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The cyanidation process has been, and still remains, a profitable and highly efficient process for the recovery of precious metals from ores. However, this process has contributed to environmental deterioration and potable water reserve contamination due to the discharge of poorly treated, or untreated, cyanide containing wastewater. The process produces numerous cyanide complexes in addition to the gold cyanocomplex. Additionally, the discharge constituents also include hydrogen cyanide (HCN) – metallic complexes with iron, nickel, copper, zinc, cobalt and other metals; thiocyanate (SCN); and cyanate (CNO). The fate of these complexes in the environment dictates the degree to which these species pose a threat to living organisms. This paper reviews the impact that the cyanidation process has on the environment, the ecotoxicology of the cyanidation wastewater and the treatment methods that are currently utilised to treat cyanidation wastewater. Furthermore, this review proposes an integrated biological approach for the treatment of the cyanidation process wastewater using microbial consortia that is insensitive and able to degrade cyanide species, in all stages of the proposed process.
An integrated biological approach for treatment of cyanidation wastewater
Lukhanyo Mekuto1SKO Ntwampe1,*, Ata Akcil2
1Bioresource Engineering Research Group (BioERG), Department of Biotechnology, Cape Peninsula
University of Technology, PO Box 652, Cape Town, 8000, South Africa
2Mineral-Metal Recovery and Recycling (MMR&R) Research Group, Mineral Processing Div., Dept.
of Mining Eng., Suleyman Demirel University, TR32260 Isparta, Turkey
* Corresponding author:; Tel: +27 21 460 9097; Fax: +27 21 460 3282
The cyanidation process has been, and still remains, a profitable and highly efficient process for the
recovery of precious metals from ores. However, this process has contributed to environmental
deterioration and potable water reserve contamination due to the discharge of poorly treated, or
untreated, cyanide containing wastewater. The process produces numerous cyanide complexes in
addition to the gold cyanocomplex. Additionally, the discharge constituents also include hydrogen
cyanide (HCN) metallic complexes with iron, nickel, copper, zinc, cobalt and other metals;
thiocyanate (SCN); and cyanate (CNO). The fate of these complexes in the environment dictates the
degree to which these species pose a threat to living organisms. This paper reviews the impact that the
cyanidation process has on the environment, the ecotoxicology of the cyanidation wastewater and the
treatment methods that are currently utilised to treat cyanidation wastewater. Furthermore, this review
proposes an integrated biological approach for the treatment of the cyanidation process wastewater
using microbial consortia that is insensitive and able to degrade cyanide species, in all stages of the
proposed process.
Keywords: biodegradation, cyanidation process, gold bearing ores, microorganisms
1 Introduction
The leaching of gold bearing ores with cyanide has been reported since 1889 in New Zealand and
South Africa (Dorr and Bosqui, 1950). Despite recent attempts to develop alternative extraction
procedures such as the use of thiocyanate, ammonia, thiourea and thiosulphate (Hilson and
Monhemius, 2006), the cyanidation process still remains the preferred method as it is economically
viable and results in rapid extraction (Mudder and Botz, 2004). Elsner (1846) studied gold
solubilisation in various cyanide concentrations and proposed the following complexation reaction:
2Au+4NaCN +2H
2NaAu(CN )
+2NaOH +H
2Au+4NaCN +H
2NaAu(CN )
Overall equation:
4Au+8NaCN +O
O 4NaAu (CN )
Equations (1) and (2) were formulated based on the realisation that an oxidant was necessary for the
process to proceed; hence, equation 1 was formulated. The second realisation was that the gold ions
combine with two cyanide anions to form a soluble gold complex, with hydrogen peroxide playing a
crucial role both as a reactant and product (Johnson, 2015, Habashi, 1966).
Since the initial discovery of the cyanidation process, there have been major variations on the gold
recovery process with various processes, which include amongst others, the introduction of carbon-in-
leach (CIL), carbon-in-pulp (CIP) and heap leaching. Additionally, there has been extensive research
and industrial application (e.g. BIOX process) on pre-treatment of refractory ores which are not
amenable to leaching by direct cyanidation (van Aswegen et al., 2007). Commercialization of such
pre-treatment technologies have resulted in direct financial gains for gold processes, as gold recovery
rates increase from ores that were not amenable to direct cyanidation. However, calamities in cyanide
storage facilities, which in the past resulted in environmental deterioration, have raised concern about
the use of cyanide in the gold mining sector (Korte et al., 2000). In most cases, cyanide entered the
environment by overflowing from storage ponds/tailings storage facilities, or through
cracks/punctures in heap leaching liners (Hilson and Monhemius, 2006). Although free cyanide is not
persistent in the environment, it does however react with chemical constituencies that are present in
the contaminated environment to form a variety of complexes, which under certain environmental
conditions dissociate to form free cyanide, which is highly toxic, thus resulting in health related
issues. Additionally, cyanide in soil has been found to form complexes with metallic species, forming
stable cyanide compounds that are less toxic. This can result in prolonged presence of cyanide in soil.
Although microbial flora contributes to cyanide degradation in soil, this however ensures gradual
release of free cyanide over extended periods. There is limited information about the gradual release
of free cyanide from its complexes resulting from microbial decomposition in contaminated soil.
Therefore, the notion that cyanide is not persistent and does not result in environment deterioration
and chronic health complications, is somewhat misleading (Logsdon et al., 1999).
Microbial species have low tolerance to free cyanide with a maximum threshold of 200 mg CN -/L
(Kuyucak and Akcil, 2013). However, in recent studies, there have been microorganisms that have
been found to exceed the stipulated threshold (Luque-Almagro et al., 2005, Mekuto et al., 2013).
Because of the toxicity of cyanide, an international body that regulates the use of cyanide was
established. The International Cyanide Management Code (ICMC) of the International Cyanide
Management Institute (ICMI) ( deals with the proper management and
approval of process certificates for cyanide-utilising industries. Additionally, the ICMC has enforced
codes of practice that compel industries to implement alternative waste management practices in order
to minimise environmental contamination by such industries (Gibbons, 2005, Akcil, 2002, Akcil,
Treatment methods have been developed to decontaminate cyanide containing process wastewater.
These methods include (1) alkaline chlorination, (2) hydrogen peroxide, (3) the sulphur dioxide
process (i.e. INCO process), (4) barren water rinse and (5) biological degradation. Processes (1) to (3)
are highly expensive and produce excess sludge that requires suitable disposal measures, while
processes such as the barren water rinse is best suited to climates that have an inexpensive source of
fresh water and a positive water balance, making it limited to implement, especially in arid regions.
Contrarily, biological degradation has proven to be a robust, environmentally benign and
economically viable process, with miniscule process input requirements (Baxter and Cummings,
2006, Akcil and Mudder, 2003). The process makes use of microbial communities that utilise
different biochemical pathways to degrade or transform cyanide and its by/end products, resulting in
decontaminated wastewater that can be recycled back to mineral leaching circuits or disposed into
surface waters.
There has been immense pressure on the mining industry to develop alternative environmentally
friendly processes that would decontaminate cyanide-containing wastewaters. The main purpose of
this review is to propose an integrated biological degradation process for cyanidation wastewater with
the purpose of recycling process water to mineral bioleaching/bio-oxidation processes using cyanide
insensitive microbial community in all the stages of the process. This would ensure improved process
efficiency even during unfavourable conditions.
2 Types of the gold bearing ores
Gold bearing ores, which also contain significant quantities of silver and base metals, are classified as
(1) free milling, (2) complex and (3) refractory, a classification that is based on the ores mineralogy.
Free milling ores are milled such that 80% passes the 75-µm screen in order to achieve 80% gold
recoveries by direct cyanidation. Complex ores are reagent-consuming ores due to the presence of
reactive matter, thus resulting in higher reagent consumption, which ultimately results in low gold
recoveries. Refractory ores are not amenable to direct cyanidation as the gold is trapped in the mineral
matrix; hence, cyanide is unable to penetrate to leach out the gold, leading to low recoveries.
Currently, refractory and complex gold ores are pre-treated via bio-oxidation, whereby iron- and
sulphur-oxidising organisms are utilised to oxidise the iron and sulphur moieties within the ore, thus
exposing the entrapped gold particles for subsequent cyanidation. Additionally, the process removes
the carboneous matter and trace elements that are responsible for reagent consumption in complex
ores. This in turn improves gold recoveries from ores.
3 Cyanidation process
Cyanidation is a process that utilizes cyanide as a leaching reagent for the extraction of precious
metals (i.e. gold and silver) from ores. The industrial practice of this technique is conducted in three
distinct ways: (1) heap leaching, (2) agitation leaching followed by carbon-in-pulp (CIP), and (3)
agitated carbon-in-leach (CIL) (Akcil and Mudder, 2003). Heap leaching is generally used to
beneficiate low-grade ores that contain low concentrations of gold (˂0.04 ounce/ton) while tank
leaching is used to beneficiate high-grade ores containing more than 0.04 ounce/ton.
3.1 Heap leaching
Heap leaching is advantageous because of its simple design, operational and capital cost, and shorter
start up times. Heaps are constructed on lined pads with ore that is received directly from the mine.
During heap leaching, ore is crushed and agglomerated prior to packing in the heap. This is done to
increase permeability, allow finer particles to attach to bigger particles (Petersen and Dixon, 2007,
Watling, 2006) and maintain high pH conditions which are necessary for the leaching process to
succeed. During agglomeration, crushed ore is mixed with lime to ensure the maintenance of a higher
pH within the heap. However, sulphidic refractory ores may need prior pre-treatment either by
roasting, autoclaving or bio-oxidation, before being heap leached (Lopes and Johnson, 1988).
The process typically has recovery efficiencies of 60 to 80% over a period of weeks or months,
depending on the permeability and size of the heap. After the process has been completed, the heap is
normally rinsed with water until the cyanide concentration is below the regulatory standard of that
particular country. This form of treatment is unfavourable especially in semi-arid or arid countries and
for countries that have a negative water balance. The spent ore is discarded in designated disposal
areas (Lopes and Johnson, 1988). The spent ore normally contains trace amounts of free cyanide but
also contains elevated concentrations of weak and strong acid dissociable cyanides. These cyanides,
under certain physicochemical conditions, dissociate to release free cyanide. Therefore, spent ore
from the cyanidation process can exacerbate environmental contamination in a similar manner as
cyanidation wastewater.
3.2 Agitated leaching
In this process, leaching takes place within a series of tanks where the ore is slurried with the leaching
solution and the resulting gold-cyanide complex is adsorbed onto activated carbon. Leaching in a
Carbon-In-Pulp (CIP) system takes places in interconnected series of tanks, while in a Carbon-In-
Leach (CIL) system it is conducted in single tanks in a batch operational mode. Easily leachable ores
such as oxide ores are milled ground to 210-µm and leached with a cyanide solution (100 mg CN -/L)
over a four- to twenty-four-hour period, with a 50% pulp density being appropriate for optimum
leaching rates. However, sulphidic ores are processed by milling the ore to 44-µm and using a higher
concentration of cyanide (200 mg CN-/L) over a 72-hour period at a pulp density of 40% (Weiss,
1985, Cummins, 1973). This process has recovery efficiencies of ≥90% over a period of hours.
Recently, higher cyanide concentrations (up to 400 mg CN-/L) are utilised for higher recovery rates (≥
During the cyanidation process, cyanide forms a variety of complexes with a number of chemical
constituents that are present within gold bearing ores. These are classified using five categories: (1)
Free cyanide, in the form of hydrogen cyanide (HCN) and cyanide ion (CN-). (2) Simple cyanide in
the form of sodium cyanide (NaCN) and potassium cyanide (KCN). (3) Thiocyanate (SCN). (4) Weak
acid dissociable cyanide (CNWAD) which is mainly metal-complexed cyanide (e.g. Zn, Cu, Ni, etc). (5)
Strong acid dissociable cyanide (CNSAD) (strong complexes with metals such as Fe, Co, Au and Ag).
The toxicity of these cyanide compounds decreases from category (1) to (5). These categories are
listed in Table 1.
Table 1: Cyanide species with their relative toxicities
Type of cyanide Species Toxicity
Free cyanide CN-
Intermediate to high (depends
on environmental conditions)4
Pt group metal complexes
Other species SCN-
High 1-2
a Presence of these species in leach solutions is unclear.
1-5Categorization of cyanide species
4 Toxicity of cyanide and its impact on active biological processes
Due to the high reactivity of the cyanide anion, it is able to form complexes with a variety of metallic
species that form part of enzymatic structures, thus inhibiting their role and cellular growth. There are
three main inhibition mechanisms: (1) Formation of cyanohydrin derivatives through reaction with
keto-compounds. (2) Formation of nitrile derivatives through reaction with Schiff-base intermediates.
(3) Chelation of di- and trivalent metals in metalloenzymes. A typical example of cyanide inhibition is
the irreversible reaction of cyanide to ferric heme iron within the structure of the cytochrome c
oxidase, thus inhibiting oxidative phosphorylation. Oxygen utilization will therefore be impaired, thus
resulting in cessation of aerobic metabolic activity (van Buuren et al., 1972, Jones et al., 1984,
Massey and Edmondson, 1970).
Cyanide contamination in surface and ground water reserves has been proven to be detrimental to the
lives of the organisms that feed from such water reserves. Additionally, the transport of cyanide and
related complexes from contaminated matrices into active biological processes such as those used in
municipal wastewater treatment plants (WWTPs), may lead to low efficiencies within such processes
because of cyanide susceptibility of the microbial communities that are employed in such systems.
Numerous studies on the inhibition of nitrification by low concentrations of cyanide have been
studied elsewhere (Han et al., 2014, Kim et al., 2011b, Sharma and Philip, 2014). The presence of free
cyanide above 0.2 mg CN-/L has been found to completely inhibit the nitrification process while
thiocyanate and ferric cyanide inhibited nitrification at concentrations of 200 and 100 mg/L
respectively (Kim et al., 2008). Additionally, Neufeld et al. (1986) observed a maximum tolerance
threshold of 0.11 mg CN-/L while in a separate study, the presence of free cyanide above a
concentration of 2.5 mg CN-/L prolonged the biodegradation of phenols and aromatic hydrocarbons
(Sharma and Philip, 2014). In contrast, the Nitrobacter and Nitrospira strains that were detected in an
activated sludge process treating coking wastewater were able to tolerate up to 50 mg CN-/L, where
the microbial community changed as a result of cyanide loading and sensitivity of some of the
microbial organisms to free cyanide, resulting in the dominance of Nitrobacter and Nitrospira strains
(Kim et al., 2011c). In a separate study, the activity of Leptospirillum ferriphillum and
Acidithiobacillus caldus, key iron and sulphur oxidising organisms during bio-oxidation of gold ores,
were inhibited by the presence of thiocyanate at a concentration of 1.25 mg/L. However, the adapted
A.caldus tolerated up to a concentration of 7 mg/L, thus demonstrating the adaptability of the
organism to changing environmental conditions, although with low sulphur oxidation rates (van Hille
et al., 2015).
The inability of microbial communities/species to perform their respective functions as a result of
cyanide presence and sensitivity proves to be disastrous. In WWTP’s, the presence of cyanide results
in limited or no nutrient removal in secondary treatment stages of the WWTP’s, where autotrophic
nitrifying and denitrifying microbial communities are employed, resulting in the failure of such
systems. Furthermore, the presence of cyanide compounds in bio-oxidation processes results in poor
refractory and complex ores pre-treatment due to the sensitivity of the organisms employed in such
systems, thus resulting in low gold recoveries during cyanidation. Therefore, adequate,
environmentally benign and cost-effective processes need to be developed to equip cyanide utilising
industries with an alternative technology for treatment of their cyanide containing effluent and
5 Treatment methods
Cyanide treatment from mining and metallurgical operations is required due to the potential toxicity
of cyanide to the environment. Cyanide treatment methods are classified as destruction-based
processes or recovery-based methods (Botz et al., 2005, Botz et al., 2015). This review focuses on
destruction-based processes where cyanide is oxidised to less toxic compounds. These destruction-
based treatment methods may include treatment or removal of cyanide from the following options: (1)
Supernatant solutions from tailings ponds. (2) Spent slurry tailings from milling operations. (3)
Seepage from ponds or tailings. (4) Solutions from heap leaching operations (Botz and Mudder,
2002). The treatment of cyanide-containing solutions is achieved through destruction or recovery
processes, with a focus on the degradation of cyanide-containing residue or wastewater. The selection
of an appropriate treatment process requires consideration of numerous factors which are based on the
chemical characteristics of the waste, the volume of the waste that is to be treated, the environmental
setting of the treatment site, the desired effluent quality, the availability of reagents or suitable process
waters and the regulations that are in place to ensure that the discharge meets the regulatory
guidelines (Akcil, 2003, Botz et al., 2015).
Cyanide treatment options are classified as (1) Natural, (2) Physical, (3) Chemical and (4) Biological
5.1. Natural Degradation
Natural degradation of cyanide is achieved by natural processes such as biodegradation, adsorption,
photodecomposition, oxidation and volatilization without human interference. However, the
mechanism of degradation is affected by physicochemical parameters such as temperature, pH, water
chemistry and dissolved oxygen concentration. The solution pH plays a major role in the volatilization
of cyanide. HCN gas has a pKa value of 9.24 at 25°C (Johnson, 2015, Kuyucak and Akcil, 2013),
meaning that below a pH of 9.24, cyanide is available as hydrocyanic acid, which volatilises as
hydrogen cyanide gas – since cyanide has a high vapour pressure. An increase in temperature, reduced
liquid-depth to surface-area ratios and turbulence increase the rate of volatilization significantly while
solution contact with carbon dioxide results in a decrease in pH, which in turn results in increased
hydrogen cyanide gas volatilization (Nsimba, 2009, Johnson, 2015).
Strong acid dissociable cyanides such as iron cyanide complexes are decomposed via ultraviolet rays’
exposure from the sun, thus resulting in the dissociation of the complex to its original constituents:
free cyanide and metal ion. However, the wastewater needs direct exposure to sunlight and this is
facilitated by pneumatic mixing of the wastewater, ensuring exposure of water to sunlight (ultraviolet
light) (Oudjehani et al., 2002). However, during natural attenuation processes, the wastewater is
normally found in stagnant form, thus limiting exposure of the water to sunlight (Botz and Mudder,
2000) and resulting in the restriction of cyanide degradation. Furthermore, natural attenuation
processes are extremely slow due to the stagnant nature of the wastewater and can take years and
sometimes, decades for the process water to meet discharge regulations.
Thiocyanate, ferricyanide and ferrocyanide were found to be the major contaminants in ground water
reserves, and remedial actions of the ground water was achieved through natural attenuation from a
period of ten years where the mechanism of degradation was proposed to be biodegradation and
adsorption. Strong acid dissociable cyanides were detected at high concentrations after the ten-year
period, suggesting that they were not degraded (Gagnon et al., 2004). In a separate study, natural
attenuation via the biodegradation mechanism was assessed in fresh (three months), and old (about
nine-year-old) tailings. Results reported free cyanide biodegradation by heterotrophic bacterial
species, while strong acid dissociable were not degraded over a period of 96 days (Oudjehani et al.,
2002). Due to slow degradation rates in natural attenuation process, alternative treatment methods
were developed and these include chemical and biological treatment methods.
5.2. Physical Methods
Physical processes that are currently utilised are mainly based on the dilution of the cyanide
containing wastewaters to meet the discharge requirements, without the utilization of any chemical
reagents. The Barren/fresh water rinse is such a method. . In this method, fresh water is utilised to
rinse the heap from barren ponds and the water is also utilised to reduce evaporative losses. The
mechanism of cyanide reduction is mainly through dilution and volatilization, and slight, microbial
degradation and complexation (Mosher and Figueroa, 1996). The engineering costs associated with
process are, amongst others; pumping capacity, line installation, instrumentation, irrigation and
equipment residence time. The practicality of such a process is virtually impossible, especially in arid
or semi-arid regions.
5.3. Chemical treatment
The following chemical treatment methods are utilised in the mining industry: the alkaline
chlorination method; ozonation; copper catalysed hydrogen peroxide; acidification-volatilization-
reneutralization (AVR) process; the sulphur dioxide process; sulfidization-acidification-recycling-
thickening (SART) process, Caro’s acid treatment methods and gas filled membrane absorption
process (GFMA). These chemical treatment methods are thoroughly discussed in a number of reviews
(Botz and Mudder, 2002, Botz et al., 2005, Mosher and Figueroa, 1996, Akcil, 2003, Baxter and
Cummings, 2006, Adams, 2013, Estay et al., 2013). However, a comparison of the different chemical
treatment methods with their advantages and disadvantages has been summarized in Table.2.
Table 2: Advantages and disadvantages of cyanide treatment methods (Adopted from Mosher and
Figueroa (1996) ).
Treatment method Advantages Disadvantages
Alkaline chlorination Well established. CNSAD, ammonia and chlorides are not
Adds hazardous metals to water.
Poor process control leads to toxic
Reacts preferentially with SCN.
Ozonation - Unable to treat CNSAD and ammonia.
Poor treatment performance.
Very expensive process.
Hydrogen peroxide Simple to operate.
High efficiencies.
Unable to treat SCN and ammonia.
Excess precipitate accumulation.
Expensive reagent.
Requires removal of catalyst after
process completion.
AVR process Highly efficient process. Very expensive process.
HCN gas might escape and cause life-
threatening situations.
Sludge handling problems.
SART High cyanide recoveries
Recovers cyanide from CNWAD
Cyanide can be regenerated
Economically viable
Difficulties with sizing of solid-liquid
separations equipment.
Use of hazardous chemicals.
GFMA Highly efficient process.
Safe, HCN escape is restricted
Applied in cyanide wastewaters from
different industries.
No secondary metabolites produced.
Simple operation.
Sulphur dioxide Inexpensive reagent use.
Faster degradation rates.
Effective at treating slurries.
Reagent use requires licence payments.
Process adds sulphates to treated water.
Excess precipitate accumulation.
Caro’s acid Faster degradation rates. Expensive reagents.
Sludge accumulation problems.
5.4. Biological treatment
Biological degradation is generally categorised into: (1) Phytoremediation and (2) Microbial
remediation. The mechanism of cyanide degradation is similar in both methods. These processes
make use of enzymes that degrade or convert cyanide species to less harmful by/end products.
5.4.1. Phytoremediation of cyanide
The remediation of cyanide by plant species has been under intense research. To date, the application
of this process has been limited to laboratory studies. A variety of plant species have been reported to
have a potential for removing cyanide in soil and solutions. These plants include Eichhornia crassipes
(Ebel et al., 2007), Hordeum vulgare, Avena sativa, Sorghum bicolor (Samiotakis and Ebbs, 2004),
Salix babylonica, Sambucus chinensis (Yu et al., 2005b), amongst other plants species. These plant
species metabolise free cyanide via two pathways; the cyanoalanine pathway and the sulphur
transferase pathways (Samiotakis and Ebbs, 2004). However, the cyanoalanine pathway is the most
common pathway due to asparagine formation – an amino acid which is critical for nitrogen supply in
most plants (Yu et al., 2012, Machingura and Ebbs, 2010, Machingura et al., 2013). The metabolism
of cyanide by Sambucus chinensis and Torilis japonica was investigated. S. chinesis was observed to
possess the highest removal rate of 8.8 mg CN- kg-1 h-1, followed by T. japonica at 7.5 mg CN- kg-1 h-1
(Yu et al., 2004); while in a separate study, Sorghum bicolor and Linum usatassimum converted the
radiolabelled Prussian blue, a CNSAD, to carbon dioxide and accumulated the majority of cyanide
within the plants 140 mg CN-/kg plant)(Kang et al., 2007). However, this process has been
hindered by the levels of phytotoxicity: a phenomenon that is attributed to the toxicity levels of
cyanide to the plant species that carry out phytoremediation. Yu et al. (2005a) investigated the
phytotoxicity of cyanide in Salix babylonica and made the following observations: Cyanide levels
below 1 mg CN-/L were not toxic, while severe signs of toxicity were witnessed when cyanide levels
were above 9.3 mg CN-/L. While Larsen et al. (2004) observed that Salix viminalis that were exposed
to 0.4 mg CN-/L demonstrated minimal signs of transpiration depression while doses of 8 to 20 mg
CN-/L were lethal to S. viminalis.
Although some of the S. bicolor species have been reported to tolerate up to 50 mg CN -/L (Trapp et
al., 2003), the majority of the reported plant species have been reported to have low cyanide tolerance
and the degradation rates are extremely low. Additionally, these process are extremely slow and
would require vast amounts of land if commercialised, as proposed by Trapp et al. (2003) . The
proposed commercialisation of the process would include, among others, the administration of the
cyanide dosage to the plants. This means, that cyanide would have to be diluted to concentrations that
can be metabolisable by the plants, which is, below a concentration of 10 mg CN -/L. The
practicability of such a process is minimal especially in arid and semi-arid regions. Additionally, the
growth of some of the proposed plants may be minimal or nonexistent in some regions. Furthermore,
the genetic and physiological traits of these plants may change in different regions due to differences
in soil type and nutrition. Therefore, this makes the commercialization of this process almost
impossible due to the reasons mentioned previously.
5.4.2. Microbial remediation of cyanide
Over the past twenty years, the application of microbial species in the form of bacteria, fungi,
protozoa, algae and yeasts has gained considerable attention due to the robustness and
environmentally friendliness of the process. These organisms use a variety of enzymatic pathways to
destruct or convert cyanide to less hazardous products. This is, however, dependent on the type of
enzyme that a particular microorganism possesses. These enzymatic pathways are published
elsewhere (Ebbs, 2004, Gupta et al., 2010, Dash et al., 2009). Briefly, these include (1) hydrolytic, (2)
oxidative, (3) reductive, (4) substitution or transfer, and (5) synthesis pathways. The first three
pathways are degradative pathways where the microbial species destruct cyanide to form ammonia,
formic acid, carbon dioxide, methane and carboxylic acid, while the last two pathways are
assimilative. The type of enzymatic pathway that a particular organism employs is dictated by the
environmental conditions (pH, temperature and presence of toxins) and gene structure. These
organisms have been observed to work efficiently in biofilms, where they are able to micro-manage
their environment to suit their needs. Within these biofilm structures, organisms are able to destruct or
convert cyanide successfully and are able to entrap cationic metals (i.e. Cu, Ni, Fe, Ni, etc) within
their structure due to the anionic properties associated with biofilms (Singh et al., 2006). A number of
studies of free cyanide, metal-complexed cyanides and thiocyanate have been reported elsewhere with
successful operational efficiencies (Jeong and Chung, 2006, Patil and Paknikar, 2000, Mekuto et al.,
2015, Van Zyl et al., 2011, van Zyl et al., 2014)
The Homestake Mine has proven that this process is robust and economically viable when the mine
commissioned a biological process that would destruct cyanidation wastewater in 1984 (Stott et al.,
2001). However, the challenge with this process is that traditional cyanide-sensitive nitrifying and
denitrifying organisms were employed and often, this resulted in process failures when cyanide
compounds overflowed to the secondary stages. More recently, Gold Fields Limited has developed a
pilot plant that treats thiocyanate containing wastewater through the Activated Sludge Tailings
Effluent Remediation (ASTERTM) process with the overall aim of recycling the process water to the
BIOX® process (van Buuren et al., 2011). However, this process is mainly limited to the destruction
of thiocyanate and relies on chemical methods to remediate the remaining contaminants.
These industrial applications demonstrate that biological processes are sustainable and economically
viable. These processes can be operated for longer periods since the microbial species are adaptable
and can be manipulated to fit under different operational and environmental conditions, thus allowing
for uptake, treatment, sorption, and/or precipitation of thiocyanate, cyanide, ammonia, heavy metals
and sulphates. Although this process has been proven to be economically viable, industry still prefers
the use of chemical methods instead of the biological technique. This may be attributed to the rapid
degradation rates that are observed in such processes and the development of biological processes can
be time consuming and requires additional research on wastewaters generated from different ore types
(Mosher and Figueroa, 1996).
6. Integrated biological treatment of cyanide
Typical industrial bioprocesses that focus on treatment of cyanidation wastewater, such as that
employed at the Homestake Mine, consist of two separate but integrated treatment stages. The first
stage consists of cyanide and/or thiocyanate degrading microorganisms where the biodegradation
products are generally dominated by the presence of bicarbonate/formate/carboxylic acid (depending
on the degradation pathway that a microbe employs), ammonia (Eq. 13) and sulphates (Eq. 14).
However, in order for these microorganisms to propagate and degrade cyanide, they need nutrients in
the form of carbon, phosphorus and nitrogen (in this case, cyanide/thiocyanate).
¿+2O2+3H2O HCO3
Due to the toxicity and eutrophication promotion by the presence of ammoniated compounds, the
second stage of the treatment process is normally incorporated with an autotrophic nitrification and
anaerobic denitrification system, where the ammonia and nitrates are oxidised to nitrogen gas.
Ammonia is oxidised to nitrate in a two-step reaction (Eq. 15 and 16) where it is firstly converted to
nitrite and thereafter to nitrate
+¿+1.5O2 NO2
¿+0.5 O
The formed nitrates are immediately reduced to nitrogen gas through the action of nitrate reductase
enzymes. The process is normally anaerobic and consists of a diverse microbial community that are
either strict anaerobes or facultative anaerobes ((Knowles, 1982, Awolusi et al., 2015b, Awolusi et al.,
2015a, Zumft, 1997). The nitrifying and denitrifying organisms are non-competitive; hence, the
wastewater can be remediated successfully. However, nitrification and denitrification using traditional
microbial communities is very slow and highly sensitive to process shifts and the presence of cyanide
compounds. The treatment of cyanide containing wastewater is hardly 100%, especially during cold
seasons, hence cyanide degradation efficiencies decrease, thus resulting in the overflow of cyanide to
secondary stages: nitrification and denitrification stages. The nitrifying and denitrifying organisms are
highly sensitive to the presence of cyanide, thus the presence of this compound on these systems
would have deleterious effects, causing decrease in effluent water quality (Wild et al., 1994, Kim et
al., 2011b). Some of the consulted literature on the effects of cyanide inhibition on nitrification and
denitrification has been discussed in Section 4.
Recently, research has focused on microorganisms that are able to conduct both nitrification and
denitrification aerobically and under heterotrophic conditions, replacing the traditional autotrophic
nitrification and anaerobic denitrification organisms. Research on these organisms gained popularity
in the late 1980s and early 1990s. Earlier studies revealed that these organisms were characterised by
the absence of nitrites and/or nitrates accumulation (Robertson and Kuenen, 1990), which
contradicted the autotrophic nitrification system that relied on the nitrites and nitrates accumulation as
proof of successful nitrification. This is due to the fact that most heterotrophic nitrifiers are also
denitrifiers, and the nitrites or nitrates produced from the nitrification process may be simultaneously
converted to nitrogen gas or one of the volatile nitrogen oxides (van Niel et al., 1992). Hence, the
concentration of oxidised ammonia may not necessarily reflect the concentration of the oxidation
products that are accumulating in the medium. However, it has been observed that these organisms
share similar enzymological characteristics and activity with the traditional nitrifiers and denitrifiers.
It has, however, been the aerobic denitrifying enzymes that have been associated with controversy
(Gupta, 1997). Although the technology received criticism in the past, it has been accepted by the
scientific community as one of the viable alternatives to traditional systems.
To date, a number of organisms that are capable of heterotrophic nitrification and aerobic
denitrification have been isolated and identified. These organisms include Pseudomonas stutzeri T13
(Sun et al., 2015), Alcaligenes faecalis C16 (Liu et al., 2015), Providencia rettgeri YL (Taylor et al.,
2009), Agrobacterium sp. LAD9 (Chen and Ni, 2012), Rhodococcus sp. CPZ24 (Chen et al., 2012),
Bacillus methylotriphicus L7 (Zhang et al., 2012) and so on. During heterotrophic nitrification and
aerobic denitrification, alkalinity is produced during aerobic denitrification, which can be used to
neutralise the acidity that is generated through the heterotrophic nitrification, thus reducing the costs
of pH-adjustment (He et al., 2016).
However, due to the susceptibility of these organisms to cyanide compounds, recent studies have
focused on the application of the same cyanide degrading organisms for heterotrophic nitrification and
aerobic denitrification. Mekuto et al. (2015) observed simultaneous nitrification and aerobic
denitrification in a cyanide degradation process using the same microbial species while in a separate
study, a cyanide-insensitive microbial community was able to nitrify and aerobically denitrify in the
presence of cyanide (Mpongwana et al., 2016), thus demonstrating the effectiveness of such microbial
communities in a cyanide biodegradation process. Hence, an overall cyanide-insensitive biological
process is proposed where all the stages, viz cyanide biodegradation stages, and nitrification and
aerobic denitrification stages, are inoculated with the same microbial community for successful
cyanide degradation as demonstrated in Fig. 1. Nitrification and denitrification stages can be
expanded by incorporating a pre-denitrification stage, to reduce the toxicity of the produced
ammonium and nitrates to the nitrification and denitrification stages. This would ensure maximum
utilisation of these contaminants in the nitrification and aerobic denitrification stages. This
phenomenon has been observed elsewhere (Kim et al., 2011a, Villemur et al., 2015), with successful
operational efficiencies.
Fig.1: A proposed cyanide degradation system
However, the successful operation of such a process will require a thorough knowledge and
understanding of the fundamental and operational aspects of the proposed process. One of the key
fundamentals is a thorough understanding and characterisation of the microbial community that
contributes to the success of the bioprocess. This information would enable development of predictive
models (Stott et al., 2001, Whitlock, 1990). The success of the Homestake Mine in treating cyanide-
containing wastewater was based on the understanding of the microbial communities that were
involved in the process. The initial assessment that was employed using continuous stirred tank
reactors (CSTRs) showed unfavourable results, owing to limited knowledge and understanding of the
microbial population used. Subsequent work using CSTRs and rotating biological contactors (RBCs)
with a different microbial consortium, showed encouraging results for cyanide degradation. Recently,
the microbial population that is involved in the ASTERTM process was observed to be far more
complex than initially reported by van Buuren et al. (2011) , and is comprised of a diverse microbial
species such as bacteria, motile eukaryotes, filamentous fungi and algae (Huddy et al., 2015).
Some of the identified culturable microbial organisms that are capable of cyanide degradation are
listed in Table 3. However, it is well known that only 1% of the environmental microbial species are
able to be cultured in a laboratory, hence the detection of these organisms through culture-based
techniques may not truly represent the microbial organisms that contribute to cyanide degradation.
Therefore, there is a need for robust and reliable techniques such as the clone library approach (Huddy
et al., 2015) and/or metanogenomic sequencing for accurate identification of microbial populations
and their metabolic roles within the process (Handelsman, 2004, Schloss and Handelsman, 2003,
Cowan et al., 2005). The microbial communities involved in the ASTERTM process and their intrinsic
metabolic contributions to the process, have been recently elucidated elsewhere (Kantor et al., 2015).
The authors observed the dominance of Thiobacillus sp. whose genomes harbour unreported operon
for thiocyanate degradation. Furthermore, the microbial community was observed to be largely
autotrophic through genome-based metabolic predictions, with a smaller portion of the community
being heterotrophic. Such fundamental knowledge of bioprocesses adds value to the effectiveness of
the process.
The operational research is equally imperative as is the fundamental research, as it also determines the
success of the bioprocess. Desirable operational parameters such as temperature, influent cyanide
concentration and aeration should be considered, as these external factors contribute to the type of
degradation pathway employed by the microorganisms (Dash et al., 2009). For example, the presence
of cyanide, either as free cyanide or as complexed cyanide, can result in the activation of one or two
enzymatic pathways for the degradation of cyanide – depending on the microbial species employed in
such a process and in such environmental conditions (Ebbs, 2004). In addition, the solubility and
bioavailability of cyanide in soil-water systems also influences the selection of a desired pathway by
the microorganisms (Dash et al., 2009). Overall, the combination of both fundamental and operational
research would ensure sustainability, reliance and robustness of the process that would effectively
eliminate cyanide presence in cyanidation wastewater through an integrated approach of using the
same microbial population for cyanide biodegradation, nitrification and aerobic denitrification. Thus,
these organisms can be referred to as cyanide degraders, nitrifiers and aerobic denitrifiers. The
generated waste sludge can be digested anaerobically to produce methane, which can assist on energy
inputs of the process. Alternatively, the waste sludge can be utilised as a source of manure for
agricultural activities.
Table 3: Culturable microbial species capable of degrading cyanide compounds.
Microorganism Operation C-source N-source Product(s) Temp (°C) pH Reference
Pseudomonas sp. Batch Whey CNWAD NH4, CO230 9.2-11.4 (Akcil et al., 2003)
Bacillus pumilus C1 Batch/fed-batch Nutrient broth NaCN - 25 10.5 (Meyers et al., 1991)
Fusarium solani Batch Glucose K2Ni(CN)4, KCN NH4, HCOOH 25 7.0 (Barclay et al., 1998)
Scenedesmus obliquus Batch NaCN NaCN NH4,CO2 - 10.3 (Gurbuz et al., 2009)
Burkholderia cepacia C-3 Batch Glucose, Fructose KCN, KSCN,
NH3, HCOOH 30 10 (Adjei and Ohta, 1999)
Azotobacter vinelandii Continuous Cassava NaCN NH3, CH430 7-8.5 (Kaewkannetra et al., 2009)
Klebsiella oxytoca Batch Glucose CN, SCN NH4, CH430 7.0 (Kao et al., 2003)
Pseudomonas fluorescens Batch Glucose Fe(CN)6NH4, CO225 5.0 (Dursun et al., 1999)
Trametes versicolor Batch Citrate KCN NH4,CO2 30 10.5 (Cabuk et al., 2006)
Trichoderma ssp. Batch Glucose CN- NH4,CO2 25 6.5 (Ezzi and Lynch, 2005)
Aspergillus awamori Continuous Citrus sinensis waste KCN NH4,HCOOH 40 8.84 (Santos et al., 2013)
Thiobacillus sp. Batch CO2SCN NH4, CO2, SO420-40 6-9.3 (Stott et al., 2001)
Klebsiella sp. Batch SCN SCN NH4, SO4, 38 7.0 (Ahn et al., 2005)
Halomonas sp. Batch Glucose, Fructose,CO2, Acetate SCN NH4,CO2, SO420-40 6-9.3 (Stott et al., 2001)
Pseudomonas pseudoalcaligenes Batch Acetate NaCN NH4, CO230 9.5-10.0 (Huertas et al., 2010)
Rhodococcus sp. Batch KCN KCN NH4, HCOOH 30 (Maniyam et al., 2013)
Pseudomonas stutzeri,
Pseudomonas putida
Batch Lactate, Sucrose KCN, KSCN NH4, SO42-, CO228-30 9.0-9.2 (Karavaiko et al., 2000)
Fusarium oxysporumCCMI 876 Continuous Czampek broth CN- NH4, HCOOH 30 8.0 (Campos et al., 2006)
Batch SCN SCN - 35 7.0-8.0 (Sorokin et al., 2014)
Pseudomonas sp.
Citrobacter sp.
Continuous Glucose
Sugarcane mollasses
- 35 7.5 (Patil and Paknikar, 1999)
Pseudomonas aeruginosa STK O3 Batch Mollasses SCN, KCN NH4, SO42- 30 8.5, 10.0 (Mekuto et al., 2016)
Pseudomonas putida Batch Glucose K2[Ni(CN)4] NH4, CO230 7.0 (Silva-Avalos et al., 1990)
Acremonium strictum Batch SCN SCN NH4, SO42- 25 - (Kwon et al., 2002a)
Cryptococcus humicolus Batch Glucose K2[Ni(CN)4] NH4, HCOOH 25 - (Kwon et al., 2002b)
Methylobacterium thiocyanatum Batch Glucose SCN NH3, CO2, KOH 30 - (Wood et al., 1998)
Klebsiella pneumoniae
Ralstonia sp.
Batch Glucose SCN NH4, SO4237 6.0 (Chaudhari and Kodam, 2010)
Thiobacillus thioparus Batch CO2SCN NH3, COS* 30-40 7.0 (Katayama et al., 1992)
Acinetobacter johnsonii
Pseudomonas diminuta
Batch SCN NH4, SO4228 7.6 (Boucabeille et al., 1994)
Paracoccus thiocyanatus - SCN SCN - 30-37 7.5-8.0 (Katayama et al., 1995)
Burkholderia phytofirmans Batch Acetate SCN NH4, SO4225 6.5 (Vu et al., 2013)
Micractinium sp. Batch NaHCO3SCN NH4, SO42- 8.2 (Ryu et al., 2015)
Thiohalobacter thiocyanaticus Batch NaHCO3SCN COS, NH330 7.3-7.5 (Sorokin et al., 2010)
Batch NaHCO3SCN COS, NH330 9.9 (Sorokin et al., 2004)
Thiohalophilus thiocyanoxidans Batch NaHCO3SCN COS, NH337 7.5 (Bezsudnova et al., 2007)
*COS – Carbonyl Sulfide
7. Conclusion
The currently utilised physical and chemical treatments for the treatment of cyanidation wastewater
present disadvantages, such as high costs and the additional environmental burden that these
processes pose. Although these processes are characterised by rapid degradation rates, however they
are unable to treat certain cyanide complexes. A biological treatment process is found to be the most
effective, robust, environmentally benign and cost effective method for the destruction of all cyanide
related compounds. However, this process has been hindered by the use of cyanide sensitive microbial
species in secondary stages that carry out the nitrification and denitrification processes, thus rendering
the process ineffective, especially during cold seasons. Hence, this review proposes the utilisation of
the same cyanide-insensitive microbial consortium throughout the process for the destruction of
cyanide compounds and its biodegradation by-products heterotrophically. This would improve
process efficiencies even in cases where cyanide compounds overflow to secondary stages.
Additionally, this would ensure process water re-use to upstream bioleaching circuits, thus preserving
the use of excessive water. This type of process would be beneficial for operations that are in areas
with negative water balances. The recent full-scale industrial projects on cyanide biodegradation has
have demonstrated the robustness and competitiveness of this process in comparison to the existing
chemical and physical treatment technologies, and in future, the process might add economic value, as
some cyanide degrading organisms produce biogas from cyanide biodegradation.
8. Acknowledgements
The authors would like to acknowledge the funding from the Cape Peninsula University of
Technology (CPUT), University Research Fund (URF RK 16) and National Research Foundation
9. References
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Geomicrobiology Journal,,%
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Environmental Science and Pollu!on Research,,%
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,= B +?>>  G*)= > 314 7 ( G*)= B  <"  ! < 7-"
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... Wastewater has been reported by experts to contain different ranges of pollutants [8][9][10][11][12] which when released directly into the environment without effective treatment, pose health threats to humans and aquatic organisms [1,10,[13][14][15] thereby causing environmental degradation [16]. Contaminants are released from both point and nonpoint sources into water bodies [17,18]. ...
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One of the biggest challenges of the 21st century is the shortage of potable water coupled with increasing water pollution, industrialization and, an ever increasing population. Water is needed to sustain all socioeconomic activities which render it vital for life support on earth. Owing to these circumstances, it has become imperatively significant to develop an active method to monitor and control these pollutants in the aquatic environment. Recently, different materials made of carbon nanomaterial have been widely used to construct different types of electrical electrodes to make biosensors and electrochemical sensors. Some of the materials made from carbon nanomaterial s include but not limited to; carbon nanotubes, graph-eme, carbon nanohorns, and carbon black. Carbon nanotubes (CNTs) have contributed to the production of active electrochemical sensors/filters as an effective alternative technique in the field of water pollution control. CNT biofilters are generally known to absorb organic and chemical pollutants as a result of their intrinsic characteristics of high flexibility, effective stability , and wide surface area. They also exhibit the quality of electro-oxidation of the adsorbed pollutants which has been attested as potential water and wastewater treatment technology in different laboratory experiments. Active electrochemical CNT has also aided in the stability, sensitivity, and selectivity of filters/sensors. In the field of nanotechnology, CNTs have been discovered to display great water and wastewater treatment potentials due to their superb physiochemical characteristics. The modern technologies that have focused on the utilization of CNTs in the area of water and wastewater treatment technology predominantly used the carbon-based material as membranes or filters, adsorbents, electrodes and catalyst to degrade pollutants in water or wastewater. This study intends to explore and make a general overview of the role of nanotechnology, based on CNTs in water and wastewater treatment.
... With the rapid development of social economy, many urban cities are faced with the problem of deteriorating water environment, which urges people to treat wastewater generated from different production activities. Phenolic compounds (phenol, nitrophenol, methylphenol, etc.) , heavy metals (Cr, Hg, Cu, Mn, etc.) (Zhu et al., 2021), cyanide (Mekuto et al., 2016), and oily compounds (Abdelnasser Abidli, 2020) are usually found in industrial wastewater. Phenol is a common phenolic pollutant, which is widely present in wastewater from various industries, such as oil refineries, coal processing, and petrochemical product manufacturing (Busca et al., 2008). ...
In this work, the removal of ammonia nitrogen and phenol by pulsed discharge plasma (PDP) and modified zeolite was investigated. The Fe-zeolite and Mn-zeolite catalysts were prepared by the impregnation method. Catalysts’ morphology, specific surface area, and chemical bond structure were characterized. Based on the pollutants removal experiments, Fe-zeolite (0.01) in the PDP system had better catalytic oxidation of phenol and adsorption effect of ammonia nitrogen. The removal efficiency of the pollutants increased with the increase of discharge voltage and solution conductivity, but decreased with the increase of discharge distance. During the plasma discharge process, the pH value in the solution decreased, and the solution conductivity gradually increased. After PDP/Fe-zeolite system treatment, the toxicity of the wastewater was significantly reduced. This study provided a new treatment method for inorganic and organic pollutants treated by PDP.
... Different degradation techniques have been proposed for the treatment of thiocyanate, including biological treatment [8][9][10][11][12][13], adsorption [14][15][16], or chemical methods [6,[17][18][19][20]. However, the above methods require long treatment times, large tank capacities, and have limitations on parameters such as concentration range, pH, temperature, and solid contents. ...
The degradation of thiocyanates (SCN⁻) by UV-C-activated persulfate (PS) in the presence of ferric ion (Fe³⁺) was investigated. As a source of monochromatic far-UV-C irradiation (222 nm), mercury-free KrCl excimer lamp was used. Results showed that compared with direct photolysis, UVC/ PS and PS/ Fe³⁺, the combined UVC/ PS/ Fe³⁺ treatment had the highest initial reaction rate ω0 and removal efficiency. 99.99% conversion of thiocyanates (100 mg/L of initial concentration) was achieved in 40 min. The addition of Fe³⁺ in the UVC/ PS treatment was found to reduce energy consumption (calculated as amount of oxidized thiocyanates per consumed electrical energy) by 4.5 times, while only a 30% difference between direct photolysis and UVC/ PS was observed. The high efficiency of the UVC/ PS/ Fe³⁺ process revealed a synergistic effect (synergy index ƒ=1.98). The effect of the initial SCN⁻, PS, and Fe³⁺ molar ratios and UV-C exposure time on SCN⁻ removal in UVC/ PS/ Fe³⁺ was further investigated. It was found that at molar ratios [S2O82–]:[SCN−] = 3:1 and [S2O82–]:[Fe³⁺] = 1:0.1, effective decomposition of SCN⁻ in a wide initial concentration range (from 50 to 500 mg/L or 0.86 to 8.6 mM) can be achieved. The strong role of •OH and SO4•− in the removal of SCN⁻ was confirmed by the addition of radical scavengers. It was demonstrated that the presence of Cu²⁺ in simulated gold mine wastewater effluents neutralizes the inhibitory effects that S2O32– and NH4⁺ have on the degradation process.
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Cyanide compounds are hazardous compounds which are extremely toxic to living organisms, especially free cyanide in the form of hydrogen cyanide gas (HCN) and cyanide ion (CN−). These cyanide compounds are metabolic inhibitors since they can tightly bind to the metals of metalloenzymes. Anthropogenic sources contribute significantly to CN− contamination in the environment, more specifically to surface and underground waters. The treatment processes, such as chemical and physical treatment processes, have been implemented. However, these processes have drawbacks since they generate additional contaminants which further exacerbates the environmental pollution. The biological treatment techniques are mostly overlooked as an alternative to the conventional physical and chemical methods. However, the recent research has focused substantially on this method, with different reactor configurations that were proposed. However, minimal attention was given to the emerging technologies that sought to accelerate the treatment with a subsequent resource recovery from the process. Hence, this review focuses on the recent emerging tools that can be used to accelerate cyanide biodegradation. These tools include, amongst others, electro-bioremediation, anaerobic biodegradation and the use of microbial fuel cell technology. These processes were demonstrated to have the possibility of producing value-added products, such as biogas, co-factors of neurotransmitters and electricity from the treatment process.
With the rapid development of society, wastewaters, such as cyanide-containing wastewaters (CBWs) have caused environmental problems. In the present work, an electrochemical approach using sacrificial Zn anode was investigated for the removal of cyanides from CBWs. The effects of operational parameters, such as current density, pH, initial cyanide concentration, and ionic strength, on the cyanides removal from synthetic solution were discussed in turn. Under the optimal conditions obtained, the treatment of industrial CBWs was considered. Subsequently, more attentions were paid to elucidate the removal mechanisms of cyanide ions and the corresponding metal-cyanide (copper and iron) complexes by a combination of cyclic voltammetry (CV), pHPZC, X-Ray diffractometer (XRD), scanning electron microscope with energy disperse spectroscopy (SEM/EDS) and X-ray photoelectron spectrometer (XPS) characterizations. Experimental results demonstrated that the removal efficiency of total cyanide (CNT), Cu, and Fe from industrial CBWs are 98%, 91%, and 96%, respectively, with an anode consumption of 1.78 kg/m³ and energy consumption of 2.50 kW·h/m³, of which 72% of Cu was collected on the cathode and almost all of Fe was in the precipitate. Removal mechanisms suggested that free cyanide (CN⁻) mainly presented as Zn(CN)2 into the electrolytic precipitate. The replacement of Zn²⁺ and electroreduction promoted 72% of Cu(I) to be reduced, while 57% of the remaining portion of Cu(I) was oxidized to Cu(II) and 43% of Cu(I) could form CuSCN into the precipitate by XPS analysis. With respect to Fe, it was mainly ascribed to the formation of Zn2Fe(CN)6 precipitate to be removed. From above, this work provided a method for enriching cyanides from wastewaters into the precipitate, while valuable metal Cu was deposited on the cathode, facilitating the separation and recovery of valuable resources.
The construction of a low-cost and green copper recovery method in gold plants is extremely important for treating CuSCN-containing refractory acidified sediments (ASs). In this work, an effective and environmentally-friendly hydrometallurgical process, consisting of thiosulfate leaching and protective electrodeposition, was proposed for selective copper recovering from ASs. The effects of leaching parameters on the leaching efficiency were systematically investigated. Subsequently, cyclic voltammetry, linear sweep voltammetry, and chronoamperometry were utilized to determine the predominant speciation and investigate the electrochemical behavior of the leachate. Eventually, the optimum electrolysis parameters for copper recovery were considered. The results demonstrated that the selective leaching efficiency of copper reached 99% under optimum leaching conditions, while no zinc and iron were leached, which coincides with the results of speciation calculations. Simultaneously, the kinetic analysis indicated that the leaching process exhibited a diffusion-controlled step with an apparent activation energy of 13.64 kJ/mol. Na2Zn3Fe(CN)6 as the main ingredient of the leaching residue was formed, which played a vital role in facilitating the copper leaching process. The electrochemical measurements indicated that the [Cu(S2O3)3]⁵⁻ complex was the dominant species in the pH range from 7.0 to 11.0, and the mass transfer diffusion was the main resistance element for the cathodic reduction of the [Cu(S2O3)3]⁵⁻ complex. Regarding the S2O3²⁻ stability, it decomposed more easily into S, S²⁻, S4O6²⁻, and SO4²⁻ in the presence of Cu²⁺. However, the presence of SO3²⁻ significantly alleviated thiosulfate decomposition. The electrodeposition experiments showed that metallic copper was obtained with a recovery efficiency of 95%, current efficiency of 50%, and energy consumption of 1.25 kW h/kg Cu. This research developed an approach for the efficiently recycling ASs, yielding great economic results and more favorable environmental performance.
Cyanide tailings(CT) are a typical hazardous waste, containing large amounts of heavy metals and highly toxic cyanide. As so far, it is difficult to solve this problem at the same time. In this study, a “two-step” process was proposed for the first time to treat CT, first using microorganisms to degrade cyanide and then using microbially induced carbonate precipitation (MICP) to solidify the CT. We isolated a bifunctional bacterium which exhibited cyanide degradation and high urease activity from CT, identified as Aneurinibacillus tyrosinisolvens strain (named JK-1). We used JK-1 bacteria to treat the CT in a “two-step” process. The results showed that the degradation of free cyanide (F–CN) and total cyanide (T-CN) in CT by JK-1 bacteria reached 94.54% and 88.13%. After the MICP treatment, the spindle-shaped CaCO3 solidified the CT into a block of calcite and sphalerite crystals, and the uncompressed compressive strength (UCS) reached 0.74 MPa; the morphology of heavy metals in the CT changed from the exchangeable state to the carbonate-bound state, and the mobility was significantly reduced. Compared with chemical treatment, the treatment of CT by the new process is highly efficient and green, which can realize the solidification of CT, degradation of cyanide and immobilization of heavy metals at the same time. Compared with chemical treatment, the new process is efficient and green, which can realize the solidification of CT, the degradation of cyanide and the immobilization of heavy metals at the same time; it is of great significance to the harmless treatment of CT and the sustainable development of gold smelting industry.
A systematic and comparative research was performed using copper, nickel and cobalt-ammonia catalyzed thiosulfate processes for environmentally friendly and efficient gold extraction from an oxide gold concentrate. 88.3% of gold was leached by thiosulfate with conventional copper-ammonia catalysis while its thiosulfate consumption was as high as 53.6 kg/t-concentrate. Thiosulfate consumption was reduced to 26.5 and 30.6 kg/t-concentrate, and 81.2% and 90.2% of gold leaching percentages were obtained with nickel-ammonia and cobalt-ammonia catalyses. The electrochemical leaching mechanism combined with XPS and SEM-EDS analyses illustrated the essential reason for the differences of three catalyses in gold leaching percentage and thiosulfate consumption. For copper-ammonia catalysis, two-stage desorption was required to desorb gold from the gold-loaded resin due to the co-adsorption of copper with gold while one-stage desorption was feasible for nickel-ammonia and cobalt-ammonia catalyses because nickel and cobalt did not co-adsorb with gold, and their gold recovery percentages were 82.4%, 98.5%, 97.7%, respectively. After 6-time cycles of barren leachate, the gold leaching and recovery percentages of copper-ammonia catalysis considerably decreased while those of nickel-ammonia and cobalt-ammonia catalyses only declined slightly. Thiosulfate leaching with cobalt-ammonia catalysis has the advantages of high gold leaching percentage, low thiosulfate consumption, simple desorption process, good leachate stability and recycle performance of barren leachate, and therefore it has a bright industrial application prospect.
Cyanide tailings are industrial hazardous solid wastes arising from gold mining industry. Hundreds of millions of tons of cyanide tailings that contain highly toxic cyanides and various valuable elements, such as gold, silver, iron, sulfur, copper, lead, and zinc are generated and discharged to tailing dams every year. Significant efforts have been undertaken to develop efficient detoxification and utilization technologies to reduce hazardous wastes and recover the valuable element in cyanide tailings. In this paper, the sources and characteristics of cyanide tailings are introduced. The technologies using various physical, chemical, and biological methods or a combination thereof to detoxify and utilize cyanide tailings are reviewed. However, the complexity of cyanide tailings and the high cost of treatment may seriously restrict their industrial application. It seems that thermal treatment with catalysts and autoclaved hydrolysis are certainly promising technologies for the detoxification of cyanide tailings with the removal rate of cyanides more than 99%, which can be good for the cleaner production of gold mining. The current research status and the obstacles in the recovery of cyanide, gold-silver, sulfur-iron, copper-lead-zinc, and “low-value content” from cyanide tailings are then reviewed in detail. These processes can be used independently or in conjunction with other treatment methods depending on the nature of cyanide tailings. The detoxification and comprehensive utilization of cyanide tailings would ultimately bring economic and environmental benefits.
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The continuous discharge of cyanide-containing effluents to the environment has necessitated for the development of environmentally benign treatment processes that would result in complete detoxification of the cyanide-containing wastewaters, without producing additional environmental toxicants. Since biological detoxification of hazardous chemical compounds has been renowned for its robustness and environmental-friendliness, the ability of the Exiguobacterium acetylicum (GenBank accession number KT282229) and Bacillus marisflavi (GenBank accession number KR016603) to co-metabolise thiocyanate (SCN⁻) and free cyanide (CN⁻) under alkaline conditions was evaluated. E. acetylicum had an SCN⁻ degradation efficiency of 99.9 % from an initial SCN⁻ concentration of 150 mg SCN⁻/L, but the organism was unable to degrade CN⁻. Consequently, B. marisflavi had a CN⁻ degradation efficiency of 99 % from an initial concentration of 200 mg CN⁻/L. Similarly, the organism was unable to degrade SCN⁻; hence, this resulted in the evaluation of co-metabolism of SCN⁻ and CN⁻ by the two microbial species. Optimisation of operational conditions was evaluated using response surface methodology (RSM). A numeric optimisation technique was used to evaluate the optimisation of the input variables i.e. pH, temperature, SCN⁻ and CN⁻ concentrations. The optimum conditions were found to be as follows: pH 9.0, temperature 34 °C, 140 mg SCN⁻/L and 205 mg CN⁻/L under which complete SCN⁻ and CN⁻ degradation would be achieved over a 168-h period. Using the optimised data, co-metabolism of SCN⁻ and CN⁻ by both E. acetylicum and B. marisflavi was evaluated, achieving a combined degradation efficiency of ≥99.9 %. The high degradative capacity of these organisms has resulted in their supplementation on an active continuous biological degradation system that is treating both SCN⁻ and CN⁻.
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Cyanides (CN(-)) and soluble salts could potentially inhibit biological processes in wastewater treatment plants (WWTPs), such as nitrification and denitrification. Cyanide in wastewater can alter metabolic functions of microbial populations in WWTPs, thus significantly inhibiting nitrifier and denitrifier metabolic processes, rendering the water treatment processes ineffective. In this study, bacterial isolates that are tolerant to high salinity conditions, which are capable of nitrification and aerobic denitrification under cyanogenic conditions, were isolated from a poultry slaughterhouse effluent. Three of the bacterial isolates were found to be able to oxidise NH4-N in the presence of 65.91 mg/L of free cyanide (CN(-)) under saline conditions, i.e. 4.5% (w/v) NaCl. The isolates I, H and G, were identified as Enterobacter sp., Yersinia sp. and Serratia sp., respectively. Results showed that 81% (I), 71% (G) and 75% (H) of 400 mg/L NH4-N was biodegraded (nitrification) within 72 h, with the rates of biodegradation being suitably described by first order reactions, with rate constants being: 4.19 h(-1) (I), 4.21 h(-1) (H) and 3.79 h(-1) (G), respectively, with correlation coefficients ranging between 0.82 and 0.89. Chemical oxygen demand (COD) removal rates were 38% (I), 42% (H) and 48% (G), over a period of 168 h with COD reduction being highest at near neutral pH.
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An alkali-tolerant bacterium, Pseudomonas aeruginosa STK 03 (accession number KR011154), isolated from an oil spill site, was evaluated for the biodegradation of free cyanide and thiocyanate under alkaline conditions. The organism had a free cyanide degradation efficiency of 80 and 32 % from an initial concentration of 250 and 450 mg CN−/L, respectively. Additionally, the organism was able to degrade thiocyanate, achieving a degradation efficiency of 78 and 98 % from non- and free cyanide spiked cultures, respectively. The organism was capable of heterotrophic nitrification but was unable to denitrify aerobically. The organism was unable to degrade free cyanide in the absence of a carbon source, but it was able to degrade thiocyanate heterotrophically, achieving a degradation efficiency of 79 % from an initial concentration of 250 mg SCN−/L. Further increases in thiocyanate degradation efficiency were only observed when the cultures were spiked with free cyanide (50 mg CN−/L), achieving a degradation efficiency of 98 % from an initial concentration of 250 mg SCN−/L. This is the first study to report free cyanide and thiocyanate degradation by Pseudomonas aeruginosa. The higher free cyanide and thiocyanate tolerance of the isolate STK 03, which surpasses the stipulated tolerance threshold of 200 mg CN−/L for most organisms, could be valuable in microbial consortia for the degradation of cyanides in an industrial setting.
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A hypothermia aerobic nitrite-denitrifying bacterium, Pseudomonas tolaasii strain Y-11, was found to display high removal capabilities for heterotrophic nitrification with ammonium and for aerobic denitrification with nitrate or nitrite nitrogen. When strain Y-11 was cultivated for 4days at 15°C with the initial ammonium, nitrate and nitrite nitrogen concentrations of 209.62, 204.61 and 204.33mg/L (pH 7.2), the ammonium, nitrate and nitrite removal efficiencies were 93.6%, 93.5% and 81.9% without nitrite accumulation, and the corresponding removal rates reached as high as 2.04, 1.99 and 1.74mg/L/h, respectively. Additionally, ammonium was removed mainly during the simultaneous nitrification and denitrification process. All results demonstrate that P. tolaasii strain Y-11 has the particularity to remove ammonium, nitrate and nitrite nitrogen at low temperatures, which guarantees it for future application in winter wastewater treatment.
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Membrane bioreactors (MBRs) are rapidly becoming the technology of choice over conventional activated sludge treatment systems due to their smaller footprint, reduced sludge production, rapid start-up of biological processes, complete removal of suspended solids and better effluent quality. The retention of sufficient amount of slow-growing nitrifiers makes it feasible for the MBRs to achieve strong tolerance against the shock loads with stable and highly efficient nitrogen removal. Various studies have focused on the ecophysiology of nitrifiers in MBRs as well as their distinctive operational parameters as well as their impact on the selection and activity of nitrifying community. Several techniques have been employed over the years to understand the nitrifying community and their interaction within the MBR system, which led to its modification from the initial design. This review focuses on the identification of optimal operational and environmental conditions for efficient nitrification in MBRs. The advantages and limitations of different techniques employed for investigating the nitrifying communities in MBRs are also emphasized.
Changes in algal and bacterial communities during thiocyanate (SCN−) decomposition in a microalga-mediated process were studied. Pyrosequencing indicated that Thiobacillus bacteria and Micractinium algae predominated during SCN− hydrolysis, even after its complete degradation. Principal components analysis and evenness profiles (based on the Pareto–Lorenz curve) suggested that the changes in the bacterial communities were driven by nitrogen and sulfur oxidation, pH changes, and photoautotrophic conditions. The populations of predominant microalgae remained relatively stable during SCN− hydrolysis, but the proportion of bacteria – especially nitrifying bacteria – fluctuated. Thus, the initial microalgal population may be crucial in determining which microorganisms dominate when the preferred nitrogen source becomes limited. The results also demonstrated that microalgae and SCN−-hydrolyzing bacteria can coexist, that microalgae can be effectively used with these bacteria to completely treat SCN−, and that the structure of the algal-bacterial community is more stable than the community of nitrifying bacteria alone during SCN− degradation.
Cyanidation tailings disposed of in a surface impoundment experience a loss of cyanide due to natural attenuation, which frequently reduces the cyanide concentration to very low levels. Quantifying cyanide losses in terms of impoundment geometry, local weather conditions and feed-solution chemistry has been largely empirical in spite of the fact that, in many cases, mining operations rely on surface impoundments to reduce cyanide to below an internally regulated concentration or below an effluent limitation. To permit a quantitative evaluation of cyanide losses in an impoundment, a computer simulation was developed to estimate the losses of free, weak acid dissociable (WAD) and total cyanide due to dissociation, photolysis and volatilization. Results of the model are compared with data collected for a North American tailings impoundment in 1998.
Cydnidation tailings disposed of in a surface impoundment experience a loss of cyanide due to natural attenuation, which frequently reduces the cyanide concentration to very low levels. Quantifying cyanide losses in terms of impoundment geometry, local weather conditions and feed-solution chemistry has been largely empirical in spite of the fact that, in many cases, mining operations rely on surface impoundments to reduce cyanide to below an internally regulated concentration or below an effluent limitation. To permit a quantitative evaluation of cyanide losses in an impoundment, a computer simulation was developed to estimate the losses of free, weak acid dissociable (WAD) and total cyanide due to dissociation, photolysis and volatilization. Results of the model are compared with data collected for a North American tailings impoundment in 1998.
Gencor has pioneered the commercialization of bioxidation of refractory gold ores. Development of the BIOX™ process started in the late 1970s at Gencor Process Research, in Johannesburg, South Africa. The successful development of the technology led to the commissioning of a BIOX™ pilot plant in 1984, followed by the first commercial BIOX™ plant at the Fairview mine in 1986 (van Aswegen et al. 1988). The BIOX™ process was fully commercialized in 1991 when the Fairview plant was expanded to treat the total concentrate production of the mine and the Edwards roasters were finally shut down.