The thiocyanate degrading microbial consortia used in this study were sampled from an active lab-scale SCN⁻ degradation system operated at the Bioresource Engineering Research Group (BioERG) laboratories, Cape Town, South Africa. The TDO’s were analyzed using the 16S rDNA amplicon gene sequencing approach to reveal the microbial species that were present within the microbial community and this data is available in Mekuto et al. [10].

For inoculum development, the reactor used had a height of 250 mm and a diameter of 104 mm with an operating volume of 1 L. Aeration was set to 0.4 L min⁻¹ and the system was operated at room temperature (20 to 22°C). Mixing was achieved using an overhead stirrer fitted with 2-x-four bladed Rushton-type impellers, set at 250 rpm. The sample was centrifuged at 14,000 rpm for 5 min to concentrate the microbial cells. The cells were washed twice with a phosphate buffer solution (pH = 7.0) and thereafter, the cells were re-suspended in sterile distilled water subsequent to inoculation (1% v/v) in a nitrogen-free minimal media (MM) (pH = 9.9) as described previously by Mekuto et al. [11], with glucose being used as a carbon source at a concentration of 1 gL⁻¹. The cells were grown for a period of 48 h at a temperature of 30°C in 250 mL Erlenmeyer flasks containing 200 mL of MM. This culture was then used to study the potential of the organisms to conduct heterotrophic nitrification and aerobic denitrification under cyanogenic conditions.

Similarly, the CN⁻ degrading consortium was sampled from a stock reactor treating CN⁻ containing wastewater operated at the BioERG laboratories, Cape Town, South Africa. The configuration of the reactor was as described in section 2.1. This mixed microbial community were identified using 16S rDNA amplicon gene sequencing and the data is available in [10]. These bacterial species were previously isolated from CN⁻ containing wastewater and were previously found to be cyanide degraders by [12], but were unable to degrade thiocyanate (data not shown). The inoculum was derived from the reactor using the procedures highlighted in section 2.1.

The collected samples were centrifuged at 14,000 rpm for 5 min and then analysed for ammonium, nitrate, nitrite, CN⁻ using Merck ammonium (00683), nitrate (14773), nitrite (14776), cyanide (09701) and sulphate (00617) test kits. The basis behind the mechanism of determination of these compounds using Merck test kits has been elucidated by Mekuto et al. [11]. Thiocyanate was determined using the ferric method [16]. The pH was measured using a Crison Basic20 pH meter, which was calibrated daily.

The degradation of SCN⁻ and CN⁻ by TDO and CDO was assessed in batch systems under alkaline conditions. The TDO were sampled from an active SCN⁻ biodegradation system and were observed to have a SCN⁻ degradation efficiency of 100 % after 96 h (Fig. 1(a)), resulting in ammonium nitrogen and sulphate formation. The sulphate concentration increased to 90 mg SO₄ ²⁻L⁻¹ at 72 h and thereafter, plateaued. Additionally, ammonium nitrogen increased logarithmically after 27 h to 104 mg NH₄ ⁺-N.L⁻¹. There was an observed pH decrease during SCN⁻ degradation, which declined from an initial pH of 9.9 to 7.53. This was due to the acid produced from the oxidation of reduced sulphur species, thus resulting in a pH decrease. This is one of the few studies reporting on the degradation of SCN⁻ under alkaline conditions since most studies are focused on the degradation of SCN⁻ under neutral pH conditions [17, 18]. Wastewaters generated from metallurgical processes are mostly alkaline [19]; hence, studies undertaken in such conditions are vital. Sorokin et al. [20] demonstrated the capacity of the Halomonas and Thioalkalivibrio species to biodegrade SCN⁻ under alkaline conditions and observed that these microbial species were able to degrade SCN⁻, producing cyanate, which was subsequently utilised by the organisms. Furthermore, Thioalkalivibrio thiocyanodenitrificans degraded SCN⁻ aerobically and anaerobically under alkaline conditions, demonstrating the effectiveness of the applied microbial organisms [21].

The CDO were maintained in CN⁻ containing media prior to preparation of the inoculum, which resulted in increased capacity of the organisms to biodegrade as they were able to achieve a BRE of 99.9% after 170 h (Fig. 1(b)). CN⁻ decreased drastically during the initial 24 h of incubation and this observation may be due to the secretion of extracellular cyanide-degrading enzymes, a phenomenon which was observed by [12]. From the same study [12], similar microbial species which were not adapted to CN⁻, resulted in a BRE of only 57% from an initial concentration of 300 mg CN⁻.L⁻¹. This observation demonstrated the effectiveness of microbial adaptation for improved capacity of the consortia to biodegrade. However, the CDO were observed to be unable to degrade SCN⁻ (data not shown). Overall, there was an insignificant decrease in pH during the biodegradation process.

Heterotrophic nitrification under non- SCN⁻ containing conditions by the members of the TDO community resulted in a nitrification rate of 1.59 mg NH₄ ⁺-N.L⁻¹.h⁻¹ while there was an observed inhibition on heterotrophic nitrification, where the nitrification rate decreased to 0.95 and 0.44 mg NH₄ ⁺-N.L⁻¹.h⁻¹ in media containing 25 and 50 mg SCN⁻.L⁻¹, respectively (Fig. 2(a)). After nitrification, SCN⁻ was not detected in the media (data not shown), suggesting its complete biodegradation. Nitrification rates under CN⁻ concentration of 5, 10 and 15 mg CN⁻.L⁻¹ were observed to be 1.14, 0.57 and 0.32 mg NH₄ ⁺-N.L⁻¹.h⁻¹, respectively (Fig. 2(b)). This demonstrated the acute toxicity and inhibition of nitrification by CN⁻. This phenomena was observed elsewhere [22], where 0.11 mg CN⁻.L⁻¹ was suggested to be an accepted concentration for nitrification to transpire. The presence of both CN⁻ (varying concentration) and SCN⁻ (25 mg SCN⁻.L⁻¹) proved to be detrimental on nitrification by TDO’s as the nitrification rates decreased drastically. The observed nitrification rates were 1.0, 0.51 and 0.301 mg NH₄ ⁺-N.L⁻¹.h⁻¹ from the initial concentrations of 5, 10 and 15 mg CN⁻.L⁻¹ while the SCN⁻ concentration was kept at 25 mg SCN⁻.L⁻¹, respectively (Fig. 2(c)). The utilisation of ammonium nitrogen by the TDO’s resulted in the production of both nitrates and nitrites within the media. The maximum concentration of nitrates produced from the media supplemented with 0, 25, 50 mg SCN⁻.L⁻¹ and 5, 10, 15 mg CN⁻.L⁻¹ was 136, 44, 1.65 (Fig. 3(a)) and 68, 32.4 and 24.7 mg NO₃ ⁻-N.L⁻¹, respectively (Fig. 3(b)). Similarly, the maximum nitrate nitrogen concentration from the dual impact of SCN⁻ and CN⁻ was 46.5, 37 and 15 mg NO₃ ⁻-N.L⁻¹, respectively (Fig. 3(c)). The nitrite nitrogen concentration in all the tested media was less than 20 mg NO₂ ⁻N.L⁻¹. In this study, it was observed that the nitrate nitrogen formation rate from ammonium oxidation was not in a 1:1 ratio as defined by Eq. (3) This was due to the utilization of the ammonium as a nitrogen source by the consortium; hence, a portion of the ammonium nitrogen was used intracellularly by the employed microbial organisms, resulting in low nitrate nitrogen detection in the media. This was also observed when Alcaligenes faecalis (A. faecalis) No.4 heterotrophically oxidised ammonium nitrogen, resulting in the lower detection of nitrates within the media [3]. The authors observed that 50% of the ammonium nitrogen was utilised intracellularly by A. faecalis No.4 and almost 39 to 48% was denitrified simultaneously and/or immediately after formation. It is imperative to note that the percentage of ammonium nitrogen that is utilised intracellularly varies between species and strains. Additionally, heterotrophic nitrifying organisms have been observed to have the ability to simultaneously nitrify and denitrify aerobically in the same media [7, 23, 24]. Therefore, the low detection of nitrates in the media was attributed to its utilisation by the microbial species.

Heterotrophic nitrification by the CDO’s proved to be robust as these species were unaffected by the presence of both SCN⁻ and CN⁻. These organisms demonstrated high nitrification rates of 1.66, 1.67 and 1.47 mg NH₄ ⁺-N.L⁻¹.h⁻¹ in media spiked with 0, 25 and 50 mg SCN⁻.L⁻¹, respectively (Fig. 4(a)). Similarly, CN⁻ concentration did not have any inhibitory effect on CDO’s as they were able to achieve nitrification rates of 1.63, 1.66 and 1.47 mg NH₄ ⁺-N.L⁻¹.h⁻¹ from media spiked with 5, 10 and 15 mg CN⁻.L⁻¹, respectively (Fig. 4(b)). Furthermore, the organisms demonstrated their effectiveness as these species were not inhibited by the dual presence of SCN⁻ and CN⁻, achieving nitrification rates of 1.67, 1.58 and 1.39 mg NH₄ ⁺-N.L⁻¹.h⁻¹ from the initial concentrations of CN⁻ and SCN⁻ of 5, 10 and 15 mg CN⁻.L⁻¹ while the SCN⁻ concentration was kept at 25 mg SCN⁻.L⁻¹, respectively (Fig. 4(c)). Although there is a scarcity of information on alkaline nitrification, it has been observed that traditional nitrifying organisms are unable to perform nitrification efficiently under highly alkaline conditions [25, 26] but high nitrification efficiencies are at pH ranges of 7.0 to 8.5 [27]. This study has shown that both the TDO’s and CDO’s were able to nitrify under alkaline conditions. The maximum concentrations of nitrate nitrogen produced from the media supplemented with 0, 25, 50 mg SCN⁻.L⁻¹ and 5, 10, 15 mg CN⁻.L⁻¹ were 135, 126.5, 107.7 (Fig 5(a)) and 115.9, 111.2 and 88.7 mg NO₃ ⁻-N.L⁻¹, respectively (Fig. 5(b)). Similarly, the maximum nitrate nitrogen concentrations in cultures containing both SCN⁻ and CN⁻ were 100.95, 98.3 and 68.9 mg NO₃ ⁻-N.L⁻¹ (Fig. 5(c)). The nitrite concentration in all the tested media was less than 20 mg NO₂ ⁻-N.L⁻¹. These organisms, i.e. CDO’s, were also observed to conduct nitrification and aerobic denitrification in a continuous CN⁻ biodegradation process elsewhere [15], demonstrating their effectiveness as cyanide degraders, nitrifiers and denitrifiers.

The aerobic denitrification potential of the TDO’s and CDO’s was evaluated under cyanogenic conditions, from an initial nitrate nitrogen concentration of 100 mg NO₃ ⁻-N.L⁻¹. The TDO demonstrated the ability to denitrify aerobically as these species were able to achieve a complete denitrification rate of 0.66 mg NO₃ ⁻-N.L⁻¹.h⁻¹ in media supplemented with 0, 25, 50 mg SCN⁻.L⁻¹, respectively (Fig. 6(a)) with minimal detection of nitrites within the media (≤ 0.04 mg NO₂ ⁻-N.L⁻¹) (data not shown). Similarly, 5 mg CN⁻.L⁻¹ did not inhibit aerobic denitrification by these species since the denitrification rate exceeded 0.64 mg NO₃ ⁻-N.L⁻¹.h⁻¹. However, denitrification was inhibited by the presence of 10 and 15 mg CN⁻.L⁻¹, resulting in denitrification rates of 0.46 and 0.22 mg NO₃ ⁻-N.L⁻¹.h⁻¹, respectively (Fig 6(b)). Additionally, the dual effect of CN⁻ and SCN⁻ proved to be detrimental, especially when the media was spiked with 25 mg SCN⁻.L⁻¹ and 10 mg CN⁻.L⁻¹, as well as 25 mg SCN⁻.L⁻¹ and 15 mg CN⁻.L⁻¹, achieving denitrification rates of 0.39 and 0.11 mg NO₃ ⁻-N.L⁻¹.h⁻¹, respectively, while denitrification was unaffected by the presence of 25 mg SCN⁻.L⁻¹ and 5 mg CN⁻.L⁻¹ (Fig. 6(c)). The denitrification trends illustrated in Fig. 6(c) were similar to those observed in Fig. 6(b), which demonstrated the inhibitory effect that CN⁻ had on the denitrification by the TDO’s.

However, the CDO’s demonstrated a high denitrification potential, as these species were not affected by the presence of SCN⁻, CN⁻ or the dual effect of SCN⁻ and CN⁻. The denitrification rates were in excess of 0.67 mg NO₃ ⁻-N.L⁻¹.h⁻¹ irrespective of toxicant concentration (Fig. 7(a) & (b)). Aerobic denitrification was slightly inhibited when the media was spiked with 25 mg SCN⁻.L⁻¹ and 10 mg CN⁻.L⁻¹, and 25 mg SCN⁻.L⁻¹ and 15 mg CN⁻.L⁻¹, achieving denitrification rates of 0.6 and 0.54 mg NO₃ ⁻-N.L⁻¹.h⁻¹ (Fig. 7(c)). Aerobic denitrification has been previously reported to produce nitrous oxide as an intermediate prior to the formation of nitrogen gas as summarised in Eq. (4). The enzymology involved in this process was explained elsewhere [28]. However, the production of nitrous oxide from aerobic denitrification systems has been largely investigated in axenic cultures [29, 30], which is not representative of microbial populations. The description of pathways and the enzymology associated with the production of nitrous oxide from aerobic denitrification systems in microbial populations has been reported to be challenging, since individual organisms within the population utilise different bio-chemical pathways for the utilisation of nitrates [31].

CDO’s were co-cultured with the TDO’s and evaluated for simultaneous nitrification and aerobic denitrification in the presence of 25 mg SCN⁻.L⁻¹ and 15 mg CN⁻.L⁻¹. The co-culture was able to simultaneously nitrify and aerobically denitrify in the presence of SCN⁻ and CN⁻, achieving nitrification and denitrification efficiencies of 100% (1.83 mg NH₄ ⁺-N.L⁻¹) (Fig. 8(a)). There was an observed nitrate nitrogen increase to 104 mg NO₃ ⁻-N.L⁻¹ between 20 h to 100 h; thereafter, the nitrate nitrogen concentration decreased.

The co-culture was further evaluated for the simultaneous bio-degradation of CN⁻ (200 mg CN⁻.L⁻¹) and SCN⁻ (200 mg SCN⁻.L⁻¹), resulting in complete degradation of SCN⁻ after 124 h while the BRE of CN⁻ exceeded 99% (Fig. 8(b)). The ammonium nitrogen accumulated in the cultures was 98.7 mg NH₄ ⁺-N.L⁻¹ after 78 h which decreased to 33.6 mg NH₄ ⁺-N.L⁻¹ at 100 h. However, there was an observed ammonium accumulation within the cultures after 100 h, escalating to 167 mg NH₄ ⁺-N.L⁻¹ at 146 h. This increase was due to ammonium production from the bio-degradation of CN⁻ and SCN⁻ as described in Eq. (1) and (2). Simultaneous nitrification and aerobic denitrification has been reported in various studies [6, 32]. However, these studies were undertaken under SCN⁻ and CN⁻ free systems. Nitrification under CN⁻ laden conditions was investigated using activated sludge obtained from a coking wastewater treatment plant and in this study, nitrification was inhibited when the media was supplemented with 0.2 mg CN⁻.L⁻¹ [9], with SCN⁻ having a minimal inhibitory effect on nitrification. While in a separate study, the Enterobacter, Yersinia, and Serratia species were able to carry out nitrification under cyanogenic conditions (66 mg CN⁻.L⁻¹), with minimal inhibition [33]. The high CN⁻ tolerance was attributed to the differences in composition, strength and overall quality of the microbial consortia used [34]. This study demonstrated the robustness of heterotrophic nitrification and aerobic denitrification under CN⁻ and SCN⁻ laden conditions, when subjected to a mixed culture of TDO’s and CDO’s.