, unfortunately, did not recognize the fundamental differences between ARUM and other passive treatment systems. It concluded that ARUM was useful for small flows, but could not be scaled up. The fundamental differences between ARUM and other passive treatment systems have been described more thoroughly (Kalin & Caetano Chaves, 2003; Kalin et al., 2005).

In Figure 9, the water quality changes are summarized for the pilot system and several other locations in which the ARUM approach was tested. The results are given as removal rates for the metal of concern in these operations, because flows and the volume of the test cells for the systems were available. The surprisingly high removal rates for sulfate are due to additions of straw to an iron-encrusted peat bog that received seepage from a waste rock pile in central Newfoundland, Canada, and the results support ecological engineering principles.

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Figure 9. A summary of water quality achieved in the field in the pilot system of the MEND program (central Ontario) and that of several other operations where we added straw or hay to test the ARUM applicability. Because flow and volume of the test areas were available, removal rates are reported.

Design criteria derived under MEND were used to scale-up a complete system for the treatment for acid mine drainage emerging from an abandoned gold mine adit flowing at 0.6 to 1 L/sec in Brazil (Kalin & Caetano Chaves, 2003). Figure 10a shows the pilot system on the upper left panel and the portal on the upper right. The precipitation ponds and below again with gravity feed the ARUM ponds. The cattails on the Styrofoam floats had a short lifespan because shortcuts were taken in float construction. Although the cattails were dying, they were replaced and they did the job as indicated by the results presented in Figure 10b.

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Figure 10. (a) The ARUM system in Brazil depicted with the portal of the abandoned gold mine (upper left), the precipitation pond (upper right), and the floating cattails being strangled to death in Styrofoam floats, which later got corrected. (b) The performance of the system after the first year of installation. Loading is given in kg/month and removal expressed as percentage.

Unfortunately, the project was terminated due to a merger with another mining company, and Boojum investigators could not track performance over more seasons. The results obtained were for the first season. The system has been running for more than 10 years as proper living floats were installed (personal communication, V. Ciminelli, University of Minas Gerais). Field work was also carried out in the highly acidic Tagebauseen, which are pit lakes from coal mining in Germany, where ARUM is referred to as “controlled eutrophication” (Fyson et al., 2006).

Other ARUM Applications

Dried Out Lake Sediments

An area of 130 ha of dried out lake sediments was studied in northern Saskatchewan. The original lakes lost their water as the groundwater table was lowered to construct an open pit that would reach the uranium ore. Most lakes did not have an inflow creek and received water from atmospheric precipitation. Organic sediments in the lake contained metals and biogenic pyrite, the natural result of ARUM, noticeable due to small acidic seeps occurring after rainfall or snowmelt. Upon reflooding at close-out of the mining operation, the rising groundwater will re-fill the lakes and with that flood the sediments. Because the buffering capacity of the groundwater was very low, the newly refilled lake would turn acidic. The environmental concern was that nickel, also present in the sediments, would be released. Hence the objective of our work was to define the mobility of the nickel and find means to counteract the acidification due to the oxidized pyrite. This appeared to be an easy job for ARUM application.

To test metal mobility and to address the means to reduce that mobility, sediments were sampled and analyzed, and subsamples of the material were placed in 2-L jars. The jars were filled with groundwater to which organics (potato peel from a French fry company) were added. Jars were kept in a fridge to provide low temperatures reflecting the northern location of the site. The porewater and overlying water were monitored for nickel concentration, pH, and Eh for several years. More than 90% of all samples to which organic carbon was added showed reduction in nickel release to the porewater and from the sediment to the overlying water (Boojum Research Ltd., 2006). Once again, evidence was provided that organics are a key component in initiating change in the waste management site.

Groundwater Plumes

ARUM was tested on a seepage plume moving away from a tailings deposit at a depth of 16 m. The plume contained close to 1 g each of iron and zinc, and about 20 g of sulfur compounds per liter. Microbiologists indicated that microbes were unlikely to inhabit the plume, given the chemical composition. Exposure of the seepage water to air or surface water led to immediate precipitation of the metals and acidification. The groundwater plume emerged into a shallow lake with a pH of 5, but, within seconds, the pH dropped to pH 2.5. Laboratory experimentation with this type of seepage is not possible, so geochemical modeling was carried out (Fleury, 1999). The model included the well-documented reactions of urea-degrading microbes:

CO(NH2)2+2H2O2NH4++CO32
(1)

CO32+H+CO3
(2)

These microbes are ubiquitous and would degrade urea to carbonate and ammonia, both of which would raise the pH.

Geochemical modeling confirmed that, with the addition of a slurry made of urea and carbon in the form of sugar the pH of the groundwater, which is between 4 and 5, is expected to increase to between 7 and 8, which would lead to precipitation of the reduced iron underground (Fleury, 1999).

The indigenous microbial flora of the seepage water were quantified and were found to be relatively rich in diversity (Lau et al., 2001), although no ureolytic microbes were documented. The microbiologists were not surprised at the absence of ureolytic microbes, as they declared that ureolytic microbes are reported not to be active below pH 4.5 which would easily prevail in the groundwater. We did not believe this report as the literature did not reveal any investigations which would confirm this restriction of their activity at or below pH 4.

A small field pilot was set up in a sandy area, which had low but reasonable flow velocities to experiment with seepage which was suspected to be the origin of the ground water seepage. It had similar chemical composition. The field tests were solely to test activity of ureolytic microbes. The seepage in the sand had a pH of 2.8. To test if additions of sugar and urea mixed in water would stimulate ureolytic microbes, a net of standpipes was installed to a depth of 0.5 m. The sugar and urea mix was injected about 15 to 20 cm into the sand. Water samples from the standpipes downstream of the injections ultimately reached pH values of 7, with concurrent increases in organic carbon and ammonia. This, and laboratory and field tests with commercial ureolytic enzymes, confirmed the activity of ureolytic enzymes at low pH.

With these confirmations a large pilot test was installed on the shore of the lake intercepting the 16-m deep seepage, which emerged into the shallow lake. The piezometer from which we obtained the contaminated seepage was an artesian well, thus no pumping was needed. The idea was to intercept the seepage and to induce, in situ or below ground, precipitation of iron and zinc, to reduce the contaminant load released into the lake. If successful, several injections would be installed to create a precipitation area below ground and the contamination of the lake gradually be reduced. In Figure 11a the schematic of the injection system is depicted.

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Figure 11. (a) Schematic of the in-situ ground water seepage treatment pilot. (b) Photographic record of the standpipes to monitor the injected plume (upper left) and the head feeding tank of the urea and sugar mixture (upper right). Lower right panel shows the view of piezometer (M60A) located in the seepage 15 m below ground, which supplies the seepage in the foreground. The straw-covered supply pipe to the injection at the T-junction, also covered with straw, is in the background with the feeder tanks. The T-junction of the injection wells is also straw-covered for the winter (lower left).

A mixture of urea and sugar was dripped into the seepage water and reinjected below ground in a gyttja-like sediment formation (mud and degraded peat of gel-like consistency). The movement of the plume is monitored through a grid of standpipes driven into the gyttja layer below a floating muskeg on the shores of the lake (Figure 11b). The upper two panels in Figure 11b depict the monitored area, with the standpipes on the left, and the head tank with the urea and sugar mixture on the right. The panel below on the left shows the piezometer from which the drainage moved toward a T-junction in the background close to the head tank, noted by the straw cover prior to the winter. The panel on the right depicts the view toward the lake, with the top of the T-junction above the injection well.

Figure 11c gives the results from one of the standpipes within the plume. On the far left, the background values are reported from summer 2000. The amount injected up to August 2002 was 529 m3 of seepage with a total urea/sugar mixture of 16.1 m3. Monitoring ceased in summer 2003 after a final injection of 2 m3 urea/sugar slurry together with 58 m3 of seepage. The monitoring data indicate that the process is working, organic carbon is used, ammonia is produced, and, as expected, the metals are removed. This was the main objective of the system. A groundwater model for the site existed, and was used to scale the system up. Although the treatment approach was scaled up using the existing groundwater model MODFLOW, due to disagreements with the regulatory agencies unrelated to the groundwater treatment, the entire R& D work at this site was abandoned in 2003. The work on the site was summarized and all measures carried out were extensively documented, as was the seepage treatment (Boojum Research Ltd., 2000; Kalin et al., 2008).

Reducing Microbial Sulfide Oxidation: Altering the Mineral Surface

In the extreme environments of sulfidic mine waste, chemolithotrophs dominate. They use oxygen to break down the minerals for energy. In contrast, healthy undisturbed ecosystems are buffered, by the activity of autotrophs, lithotrophs, and heterotrophs each group balancing one or more others. This is an oversimplification, but it is certain that the growth rates of microbes using rocks to gain energy are higher than those that break down organic material. Heterotrophic microbes will grow much faster. Clearly degrading organic matter requires less energy than breaking down minerals. Heterotrophs degrade organics but consume oxygen. If the heterotrophs on the mineral surface could be fostered, their growth would out-compete the chemolithotrophs for surface area, effectively sealing the rocks. So can heterotrophs be found in enough numbers to effectuate a change in the oxidation state on the mineral surface?

To test if heterotrophs are present and viable after blasting surviving the milling circuit and alive in the tailings, water was sampled from the mill’s intake pipe to its discharge into the tailings pond and in ditches collecting acid drainage. Samples were plated on media supporting growth of heterotrophs and the growing colonies were counted (Table 9).

Table 9. A Summary of the Three Phases Used to Develop Decommissioning Approaches

Phase

Description

► Site history, mine waste management area

► Physical layout

Phase I

► Climate, hydrology, surface water and groundwater quality

Site Characterization

► Contaminant loadings

► Waste mineralogy, geochemistry contaminant paths and fate

► Contaminant paths and fate

► Process selection

► Existing terrestrial and aquatic ecosystems

► Biological system selection

Phase II

► Geochemical and biological reaction rates

Field Testing of Selected Strategy

► Define site-specific design criteria for treatment strategies

► Assess feasibility of strategies

► Decision to proceed to Phase III or address missing data

Phase III

► Full-scale design, construction and monitoring of treatment approach

Scale-up of Treatment System

► Modify if necessary, fine-tune system

Source: Kalin (2004).

The high population of heterotrophs found at the tailings beach was likely due to the organic metal flocculating agents used in separating the mineral from the ground rock slurry. In comparison, samples taken in acidic seepage ditches showed that heterotroph numbers were very low, but present. These findings supported the hypothesis that heterotrophs are ubiquitous. Although it was uncertain, it was assumed that if they can survive milling, they can also survive blasting, so they will also be present on mineral waste rock surfaces. If they are there and could be stimulated to grow on the mineral surfaces, overgrowing the chemolithotrophs, less oxygen would be available for weathering and it would slow or stop the oxidation of minerals in the waste material.

How can the habitat on the mineral surface be altered to stimulate heterotroph growth? Some basic guidance can be obtained from the literature. Oxidation of minerals in mining wastes is similar to the corrosive process on steel and other industrial materials (Wakefield & Jones, 1998). In this area, when chemicals are used to reduce or prevent corrosion, it is referred to as phosphating (Schweitzer, 1988). Since the early 1990s, the involvement of microbes in the deterioration of these materials have been recognized, and it is referred to as microbial-induced corrosion (MIC; Beech et al., 2000; Little & Wagner, 1992). Phosphate has also been thought to serve as a means to slow corrosion in mining wastes. In the presence of oxidized iron, phosphate forms a precipitate, iron phosphate. This coat could possibly prevent access by oxygen and reduce oxidation rates. One set of field experiments using carbonaceous phosphate ore in coal wastes was somewhat successful and produced intriguing results (Meek, 1991; Spotts & Dollhopf, 1992), although Evangelou (1995) published a comprehensive book about phosphate coatings oxidation. Numerous publications described experiments where phosphate fertilizer was added to sulphidic wastes, but this water soluble phosphate produced mixed results. The industry did not carry the approach forward, for many reasons, including economic and operational (i.e., application) difficulties.

Boojum pursued the intriguing results obtained when the carbonaceous marine ore from the phosphate mine in North Carolina was added to coal mine wastes. We found that lowest dosage of ore added produced the best results, that is, effluents with the highest alkalinity and pH. This was a strong hint that microbes are at work. It was material which had carbonate neutralizing capacity which would be useful to be transpored to mineral surfaces and, to an ecologist, where there is phosphate, there are microbes, a possibly a source of organics and several types heterotrophs.

Further, a mining magazine reported that a copper heap leach dump in British Columbia had stopped leaching copper (Scott, 1991). Why had it stopped working? Was it due to microbial or chemical inhibition as there might have been some phosphate in the ore? Rocks from the section of the heap leach were obtained (Figure 12). If the results of the work with the carbonaceous ore were correct, the surface coating would have either shown an iron phosphate coating, or a coating with microbes, or both. When the surfaces were inspected with SEM (scanning electron microscopy) and XRD (X-ray diffraction), very few peaks reflected phosphate, but a considerable number of microbial colonies were evident─findings that were a “Eureka!” for Boojum staff.

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Figure 12. Rock from heap leach that ceased generating acid. The rock will undergo microscopic inspection for iron coating.

Since Texas Gulf (Raleigh, North Carolina) had supplied phosphate ore for the test with the coal wastes in West Virginia, they were asked to supply below-grade phosphate ore or their wastes. It would be ideal if one mining waste could solve other mining wastes problem. The material, initially called natural phosphate rock (NPR), was washed and screened at the site, and then was shipped to Ontario and Quebec in rail cars for field and laboratory experiments. The gravel contained seashells and sharks’ teeth in abundance and had a size distribution of 0.001 m to 0.01 m in diameter. It was composed of 36% calcium phosphate and 48% calcium carbonate and some other material (Kalin et al., 2003). We carried out a controlled experiment outdoors in 75 liter plastic drums to which about 8 L of NPR was added on top. We assumed that the rain would transport smaller particles from the gravel into the drums with the rain or snowmelt. The effluent from the drums was collected periodically and was analyzed for acidity, pH, Eh, and electrical conductivity. After 2.7 years outdoors and after three winters, the final effluents collected were filtered and analyzed for elemental composition, and the drums were dismantled. About 50% of the 8 L of the NPR remained. The NPR was later renamed Carbonateous Phosphate Mining Wastes (CPMW) to redirect the general belief that a phosphate addition is a potential source of eutrophication. Selected rocks from the drum experiment were subjected to SEM/EDS (scanning electron microscopy/energy dispersive X-ray spectroscopy; Figure 13a and 13b).

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Figure 13. (a) Organic coating on rocks from the outdoor drums with waste rock and CPMW added after 2.7 years of exposure. (b) Organic coating after indoor storage for 11 years under various storage conditions.

An organic coating was evident on the rocks from the drums to which NPR was added, although it was somewhat dried out (Figure 13a). However, an organic coating alone made microbial involvement in forming a biofilm likely, but it was not definite proof. The rocks from the experiment were stored in an industrial basement 4.5 years before they were re-exposed outdoors to test the longevity of whatever coating had developed. Kalin and Harris (2005) concluded that neutralization of the effluent due to the carbonaceous nature of the NPR could not be supported, but essentially the improved effluent could not be explained, because there was limited microbial evidence. Funding was obtained for a postdoc at the University of Ottawa to investigate the surfaces of the rocks and to carry out detailed experiments with the same NPR and biofilm formation was confirmed Ueshima et al. (2003, 2004). The biofilms also appear to be long-lived, because they were found to persist for 11 years, as determined by microscopic scanning of the surfaces of the rocks again after storage (Figure 13b; Kalin et al., 2010).

Rock surfaces from the same outdoor drum experiment were shipped to Australia along with some NPR, the same which was used in the drums for further tests under the supervision of R. Smart of AMIRA (Australian Mineral Industries Research Association) at the University of South Australia. An organic coating containing inorganic precipitates was still present after storage indoors, a second outdoor exposure for a complete season, and shipment to Australia (AMIRA, 2017). The work in Australia also repeated experiments indoors with NPR with wastes from Red Dog, a Canadian zinc mine, obtaining again the formation of an organic coating.

As the investigations at the University of Ottawa and in South Australia continued, Boojum set up field experiments in larger and smaller plots with the NPR obtained in train loads. The systematic controlled experimentation in outdoor drums was completed and published in 2005 (Kalin & Harris, 2005). Field plots were implemented on fresh pyrrhotite and on acidic uranium mill tailings, as well as in a concentrate spill area of a polymetallic concentrator in Newfoundland. The field plots were sampled after having been exposed outdoors for 2 to 5 years. The field samples were slurred 1:5 (weight to volume) with tapwater and were monitored during 22 months of exposure to oxidizing conditions in the laboratory. The supernatant from 23 tailings samples of the field plots was analyzed, but only a few samples produced a higher pH and lower acidity and showed traces of phosphate. Considering that sampling the old plots was not easy, because the markings of the dosages applied in the field were partially destroyed, this was a relatively small success, but some evidence of improved porewater was found in some samples (Kalin et al., 2003).

One of the important functions of NPR is its neutralization capacity on a very local level once a particle reaches the mineral surface. The characteristics of NPR were determined by leaching the material in 0.1 N sulfuric acid (Kalin, 2009; Kalin et al., 2012). It showed very limited solubility in rainwater and distilled water, but in acid solution it released a rich assortment of nutrients for microbial growth (phosphate, calcium, potassium, and micronutrients).

Although an organic coating was now documented by independent researchers and repeated experimentation, the presence of actual heterotrophic populations and their growth with NPR had yet to be proven. However, it was now possible to hypothesize that rain/snowmelt can carry small carbonaceous particles into the wastes, altering the acid conditions on the mineral surfaces so that heterotrophic biofilms could develop, as we collected about 4l of the NPR gravel at the bottom of the drums during dismantling. To us it was clear with the repeated organic coating observed by independent groups of researchers that the same could happen in tailings, although only observed indirectly by some improved effluent. The biofilms could cover the oxidation sites, thereby reducing the activity of the oxidizing microbes and a hypothesis was formulated but it remains to be verified (Kalin & Wheeler, 2011; Kalin et al., 2012).

Professor Wolfgang Sand at the Biofilm Center at the University of Duisburg Essen, together with S. Bellenberg, tested NPR on German lignite coal. Sterilized pyritic coal columns were set up in the laboratory. Chemolithotrophs were inoculated into the columns to induce oxidation. Others columns were left sterile as controls. Leaching started at a pH of 1.5. The effluent pH from the columns with CPMW increased and a biofilm of acidophilic, heterotrophic microbes developed and was documented by counting and identifying the microbial groups. The reduction of viable sulfate oxidizers over the period of the experiment was documented and the experiment was terminated after 213 days. The effluents from the final flushing are shown in Figure 14.

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Figure 14. Photographic record of effluent color after 213 days in microbial colonization experiment with CPMW added to German lignite coal. B2 and B4 CPMW added a clear solution, free of iron, in contrast to B1 and B3 without CPMW. Unfortunately, there is a red box behind B4.

Due to lack of funding, only a photographic record of the clear solutions from the NPR-treated columns can be presented, but they had improved pH and conductivity as well as lower Eh values (Bellenberg et al., 2013). A summary of all publications about CPMW can be found in Kalin et al. (2015). It is now possible to rename the NPR (Natural Phosphate Rock) wastes as CPMW (carbonaceous phosphate mine waste). Indeed this is justified, as we analyzed and leached in sulphidic acid phosphate mine tailings along with other phosphate ores occurring in Ontario. Only the tailings displayed a solubility and elemental composition similar to NPR, although it was a mine in Northern Ontario not in North Carolina. Hence it is likely that many alternatives can be identified and the original material characteristics used in all the tests to date are not unique to the North Carolina mine.

Biological Polishing: Removing Metals/Contaminants From the Water Column

Mine waste water contains both inorganic and organic particles. They are measured as total suspended solids (TSS), in contrast to total dissolved solids (TDS). Suspended solids are defined by their size: they pass through a filter with a pore size of 2 µm, but are blocked by a filter of 0.45 µm. The compounds passing through the 0.45- µm filter are considered dissolved solids (or TDS), as also reflected in the measurement of electrical conductivity, and those larger than the filter size will be blocked on the filter paper are suspended solids (or TSS). Both TDS and TSS form colloids, gels, or emulsions like particles, which do not settle out of the liquid until they are large enough and would not pass the filter paper (depending on pore size) or they are not heavy enough to sink to the bottom. The particles “play a commanding role in regulating the concentrations of most reactive elements and of any pollutants in soils and natural waster systems and in the coupling of various hydro chemical cycles” (Stumm & Morgan, 1996, p. 818).

The complex world of environmental particles and their characteristics are described in detail by Buffle and van Leeuwen (1992). Particles are essential in aggregating elements or substances that occur in dissolved form. They assist in coagulation, aggregation, and complexation of dissolved elements onto or into aggregates, facilitated to a large degree by changing the surface charges of particles. Removal of elements (dissolved or suspended) by conventional methods is achieved by altering the chemistry of the waste water by adding chemical agents, clarifiers, or flocculants so that the elements or compounds precipitate. The chemicals added lead to a fast reaction and produce a sludge that, after dewatering, is mostly relegated to landfill sites. Inorganic particulates, such as clay and metal-hydroxides (Raoul et al., 1967), are also adsorbents, and are often present in abundance in mine waste water, and they accumulate as a sludge containing metals that is referred to as yellow boy. However, with aging, this sludge releases the metals again. Thus, stabilization is needed, possibly with an ARUM approach.

Periphyton and Phytoplankton as Biopolishers

Phytoplankton, free-swimming algae, live in the water column, as opposed to periphyton, which are algae growing attached to substrates. Both groups photosynthesize, which raises the pH near the cell, facilitating coprecipitation of metals (van de Vossenberg et al., 1998). Some algae produce carbonates on their outside cell wall onto which elements coprecipitate or they take up metals into their cells. Finally, some algae generate mucilage shells around their cells, which can carry various surface charges. Together, these organisms constitute “biological polishing agents,” effective not only at adsorbing contaminants in the water column, but also at aggregating inorganic particulates, hydroxides, and clay particles to facilitate settling. Algae and microbes have been reported from extreme environments, with pHs as low as 0.7 and as high as 8.5 (Johnson, 1998; Kalin et al., 2004).

It is not surprising, then, that periphyton populations have been found growing profusely in a seeps, ponds, and ditches in mine waste management areas (Figure 15a). These periphyton associations contain several algal species, primarily benthic algae from the order Ulothricales, diatoms, fungi, and microbes, and sometimes also mosses (Olaveson & Stokes, 1989). Mixed periphyton populations have been found growing in mine effluent with pH values as low as 1.3 (Hargreaves & Whitton, 1976), with zinc concentrations as high as 450 mg/L, and with aluminum concentrations as high as 95 mg/L. When mine drainage runs over rocks or organic debris, also found often in underground workings, colonies of these organisms form structures that sometimes resemble stromatolites (Figure 15b).

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Figure 15. (a) Seepage creek with attached Ulothrix algal growth (Iron Mountain in California; pH around 2.0). (b) Stromatolite-like formation in acid mine drainage stream from a coal mine adit in China, running through a garbage heap with waste coal.

The role of algae in particle aggregation was documented in an open pit in Saskatchewan. The pit was force-flooded with water from an adjacent lake at decommissioning, and was filled with 5 million cubic meters of neutral pH water from the adjacent lake over the winter. However the arsenic and nickel concentrations of 0.22 mg/L (S.D. 0.1) and 0.26 (S.D. 0.08) mg/L, respectively, were considered contamination by the regulatory agencies, released from the pit walls. It needed to be treated in the chemical treatment plant before the pit water could be released by overflow into the adjacent lake. The company decided to fund an investigation into biological polishing within the pit, as sufficient time existed before the water level in the pit would reach overflow. Only atmospheric precipitation entered the pit, not groundwater.

The force-flooding produced high concentrations of total suspended solids (TSS) in the first 2 years. Water quality and phytoplankton populations were monitored, and sedimentation traps collected the settling particulates at various depth. Statistically significant correlations between phytoplankton biomass increases and reductions of arsenic and nickel concentration occurred gradually over 7 years. As time progressed, phytoplankton diversity increased. The composition of the organic-inorganic particles that were formed and that were collected in the sedimentation traps was investigated with SIMS electron microscopy. A schematic was derived to show the composition of the particles (Figure 16; Boojum Research Ltd., 1994; Cao & Kalin, 1999; Kalin et al., 2001).

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Figure 16. Schematic of particle aggregation based on SIMS microscopy, depicting algae as flocculating agent.

Although the particles’ composition showed clearly that algae act as flocculants or particle aggregators, in the engineering world, no treatment process is accepted without a mass balance of the elements removed, also called a flowsheet. Hence, the sediments that had accumulated at the pit bottom had to be sampled and the mass of elements in the sediment had to be estimated and compared with the mass of the cumulated material in the sedimentation traps (Table 10).

Table 10. Mass Balance of Settling Particles and Accumulation in Sediments in a Pit Lake with Biological Polishing

Sediment Rate (g/m2/day)

Per Year (g/m2)

Arsenic (As) Content μ‎g/g (g/m2)

Nickel (Ni) Content μ‎g/g (g/m2)

Iron (Fe) Content μ‎g/g (g/m2)

Aluminum (Al) Content μ‎g/g (g/m2)

Phosphate (PO4) Content μ‎g/g (g/m2)

1992

28.69

10,472

144

(1.51)

166

(1.74)

17,600

(184.30)

15400

(161.27)

2,319

(24.29)

1993

28.69

10,472

363

(3.80)

450

(4.71)

9,480

(99.27)

7360

(77.07)

2,100

(21.99)

1994

11.58

4,227

1,880

(7.95)

1,070

(4.52)

14,300

(60.44)

8,980

(37.96)

2000

(8.45)

1995

13.04

4,758

2,695

(12.82)

723

(3.44)

13,000

(61.85)

6,385

(30.38)

1,900

(9.04)

1996

3.52

1,285

2,090

(2.69)

715

(0.92)

20,200

(25.95)

14,300

(18.37)

1,885

(2.42)

1997

2.40

876

7,600

(6.66)

1,300

(1.14)

47,000

(41.17)

16,050

(14.06)

3,060

(2.68)

1998

2.55

931

4,300

(4.00)

1,000

(0.93)

46,800

(43.56)

17,800

(16.57)

2,387

(2.22)

1999

2.02

737

3,600

(2.65)

540

(0.40)

51,600

(38.04)

14,100

(10.40)

2,142

(1.58)

Total 1992−1999 1999

33,757

42.08

17.80

554.60

366.07

72.68

In sediment (Depth 0−7 cm)

30.30

17.10

566.00

619.10

57.40

Source: Boojum Research Ltd. (2000) CA105.

The next step was to support the growth of biomass in the pit lake, as it had no shorelines with semiaquatic vegetation and did not receive any runoff from a drainage basin, which would provide nutrient input to establish a healthy limnological system. An organic-rich sediment was created during force-flooding, as the muskeg at the pit edges eroded into the pit. The phytoplankton population density was relatively low, and a nutrient limitation was suspected. The method for determining such a limitation is given by the Redfield ratio, the atomic ratio of carbon, nitrogen, and phosphorus, described originally in 1934. To date, this ratio remains a guide for determining nutrient limitation, which is confirmed if the ratio deviates significantly from a C:N:P ratio of 106:16:1 (Falkowski, 2000). In the pit water, the elemental composition of the water indicated that nitrogen was likely limiting growth. It was estimated that an addition of 720 kg of calcium nitrate to the 5 million cubic meters of water would likely serve to maintain or increase the biomass, thereby increasing the polishing capacity (Boojum Research Ltd., 1997; Kalin et al., 2002; Kalin & Wheeler, 2013). One to two fertilizer applications over time would suffice, since growth and decay would adjust the nutrient balance.

The situation is different in acidic water, because inorganic carbon is always limiting. Carbon dioxide is necessary for photosynthesis. As the pH of water drops, the concentration of dissolved carbon dioxide decreases. Below a pH of 4.5, very little carbon dioxide is dissolved in water. This severely limits the growth of algae in deeper water. Algae can still survive, but only relatively close to the surface of the water where diffusion from the air will provide some carbon. At pH at or below 3, algae are usually found only on rocks or surfaces with running water in creeks and streams in equilibrium with the atmosphere, as shown in Figure 15a and b.

Hence, in addition to adjusting the Redfield ratio in the water, surfaces can be created for algal attachment, adding extra renewable biological polishing capacity and organic supply to the sediment.

In one of the ecological engineering R&D projects (an abandoned copper/zinc mine), a lake enclosed in the mine waste management areas acidified with runoff and groundwater plumes from the mill site. Here, measures were implemented gradually over a 17-year period (Boojum Research Ltd., 1995). In the early years of the project, the zinc concentrations in the lake rose from 10 mg/L to 35 mg/L, while the pH decreased from 4.5 to 3. During this time, the diversity of algae growing in the lake declined from 52 to 42 different groups. The continued decrease in pH was the result of reduced iron on the sediment surface being oxidized, a cycle proceeding each year with the spring turnover. During winter months, under the ice the pH increased by about 0.5 units, only to decrease again in spring, a perfect example of the iron cycle in a lake system (Stumm & Morgan, 1996).

Acid-tolerant moss was introduced to the iron-rich lake bottom to stabilize the sediments, preventing re-oxidation of the iron. The moss consumed oxygen above the sediment surface as the biomass extended about 0.5 to 1 m into the water column. Iron precipitation occurred on the biomass above the sediment column (Figure 17; Cao & Kalin, 1999; Kalin, 1998; Kalin et al., 1989, 2001, 2006).

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Figure 17. Moss biomass covering the iron-rich sediments in the acid lakes, showing the upper green photosynthesizing portion of the moss that consumes oxygen and below, pulled up, the iron hydroxide precipitate on the biomass.

Benthic moss could survive on the lake sediments in the low pH because organic sediments contained heterotrophic microbes, which respire carbon dioxide. A large-scale pilot test with ground CPMW to fertilize the sediment resulted in rapid spread of a complete underwater meadow of moss in the lake that protected the sediment from oxygen at annual turnover. The underwater moss meadow also provided a large surface area for iron to precipitate.

The next challenge was to support the phytoplankton population to increase the polishing capacity. A multivariate analysis of phytoplankton population composition and water quality showed that pH and nickel were the main controlling factors in the years 1986 to 1992, and for one-third of the algal groups, the low pH alone was the controlling factor (Kalin et al., 2006). An increase in pH was needed, at least to pH 4.5 from the prevailing value of 3. Liming of the lake was carried out several times during mine operation, but, each time, within months, the pH and metal concentrations rebounded. Further using, lime to neutralize the lake would destroy the existing living systems, both the remaining algae in the water as well as the moss cover over the sediment, because the rapid pH change would adsorb the phytoplankton and smother the underwater meadow. It was decided to suspend magnesium scraps obtained from a refinery in the lake water (Figure 18a) because this had the potential to neutralize the lake water gradually and to raise the pH.

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Figure 18. (a) Magnesium scrap pieces covered in iron after being suspended from a barge in an acid lake. (b) Close-up photograph of corrosion channels with hydrogen bubbles, reflecting the continued corrosion.

As magnesium corrodes, it forms hydrogen gas, which consumes hydrogen ions and raises the pH. The following reaction is expected: Mg 2 + + H 2 + → Mg(OH)2 (solid) MgSO 4 (water soluble) + H2(gas); which forms hydro magnesite [MGk3.Mg(OH)2.9H2O], Nesquehonite [MgCO3.3H2O], and/or Lansfordite [MgCO3].

Most importantly, the reaction is gradual, the metal does not become coated with iron, and corrosion continues as long as the pH is acid (Song & Atrens, 1999).

In 1996, six tons of scrap magnesium were suspended from barges in the one million cubic meter lake with a pH of 3.0 (Kalin, 2001). The barges moved with the wind during the summer and the hydrogen bubbles are released to the air. In the winter of 2002, water from below the ice near the barge was recovered. The water had a pH of 8, and no bubbles were noted, as no movement of the barge had occurred.

Characeae or Stoneworts as Biopolishers in Alkaline Water

Alkaline effluents from waste rock piles and the mill at the top of a drainage basin in Saskatchewan contained elevated concentrations of radium and uranium. These effluents passed through two lakes separated by a narrow wetland. The concentration of radium was increased further by contaminated groundwater emanating through the sediments in the upper of the two lakes, which was void of any underwater vegetation. Regulatory agencies requested the construction of a chemical treatment plant. The company noted drastic decreases in radium concentrations as water passed through the wetland with floating muskeg between the two lakes (Figure 19a). Boojum was requested to assess if one of the “Boojum weeds” might be growing in the wetland.

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Figure 19. (a) Upper lake in the background right corner (with the mill building on the horizon), showing the floating muskeg area populated by the stoneworts that are removing radium 226. The lower lake was also invaded by stoneworts (not visible) and the biomass from it was transplanted by helicopter. (b) The containers with biomass harvested from the lower lake that were to be transplanted to the upper lake in transects marked by floats for monitoring under water. After several years, the upper lake was completely populated and no chemical treatment plant was constructed for the removal of radium 226.

Boojum had been evaluating the use of macrophytic algae, stoneworts (members of the Characeae family), for the removal of radium from Elliot Lake during the studies of the ecology of the uranium tailings at the Institute for Environmental Studies (Kalin & Smith, 1986).

Indeed, a stonewort, Nitella flexilis, displayed extensive growth in the narrows and the lake below. A project was launched to test the possibility that the stoneworts could replace a chemical treatment plant. Biomass in significant quantities was transplanted to the upper lake (Figure 19b).

The transplants took hold and removed uranium and radium (Kalin et al., 2002). The project was completed by 2004, when a complete underwater meadow covered the lake sediments. As of 2014, no chemical treatment plant has been constructed (J. Jarrell, personal communication, 2014).

Stoneworts were also used in Germany in a constructed pond system for effluents from the adit of an abandoned uranium mine to remove radium and arsenic (Kalin et al., 2002).

Modeling the Biological Polishing Process

The object of the dynamic model of biological polishing is to simulate the biogeochemical processes operating in the removal of contaminants from mine drainage by means of attached and free-swimming algae. The model describes the growth of the algae and the interactions the between mine drainage and any fertilizer added to support or promote the process. Eventually, it is hoped that the model will serve as a management tool for the biological polishing system, in that it will allow the user to determine the amount and time for adding an attachment area to the pond and the amounts of fertilizer needed until the system has reached a self-sustaining state. The model constructed is an attempt to bridge between the “top down” and “bottom up” approaches, combining the mechanistic theoretical perspective of the key biogeochemical processes operating in polishing ponds and the empirical approach for quantifying the complex ecological growth processes. Finally, the mechanistic/empirical model constructed must be calibrated and verified in the field. The first attempt to use the model on a 1 million cubic meter acidic lake with a turnover of 3 years had limited success (Romanin, 1994). Figure 20 gives an overview of the model with a flow diagram showing the model’s feedback loops (Kalin et al., 1993).

Mining, Ecological Engineering, and Metals Extraction for the 21st CenturyClick to view larger

Figure 20. Flow chart of biological polishing model that integrates biogeochemistry and algal growth and decay dynamics to determine the contaminant-removal efficiency that might be expected from the process.

The algal input parameters for growth are given in Box 1 (copied from Romanin, 1994). These parameters are combined with the water characteristics of the biological polishing pond. The changes that can be brought about by algal growth in the water are evaluated with PHREEQC, a model generating speciation of compounds in the lake chemistry and at the sediment–water interphase.

Toward a Sustainable Metals Extraction Technology

The metals extraction industry is now facing possibly its greatest-ever challenge, with the need to demonstrate “sustainability” in the face of dwindling reserves and grades, increased restrictive legislation, and increasing costs. To even entertain the idea of being “sustainable” in the face of being essentially nonrenewable, the industry theoretically can no longer afford to throw away up to 99% of the material it mines, the largest cost of getting the mineral-bearing rocks. Mining, as opposed to processing (grinding and concentrating), represents generally the main cost associated with any metals extraction project. There are some polymetallic ore bodies, but at most, only one or two metals are extracted, dictated by the market value of the metal.

Of the various extraction technologies, a brief mention should be made of bacterial leaching for base metals, as it is often promoted as being the “environmental solution.” This is because it uses “natural processes,” but it is in many ways worse than conventional approaches. While bacterial leaching has seen some measure of success in uranium and gold plants, attempts to apply the technology to base metals have been largely unsuccessful. An initial pilot project in Chile, comprised of a joint venture between BHP Billiton and Codelco, known as Alliance Copper, ultimately resulted in the building of a 20,000 tpa (tons per annum) copper plant (Batty & Rorke, 2006). However, the project was terminated in October 2006, having not achieved its objectives. Similar processes were tried for nickel and zinc, with similar results. The failures likely had similar causes, given that an understanding of microbial processes was lacking or was not utilized. Microbial systems are considered difficult to control.

In terms of metal extraction and recovery, the industry has a well-deserved image of being slow and conservative, although it is not totally averse to change and innovation. There have been some exciting and innovative processes developed since the end of World War II, notably:

  • Pressure leaching (leaching at higher than atmospheric pressures)

  • Ammonia-based processes for Ni, Co, and Cu pioneered by the Canadian company Sherritt Gordon (now known simply as Sherritt)

  • Oxidation of zinc sulfide concentrates in sulfuric acid, also pioneered by Sherritt

  • Oxidation of refractory gold concentrates in sulfuric acid to make the gold amenable to subsequent cyanidation leaching, pioneered by Barrick and others

  • Nickel laterites in sulfuric acid (although this has not yet been proven as being generally viable)

  • Copper solvent extraction in various forms using “designer” complex organic carbon molecules

  • CIP/CIL (carbon-in-pulp or carbon-in-leach) for gold and silver recovery, using activated (usually coconut shell) carbon to preferentially absorb the gold or silver

  • Falconbridge chlorine leach process (Falconbridge is one of a few plants that make use of chloride chemistry)

These, while in themselves highly commendable and successful, have unfortunately not addressed the basic issues confronting the industry today, namely sustainability and environmental liability. It is considered, therefore, that the industry needs to completely change its mindset and how it operates if it is to remain both competitive and at the same time to reduce or possibly to eliminate environmental liability, and be “sustainable.” Given the actual costs of mining itself (getting to the ore body and breaking the rock), and that large, rich ore bodies are no longer being found, then it surely makes both economic and sustainable sense to maximize the recovery of all metals that have value and have been mined. The question, therefore, is obvious: why is the industry not extracting more out of the mined rocks?

Throughout the 1970s and 1980s there was a great deal of almost evangelical interest in and enthusiasm for, and research into, developing hydrometallurgical processes for the treatment of (primarily) copper sulfide concentrates (Flett et al., 1983; McLean, 1982; Paynter, 1973; Wadsworth, 1984). An astonishing number of different processes emerged during this time, with so-called “out of the box” thinking. Great emphasis was placed on chloride-based processes, although several sulfate-based circuits were also conceived, together with one notable ammonia-based plant, Anaconda’s Arbiter Plant (Kuhn et al., 1974). Unfortunately, only two of the processes achieved commercial operation (CLEAR and Arbiter), and then only for a limited time, thus leading to a general suspicion of “new technologies” that continues to exist.

In 2003, at a major international hydrometallurgy symposium, the keynote paper addressed why new hydrometallurgical processes failed (Halbe, 2003). Four important aspects were highlighted, namely, (i) if any pilot-scale testing was conducted, it was to generate product, not to confirm process parameters; (ii) equipment was downsized or design criteria were made less conservative in response to projected cost overruns; (iii) process flowsheets were unusually complex, with prototype equipment in two or more critical unit operations; and (iv), somewhat surprisingly, there was a lack of understanding of the process chemistry. Consequently, very few of the processes even reached the pilot stage, with the unfortunate result that there is now an inherent distrust of any new extraction processes or technology. This distrust has only been enhanced by the failure of four HPAL (high-pressure acid leach) projects in Western Australia since the mid-1990s, and more recently, the failed hydrochloric acid regeneration plant of SMS Siemag at the (formerly) Thyssen Krupp steel plant in Calvert, Alabama.

Estimating the Full Extraction Potential of Mined Rock

Referring to the fact that mining costs represent a large and significant part of any overall metals or industrial minerals project cost, a hypothetical example is given below where it would make both economic and environmental sense to maximize the recovery of all metals that have value and have been mined, a desirable step toward sustainability of the mining industry. Consider that a nickel laterite, with a composition of 1.2% Ni, 0.1% Co, 5% Al, 15% Mg, and 30% Fe is processed with 90% recovery of Ni, Co, and Fe, and 75% recovery of Al and Mg. Taking prices (in dollars) of $5/lb for Ni, $10/lb for Co, $0.2/lb for Al2O3, $40/ton for Fe2O3, and $50/ton for MgO, the following revenues are generated for a plant nominally producing 50,000 tons of LME (London Metal Exchange) grade Ni:

  • Ni—$550 million

  • Co—$90 million

  • Al2O3—$150 million (350,000 tons)

  • Fe2O3—$70 million (1.8 million tons)

  • MgO—$30 million (610,000 tons)

By this analysis, the revenues of the project could be increased significantly over those generated simply by nickel (and cobalt). Furthermore, there are additional benefits in that there are close to 2 million tons of residues (equivalent to approximately 50% of the material originally mined) that will not have to be disposed of, and hence an appreciable reduction in mining wastes. There are also indirect savings and benefits, particularly from an environmental viewpoint, in that water is saved because it is not used for mining the equivalent tonnage of Al, Fe, or Mg from a primary ore body, such as bauxite, iron ore, or magnesite/dolomite mines, and the tailings that would necessarily be produced from such mining would no longer be generated. Furthermore, there is always a premium for high-grade hematite, which does not need further processing, which also occurs in nickel laterite deposits, so that the revenues to be derived for this hypothetical mine are probably significantly understated.

However, in the context of the illustration, the actual prices are irrelevant, since the objective is simply to demonstrate the points that these values have been mined, but, with traditional processing methods and especially mindsets, they are not only not being realized, but also are being disposed of, thereby creating an environmental problem that can, and should, be avoided (Dry, 2015).

Barriers to Higher Recovery of Metals From Mined Rock

The industry, because of the past failures noted above, and being generally reluctant to embrace any sort of risk or major change, has standard arguments formulated against such an approach for all or more metal extraction.

  1. 1. The technology to achieve the recoveries in sufficiently pure form does not exist.

  2. 2. If it did exist, then it would be too expensive and difficult to implement, especially as a retrofit, that is, into the existing process equipment.

  3. 3. The existing markets could not absorb additional tonnages.

For the case of iron and aluminum, the amounts generated from the hypothetical laterite project are sufficient to operate a stand-alone steel mini-mill and aluminum smelter. This ought to be attractive in an established and diverse mining area, such as Western Australia. There is clearly sufficient aluminum associated with the Western Australian laterites to sustain the existing aluminum smelters, resulting in less bauxite needing to be mined.

For environmental technologies, that is, technologies that deal with existing tailings ponds, and especially those based on ecological engineering principles, points 1 and 2 apply. The fact that microbial processes can be controlled is generally not understood, and hence not accepted, by the industry, largely because it is not within “normal” paradigms, which are largely based on the principles of classical inorganic chemistry. Such ecologically based technologies, which would reduce or completely halt the weathering rate at source, are generally either ignored or are declared to be impossible or uneconomic, unfortunately with no real basis for such declarations. Further, without solid geomicrobiological knowledge, bioleaching and biological oxidation/corrosion control cannot be implemented. This know-how is often lacking in bioleach operations. There is, nevertheless, in this respect an opportunity for the mining industry to embrace a technology (biological or ecological engineering) that could have far-reaching benefits. The industry needs to take the risk.

In the contexts of both efficiently recovering more value from the material mined, and at the same time appreciably reducing the amounts of toxic wastes generated, revisiting chloride processing as it was originally conceived in the 1970s has merit. In recent times, chloride-based extraction has been further developed and refined, with a greater understanding of the parameters involved, and with the objectives discussed above in mind, to the extent that it can now be considered a viable option.

Modern Chloride Extraction

Despite a track record of “nonsuccess” of the many chloride-based processes for sulfide feeds, there remain many compelling reasons why the application of chloride chemistry not only can result in improved processing, but also can contribute greatly to achieving sustainability and improved environmental performance. There have been, and still are, a few very successful chloride-based base metal operations. Glencore (formerly Falconbridge) has operated a chloride process for many years at its nickel-cobalt refinery in Kristiansand, Norway, which is arguably the best base metal recovery plant in the world today. It was initially a hydrochloric acid leach, but more lately has used chlorine as the main oxidant/lixiviant, largely because chlorine is available, being generated from the subsequent electrowinning circuits for the recovery of nickel and cobalt metals (Stensholt et al., 1986a, 1986b, 1988; Thornhill et al., 1971). Noranda (as it then was) operated the Brenda Leach Process, which employed a high temperature (105°C to 110°C), high-strength chloride (30% CaCl2 + NaCl + HCl) atmospheric leach of copper-molybdenum concentrates until the mine shut down in the 1990s (Jennings et al., 1973). This process was highly efficient, and essentially leached out all the copper, lead, and calcium from molybdenum concentrates to allow further, conventional processing of molybdenum to take place, without generating large quantities of toxic residues.

Most of the advantages that were originally expected from the use of chloride (Harris, 2014) with the processes developed in the 1970s and 1980s have remained, and are briefly discussed below.

Leaching and Intrinsic Energy Content of Sulfides

Intrinsic energy can be substantially recovered, especially if the feed contains appreciable levels of pyrrhotite (Harris et al., 2007). The presence of pyrrhotite in an ore or concentrate of all the iron sulfide minerals is generally regarded as a major disadvantage, as it can be a major factor in acid mine drainage. This is equally true for sulfate leaching, but the advantage that chloride has is that the acid used to effect leaching can be recovered and recycled (see iron below) which is not the case with sulfate. Processing sulfide minerals in this way to recover energy as heat is environmentally advantageous, since for every GJ of heat recovered, an equivalent amount of burning carbon is prevented, and a toxic, acid-generating waste is eliminated. Chloride circuits can be operated at atmospheric pressures, and are more readily adjusted to ensure that the sulfide-sulfur that accompanies the mineral either ends up as H2S gas, or as elemental sulfur, a product that can also be sold, and both forms can be converted to sulfuric acid.

Chloride extraction circuits are more aggressive than their sulfate counterparts. This has the advantage that higher metal recovery can be achieved, along with a residue that is often easier to filter, that has lower volume, and that is less prone to metal/acid leaching into the environment. Indeed, most chloride leach residues are predominantly benign alumino-silicate gangue.

Iron and Hydrochloric Acid

Iron is the major contaminant in virtually every hydrometallurgical processing circuit, and has been deemed worthy of five international conferences devoted entirely to its control and disposal. Sulfate chemistry is such that iron must be precipitated via the use of some form of a base or neutralizing agent, generating large volumes of sludges, whether they be jarosite, goethite, hematite, or “ferric hydroxide.” However, chloride chemistry affords the possibility not only of recovering the associated acid for reuse, but also of generating a marketable iron product, namely hematite. At the very worst, this hematite is easy to filter, has a low volume, and is environmentally benign.

Aluminum and Magnesium

Because of the highly aggressive nature of the chloride leaching operation, both aluminum and magnesium tend to report to the resultant leach filtrate in significant concentrations. During acid recovery, through hydrolysis of the iron chloride, the aluminum reports virtually 100% along with the hematite. However, the different crystal structures of the two oxides result in discrete compounds, allowing easy separation of the aluminum. Magnesium, on the other hand, remains in the liquid phase when either iron or aluminum is present, thus affording an efficient and simple separation. It can be recovered in a subsequent hydrolysis step as a magnesium oxychloride, which can be heated and results in a marketable magnesia.

Environmental Aspects

Because chloride is so aggressive, as noted above, it tends to dissolve all metals from the ground rock. Thus, leach residues, which in conventional processing are generally voluminous and must be ponded as tailings, are low in volume, are generally crystalline, and, most importantly, are no longer reactive and hence are environmentally nonthreatening.

Current State of Development

Considerable development work has been undertaken on the chloride-based process over the past decade. Ideally, the mining and metals extraction industry will consider embracing what is essentially a quantum change in how it goes about its business, and at the same time will overcome the negative perceptions created in the past due to the many failed processing routes.

The Way Forward

Ecological and hydrological processes can serve as cornerstones for mine waste and waste water management. Treating mine waste management areas as ecologically primitive ecosystems provides an alternate view to address mine waste management. Applying existing, but neglected, hydrometallurgical processes will lead to more efficient mineral extraction, reducing the land requirement for mining wastes. The urgent need to change paradigms is highlighted by global estimates of water usage and mine waste generation, along with the anticipated collision with competing requirements for available water and arable land. Water, mining, and agriculture are on this collision course (Sudbury, 2013).

The authors of this article all have lifelong experiences in mining, waste management, and hydrometallurgy, a “sine qua non” for contributing to an encyclopedia. The authors have built the needed bridges between these three areas—mining with its waste generation, water use, and extraction technologies. In all three areas, a paradigm shift is needed if mining is to become a more efficient and sustainable industry.

Our approach to mine waste management, regarding the areas as extreme environments dominated by microbes, fungi, and algae, has been met with considerable skepticism both by academe and by the engineering fraternity. Environmental regulations focus on restoring these sites to productive ecosystems, grassland, and forests. This is possible in some cases, but long-term sustainability of these efforts is not achieved, mainly because of a lack of attention to indigenous biota and hydrological constraints. In principle, any rehabilitation measures that ignore the activity of microbes on rock surfaces and particles in water, soil, and air are relatively short-lived, as microbes play a pivotal role in weathering and altering of rock and ecosystems.

The first step with mine wastes is to establish a microbiological balance in the rudimentary primitive ecosystems between oxidative and reductive processes, particularly with regard to sulfides and iron. Lewontin (1983, p. 280) postulated: “Organisms do not adapt to their environments; they construct them out of the bits and pieces of the external world.” This thesis has been long debated and was defended by Odling-Smee (1988) and Odling-Smee et al. (2003) after it had been defined as niche construction. It is an essential part of evolutionary biology, and it is relevant for ecological functions or processes.

The recent trends in genomics promise panaceas for the afflictions of many industries and the planet, including pollution control. Weigel and Tautz (2015) have provided an overview of areas where genomic methods are applicable. For ecosystems, they predict it will take at least a decade before any reasonable understanding of ecosystems and their microbial interactions is achieved. Such an understanding is required to support ecosystem services and stability. This will certainly apply to extreme ecosystems. The application of genomics methods needs to integrate the understanding of niche construction, or the functions of biofilms, as they are part of an ecological and evolutionary process. After several decades, the empirical ecological engineering approach used by Boojum Research has alternated between field and laboratory tests. These empirical results, along with progress in our understanding of the geomicrobiology of extreme ecosystems, are replacing the old skepticism with a new understanding of the roles of microbes. This is also gradually gaining acceptance in the industry. Ecologists are now promoting empirical scientific approaches to understand and solve urgent environmental problems (Barley & Meeuwig, 2017). Bridging the gap between engineering and ecology has been a challenge that Boojum has failed to achieve, but bridges are being built (Madsen, 2011). Madsen differentiates five stages in the development of biotechnological applications. These stages need to be carried out to understand complex processes in a natural system. Stages 1 to 3 consist of finding, confirming, and characterizing the microbial activity, with the fourth stage being field testing, the stage which most of our processes have reached. It is now up to industry to approach the fifth stage, a biotechnological approach that includes ecological engineering.

This could be realized in part by an evaluation of the R&D field demonstration projects of Boojum. Barley and Meeuwig (2017) claimed that formulating and testing a hypotheses at an ecologically realistic scale with large-scale unreplicated natural experiments (LUNEs) had a disproportionately positive effect on conservation policy, providing powerful insights into cosmology, evolution, and geology. This, despite some initial resistance.

The empirical approach of ecological engineering summarized herein essentially defines processes that assist in stabilizing mine waste and water management areas. The success of this approach has been achieved by confirming results at different scales, in the laboratory and simultaneously with small field pilots. The field tests have been gradually scaled-up, ultimately combining the processes as dictated by the site topography. Our ecological approach focuses on retention of contaminants within waste management areas, detoxifying the areas, and returning them to something approaching their original state in the environment.

Hydrometallurgical methods can and should extract more metals out of the mined rocks, which would drastically reduce the global footprint of land consumption. We alert both industry and academe not to use the “more research is needed” excuse to delay, change, or slow the needed paradigm shift. Already, the global consequences of inaction are evidenced in the dramatic changes in weather, frequent floods, droughts, and water shortages, all factors affecting mining operations intensely. “More research is needed” has its place, but only when integrated with a LUNE or large-scale waste management areas, which need the support of the mining industry.

More effective freshwater reclamation methods should be applied in mining operations, possibly using the example of the clean-up of the Rhine River, where the same water quality that is taken in must be returned to the environment (Malle, 1996). To resolve the global issues facing the mining industry in waste and water management, the mainstream science and engineering approaches, which emphasize statistics and replication, will not support the way forward. The answer to achieving a sustainable decommissioning methodology lies in large-scale testing of processes that will lead to a self-designing ecosystem in which elements are recycled. Perhaps this could be achieved with the creation of several LUNEs similar to the Experimental Lakes Area in Ontario (Schindler et al., 1980), or the Experimental Forest Areas l Hubbard Brook Agricola’s dilemma could be addressed.

We hope the mining and metals extraction industry will consider embracing what is essentially a quantum change in approaching the future, will bury the conventional approaches to waste and water management, and will be open to ecological engineering as a means of dealing with the remaining tailings and waste rock. We believe this will improve the public perception and provide a more socially and environmentally acceptable horizon for what is essentially the world’s oldest industry. By combining all three facets of mining, an increase in efficiency of mineral extraction, water use, and reclaiming mine waste management areas, a synergy could be created to lead to a more sustainable and economical mining industry in the 21st century.

Acknowledgment

For the ecological field work on the mine waste management areas, many mine managers, too numerous to name, have been instrumental and we are grateful for their support and insight provided to work safely on the sites. Knowledge about mine operations generating the waste rock piles and the hydrometallurgical processes which reject barren liquors and tailings is necessary, and we received much guidance. In the past decades both on the scientific understanding and the practical engineered handling of mine wastes has progressed dramatically. However, between the many disciplines needed in mine waste and water management (e.g., geology, hydrology, mineralogy and microbiology and others) a chasm exists to the engineering approaches. This chasm is hindering progress toward a more sustainable approach. With the contributions of Michael P. Sudbury and Dr. Bryn Harris we have provided a long-term perspective, integrating the industries water consumption and waste production globally with its land consumption. The ecological engineering approach, fundamentally nothing revolutionary from an ecologist perspective, gains relevance given the scale, dimensions and urgency of the potential collisions of the natural resource industry. Although we are co-authors, we do not claim to have the expertise all three areas, insight in the industries resource use, the ecological processes and the hydrometallurgy.

Further Reading

Cairns, J., Jr. (2005). Ecological overshoot and ecological restoration. Asian Journal of Experimental Science, 9(2), 1–12.Find this resource:

Eisenstein, C. (2011). Sacred economics: Money, gift, and society in the age of transition. North Atlantic Books.Find this resource:

Jorensen, S. E. (2008). Ecological engineering: Overview. Encyclopedia of Ecology (pp. 1024–1027). Academic Press.Find this resource:

Leland, K., Mathews, B., & Feldman, M. W. (2016). An introduction to niche construction theory. Evolutionary Ecology, 30, 191–202.Find this resource:

Mudd, G. M. (2007). Global trends in gold mining: Towards quantifying environmental and resource sustainability? Resources Policy, 32(1–2), 42–56.Find this resource:

Pattern, C. B. (2014). Systems ecology and environmentalism: getting the science right. Ecological Engineering, 65, 15–23.Find this resource:

Quatrini, S., Barkemeyer, R., & Stringer, L. (2016). Involving the mining sector in achieving land degradation neutrality. Solutions Journal, 7(5), 55–63.Find this resource:

Sienkiewicz, E., & Gasiorowski, M. (2016). Evolution of a mining Lake—From acidity to natural neutralization. Science of the Total Environment, 557–558, 343–354.Find this resource:

Steinberg, C. E. W., & Paul, A. (2008). Photolysis: Ecological processes. In: S. E. Jorgensen & B. S. Fath (Eds.), Ecological processes (pp. 2724–2732). Volume 4 of Encyclopedia of ecology. Oxford: Elsevier.Find this resource:

Vitousek, P. M., Mooney, H. A., Lubchenco, J., & Melillo, J. M. (1997). Human domination of Earth’s ecosystems. Science, 277(5325), 494–499.Find this resource:

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