Remediation Opportunities for Arsenic-Contaminated Gold Mine Waste

Abstract

Arsenic (As)-contaminated gold mine waste is a global problem and poses a significant risk to the ecosystem and community (e.g., carcinogenic, toxicity). Arsenic concentrations of 77,000 mg/kg and 22,000 mg/kg in mine waste have been reported for Canada and Australia, respectively. Research is investigating environmentally sustainable techniques to remediate As-rich mine waste. Biological techniques involving plants (phytoremediation) and soil amendments have been studied to bioaccumulate As from soil (phytoextraction) or stabilise As in the rhizosphere (phytostabilisation). Identified plant species for phytoremediation are predominately fern species, which are problematic for arid to semi-arid climates, typical of gold mining areas. There is a need to identify native plant species that are compatible with arid to semi-arid climates. Arsenic is toxic to plants; therefore, it is vital to assess soil amendments and their ability to reduce toxicity, enhance plant growth, and improve soil conditions. The effectiveness of a soil-amending phytoremediation technique is dependent on soil properties, geochemical background, and As concentrations/speciation; hence, it is vital to use field soil. There is a lack of studies involving mine waste soil collected from the field. Future research is needed to design soil-amending phytoremediation techniques with site-specific mine waste soil and native plant species.

1. Introduction

Arsenic (As) is an element with metalloid characteristics which occurs naturally and is the 20th most abundant element in the Earth’s crust [1]. Due to anthropogenic activities, such as gold mining, As concentrations in the environment are elevated above naturally occurring levels [2]. In legacy gold mining areas, abandoned As-rich mine waste dumps act as ongoing sources of As to the surrounding environment. Mine waste such as tailings and grey sands have polluted landscapes through mechanisms such as soil erosion as a result of a lack of vegetation cover and As leaching [3]. This is a concern because arsenic is toxic to all living organisms, is carcinogenic, and disrupts ecosystem functions, for example those involved in soil properties, geochemical cycles, and water acidification [4,5]. As such, to reduce the burden posed by arsenic from mine waste, there is a need to identify and explore sustainable remediation techniques that alleviate environmental pollution whilst also restoring ecosystem functioning. It is vital to restore ecosystem functions because natural fertile soil takes thousands of years to form [6]. Currently, in some regions of the world, 50 tonnes (t) per hectare of fertile soil has been lost to degradation [6]. In India, mined land equates to approximately one-third of the agricultural land [4]. It is essential to restore degraded soil with increased population growth and increasing demand for fertile agricultural soil [6]. In addition, soil organic matter stores more organic carbon (C) than the atmosphere and global vegetation combined [6,7]. Increasing soil organic matter and sequestering soil carbon in degraded soils will improve key soil fertility parameters, such as nutrients, microorganisms, and soil structure, which supports living organisms and biodiversity [7,8,9,10,11,12]. Biological remediation techniques involving plants and soil amendments are environmentally sustainable methods; revegetation of barren mine waste sites also reduces soil erosion, mitigating further contamination. Phytoremediation is the use of plants to remediate soil. Plants can either bioaccumulate As (phytoextraction) or stabilise As in the rhizosphere (phytostabilisation) [13,14]. However, As is toxic to plants; for example, arsenic-contaminated soil and pore water can have detrimental impacts on the agriculture sector, such as rice production [15]. Rice is a staple food for many. A study by Muehe et al. [15] investigated rice production under climatic and As stress. Climatic stress (38 °C/850 ppmv CO₂) and As stress caused a 16% and 39% yield reduction, respectively [15]. Soil amendments are vital to reduce toxicity and improve plant productivity [13]. Discovering new and effective remediation techniques for legacy mine waste has never been more important for the environment and world food security.

This review aims to identify plant species compatible with an arid to semi-arid climate, which can survive elevated As concentrations ranging from 1000 to 2300 mg/kg, a common As concentration for grey sands [16,17]. This review also investigates soil amendments and their abilities to enhance plant survival, phytoremediation and their soil-ameliorating capabilities. A novel approach to this review is to also investigate the role of As speciation in combination with arsenic remediation, soil amendments, toxicity, and restoring ecosystem functions. There are many reviews addressing total reclamation of mine waste sites, but additional assessment of As speciation is limited. The information obtained from this review will be used to identify the knowledge gaps in soil amendment phytoremediation studies and future research opportunities.

1.1. Gold Mining and Refining

In 1848, the great gold era was born in California, following further discoveries primarily in North America, Australasia, and South Africa [18,19]. The early 19th century began with a yearly output of 10–20 t of gold, but by the mid-19th century it was 180 t, and by the end of the century it reached 450 t. Gold oxides are rare, and gold is usually found in the form of a mineral and therefore requires extraction [20]. For example, arsenopyrite (FeAsS) and arsenian pyrite (FeS₂) host gold in the crystal structure, and extraction techniques are required to isolate gold [21,22]. The most effective extraction operation, where approximately 95% of gold is recovered, is cyanidation of free milling ores such as quartz vein gold ores (Figure 1) [20]. However, free milling ores quickly became exhausted, and mining operations focused on refractory ores [23], which required new technology because cyanidation alone was not adequate for effective gold extraction [20]. Refractory ores are characterised by carbon, sulphide, and tellurium ores or a combination, sulphide being the most common (Figure 1) [20,23]. A primary sulphide mineral that hosts gold is arsenopyrite; extraction is complex and requires additional treatments [24]. Additional treatments include fine grinding for semi-refractory ores and thermal treatments for refractory ores such as roasting and pyrolysis (Figure 1) [24]. Roasting was common because it was cost effective. Sulphides are oxidised, and gold is rendered into a leachable state [20,24]. However, this process creates arsenic emissions such as arsenic trioxide (As₂O₃) dust through precipitation of arsenic vapours [21]. Post-precipitation, arsenic trioxide dust is deposited on the surrounding environment, can be found in soil particle sizes as low as <20 µm, and is highly bioavailable [21]. Relative bioavailability is the amount of a toxic substance that can enter the human/animal circulatory system relative to a reference compound, such as sodium arsenate [17,25]. Bioavailability is vital in understanding and evaluating toxicity effects of arsenic-rich soil if it is consumed by a living organism [25].

Future mining prospects are led by the worlds demand for new technology such as phones, computers, electric vehicles, and solar panels, all of which require metals. In 2017, there were approximately 7.7 billion mobile phone subscriptions, 48% of global households owned a computer, and global commitments to mitigating climate change focused on electric cars [26]. Advancements in electric technology and batteries heavily rely on mining of raw elements. For example, cobalt is an element mined to produce batteries, e.g., nickel–cobalt oxide batteries and lithium–cobalt oxide batteries [27]. Global climate change agreements could see cobalt demand increase 500% by 2025 for the production of 100 million electric cars by 2030 [26]. The majority of cobalt deposits were discovered in the Democratic Republic of the Congo (DRC), and a mining boom, resembling the gold rush, has developed [26,27]. Approximately 60% of global cobalt originates from the DRC, and cobalt is considered the “oil” of a modern-day low-carbon future [26]. There are numerous concerns regarding cobalt mines in the DRC, such as land degradation, humanitarian rights, human health, and pollution [27]. It is important to identify remediation techniques not only for legacy mine waste sites but also to provide sustainable solutions for the future mine waste that is to come because of our global demand for technology.

1.2. Environmental Impacts

Gold mining operations transformed landscapes and devastated local environments [28]. Deforestation was widespread, and open pits left crater-like features. Arsenic and other potentially toxic elements (PTEs) leached into fresh water sources, and As-rich waste in the form of slag (a byproduct of smelting) littered the landscape (Figure 1) [24,25].

Gold mining resulted in mine waste, characterised as tailings, grey sands, and calcinated soil [17]. Tailings are enormous congregations of crushed rock derived from crushing ore to release minerals [17,28,29,30]. Tailings are typically yellow in colour, can be dry soil mounds or in the form of a settling pond, and can increase arsenic concentrations in the environment [31]. The term “tailings” is commonly used to describe generic mine waste material, and therefore these tailings can vary in As concentrations; from 318 to 76,500 mg/kg [32]. Grey-coloured mine waste is called grey sands, or grey slimes/sludges, and are a fine silty texture [4,17]. Grey sands are a byproduct from fine crushing of ore and accumulate at the bottom of settling tanks during mineral separation (Figure 1) [17,28]. Arsenic concentrations in grey sands usually range from 1000 to 2300 mg/kg [17]. Calcinated soils are maroon in colour and are the byproduct of thermal treatment (roasting at 650–700 °C) and have high As concentrations, e.g., 15,200 mg/kg (Figure 1) [17,24,33].

2. Arsenic Speciation and Toxicity

2.1. Speciation

Arsenic is the 20th most abundant element found in the Earth’s crust, at concentrations of up to 2.5 mg/kg [1]. Elements in the soil are divided into two groups, essential and non-essential elements [34]. Essential elements include manganese (Mn) and zinc (Zn) and are known as micro-nutrients [34]. Non-essential elements such as As and lead (Pb) are toxic to plants, animals, and humans at high concentrations (>100 mg/kg As) and are referred to as toxic elements [5,34]. Arsenic commonly occurs in four oxidation states, As⁵⁺, As³⁺, As⁰, and As⁻³. The two most common oxidation states are the pentavalent form As⁵⁺ (arsenate, H₃AsO₄ in solution) and trivalent form As³⁺ (arsenite, i.e., arsenic trioxide as a solid), with arsenite being the most toxic to humans [4,35,36]. The microbial community influences As speciation through reduction, oxidation, and methylation processes (Figure 2) [37]. Inorganic arsenate is microbially reduced to arsenite through phosphate transporters, and arsenate-respiring microbes (anaerobic respiration) release arsenite from arsenate-rich sediments, thus increasing As toxicity [35]. Microbes can also methylate arsenic to methylarsonic acid (MMA) and a variety of dimethylated and trimethylated arsenic forms (DMA, TMA) (Figure 2) [35]. Heterotrophic and chemoautotrophic microorganisms such as Alcaligenes faecalis and Hydrogenophaga sp. are known to oxidise arsenite to a less toxic state, i.e., arsenate [35]. Microbial activity releases As atmospheric emissions, such as arsines (AsH₃), which are then oxidised to a non-volatile species allowing for As-rich sediment deposition [38]. Microbial reduction, oxidation, and methylation influence toxicity of (in)organic arsenicals. For example, arsenate reduction increases toxicity, DMA⁵⁺ is approximately 100-fold less toxic than As³⁺, and oxidation of arsenite decreases toxicity [35].

Speciation has gained attention because the oxidation state of arsenic can influence its behaviour in the environment, such as solubility, reactivity, and bioavailability (i.e., toxicity) [22,39,40,41]. As stated earlier, arsenic trioxide is highly bioavailable and the most toxic; it is a fine dust that can be easily inhaled or accidentally ingested [21]. Soil characteristics such as pH also effects As bioavailability; for example, increasing soil pH to alkaline levels (pH > 7) induces As mobility and pore water leaching [13,36,42,43]. Identifying soil characteristics, soil microbes, and As species is crucial for developing appropriate remediation techniques [22,39,40,41].

2.2. Arsenic and Human Health

Arsenic is a toxic element that has carcinogenic capabilities which is dependent on the As species [38,44,45,46,47]. Arsenic is a Category 1 carcinogen according to the International Agency for Research on Cancer and is associated with bladder, skin, and lung cancer [48]. Other species, depending on exposure, may cause skin lesions, cardiovascular implications, gastrointestinal complications, reproductive difficulties, nervous system disorders, and neurological abnormalities [1,38,44,45,46,47,49,50]. Pearce et al. [51] investigated the correlation between As-contaminated soil and cancer in the Victorian goldfield regions. Lower socioeconomic regions of the goldfields had higher cancer probability in relation to elevated soil As concentrations [51]. The most common cancers associated with soil As were colon, prostate, leukaemia, and melanoma for males and melanoma for females [51]. Arsenic exposure is of concern because gold mine waste still exists near residential land and poses a risk to human and ecosystem health.

3. Mitigating Arsenic-Contaminated Mine Waste

3.1. Global Arsenic Concentrations

Increasing arsenic concentrations are a global problem; countries such as Canada, USA, UK, Poland, Australia, New Zealand, China, Bangladesh, and India report high arsenic contaminations (Table 1) [39]. Arsenic contamination worldwide is present in soil, water, drinking water, groundwater, animals, food products, and more [39]. Arsenic background levels (uncontaminated soil) in soils worldwide have been reported to range from 0.4 to 40 mg/kg [2].

In the Canadian Northwest Territories, the Giant Mine, which was operational from 1949 to 1999, roasted Au-bearing arsenopyrite ore [21]. Poorly monitored roasting of arsenopyrite caused the release of approximately 20,000 t of As-contaminated emissions and the production of dust containing approximately 80% arsenic trioxide [21]. Bromstad et al. [21] analysed soil outcrops and reported average As concentrations of 1550 mg/kg with a maximum of 5760 mg/kg; only 3 out of 40 samples met the Canadian soil quality guidelines for industrial soils (Table 1). Bromstad et al. [21] also analysed soil cores (0–37 cm) and reported that surface soil (0–6 cm) presented the highest As concentrations of 7680 mg/kg, and soil at 6–15 cm depth had 323 mg/g of As. On the east coast of Canada in Nova Scotia, there were 64 mining districts that produced 34 t of gold and 3 million t of tailings during 1861–1942 [32,52]. Meunier et al. [52] and Walker et al. [32] reported similar As concentrations in Nova Scotia tailings, ranging from 200 to 77,000 mg/kg and 318 to 76,500 mg/kg, respectively (Table 1). The Canadian Council of Ministers of the Environment (CCME) developed the Canadian Soil Quality Guidelines for the Protection of Environment and Human Health 1997 document for inorganic arsenic. The soil quality guideline for agricultural, residential/parklands, and commercial/industrial soils are 12 mg/kg (Table 1) [53]. Mine tailings have been a problem for decades, possibly longer, and are significantly exceeding the soil quality guidelines in Canada.

In England, Cornwall was a major producer of arsenic, tin, and copper during 1860–1900 and smelted imported arsenical ores [54]. Caille et al. [55] conducted a pot experiment with Cornwall soils containing elevated As concentrations from mining and natural sources such as As-rich bedrock. Soil samples were collected from six sites at Camborne at a 20 cm depth then dried and sieved to <2 mm. The six sites had As concentrations of 67 ± 2357 ± 15,417 ± 2361 ± 2475 ± 2, and 4550 ± 52 mg/kg (Table 1) [55]. Rieuwerts et al. [56] also investigated As concentrations and human health exposure in Cornwall by analysing household garden soil and dust. Arsenic concentrations in household garden soil and dust ranged from 23 to 471 and 43 to 486 mg/kg, respectively (Table 1) [56]. The Department for Environment, Food and Rural Affairs and the Environment Agency (DEFRA EA) have developed a Soil Guideline Value to limit exceedances and human exposure [57]. The soil guideline value for As is 20 mg/kg, and all As concentrations at Cornwall reported by Caille et al. [55] and Rieuwerts et al. [56] exceed 20 mg/kg. In Portugal, Da Silva et al. [58] investigated As concentrations at Castromil mine because of concerns for local communities living in the city of Porto, 23 km east from Castromil mine. One hundred and six soil samples were collected at a depth of 15 cm, and As concentrations ranged from 31 to 6909 mg/kg (median: 273 mg/kg As) (Table 1) [58].

Gold mine waste with elevated As concentrations is detrimental to human and ecosystem health and has been abandoned, leaving it for future generations to restore or mitigate the toxic effects. In Southwest Poland, the Zloty Stok mine was studied by Antosiewicz et al. [59]. Soil and plant material were collected and analysed for elemental concentrations to identify phytoremediation characteristics [59]. Three sites were chosen for sampling, a former mining site, a smelter site, and a sludge deposit site that has naturally been covered with a soil layer [59]. Soil samples resulted in As concentrations ranging from 860 to 5751 mg/kg (Table 1). The Czech Republic is also greatly affected by historic mining activities in the northwest; Jáchymov is home to the Giftkies arsenic deposit [60]. Drahota et al. [60] investigated contaminated soils and reported As concentrations ranging from 74 to 3697 mg/kg (Table 1). The European Union recommends that As soil concentrations should not exceed 20 mg/kg for agricultural soils [61,62].

In Australia, Martin et al. [63] studied Au mining sites in Bendigo, Ballarat, and Mount Egerton; samples taken from the three sites were combined and homogenised. Martin et al. [63] reported As concentrations ranging from 1210 to 22,000 mg/kg (Table 1). The National Environment Protection Measures (NEPMs) for the Assessment of Site Contamination are national guidelines used to investigate soil and groundwater contamination in Australia for human and ecosystem health [5,64]. The NEPM Health-Based Investigation Level A (HIL A) provides a standard for residential land with a garden and easily accessible soil [5,64]. The NEPM HIL A reports that As concentrations should not exceed 100 mg/kg, and, if this limit is reached, contaminated soil requires further investigation [5,64]. Arsenic concentrations found in regional Victoria at some mine waste sites exceed the HIL A by a minimum of 10-fold.

Table 1. Global arsenic concentrations in mine waste and mine-impacted soil.

Site	Soil Type	As (mg/kg)	Legislated Values (mg/kg)	References
Giant Mine, Yellowknife, Northwest Territories, Canada	Soil outcrops	156–5760 (mean 1550)	12	[21,53]
Giant Mine, Yellowknife, Northwest Territories, Canada	Soil cores	11–7680 (mean 177)	12	[21,53]
Nova Scotia Gold Mining District, Canada	Tailings	200–77,000	12	[52,53]
Meguma Terrane, Nova Scotia, Canada	Tailings	318–76,500	12	[32,53]
USA	Soil	162–12,483 (median 73)	5–20	[65,66]
Morro do Ouro, Minas Gerais State, Brazil	Soil	26–699		[67]
Delita, Cuba	Tailings	1085–35,372		[68]
Cornwall, England	Soil	67–4550	20	[55,57]
Cornwall, England	Household garden soil	23–471 (mean 262)	20	[56,57]
Cornwall, England	Household dust	43–486 (mean 149)	20	[56,57]
Southern, Tuscany, Italy	Soil	5–2035		[69]
Castromil Mine, Portugal	Soil	31–6909 (mean 820)		[58]
Zloty Stok Gold Mine, Poland	Soil	860–5751		[59]
Giftkies Mine, Jáchymov, Czech Republic	Soil	74–3697		[60]
Hillgrove Mine, New South Wales, Australia	Soil	826–1606	100	[5,70]
Victoria, Australia	Soil	9–9900	100	[5,49]
Victoria, Australia	Mine waste	1210–22,000	100	[5,63]
Macraes Mine, South Island, New Zealand	Mine waste	65–98,300		[71]
Dandong, Liaoning Province, China	Soil	1944		[72]

Table 1. Global arsenic concentrations in mine waste and mine-impacted soil.

Site	Soil Type	As (mg/kg)	Legislated Values (mg/kg)	References
Giant Mine, Yellowknife, Northwest Territories, Canada	Soil outcrops	156–5760 (mean 1550)	12	[21,53]
Giant Mine, Yellowknife, Northwest Territories, Canada	Soil cores	11–7680 (mean 177)	12	[21,53]
Nova Scotia Gold Mining District, Canada	Tailings	200–77,000	12	[52,53]
Meguma Terrane, Nova Scotia, Canada	Tailings	318–76,500	12	[32,53]
USA	Soil	162–12,483 (median 73)	5–20	[65,66]
Morro do Ouro, Minas Gerais State, Brazil	Soil	26–699		[67]
Delita, Cuba	Tailings	1085–35,372		[68]
Cornwall, England	Soil	67–4550	20	[55,57]
Cornwall, England	Household garden soil	23–471 (mean 262)	20	[56,57]
Cornwall, England	Household dust	43–486 (mean 149)	20	[56,57]
Southern, Tuscany, Italy	Soil	5–2035		[69]
Castromil Mine, Portugal	Soil	31–6909 (mean 820)		[58]
Zloty Stok Gold Mine, Poland	Soil	860–5751		[59]
Giftkies Mine, Jáchymov, Czech Republic	Soil	74–3697		[60]
Hillgrove Mine, New South Wales, Australia	Soil	826–1606	100	[5,70]
Victoria, Australia	Soil	9–9900	100	[5,49]
Victoria, Australia	Mine waste	1210–22,000	100	[5,63]
Macraes Mine, South Island, New Zealand	Mine waste	65–98,300		[71]
Dandong, Liaoning Province, China	Soil	1944		[72]

3.2. Remediation Techniques

Mine waste is often unrehabilitated, and low-grade ore waste, dumped leachates, residues, and acid mine drainage are left to contaminate the environment [64]. Little, if any, environmental policies and regulations existed during the gold mining boom [64]. Today, there are physical, chemical, and biological remediation techniques (Table 2).

3.3. Arsenic Remediation Techniques

3.3.1. Physical and Chemical

Traditional methods for remediating potentially toxic elements (PTEs) from contaminated soil involve physical and chemical techniques. Physical techniques include excavation, soil washing, diluting with clean soil, deep ploughing to mix surface soil with subsurface soil, and extracting soil for relocation to landfill (Table 2) [16,73,74]. Currently, physical mine waste disposal, i.e., tailings dams, is common practice. For example, in Turkey, 71.2% of total mine waste was sent to spoil locations, tailings dams, and designated long term storage [75]. Chemical treatments of contaminated soil involve ion exchange, chemical precipitation, adsorption, and encapsulation [76]. Some common chemicals used for the oxidation of arsenite to arsenate are hydrogen peroxide, potassium permanganate, and sulphur dioxide [31]. One chemical treatment that has gained attention is containment of mine waste with repositories to limit oxidation and achieve cement encapsulation [31,77]. Encapsulation aims to stabilise toxic elements and pollutants by immobilising and/or insolubilising elements to prevent leaching and bioavailability. Some examples are solidifying mine waste onsite in a monolithic structure and utilising mine tailings to produce cement for building infrastructure [31,77,78]. Recently, a study by Garcia-Troncoso et al. [78] proposed cement encapsulation by replacing sand with mine tailings in the cement matrix for the construction industry. This cement would be used to build infrastructure such as housing materials, roads, and sidewalks [78]. Results stated that ecological toxicity (ecotoxicity) and human health toxicity decreased, and encapsulation was achieved; however, PTEs like As were not quantified or discussed. Additionally, no toxicity tests were conducted on cement dust such as the Microtox bioassay. The primary carcinogenic elements in cement dust are As, silica (Si), and chromium (Cr) [79]. A study by Kamaludin et al. [79] addressed the toxic effects of exposure to As-, Si-, and Cr-rich cement dust for cement workers and administrative staff in the cement industry. Administrative staff had higher risk of exposure and health impacts such as lung damage because of the absence of control measures (e.g., N95 masks) [79]. Ledda et al. [80] researched dust sediments from the Mount Etna region in Sicily to study the toxic effects using the Microtox method. Dust sediments included basalt, volcanic ash, basalt + cement, and cement; the Microtox results were 19.31%, 31.22%, 99.05%, and 98.95% toxicity, respectively, which confirmed that cement dust sediments had a significantly higher toxicity percentage and risk to the ecological environment and human health [79,80]. It is dangerous to substitute sand with As-rich mine tailings to produce cement/bricks as there is currently a toxicity problem with cement dust without the addition of mine tailings. Garcia-Troncoso et al. [78] performed a compressive strength analysis on varying sand to tailings ratios over twenty-eight days. Tailings that replaced 5% of sand had slightly weaker strength compared to the control; however, a tailings ratio beyond 10% decreased strength. Mine tailings were solidified in the cement, but compressive strength was compromised, causing concern for eroding cement containing PTEs in households and recreational spaces. Physical and chemical techniques are costly, labour-intensive, relocate toxins, introduce future toxicity problems, and the in situ soil microbial community is neglected [73,74,81].

3.3.2. Biological

Remediation techniques that are environmentally sustainable and cost-effective have been investigated to minimise the impacts of traditional techniques [59,82]. Biological techniques are environmentally sustainable because they utilise biological processes, such as microorganisms and plants to remediate PTEs (Table 3) [82,83,84,85]. Microorganisms, such as fungi Penicillium sp. and sulphur- and iron-oxidising bacteria such as Acidithiobacillus thiooxidans, can induce bioleaching and reduce toxic elements in soil [86]. Plants can be used to remediate contaminated soil through physiochemical mechanisms, called phytoremediation [87]. There are many phytoremediation techniques (Table 3), with phytoextraction and phytostabilisation being the most studied [73,82,88,89]. Phytoextraction and phytostabilisation are dependent on the plant species, soil microbes, and the associated PTEs [84,90]. Phytoextraction aims to reduce the concentrations of PTEs in contaminated soil by plant uptake mechanisms and storing PTEs in the above ground biomass, which is then harvested and appropriately disposed of (Figure 3) [31,59,90,91]. There are various disposal options for harvested biomass; some include landfill, composting, incineration, and pyrolysis [90,92]. Composting has concerns with leaching when toxic elements are still present in the final product; therefore, it is common to apply hydrated lime to reduce leachate [92,93,94]. Arsenic leaching occurs when soil pH is below 6.0, and the aim of the application of hydrated lime is to increase soil pH to mitigate arsenic leaching [92,93,94]. However, lime amendments have been shown to inhibit plant growth [93,94,95]. Pyrolysis and incineration results in ash biomass and is currently being investigated as a solution to harvested biomass post-phytoextraction [92]. Phytostabilisation aims to accumulate PTEs in plant roots to mitigate bioavailability, reduce mobility, and limit exposure within the ecosystem (Figure 3) [13,96]. However, As is a non-essential element for plants and can induce plant toxicity (phytotoxicity) and prevent plant function [16,97]. Phytotoxicity must be considered, and a plant’s ability to survive As or multi-elemental contamination must be assessed to ensure a successful phytoremediation application. Phytotoxicity can prevent high biomass, which is required for phytoextraction and can delay overall plant growth [16,87,97]. To mitigate phytotoxicity, the identification of metallophytes (metal(loid) tolerant plants) and hyperaccumulators is crucial. Hyperaccumulators translocate more toxins from soil via the root to the shoot by 2 or 3 orders of magnitude compared to other plant species [98]. Currently, there is a lack of identified metallophytes and As hyperaccumulators [16,87,97].

Biological remediation techniques are considered to be low-cost compared to the traditional physical and chemical methods [13,16,97]. In Europe, landfill disposal is estimated to cost USD 88.6 per t of mine waste [99]. Globally, approximately USD 1.5 billion is spent annually on acid and metalliferous drainage (AMD) management of mine waste [100]. Management of AMD mine waste includes physical and chemical treatments such as layering and mixing of waste rock, encapsulation, backfill disposal, and barriers to prevent leaching into groundwater [100]. The more cost-effective biological remediation methods, however, are slow and require longer application time which may take years or decades to reduce the toxicity of sites to safe levels [16,86,97].

Table 2. Physical, chemical, and biological remediation techniques for elemental contaminated soil.

Treatment	Process	Elements	Advantages	Disadvantages	References
Physical	Backfilling (in-pit disposal)	Toxic elements	Eliminates new tailings dams/ponds. No erosion, spillages, or infrastructure failures. Deep underground disposal suppresses oxidation processes and leaves the surface for revegetation.	Groundwater pollution. Chemical and physical changes. Surface/groundwater and aquifer pollution. Mobility of toxic elements.	[4]
Physical	Wet and dry covering of tailings dams/ponds	Toxic elements	Limited oxidation and acidification. Mitigates wind and water erosion. Reduced permeability.	Erosion, instability, contamination, leachates, groundwater pollution, aesthetically unpleasant.	[4]
Physical and Chemical	Cement stabilisation/ solidification without/with chemical amendments (e.g., Fe, fly ash, and lime)	Pb, Zn, As, P, Cd	Elemental stabilisation. Reduced leachates. Immobilisation.	Long-term deterioration. Production of carbon dioxide. High energy consumption. Air pollution (cement dust).	[77,101,102,103,104,105,106]
Chemical	Fe-based sorbent	As	As stabilised Risk to human health decreased. pH increased to neutral (6.9).	Ecotoxicity still present. Available phosphorus significantly decreased, causing plant malnourishment.	[107]
Biological	Phytoremediation	As	Hyper-accumulator species identified for As in highly contaminated soils (1603 mg/kg).	Lack of hyper-accumulating species for arid to semi-arid climates.	[14,16,108]
Biological	Phytoremediation	Cu, As	Environmentally friendly. Stabilisation of PTEs in the rhizosphere. Removal via phytoextraction of PTEs. Cost effective. Can be applied in situ.	Lack of metallophytes and hyperaccumulating plants. Plant biomass can be reduced under toxic conditions. Long application time, possibly decades. Leachates.	[16,87,97]
Biological and Chemical	Phytoextraction with ethylene diaminetetraacetic acid (EDTA)	As	Increased plant shoot uptake of As. Increased organic matter, EC values, extractable As. Soluble As and As root uptake increased with high EDTA (5 mmol/kg). Root–shoot translocation factor increased.	High dose of EDTA (5 mmol/kg) inhibited plant growth. Soil pH decreased. Bioaccumulation factor decreased. Synthetic and does not biodegrade.	[109,110]
Biological and Chemical	Combinations of plant, organic matter, and inorganic lime	As	Organic matter + plant treatment stimulated the microbial community. Compost + plant was the optimal treatment.	Plant growth was inhibited by all treatments except for the compost amendment.	[93]
Biological	Phytostabilisation with biochar	Pb, Zn, Cu, Cd, As	Environmentally friendly. Stabilisation of PTEs in the rhizosphere. Cost effective. Improve soil nutrients. Biochar improved plant growth. Biochar immobilised elements. pH, TC, and TN increased.	Little immobilisation of As, Zn and Cd. Possible increase in As bioavailability. Long application time, possibly decades. Leachates.	[13,16]

Copper (Cu); cadmium (Cd).

Table 2. Physical, chemical, and biological remediation techniques for elemental contaminated soil.

Treatment	Process	Elements	Advantages	Disadvantages	References
Physical	Backfilling (in-pit disposal)	Toxic elements	Eliminates new tailings dams/ponds. No erosion, spillages, or infrastructure failures. Deep underground disposal suppresses oxidation processes and leaves the surface for revegetation.	Groundwater pollution. Chemical and physical changes. Surface/groundwater and aquifer pollution. Mobility of toxic elements.	[4]
Physical	Wet and dry covering of tailings dams/ponds	Toxic elements	Limited oxidation and acidification. Mitigates wind and water erosion. Reduced permeability.	Erosion, instability, contamination, leachates, groundwater pollution, aesthetically unpleasant.	[4]
Physical and Chemical	Cement stabilisation/ solidification without/with chemical amendments (e.g., Fe, fly ash, and lime)	Pb, Zn, As, P, Cd	Elemental stabilisation. Reduced leachates. Immobilisation.	Long-term deterioration. Production of carbon dioxide. High energy consumption. Air pollution (cement dust).	[77,101,102,103,104,105,106]
Chemical	Fe-based sorbent	As	As stabilised Risk to human health decreased. pH increased to neutral (6.9).	Ecotoxicity still present. Available phosphorus significantly decreased, causing plant malnourishment.	[107]
Biological	Phytoremediation	As	Hyper-accumulator species identified for As in highly contaminated soils (1603 mg/kg).	Lack of hyper-accumulating species for arid to semi-arid climates.	[14,16,108]
Biological	Phytoremediation	Cu, As	Environmentally friendly. Stabilisation of PTEs in the rhizosphere. Removal via phytoextraction of PTEs. Cost effective. Can be applied in situ.	Lack of metallophytes and hyperaccumulating plants. Plant biomass can be reduced under toxic conditions. Long application time, possibly decades. Leachates.	[16,87,97]
Biological and Chemical	Phytoextraction with ethylene diaminetetraacetic acid (EDTA)	As	Increased plant shoot uptake of As. Increased organic matter, EC values, extractable As. Soluble As and As root uptake increased with high EDTA (5 mmol/kg). Root–shoot translocation factor increased.	High dose of EDTA (5 mmol/kg) inhibited plant growth. Soil pH decreased. Bioaccumulation factor decreased. Synthetic and does not biodegrade.	[109,110]
Biological and Chemical	Combinations of plant, organic matter, and inorganic lime	As	Organic matter + plant treatment stimulated the microbial community. Compost + plant was the optimal treatment.	Plant growth was inhibited by all treatments except for the compost amendment.	[93]
Biological	Phytostabilisation with biochar	Pb, Zn, Cu, Cd, As	Environmentally friendly. Stabilisation of PTEs in the rhizosphere. Cost effective. Improve soil nutrients. Biochar improved plant growth. Biochar immobilised elements. pH, TC, and TN increased.	Little immobilisation of As, Zn and Cd. Possible increase in As bioavailability. Long application time, possibly decades. Leachates.	[13,16]

Copper (Cu); cadmium (Cd).

Table 3. Phytoremediation techniques and characteristics.

Techniques	Characteristics	References
Phytoextraction	Arsenic uptake in above ground biomass.	[14,108]
Phytostabilisation	Immobilising As in the rhizosphere to prevent leaching into water and bioavailability.	[13,16]
Phytovolatilisation	Plant uptakes As through the xylem and transforms As to a less toxic and volatile form. Plant then releases As emissions.	[82,111]
Rhizodegradation	Microbial community in the rhizosphere degrade organic contaminants.	[111]
Rhizofiltration	Filtering of groundwater, surface water, and wastewater with plant roots.	[82,112]

Table 3. Phytoremediation techniques and characteristics.

Techniques	Characteristics	References
Phytoextraction	Arsenic uptake in above ground biomass.	[14,108]
Phytostabilisation	Immobilising As in the rhizosphere to prevent leaching into water and bioavailability.	[13,16]
Phytovolatilisation	Plant uptakes As through the xylem and transforms As to a less toxic and volatile form. Plant then releases As emissions.	[82,111]
Rhizodegradation	Microbial community in the rhizosphere degrade organic contaminants.	[111]
Rhizofiltration	Filtering of groundwater, surface water, and wastewater with plant roots.	[82,112]

Figure 3. Phytoextraction and phytostabilisation mechanisms for phytoremediation of potentially toxic elements in soil. Orange arrows show direction of potentially toxic element uptake [31,59,73,82,88,89,90,91,113].

Figure 3. Phytoextraction and phytostabilisation mechanisms for phytoremediation of potentially toxic elements in soil. Orange arrows show direction of potentially toxic element uptake [31,59,73,82,88,89,90,91,113].

4. Phytoremediation

Many studies have investigated the impacts of As-rich soil on vegetable plants, crops, and various methods involving low As concentrations (e.g., 30 mg/kg), As-spiked soil, and hydroponics [114,115,116,117]. However, phytoremediation is dependent on soil characteristics and the geochemical background such as pH, soil texture, organic matter, inorganic/organic carbon, and speciation [21,22,118,119]. It is essential to test the ability of a plant in phytoremediation studies using site-specific soil to account for physical and chemical characteristics as well as biological factors [22,118].

4.1. Identified Plant Species

Identifying plant species that bioaccumulate, hyperaccumulate, or stabilise As is vital for phytoremediation and has been widely studied (Table 4) [98]. It is stated that to confirm a As hyperaccumulating species, the above ground biomass must accumulate >1000 mg/kg of arsenic [120,121,122,123]. The first As hyperaccumulator was discovered by Ma et al. [14]; the Chinese brake fern, Pteris vittata, grew on chromated copper-arsenate-contaminated soil. The fronds of P. vittata contained 1442–7526 mg/kg of As, and contaminated soil had As concentrations ranging from 18.8 to 1603 mg/kg [14]. P. vittata growing on uncontaminated soil (0.47–7.56 mg/kg, As) was also analysed, and fronds bioaccumulated 11.8–64.0 mg/kg of As [14]. In general, the expected bioaccumulation of As on uncontaminated soil is 3.6 mg/kg, almost half the bioaccumulation reported for P. vittata [14]. The search for additional As hyperaccumulators is currently underway, and reports have identified Pteris umbrosa as a possible candidate [98,124]. A study by Zhao et al. [124], aimed to identify additional Pteris species for As hyperaccumulating characteristics. P. vittata, Pteris cretica, Pteris longifolia, and P. umbrosa were exposed to 0–500 mg/kg of As and results showed similar As hyperaccumulating capabilities to those displayed by P. vittata [124]. For example, both P. vittata and P. umbrosa were efficient in bioaccumulating As in their fronds, 57.1 ± 9.0% and 56.8 ± 4.8%, respectively, a characteristic essential for phytoextraction [125]. Pteris species are fern plants typically found in rainforests, along coasts, dry regions, and tropical and subtropical landscapes [126]. The habitat of P. umbrosa is the east coast of Australia, which might have implications for phytoextraction in non-coastal and arid to semi-arid regions such as Bendigo, a gold mining town in regional Victoria [98]. Hyper/accumulating plant species currently identified for As are primarily fern species, which have specific climate requirements; this causes concern for their use in phytoremediation projects for gold mine waste locations [16]. Fern species are unsuitable for arid to semi-arid climates, which are the common environments for gold mining towns. There is a need to identify plant species for phytoremediation projects that are compatible with arid to semi-arid climates, such as tussock grasses, and conducting studies with site-specific soil.

Table 4. Identified plant species and their abilities to phytoextract or phytostabilise arsenic in soil.

Plant Species	Common Name	Soil As (mg/kg)	As Accumulation	As Stabilisation	Characteristic	Reference
Pteris vittata	Chinese brake fern	18–1603	1442–7526 mg/kg (fronds)		Hyperaccumulator	[14]
Pteris vittata	Chinese brake fern	0.47–8	11–64 mg/kg (fronds)		Hyperaccumulator	[14]
Pteris vittata	Chinese brake fern	500	6200–7600 mg/kg		Hyperaccumulator	[124]
Pteris cretica	Cretan brake fern	500	6200–7600 mg/kg		Hyperaccumulator	[124]
Pteris longifolia	Longleaf brake	500	6200–7600 mg/kg		Hyperaccumulator	[124]
Pteris umbrosa	Jungle brake fern	500	6200–7600 mg/kg		Hyperaccumulator	[124]
Pteris vittata	Chinese brake fern		57.1 ± 9%		Hyperaccumulator	[125]
Pteris umbrosa	Jungle brake fern		56.8 ± 4.8%		Hyperaccumulator	[125]
Cassia alata	Ringworm bush	1587	ca. 25 mg/kg (shoots)	ca. 130 mg/kg (roots)	Phytostabiliser	[13]
Pityrogramma calomelanos	Silver fern	135–510	5130–5610 mg/kg (young fronds) 2760–8000 mg/kg (old fronds)	88–370	Hyperaccumulator	[127]
Pityrogramma calomelanos	Silver fern	20–8800	3820–8350 mg/kg (frond)	88–370 mg/kg (root)	Hyperaccumulator	[128]
Pteris vittata	Chinese brake fern	20–8800	4240–6030 mg/kg (frond)	103–330 mg/kg (root)	Hyperaccumulator	[128]
Athyrium filix-femina	Lady fern	74 ± 20.1	1.95 mg/kg (leaves)	7.56 mg/kg (roots)	Metallophyte	[129]
Geranium robertianum	Herb Robert	74 ± 20.1	1.95 mg/kg (leaves + stems)	18.3 mg/kg (roots)	Metallophyte	[129]
Rhizomnium punctatum	Dotted thyme-moss	218 ± 53.8	4.66 mg/kg (whole plant)		Metallophyte	[129]

Table 4. Identified plant species and their abilities to phytoextract or phytostabilise arsenic in soil.

Plant Species	Common Name	Soil As (mg/kg)	As Accumulation	As Stabilisation	Characteristic	Reference
Pteris vittata	Chinese brake fern	18–1603	1442–7526 mg/kg (fronds)		Hyperaccumulator	[14]
Pteris vittata	Chinese brake fern	0.47–8	11–64 mg/kg (fronds)		Hyperaccumulator	[14]
Pteris vittata	Chinese brake fern	500	6200–7600 mg/kg		Hyperaccumulator	[124]
Pteris cretica	Cretan brake fern	500	6200–7600 mg/kg		Hyperaccumulator	[124]
Pteris longifolia	Longleaf brake	500	6200–7600 mg/kg		Hyperaccumulator	[124]
Pteris umbrosa	Jungle brake fern	500	6200–7600 mg/kg		Hyperaccumulator	[124]
Pteris vittata	Chinese brake fern		57.1 ± 9%		Hyperaccumulator	[125]
Pteris umbrosa	Jungle brake fern		56.8 ± 4.8%		Hyperaccumulator	[125]
Cassia alata	Ringworm bush	1587	ca. 25 mg/kg (shoots)	ca. 130 mg/kg (roots)	Phytostabiliser	[13]
Pityrogramma calomelanos	Silver fern	135–510	5130–5610 mg/kg (young fronds) 2760–8000 mg/kg (old fronds)	88–370	Hyperaccumulator	[127]
Pityrogramma calomelanos	Silver fern	20–8800	3820–8350 mg/kg (frond)	88–370 mg/kg (root)	Hyperaccumulator	[128]
Pteris vittata	Chinese brake fern	20–8800	4240–6030 mg/kg (frond)	103–330 mg/kg (root)	Hyperaccumulator	[128]
Athyrium filix-femina	Lady fern	74 ± 20.1	1.95 mg/kg (leaves)	7.56 mg/kg (roots)	Metallophyte	[129]
Geranium robertianum	Herb Robert	74 ± 20.1	1.95 mg/kg (leaves + stems)	18.3 mg/kg (roots)	Metallophyte	[129]
Rhizomnium punctatum	Dotted thyme-moss	218 ± 53.8	4.66 mg/kg (whole plant)		Metallophyte	[129]

4.2. Co-Application of Soil Amendments

As discussed, plants can exhibit phytotoxicity and commonly require soil amendments to alleviate toxic conditions and improve plant growth and phytoremediation mechanisms [16,110]. Soil amendments also improve physical characteristics such as water holding capacity, permeability, infiltration, drainage, aeration, soil structure, and the microbial community [95,110]. There are a variety of inorganic and organic soil amendments that have been studied to enhance plant growth and element bioavailability (Table 5).

4.2.1. Organic Amendments

A study by Huang et al. [13] investigated phytostabilisation and the impacts of three different biochar amendments on mine tailings. Biochar is a carbonaceous material derived from pyrolysis methods that involve heating organic material (in a range of 500–900 °C in a limited oxygen environment) [13,130,131]. The organic matter used to produce biochar is typically biomass, sewage sludge, agricultural waste, and compost [13,131]. Biochar has a porous structure, high surface area, and other physiochemical properties that improve soil characteristics such as carbon and water holding capacity (Table 5) [131]. The mine tailings had high As and Pb concentrations (1587.1 and 3642.7 mg/kg, respectively), a pH of 6.5, and were low in nutrients such as total carbon (0.3%) and total nitrogen (0.01%) [13]. A greenhouse pot experiment was implemented over 100 days involving 10 treatments, three biochar types (Hibiscus cannabinus core, sewage sludge, and chicken manure) at three biochar concentrations (0.4%, 1%, and 3% w/w) and one control (mine tailings only) [13]. The results showed that overall plant shoots and roots increased in biomass and As shoot concentrations decreased by 54.9–77.5% [13]. Sewage sludge biochar applied at 3% concentration provided the best results for plant growth [13]. Plant height was 12.1 cm, and root length was 370.9 cm, compared to the control values of 8.1 cm and 112.4 cm for plant height and root length, respectively [13]. Sewage sludge biochar did not immobilise As; further focus is needed on the potential As release during increased biochar application rates [13]. Arsenic leaching as a result of organic soil amendments was also reported by Beesley et al. [42]. Beesley et al. [42] investigated mine waste soil amended with compost and biochar then monitored the effects on pore water and aqueous soil extracts. It was reported that As leaching was a concern and potential risk for further environmental pollution [42]. Arsenic leaching into pore water was >2500 µg 1⁻¹ (World Health Organisation drinking water standard is 10 µg 1⁻¹) and was found to be related to pH, dissolved organic carbon, and soluble P [42]. However, arsenic toxicity was reduced most effectively when compost and biochar were combined, and soil nutrients increased [42]. A study by Simiele et al. [132] researched the effects of combined biochar and iron-sulphate-assisted phytoremediation. Simiele et al. [132] also reported an increase in soil nutrients such as increased pH, electrical conductivity (EC), and reduced metal(loid) concentrations in soil pore water. However, there was no improvement in plant growth, which is essential for phytoremediation projects. Beesley et al. [42] concluded that field trials to further understand As leaching and effects of organic amendments are needed. Biochar should be investigated further to assess As leaching and its ability to improve soil properties and plant growth.

4.2.2. Phosphate and Organic Amendments

Barbafieri et al. [16] investigated the use of phosphate as a soil amendment to improve phytoextraction using crop species suited to the Mediterranean climate and P. vittata. Their results showed that phosphate was successful at improving As bioaccumulation, and plants showed no signs of phytotoxicity; however, biomass did not increase. A similar study by Niazi et al. [133] reported that phosphate increased As phytoextraction and improved biomass for Brassica species under spiked As concentrations of 0–75 mg/kg. Barbafieri et al. [16] used soil collected from the field, an industrial area, with average As concentrations of 25–2595 mg/kg as a result of chemical plant productions. These As concentrations are considerably higher than the Niazi et al. [133] study, and the use of As-rich field soil (non-spiked soil) introduced other soil As characteristics (e.g., speciation) that influence mobility and toxicity. For example, in Canada at Giant Mine, the dominant As species present was arsenic trioxide, which is the most bioavailable form of As for humans [21,134]. Additionally, a study by Cao et al. [135] investigated As uptake by P. vittata with chromated-copper-arsenate (CCA)-contaminated soil (135 mg/kg) and As-spiked soil (126 mg/kg). The study included soil amendments, which were phosphate rock, municipal solid waste, and biosolid compost. Cao et al. [135] stated that phosphate enhanced As bioaccumulation and removed 8.27% of As from CCA soil and 14.4% from the spiked soil. Compost amendment improved As bioaccumulation in CCA soil but decreased in spiked soil, and it was concluded that soil characteristics influenced the compost treatment [135]. Phosphate and compost treatment increased As leaching, but this was aided by P. vittata As uptake. A study by Caille et al. [55] investigated the use of co-application of phosphate and lime for As bioaccumulation, also using P. vittata. This study used soil collected from mining and smelting locations in Cornwall, England, and reported that there was no increase in As bioaccumulation. It was also stated that natural As in legacy mine waste soils may not be as easily bioavailable as other sources of As, e.g., industry, CCA, and spiked, because of soil properties. There is a need to assess phytoremediation techniques and soil amendments using mine waste soil collected from the field as results vary depending on the As source and soil characteristics.

4.2.3. Lime and Organic Amendments

Hydrated lime and organic amendments have been studied to assist phytoremediation projects (Table 4). A study by Clemente et al. [94] conducted a field experiment 10 km from the Aznalcóllar mine in Spain to assess two crops of Brassica juncea and soil amendments for phytoremediation. Field plots were initially fertilised with inorganic fertiliser (8:15:15 N:P:K). Seeds were planted and after the first crop, and the treatments were applied. Treatments included sugar beet lime and organic amendments including fresh cow manure and olive leaves/olive mill-waste mature compost. For the second crop, soils with a pH < 5 were limed before amendments were applied. It was identified that pH affected bioavailability and bioaccumulation over time as well as plant growth and biomass production. Organic matter improved plant survival and growth because soil conditions were enhanced. Manure treatment had the largest biomass production; however, it introduced weeds in the second crop. Compost treatment did not introduce weeds but was less effective at improving biomass and metal uptake. The soil amendments olive mill-waste compost and lime were also studied by Pardo et al. [93]. Pardo et al. [93] implemented a 2.5-year field phytostabilisation experiment using Atriplex halimus and soil amendments, mature olive mill-waste compost, pig slurry, and hydrated lime. Plants assisted with organic amendments ameliorated essential soil nutrients, which improved the soil microbial community. For example, soil enzyme activities, β-glucosidase, urease, and arylsulphatase significantly increased when compost and pig slurry were applied compared to the control. Additionally, dehydrogenase activity and basal respiration also significantly increased for compost and pig slurry amendments compared to the controls. Overall, soil enzyme activity improved with the combination of plant and soil amendment and stimulated the soil microbial community. Phytostabilisation was achieved, ecotoxicity decreased, and water-soluble organic C increased notably with the compost and plant combination. Organic matter was lower in pig slurry, but the plant compensated for organic matter with leaf litter and plant roots. However, the lime treatment did not benefit from plant features because plants did not grow well. Organic amendments improved water-soluble N, and only compost enhanced available P. Water-soluble N was greater in limed soils, which may be a result of poor plant and microbial growth because N was mainly inorganic N and highly bioavailable. Additionally, Pardo et al. [93] stated that the compost and plant treatment was most effective at decreasing ecotoxicity. A similar study by Clemente et al. [95] also used A. halimus and soil amendment compost, pig slurry, and hydrated lime for a field phytoremediation project in a semi-arid mine waste location at El Llano del Beal, Cartagena, Spain. Clemente et al. [95] also reported that the compost treatment had soil-ameliorating capabilities by improving organic matter, total organic carbon, and soil microbial biomass -C and -N. Plant growth was enhanced by the hydrated lime by 58%, compost by 79% and pig slurry by 89%. Organic soil amendments outperformed the hydrated lime for plant growth, which is also supported by Pardo et al. [93] and Clemente et al. [94] who reported poor plant growth in lime treatments and greater biomass production in manure treatments, respectively. Compost as a soil amendment to assist phytoremediation has proven to enhance soil conditions, organic matter, the microbial community, and plant growth and avoid the input of weeds.

Table 5. Phytoremediation combined with organic and/or inorganic soil amendments.

Treatment	Soil Amendment	Plant	Soil As (mg/kg)	Advantages	Disadvantages	Reference
Organic	Hibiscus cannabinus core biochar (HB), sewage sludge biochar (SB) and chicken manure biochar (MB) (500 °C 3 h).	Cassia alata	1587	Plant height and root length increased for SB 3%. SB immobilised elements. HB significantly improved total C. Arsenic availability increased with SB at 3%.	Conflicting results for As mobility and solubility as well as speciation may play a role.	[13]
Organic	Compost (olive mill-waste + 10% cow manure) Biochar (Orchard prunings pyrolysed at 500 °C		7490	Soil + compost + biochar treatment had largest decrease in ecotoxicity using the Microtox method. Soil nutrients and fertility improved.	As leaching Unknown long-term effects of increased As bioavailability.	[42]
Inorganic and organic	Hardwood biochar (500 °C, 3 h), iron sulphate	Populus Euramericana clone I45/51, Salix pupurea, Salix vimnalis	297 ± 30	Biochar + iron sulphate increased pH, EC, reduced metal(loid) concentrations in soil pore water.	Biochar + iron sulphate had no effect on plant growth.	[132]
Inorganic	Phosphate	Lupinus albus, Helianthus annuus, Brassica juncea, Pteris vittata	25–2595	No signs of phytotoxicity e.g., yellowing of leaves. Increased As bioaccumulation in roots and shoots. P. vittata and H. annuus were most successful at phytoextraction.	No significant change in biomass growth.	[16]
Inorganic	Phosphate	Brassica napus, Brassica juncea	0–75	As concentrations increased in shoots. Biomass increased. B. napus was most successful at As uptake. Improved physical and photosynthetic characteristics.	Results varied between plants.	[133]
Inorganic and organic	Phosphate rock, municipal solid waste and biosolid compost	Pteris vittata	135 and 126	Phosphate and compost increased As bioaccumulation. Plant mitigated leaching.	Field soil and spiked soil had different results.	[135]
Inorganic	Phosphate and lime	Pteris vittata	361	Lime balanced soil pH after phosphate application.	Phosphate and lime had no significant effect on As bioaccumulation.	[55]
Inorganic and organic	Fresh cow manure. Olive leaves/olive mill-waste mature compost. Sugar beet lime.	Brassica juncea	86–634	Manure had highest plant biomass production. Organic matter improved plant survival, growth, and soil conditions.	Manure introduced weeds. As uptake was low, Brassica juncea was not suited for phytoextraction	[94]
Inorganic and organic	Mature olive mill-waste compost Pig slurry Hydrated lime	Atriplex halimus	664 ± 28	Organic amendments with plants provided soil with essential nutrients, which improved the soil microbial community and were effective at phytostabilisation. Ecotoxicity decreased most notably with compost + plant. Limed soil had highest dissolved N in pore water, mainly inorganic N.	Plants did not grow well in lime amendment.	[93]
Inorganic and organic	Compost, pig slurry (PS), hydrated lime (HL).	Atriplex halimus	664 ± 28	Compost improved organic matter, total organic carbon, and soil microbial biomass -C and -N. Plant growth improved with amendments HL (59%), compost (79%) and PS (89%). As leaf concentrations decreased = phytostabilisation.	HL did not significantly decrease bioaccumulation of As in fruits compared to organic amendments after 12 months	[95]

Table 5. Phytoremediation combined with organic and/or inorganic soil amendments.

Treatment	Soil Amendment	Plant	Soil As (mg/kg)	Advantages	Disadvantages	Reference
Organic	Hibiscus cannabinus core biochar (HB), sewage sludge biochar (SB) and chicken manure biochar (MB) (500 °C 3 h).	Cassia alata	1587	Plant height and root length increased for SB 3%. SB immobilised elements. HB significantly improved total C. Arsenic availability increased with SB at 3%.	Conflicting results for As mobility and solubility as well as speciation may play a role.	[13]
Organic	Compost (olive mill-waste + 10% cow manure) Biochar (Orchard prunings pyrolysed at 500 °C		7490	Soil + compost + biochar treatment had largest decrease in ecotoxicity using the Microtox method. Soil nutrients and fertility improved.	As leaching Unknown long-term effects of increased As bioavailability.	[42]
Inorganic and organic	Hardwood biochar (500 °C, 3 h), iron sulphate	Populus Euramericana clone I45/51, Salix pupurea, Salix vimnalis	297 ± 30	Biochar + iron sulphate increased pH, EC, reduced metal(loid) concentrations in soil pore water.	Biochar + iron sulphate had no effect on plant growth.	[132]
Inorganic	Phosphate	Lupinus albus, Helianthus annuus, Brassica juncea, Pteris vittata	25–2595	No signs of phytotoxicity e.g., yellowing of leaves. Increased As bioaccumulation in roots and shoots. P. vittata and H. annuus were most successful at phytoextraction.	No significant change in biomass growth.	[16]
Inorganic	Phosphate	Brassica napus, Brassica juncea	0–75	As concentrations increased in shoots. Biomass increased. B. napus was most successful at As uptake. Improved physical and photosynthetic characteristics.	Results varied between plants.	[133]
Inorganic and organic	Phosphate rock, municipal solid waste and biosolid compost	Pteris vittata	135 and 126	Phosphate and compost increased As bioaccumulation. Plant mitigated leaching.	Field soil and spiked soil had different results.	[135]
Inorganic	Phosphate and lime	Pteris vittata	361	Lime balanced soil pH after phosphate application.	Phosphate and lime had no significant effect on As bioaccumulation.	[55]
Inorganic and organic	Fresh cow manure. Olive leaves/olive mill-waste mature compost. Sugar beet lime.	Brassica juncea	86–634	Manure had highest plant biomass production. Organic matter improved plant survival, growth, and soil conditions.	Manure introduced weeds. As uptake was low, Brassica juncea was not suited for phytoextraction	[94]
Inorganic and organic	Mature olive mill-waste compost Pig slurry Hydrated lime	Atriplex halimus	664 ± 28	Organic amendments with plants provided soil with essential nutrients, which improved the soil microbial community and were effective at phytostabilisation. Ecotoxicity decreased most notably with compost + plant. Limed soil had highest dissolved N in pore water, mainly inorganic N.	Plants did not grow well in lime amendment.	[93]
Inorganic and organic	Compost, pig slurry (PS), hydrated lime (HL).	Atriplex halimus	664 ± 28	Compost improved organic matter, total organic carbon, and soil microbial biomass -C and -N. Plant growth improved with amendments HL (59%), compost (79%) and PS (89%). As leaf concentrations decreased = phytostabilisation.	HL did not significantly decrease bioaccumulation of As in fruits compared to organic amendments after 12 months	[95]

5. Ecotoxicological Assessment

It is vital to perform ecotoxicity tests to identify toxic effects on the ecological community and soil biota; ecotoxicity tests have also been used to determine the success of a remediation treatment [93,136]. There are many ecotoxicity tests, but there are two that are prominent, the earthworm acute toxicity test and the Microtox test [136,137,138,139]. Earthworms are exposed to contaminated soil, and bioaccumulation of toxic elements occurs via dermal absorption and/or ingestion [136]. The “Organisation for Economic Cooperation and Development” (OECD) and “International Organisation for Standardisation” (IOS) developed standards and determined the lethal concentration 50 (LC₅₀), the known concentration at which a substance kills 50% of test subjects (earthworms); it is considered the median lethal concentration [136,140,141]. The Microtox test is commonly used to assess the ecotoxicity of contaminated soils or sediments [142,143,144,145]. The Microtox test method follows a standard test method of the American Society for Testing and Materials (ASTM) [144,146]. The intensity of light emitted by the bioluminescent marine bacterium Aliivibrio fischeri, previously known as Photobacterium phosphoreum, is measured [146]. Effective concentration 50 (EC₅₀) is when light emitted has decreased by 50% relative to the control sample [144,147]. When the EC₅₀ is observed, it is reported as toxic and is often measured at 0-, 5-, 10-, and 15-minute intervals [43,143,144,145].

Recent studies have employed the Microtox bioassay to assess ecotoxicity in phytoremediated soils [42,93]. Pardo et al. [93] assessed soil recovered post-phytostabilisation with soil amendment application by evaluating soil physicochemical characteristics, extracted soil solution and the soil microbial community (e.g., biomass -C and -N) and performing ecotoxicity bioassays. Phytoremediation with co-application of compost outperformed all other treatments (pig slurry and hydrated lime) when assessed with the A. fischeri Microtox bioassay. The results reported the lowest EC₅₀ values of 47.9% and 45.8%, at 15- and 30-minute intervals, respectively. The control had EC₅₀ values of 61.7% at 15 min and 61.5% at 30 minutes. Relative to the control, the compost assisted phytoremediation application decreased in toxicity. The decreased toxicity and input of soil nutrients supported plant life, which then added more soil nutrients e.g., leaf litter. The enhanced soil nutrients supported microbial communities, which also increased in microbial growth and function and improved the soil ecosystem. Many other studies have investigated phytoremediation with a variety of plant species and soil amendments (Table 3 and Table 4), but there is a lack of ecotoxicological analysis. Clemente et al. [95] proposed that there is a need for long-term analysis of plant bioaccumulation, ecotoxicity assessments of remediated soils, greater interest in studying soil pore water to further understand leaching, and more in situ studies.

6. Conclusions

Remediating legacy, current, and future mine waste sites is our humanitarian responsibility towards the environment, local community, and human health. Legacy gold mine waste remains an environmental disaster. Sending mine waste to landfill unfortunately was, and potentially still is, the most adopted mitigation method. Landfill dumps are common because there were no environmental regulations during the 19th century gold rush. This started a trend of poor mine waste management practices which continues today. Modern mines, such as the cobalt mines in the Democratic Republic of the Congo, are resembling the gold rush era with poor environmental regulations and will only amplify the mine waste problem. Advancing technology and climate change solutions, i.e., electric cars, require extraction of raw elements. Mines are inevitable; however, society must do better to protect the land, soil, air, water, biodiversity, local rights, and human health.

Future research is needed to identify native plant species suited to mining locations (e.g., arid to semi-arid climate) for phytoremediation of As-rich mine waste. The current fern species that have been identified as hyper/accumulators are not suited to typical regional mining locations and pose problems for plant survival. Soil amendments are required to alleviate toxic effects on plants, and although this has been studied, more work is required. There a is need to conduct phytoremediation studies with mine waste soil collected in situ to factor in site-specific soil physiochemical properties and speciation. Mature olive leaves/mill-waste compost has been shown to enhance soil conditions, organic matter, microbial community, and plant growth; avoid the input of weeds; and decrease ecotoxicity. Sewage sludge biochar also enhances plant growth and soil properties; however, the primary concern for biochar was increased As leaching, and this should be investigated further. Arsenic leaching could be mitigated by the presence of a native plant species. There is a lack of phytoremediation projects assessing ecotoxicity and pore water. Ecotoxicity assessment is essential to evaluate the effectiveness of the remediation technique, and monitoring pore water will help to understand the risk of As leaching. Finally, field studies are required to further validate a remediation technique design and confirm its effectiveness and practicality.