Insights into the Physicochemical Parameters, Microbial Community Structure, and Functional Variations in Biodegradation of N-Alkane Derivatives from Fischer–Tropsch Wastewater

Abstract

This study provides a theoretical baseline on the application of chemical and microbiological indicators as rapid system performance monitoring tools that will allow for timely corrective measures to maintain and improve the bioremediation performance of the Fischer–Tropsch wastewater (FTWW) treatment plants. Microorganisms isolated from the sediments and water samples collected from site 1 of Blesbokspruit wetland exhibited the highest biodegradation efficiency of up to 98.04% and 92.85%, respectively, in 96 h reaction time using batch culture media spiked with 300 ppm short chain n-alkane derivatives. The highest COD reduction rate was observed during the first 24 h of biodegradation, and it steadily declined thereafter. The decline in pH from 7.0 to 6.3 was observed in the 96 h reaction time and was attributed to the production of acidic secondary metabolites and the entrapment of the produced CO₂ within the batch media. The ORP also declined from the aerobic zone to the anaerobic zone within 24 h (day 1) reaction time. The EC and TDS results were also indicative of the rate of consumption of essential nutrients during the biodegradation process, which could be related to biochemical reactions involved in biodegradation of n-alkane derivatives. Proteobacteria and Firmicutes were the prevalent phyla during the biodegradation of the n-alkane derivatives. Enterococcus and Escherichia genera were more dominant on most days of biodegradation, therefore, indicating that these genera were actively involved in the biodegradation process of the n-alkane derivatives. These genera displayed a positive correlation with EC, ORP, pH and TDS in the four days of biodegradation for batch cultures inoculated with microorganisms from the water and sediments samples collected from the Blesbokspruit wetland. The results obtained demonstrated that physicochemical and microbiological indices can be used to infer the biodegradation rates, patterns and system operations in FTWW bioremediation.

1. Introduction

The rapid and continuous depletion of natural energy sources has become a global catastrophe within the energy sector. As a result, humankind has been compelled to derive alternative methods of renewable energy production to sustain their economy and basic livelihood [1,2]. As a result, technology-based processes such as Fischer–Tropsch have been designed and used to produce liquid fuels from hydrogen (H₂) and carbon monoxide (CO) gases [3,4,5]. Recently, this catalytic process has become the major source of transportation fuel and has contributed tremendously to the global reduction of fossil carbon dioxide (CO₂) emissions, especially in the transport sector [5]. However, the large amounts of liquid fuels produced by this gas-to-liquid process are accompanied by relatively higher volumes of toxic wastewater that is characterized by very high COD and TOC contents and low pH [6,7].

Fischer–Tropsch wastewater (FTWW) consists mainly of alcohols, ketones, and organic acids, which results in a high organic content. For instance, for every ton of synthetic liquid fuel produced, about 1.3 tons of FTWW is generated [8,9]. This significant wastewater generation poses a serious environmental threat to both terrestrial and aquatic ecosystems and has, therefore, become a major concern to environmental management and monitoring agencies worldwide. Physicochemical treatment processes such as distillation have been used to recover and reuse FTWW within the processing plant, but the high energy requirement of this treatment method negates its economic feasibility [10]. Moreover, the wastewater treated using this method often has COD levels that are still at unacceptable levels and may require further post-treatment. Biological methods have proven to be less costly, more environmentally friendly and effective approaches in the treatment of FTWW to achieve allowable COD levels [11].

Biological treatment technologies employed in the decontamination of FTWW and petroleum wastewater involve the use of hydrocarbon-degrading microorganisms to break down the hydrocarbons into carbon dioxide, water, and innocuous metabolites that have less negative impact on both the terrestrial and aquatic ecosystems [12]. The interspecific interactions of different strains in microbial consortia have resulted in enhanced biodegradation to complete the mineralization of petroleum hydrocarbons [13]. FTWW comprises recalcitrant complex nonoxygenated and oxygenated hydrocarbons that require various selective genes and enzymes from different microorganisms to break down these hydrocarbons in a stepwise approach [14]. Hydrocarbons are biodegraded by both aerobic and anaerobic microorganisms through different pathways. Microorganisms use hydrocarbons as their source of carbon. The biodegradability of FTWW depends on the metabolic and catalytic functions of the microbial communities, chemical complexity, bioavailability, bioaccessibility, and concentration of the hydrocarbons, and the prevailing environmental factors such as pH, salinity, temperature, dissolved oxygen, oxidation–reduction potential (ORP), etc. [15,16]. Based on previous literature, the susceptibility of petroleum wastewater hydrocarbons towards microbial degradation has been ranked as n-alkanes > low molecular weight aromatic hydrocarbons > polyaromatic hydrocarbons in decreasing degrees of susceptibility. Straight-chain alkanes are also more biodegradable than branched alkanes. Previous studies have also shown that the longer the hydrocarbon chains, the lower the biodegradability; the degree of biodegradation susceptibility decreases with the number of carbon chains in the hydrocarbon [14].

Aerobic degradation of n-alkanes, which is considered a relatively faster and less complex degradation mechanism, involves the use of molecular oxygen (O₂) for the activation of the n-alkane through either the terminal or the subterminal oxidation. In both pathways, the alkane-degrading microorganisms use monooxygenase enzymes to overcome the low chemical reactivity of the n-alkanes through the generation of reactive oxygen species. The terminal activation pathway starts with the oxidation of the terminal methyl functional group of the alkane to form a primary alcohol using alkane-hydroxylase. The produced primary alcohol is subsequently converted to a primary aldehyde (using alcohol dehydrogenase), which is further oxidized to form the corresponding fatty acids (using aldehyde dehydrogenase). The generated fatty acids then undergo β-oxidation to produce Acetyl-CoA, which is further metabolized in the TCA (tricarboxylic acid) cycle to generate energy, carbon dioxide, and water [17]. In some cases, both primary methyl functional groups on both ends of the alkane chain are oxidized through ώ-hydroxylation to form ώ-hyroxy fatty acids, which are subsequently converted to dicarboxylic acids. The dicarboxylic acids further undergo β-oxidation and complete mineralization. Some microbial species prefer to attack the subterminal functional group of n-alkanes to form a secondary alcohol which is further oxidized to a ketone and then an ester. The generated ester is then hydrolyzed into acetic acid and primary alcohol that further undergoe terminal oxidation to also produce energy, carbon dioxide, and water via the TCA cycle [15,18,19].

Anaerobic biodegradation of n-alkanes also follows two pathway routes where the fumarate molecule either attaches to the subterminal or terminal position to yield an alkyl-succinate derivative. The organic intermediate is subsequently bonded to CoA to form an Acyl-CoA that further undergoes β-oxidation. This is the common biodegradation pathway by which most hydrocarbons are broken down under anaerobic conditions. Microorganisms facilitate the anaerobic biodegradation of alkanes in the presence of nitrates, sulfates, manganese, iron, and organic intermediates as electron acceptors. The fumarate has also been reported to bind on the terminal position of the alkane before the formation of acetyl-CoA and its subsequent β-oxidation [20,21]. This biodegradation pathway has only been reported in the anaerobic biodegradation of propane [17]. The aerobic and anaerobic biodegradation of alkane derivatives such as short-chain alcohols, fatty acids, aldehydes, and ketones can be deduced from the biodegradation pathways of alkanes mainly because these derivatives are formed as intermediates in the biodegradation of the parent alkanes. Anaerobic biodegradation of hydrocarbons is often slower than aerobic biodegradation because of less favorable energetics and the requirement of alternate electron acceptors.

FTWW comprises of numerous hydrocarbons, with alkanes and their derivatives being the most dominant species in this carbon-rich wastewater. Although possible biodegradation pathways of alkanes and their derivatives have been well studied and deduced, monitoring of their microbial degradation performance in FTWW treatment plants remains a very scarce practice. Based on the much clearer and well-understood biodegradation pathways of n-alkanes and their derivatives (i.e., short-chain alcohols and fatty acids), performance-indicating parameters can be used to monitor and assess their degradation efficiency in the FTWW treatment process [22]. Both biotic and abiotic system parameters have been used to monitor the biodegradation performance of most biological treatment systems and have evolved into the use of online sensors with very high accuracy and reliability. As a result, the use of biosensors and physicochemical parameter meters has gained more popularity as reliable and rapid performance assessment tools in biological treatment processes [23,24]. For instance, physicochemical parameters such as pH, COD, DO, and ORP can be used as biodegradation process indicators to monitor and assess the biodegradation mechanics and efficiency in biological treatment processes [25].

These chemical parameters change as organic pollutants are biodegraded into intermediate and dead-end metabolites and possibly mineralized into carbon and water. Wang et al. [26] (2022) established a direct correlation between ORP and other process parameters such as COD, DO, and nitrate content in an anoxic denitrification process. The results obtained indicated that under low nitrate concentrations of about 3 mg/L, there was a direct relationship between ORP and COD, and this relationship could be used in process surveillance and control of the carbon dosage in anoxic denitrification processes. Howard et al. [24] also investigated the use of physicochemical parameters such as BOD, COD, conductivity, suspended solids and total solids, nitrates and phosphates, total nitrogen and phosphorus, among others, as performance indicators in a municipal wastewater treatment plant in Granada, Spain. The results obtained confirmed that BOD, COD, suspended and total solids, and fats were statistically accurate and reliable indicators for process monitoring and control of the treatment process. Although some direct correlations between these physicochemical indicators have been established, their long-term use in biological treatment processes is limited by the frequent sensor drifting and breakage, which results in high maintenance costs. The physicochemical indicators are often used in conjunction with biological indicators to portray a more comprehensive analysis of the operational treatment functions in biological treatment systems.

Recently, advanced microbial profiling techniques such as high throughput metagenomic sequencing have gained popularity as a more accurate and reliable approach for microbial analysis of both culturable and unculturable microorganisms. The system chemistry, microbial biodiversity, and metabolic functions change at each stage of the treatment process as pollutants are transformed and/or broken down into less toxic and simpler fragments. Therefore, tracing these changes paves the way for an integrated approach that projects the true reflection of the biochemical processes involved in biological wastewater treatment. To the best of the authors’ knowledge, monitoring and evaluation of these changes to elucidate the biochemical processes and biodegradation patterns involved in the degradation of n-alkane derivatives from FTWW remains untapped. In light of this, this study seeks to investigate the applicability of both physicochemical and microbial performance indicators to monitor and assess the biodegradation patterns and dynamics involved in microbial degradation of short-chain hydrocarbons in FTWW. This was achieved by first identifying microbial consortia that possess high biodegradation capabilities for n-propanol, butyric acid, and valeric acid, which are dominant short-chain hydrocarbons found in FTWW. These microbial consortia were screened by monitoring the COD reduction capacities of microorganisms isolated from six sampling sites with varying pollution gradients. Secondly, the identified ultrafunctional microbial consortia were used in the biodegradation of these short-chain hydrocarbons in batch cultures, and the biodegradation patterns and efficiencies were assessed by monitoring the physicochemical and microbial diversity and functional profile changes. This study, therefore, provides a theoretical baseline for the application of rapid performance indicator monitoring tools that could be used to assess and evaluate the operation performance of FTWW treatment plants and thus allow for highly responsive system control measures for effective FTWW bioremediation.

2. Materials and Methods

2.1. Biogeochemical Sites Description and Sample Collection

Water and sediment samples for microbial isolation were collected from the Blesbokspruit wetland located in southeast Johannesburg, South Africa. It is one of the largest wetlands covering the gold Witwatersrand basin, and it covers about 1858 ha and stretches 1855 m along this basin. Blesbokspruit wetland was once a Ramsar-accredited wetland of global significance but was later enlisted on a Montreux Record of Ramsar sites of degradation in 1996 owing to its intensified ecological degradation brought about by a myriad of contaminants that continuously enter this wetland. The water and sediment microbial samples were collected from six selected biogeochemical sites based on their proximity to various anthropogenic activities and land use practices, as described in our previous study [27]. The physicochemical and microbial structure and functional variations at each respective site along the Blesbokspruit wetland catchment were also reported in the previous study.

2.2. Enrichment of the Isolated Microbial Consortia

First, 10 mL of the water and 10 mg of sediments collected from the selected biogeographical sites were transferred into 200 mL of nutrient broth and incubated for three days at 28 °C with shaking at 120 rpm to allow for sufficient cell growth.

2.3. Preparation of the Synthetic Fischer–Tropsch-Derived Wastewater

Four synthetic culture media spiked with 1-propanol, butyric acid, valeric acid, and the mixture of these three n-alkane derivatives (PBV) were prepared in minimum salt media (MSM) at the resultant concentrations of 200, 300, and 600 ppm for each respective n-alkane derivative and their mixture (PBV). Minimum salt media (MSM) was prepared in 1 L Schott bottles by adding the following chemicals: 1 M sodium phosphate (Na₃PO₄), 1 M potassium phosphate (KH₂PO₄, monobasic), 1 M sodium chloride (NaCl), and 1 M ammonium chloride (NH₄Cl). The bottle was then filled up with distilled water and thoroughly mixed by shaking several times. The resultant solution was then autoclaved at 121 °C for 15 min, with subsequent addition of 1 M filtered MgSO₄ and CaCl₂. All chemical were purchased from Merk Life Science (Pty) Ltd., Johannesburg, South Africa.

2.4. Inoculation of Synthetic Solutions

The synthetic 1-propanol solution, butyric acid solution, valeric acid solution, and PBV solution were inoculated with the microorganisms isolated from the water and sediment samples collected from Sites 1–6 of the Blesbokspruit wetland. The synthetic solutions and controls setup were prepared as described in

Table 1. Composition of reaction solutions controls.

Reaction Solutions	MSM	Synthetic Solution (mL)	Inoculum (mL)
ISC	Present	-	-
SC	Present	-	5
IC	Present	500	-
Test	Present	500	5

Note: ISC: Inoculum and substrate control, SC: Substrate control, IC: inoculum control.

.

2.5. COD Reduction Analysis

The initial COD measurements in each reaction solution were taken before inoculation (in triplicates) using the Lovibond M160 photometer (Tintometer, Dortmund, Germany) and were subsequently measured every 24 h for four days. Equation (1) was used to calculate the COD reduction at a 24 h reaction interval.

$$\mathrm{C}\mathrm{O}\mathrm{D} \mathrm{r}\mathrm{e}\mathrm{d}\mathrm{u}\mathrm{c}\mathrm{t}\mathrm{i}\mathrm{o}\mathrm{n} (\%)=\frac{Ci-Cf}{Ci}\times 100$$

(1)

where Ci represents the initial COD measurement (mg/L) of the reaction solution before inoculation at time 0 h and Cf represents the final COD measurement (mg/L) of the reaction solution at each 24 h interval.

2.6. Physicochemical Parameter Changes

Physicochemical parameters measurements for pH, oxidation–reduction potential (ORP), electrical conductivity (EC), total dissolved solids (TDS), temperature, and pressure were measured at 24 h intervals for four days using a H19829 HANNA multiparameter meter (HANNA Instruments, Johannesburg, South Africa). All measurements were performed in triplicates and reported as mean ± standard deviation.

2.7. Microbial Community Structure Analysis

DNA was extracted from the collected water and sediment samples according to the manufacturer’s protocol using the Qiagen DNeasy PowerWater extraction kit and DNeasy Powersoil Pro extraction kit, respectively (Qiagen, Hilden, Germany). The extracted genomic DNA was quantified using the Qubit 2.0 fluorometer (Thermofisher, Waltham, MA, USA) and stored in cryovials at −80 °C until it was sent to Inqaba Biotechnical Industries Pty (Ltd.) in Pretoria, South Africa, for Illumina shotgun metagenomic analysis.

2.8. Bioinformatics and Statistical Analysis

Demultiplexed paired-end reads obtained from the Inqaba Biotech sequencing facility were quality-checked using FastQC software version 0.11.5. Next, Trimmomatic software (version 0.38) [28] was used to quality-trim paired reads, including clipping off any Illumina barcodes and eliminating reads with an average quality score (Phred Q score) lower than 20. Quality-filtered paired reads were then analyzed in the Quantitative Insights into Microbial Ecology (version 2) (QIIME2) software [29]. DADA2 denoiser version 1.14 was used to merge pair-end sequences into full-length sequences and remove chimeras [30]. USEARCH version 7 was used to cluster similar sequences into operational taxonomic units (OTUs) at 97% similarity [31]. Taxonomic classification of the clustered OTUs was performed against the RDP classifier [32]. The obtained OTU table was used for further analysis involving visualization of the data using RStudio version 2023.03.0-daily+82.pro2, STAMP version 2.8, and Microbiome tools. The OTU table in BIOM format and representative sequences obtained from QIIME2 were used as input to the Nephele platform from the National Institute of Allergy and Infectious Diseases (NIAID) Office of Cyber Infrastructure and Computational Biology (OCICB) in Bethesda, MD, to predict metabolic functions using PICRUSt2. The PICRUSt2 output was also visualized using STAMP. SPSS 22.0 software was used to calculate the mean and standard deviation of all measurements, which were recorded in triplicates. Kendall correlation plot was used to establish the correlations between the significant environmental variables (i.e., EC, ORP, pH, and TDS) and the microbial communities at the genus level.

2.9. Sequence Accessions

The original sequencing data obtained in this study were submitted to the National Center for Biotechnology Information (NCBI) Sequence Read Archive (SRA) database with accession number: PRJNA1050046.

3. Results and Discussion

3.1. Biodegradation Studies

The COD reduction rates of the microorganisms isolated from the water and sediments collected from the six selected biogeochemical sites were compared to identify the ultrafunctional microorganisms with the highest biodegradation efficiency for the aliphatic hydrocarbons in Fischer–Tropsch wastewater. The COD reduction rates of the collected microorganisms were tested at 200, 300, and 600 ppm initial concentrations of the Fischer–Tropsch short-chain hydrocarbons components. As shown in Figures S1–S6, the COD reduction rate was faster within the first 24 h reaction time when using the different microbial consortia isolated from the water and sediments collected from all the selected biogeochemical sites. This trend was observed within the three initial concentrations, i.e., 200, 300, and 600 ppm, in all batch culture reaction setups except for the controls. Propanol biodegradation reached the highest COD reduction for most batch culture reactions, followed by butyric acid, valeric acid, and the mixture of the three hydrocarbon components (PBV) (Figure 1). Microorganisms collected from the sediments from all six selected sites exhibited relatively higher COD reduction rates than those collected from the water samples for the three initial concentrations (200, 300, and 600 ppm). The effect of hydrocarbons’ initial concentrations on biodegradation efficiency was monitored across the three initial concentrations (200, 300, and 600 ppm), with the highest biodegradation efficiency reported at 300 ppm initial concentration, followed by 200 ppm initial concentration and lastly at 600 ppm initial concentrations for propanol, butyric acid, valeric acid and their mixture (PBV). This trend was observed for microorganisms isolated from both the water and sediment samples from all six biogeochemical sites. The highest COD reduction rates of 98.04%, 96.34%, 94.34%, and 90.34% for propanol, butyric acid, valeric acid, and their mixture (PBV), respectively, was recorded for microorganisms collected from the sediments samples at site 1 at the initial concentration of 300 ppm for all the n-alkane derivates. The second highest COD reduction rates were recorded at an initial concentration of 300 ppm for all the batch cultures inoculated with microorganisms isolated from the surface water collected from site 1 and were 92.85%, 84.85%, 98.04%, and 90.34% for propanol, butyric acid, valeric acid and their mixture (PBV), respectively. Microorganisms collected from the water and sediment samples from site 6 also showed relatively high biodegradation efficiencies for propanol, butyric acid, valeric acid, and their mixture (PBV) at 300 ppm initial concentrations. The results obtained revealed that the biodegradation rates of both the water and sediment microorganisms decreased at a low initial concentration of 200 ppm and a higher initial concentration of 600 ppm. At low hydrocarbon concentrations, there exists a minimal contact time between the hydrocarbons and the microbial hydrocarbon degraders, thus limiting the biochemical interactions required for the biodegradation process. However, very high hydrocarbon concentrations may result in the reduced biodiversity of the active hydrocarbon degraders and induce inhibitory effects that suppress the biodegradation performance of the microorganisms due to the inherent toxicity of the hydrocarbons. Therefore, sufficient amounts of the n-alkane substrates are required for cell proliferation and metabolic activity of the microbial hydrocarbon degraders while allowing for enough contact time for the biochemical interactions responsible for the biodegradation process to occur. The optimal initial concentration of the n-alkane derivatives was therefore found to be 300 ppm. As demonstrated by Bacosa et al. (2021) [33], the initial concentration of n-alkanes and polycyclic hydrocarbons results in a shift in the microbial community structures of the oil-degrading microorganisms, which in turn affects the biodegradation capabilities of these microbial communities at the varying concentration gradient. Different hydrocarbon-degrading microorganisms have a varying degree of tolerance towards varying initial concentrations of specific hydrocarbons. Therefore, very high or low initial concentrations may enhance or suppress their relative abundances and activity in biodegradation processes or environmental habitats with varying hydrocarbon pollution gradients. Liu et al. (2018) [34] also confirmed that varying n-alkane contamination levels influenced the presence and abundance of dominant alkane degraders and degrading genes in contaminated soil microbial communities, which directly affected the biodegradation performance at 1%, 3%, and 5% n-alkane contamination levels. The results obtained showed a rapid COD reduction in the first 24 h of reaction time, which declined afterward. The fast biodegradation rates observed in the first 24 h could be attributed to the high susceptibility of the short-chain n-alkanes to microbial biodegradation due to their less complex chemical structures and properties. Microorganisms isolated from the water and sediments from site 1 at an initial n-alkanes concentration of 300 ppm displayed the highest biodegradation potential and were, therefore, further used for downstream studies to monitor the physicochemical parameters, microbial diversity variations, and metabolic functional changes to assess the biodegradation rates and patterns involved in the breaking down of the aliphatic n-alkane derivatives in Fischer–Tropsch derived wastewater.

3.2. Physicochemical Parameter Changes

3.2.1. pH Variations

The pH changes during the biodegradation of the short-chain hydrocarbons were monitored at 24 h intervals for propanol, butyric acid, valeric acid, and PBV batch cultures over a period of 96 h of reaction time (Figure 2). The pH changed from 7.0 to 6.3 over the 96 h reaction time, indicating a slight shift from neutral to slightly acidic media. However, there were no significant changes in the inoculum and substrate control (ISC), substrate control (SC), and inoculum control (IC). In aerobic biodegradation of short-chain alkane derivatives, the alkane derivates are oxidized into fatty acid metabolites that could contribute to the decline in pH and create an acidic microenvironment [35]. As observed in Figure 2, the pH of the reaction media during propanol biodegradation showed a relatively steep gradient, from neutral (7.1) to acidic (6.3), which could be attributed to the oxidation of the primary alcohol to form slightly acidic aldehydes, fatty acid, acetyl CoA, propionyl CoA metabolites. The fatty acids undergo mineralization through β-oxidation to form acetyl CoA or propionyl CoA, which are further oxidized in the TCA cycle (tricarboxylic acid cycle) to form carbon dioxide and water. It is highly possible that the carbon dioxide produced could be trapped within the reaction media due to the limited agitation in the batch culture media, thus rendering it acidic. Microbial growth and metabolic activity of hydrocarbon-degrading microorganisms are pH dependent. Most hydrocarbon-degrading bacteria achieve their optimal biodegradation performance at neutral pH (ranging from 6.5 to 7.5) [36,37]. Liu et al. (2022) [38] also observed a similar trend where the pH of the inoculated media declined from 7.0 to 4.2 in 7 days due to a large number of acidic metabolites produced during the degradation of hexadecane by Gordonia sihwaniensis species. The presence of acid emulsifiers may also contribute to the decrease in pH during the biodegradation of petroleum hydrocarbons such as alkanes and polyaromatic hydrocarbons. A shift in pH during biodegradation may also influence the nutrients and hydrocarbon availability as well as the adsorption–desorption kinetics crucial for enhanced biodegradation [39].

3.2.2. Oxidation–Reduction Potential (ORP) Variations

Oxidation–reduction potential is a chemical parameter that indicates the oxidizing or reducing ability of water based on the availability of electron-donating and accepting chemical species in solution [40]. The presence of these chemical species, especially in wastewater effluents, determines the biodegradation pathways and rates that rely on electron transfer processes (i.e., oxidation–reduction reactions) involved in the bioremediation processes. The interspecific interactions of the degrading microbial assemblages with these reducing and oxidizing chemical species facilitate the breakdown and/or transformation of organic and inorganic pollutants [41]. As in a biological wastewater process, mineral batch media also contain minerals that possess oxidizing and reducing capabilities to promote microbial oxidation–reduction reactions responsible for the breakdown of hydrocarbons. The results obtained in this study indicated that the ORP decreased drastically in the first 24 h of the biodegradation process at which it started to rise gradually thereafter (Figure 3). The largest ORP reduction gradient was recorded in 1-propanol biodegradation from 156.4 to −184.4 mV when using microorganisms isolated from sediments (from site 1). Moreover, 1-propanol biodegradation also recorded the largest ORP reduction of 156.4 to −177.2 mV for microorganisms isolated from the water (from site 1). A more positive ORP indicates a more oxidizing environment, while a more negative ORP indicates a more reducing environment. Therefore, for biodegradation of 1-propanol, butyric acid, valeric acid, and their mixture (PBV), the reaction media was more oxidizing initially and became more reducing within the 24 h reaction time. The positive ORP at the initial stage of the biodegradation process could be attributed to the high levels of dissolved oxygen (O₂) as an oxidizing agent, which got depleted rapidly in the first 24 h of the biodegradation due to the aerobic degradation of the short chain hydrocarbons to produce energy for cell growth (i.e., maintenance energy) and some intermediate metabolites. These hydrocarbons (i.e., the electron donors) were rapidly broken down through the oxygen-mediated respiration process, which led to a decrease in the ORP of the batch cultures at the initial stage of the biodegradation process. Moreover, the agitation in batch culture media does not allow sufficient atmospheric oxygen to be dissolved in the media to replace the oxygen consumed in the biodegradation process, thus resulting in an oxygen deficit in the reaction media. This oxygen deficit and the presence of the reducing agents such as ammonium and sulfate ions in the minimum salt media culture could, therefore, be attributed to the steep gradient in ORP reduction in the first 24 h, which led to the reduced biodegradation rates afterward [42]. Previous studies have shown that ORP measurement is more accurate and sensitive than the direct measurement of dissolved oxygen (DO) to monitor the oxidizing and/or reducing the strength of wastewater effluents and, thus, can be used to determine the onset of the aerobic to anaerobic shifts and vice versa in biological processes [26]. The ORP slowly increased towards the positive side after 24 h, and this could be explained by the fact that the log phase, which is associated with high oxygen consumption, had ended, and the remaining oxygen-demanding process was the catabolic breakdown of the short chain hydrocarbons, which required relatively lesser oxygen consumption. As a result, the oxygen deficit gradually decreased as the dissolved oxygen was slowly utilized, while the dissolution of the atmospheric oxygen into the batch culture media was steadily maintained. The results obtained, therefore, suggest that external aeration should be included in such batch cultures to maintain sufficient dissolved oxygen required for oxidizing conditions in the reaction media to promote or maintain biodegradation efficiency.

3.2.3. Electrical Conductivity Variations

The variations of EC values during the biodegradation of 1-propanol, butyric acid, valeric acid, and their mixture (PBV) are shown in Figure 4a,b. The EC values of the four chemical streams decreased dramatically in the first 24 h of the process reactions, with the EC value for 1-propanol dropping from 11.72 to 10.72 mS/cm, butyric acid from 11.50 to 10.52 mS/cm, valeric acid from 11.48 to 10.45 mS/cm, and PBV mixture from 11.35 to 10.24 mS/cm for biodegradation reactions inoculated with microorganisms from the water samples. The EC values increased drastically after 24 h for the remaining 72 h of reaction time. A similar trend was observed for the biodegradation reactions inoculated with microorganisms isolated from the sediment samples. However, the EC variations in the batch cultures inoculated with microorganisms isolated from sediment samples were greater than those inoculated with microorganisms isolated from water samples. Electrical conductivity (EC) of water is the measure of the amount of dissolved minerals and other chemicals soluble in water. The initial EC levels reported in this study (up to 11.72 mS/cm) are relatively higher than the WHO acceptable EC limit of water, which is 400 µS/cm. The high initial EC value can be attributed to the dissolved salts used to prepare the MSM in the batch cultures [43]. The sharp EC decline in the first 24 h of the reaction process could be ascribed to the accelerated consumption of salt minerals such as sodium phosphate (Na₃PO₄), potassium phosphate (KH₂PO₄), sodium chloride (NaCl), ammonium chloride (NH₄Cl), magnesium sulfate (MgSO₄) and (CaCl₂) in the MSM which are used up by the microorganisms for their cell growth and metabolic activities responsible for the breaking down of the short chain hydrocarbons. Although hydrocarbons have very small electrical conductivity, it is possible that their derivatives that are formed as metabolites during biodegradation may have higher electrical conductivity, which resulted in the EC spikes after the 24 h reaction time. In addition to this, the cell lysis of the microorganisms, which are less tolerant to the produced metabolites, may also result in the dissolution of the mineral salts back into the culture media, thus increasing the EC of the reaction media. EC measures the ability of the media to conduct electricity in aqueous solution, which influences electron transfer and can, therefore, be used to monitor and control the electrochemical processes involved in the biodegradation of hydrocarbons. This has led to the use of electroconductive carbonaceous biofilters to enhance the microbial activity of electroactive microorganisms, thus resulting in improved organic biodegradation. Wang et al. (2023) [44] reported on the use of electroconductive coke and quartz fillers to enhance microbial degradation of sewage wastewater in a microbial electrochemical-based constructed wetland. The incorporation of these fillers resulted in enhanced removal rates; quartz fillers exhibited higher electrical biostimulation than coke fillers, which resulted in increased COD reduction in the sewage wastewater. Wang et al. (2017) [45] also investigated the use of biochemical systems to enhance FTWW biodegradation in an upflow anaerobic sludge blanket bioreactor. Apart from the induced electrical field, the biochemical system resulted in an increment in sludge conductivity, which could have also improved the direct cell-to-cell electron transfer between the Geobacter and Methanosarcina species, thus promoting the syntropic biodegradation of propionate and butyrate which resulted in an enhanced COD removal in the FTWW. Li et al. (2016) also discovered that due to high osmotic pressure brought about by high salinity, the high ion concentrations in saline environments limit the microbial activity of hydrocarbon-degrading microorganisms, thus suppressing their biodegradation abilities in bioelectrochemical biodegradation. However, low electrical conductivity, especially in soil, also tends to inhibit the bioelectrochemical biodegradation of hydrocarbons in contaminated soil [46].

3.2.4. Total Dissolved Solids (TDS) Variations

Total dissolved solids (TDS) is a parameter used to measure the amount of dissolved substances such as cations, anions and dissolved organics in water. Unlike EC, TDS measures both the conductible and nonconductible dissolved chemicals in water. Figure 5a,b shows the TDS variation patterns of the batch culture reactions during the 96 h biodegradation process of the hydrocarbon derivates found in FTWW. The TDS decreased dramatically in the first 24 h of the batch reactions, with the highest TDS reduction from 5720 to 5134 ppm recorded for biodegradation of PBV mixture inoculated with microorganisms isolated from the water samples. The second highest TDS reduction from 5890 to 5460 ppm was recorded for biodegradation of 1-propanol inoculated with microorganisms isolated from sediment samples. The TDS values increased after the 24 h reaction time, and 1-propanol displayed the largest increase in the remaining 72 h of reaction time. The trend observed in the variations of TDS (Figure 5) is similar to the one shown in the EC variation plots in Figure 4. This is mainly because TDS and EC are directly related and form a linear relationship in the case that there are very low concentrations of dissolved organics, and the water is comprised mainly of ionic species [47]. As explained in the discussion for EC variations, the sharp decrease in TDS during the first 24 h could also be attributed to the rapid consumption of ionic minerals and dissolved hydrocarbons in the MSM culture during cell proliferation of the microbial inoculum and the increase in TDS in the remaining 72 h could be ascribed to the produced metabolites during the biodegradation process. The TDS increase in the remaining 72 h of reaction time can also be attributed to the soluble biomolecules such as cationic, anionic, and nonionic surfactants that are secreted by the microorganisms during their interactions with the hydrocarbons in the culture media. Among other factors, the limitation of nutrients in biological wastewater treatment processes inhibits the microbial growth and metabolic activities responsible for the biodegradation of organic pollutants such as hydrocarbons and thus compromises the biodegradation efficiency of the wastewater treatment system. TDS and EC can be used to monitor the nutrient consumption in these biological wastewater treatment systems, which could assist in determining the appropriate nutrient dosage required to maintain and/or improve the biodegradation efficiency of these treatment systems. Sinlapacheewa et al. (2022) [48] investigated the effect of TDS and organic dosages on COD reduction in a yeast-activated sludge system. The results obtained indicated that a 50% TDS loading rate resulted in maximum biomass production and, therefore, enhanced COD removal efficiency. It was also observed that for every increase in organic load, there should be a proportional TDS increase to maintain the COD removal efficiency of the yeast-activated sludge system. However, a TDS loading rate higher than 50% was found to inhibit biomass production. Hanson et al. (2019) [49] also discovered that high TDS levels inhibited the biodegradation of alkyl and nonylphenol ethoxylate surfactants in shale gas wastewater. This implies that there is a TDS dosing rate threshold value beyond which the COD removal efficiency of the biological wastewater treatment system becomes compromised.

3.3. Microbial Variations

3.3.1. Alpha and Beta Biodiversity

Figure 6 displays the alpha biodiversity variations (within a single sample) and the beta biodiversity variations (among the distinct geoclusters) of the microbial communities involved in the biodegradation of n-alkane derivatives from FTWW. The alpha biodiversity variations of the microbial communities were represented by observed richness, Chao1, and Shannon diversity estimator indices shown in Figure 6. The observed alpha diversity index revealed a dramatic increase in species richness from day 1 to day 2 for microorganisms isolated from the sediment samples (22.2 to 27.2) and water samples (16.4 to 32.6). A significant decline in species richness, as indicated by the observed index, was observable for microorganisms isolated from sediment samples (from 27.2 to 24.2) and those isolated from water samples (from 32.6 to 22.6) from day 2 to day 4. The chao1 diversity index also indicated similar species abundance trends observed in the observed/richness diversity index for both the water and sediment samples. The Shannon alpha diversity index also showed an increase in species richness and uniformity in the microorganisms isolated from sediment samples (from 21.2 to 24.4) and water samples (13.3 to 29.1) from day 1 to day 2 reaction time. A significant decline in species richness and evenness was observed in microorganisms isolated from the sediment samples (24.2 to 19.6) and water samples (from 29.1 to 24.6) from day 2 to day 4. The increase in species richness of the microorganisms isolated from the sediment and water samples from day 1 to day 2 could be attributed to the presence of essential nutrients required for the cell proliferation and metabolism of the microorganisms and possibly the minimum inhibition caused by trace amounts of the produced secondary metabolites. From day 2 to day 4, the essential nutrients could have been depleted during the microbial biodegradation activity and the inhibition effect caused by the increasing number and concentrations of the secondary metabolites that depress the reproduction rate of some of the microorganisms [50]. The principal component analysis (PCoA) plots indicated that there were significant differences in beta diversity within the microorganisms isolated from the water and sediments within four days of biodegradation. The prediction ellipses observed showed that both the microorganisms isolated from the water and sediments were clustered into different ellipses within the four days of biodegradation with the PCoA1 (x-axis variances) of 37.3% and PCoA2 (y-axis) of 14.9%.

3.3.2. Relative Abundance at Phyla Level

Classified sequence reads at the phyla level revealed the four dominant phyla as Proteobacteria, Firmicutes, Bacteroidetes, and Actinobacteria, which have all been reported to possess biodegradation capabilities of hydrocarbons [51]. For both water and sediment microorganisms, Proteobacteria displayed a relatively high abundance of up to 81.04% (in propanol batch culture) for water microorganisms and 78.44% (in butyric acid batch culture) for sediment microorganisms after the first day of biodegradation (day 1). Proteobacteria have been reported as prominent hydrocarbon-utilizing bacteria in previous studies [52,53,54]. Proteobacteria species contain the alkB-encoded alkane hydroxylase enzyme responsible for the biodegradation of alkanes and their derivatives [55]. Bacteroidetes also exhibited a higher relative abundance of up to 23.43% (in valeric acid batch culture) for water microorganisms and 24.56% (in mixed batch culture, PBV) for sediments microorganisms on the first day of biodegradation (day 1). Actinobacteria showed a relative abundance of up to 14.04% for water microorganisms (in the butyric acid batch culture) and 14.02% (in the mixed batch culture) for sediment microorganisms on day 1. Firmicutes were more abundant (15.55% relative abundance) in the mixed batch culture inoculated with water microorganisms and also in the butyric acid batch culture inoculated with sediments microorganisms exhibiting 12.22% relative abundance on day 1. On day 2 of biodegradation, there was an apparent decrease in the abundance of Bacteroidetes and Actinobacteria up to the point of extinction in all batch culture media for both the water and sediments microorganisms. This was associated with a successional decrease in the abundance of Proteobacteria and an increase in the abundance of Firmicutes in the remaining 3 days of biodegradation in both the water and sediment microorganisms. Actinobacteria was revived again on day 4 in the valeric acid batch culture for water microorganisms and in butyric and the mixed batch culture (PBV) for sediment microorganisms. It is evident from Figure 7 that the dominance of Proteobacteria and Firmicutes throughout the biodegradation process indicates that these two phyla played a crucial role in the biodegradation of the n-alkane derivates in the batch cultures, possibly through symbiotic interactions.

3.3.3. Relative Abundance at the Genus Level

On day 1 of the biodegradation process, the valeric acid batch that was inoculated with sediment microorganisms was enriched with Actinomyces with relative abundance of up to 65.42% (Figure 8). Stenotrophomonas was also detected at a significantly high abundance of up to 20.66% in butyric acid batch culture inoculated with water microorganisms. Dysgonomonas, Enterococcus, and Escherichia growth was also observed on the first day of the biodegradation process at the maximum relative abundance of 22.3% (in valeric acid batch culture inoculated with water microorganisms) and 7.82% (in the mixed batch culture inoculated with water microorganisms), 3.83% (in valeric acid batch culture inoculated with water microorganisms), respectively. Enterococcus and Escherichia genera were prevalent in the remaining days of biodegradation, therefore, indicating that these genera were actively involved in the biodegradation process of the n-alkane derivatives. Various strains of Enterococcus genera, such as Enterococcus mundtii SS 1 [56], Enterococcus casseliflavus [57], Enterococcus faecalis [58], and Escherichia species, such as Escherichia coli [59], Escherichia fergusonii [60], have been reported to be potent hydrocarbonoclastic microorganisms. Actinomyces abundance was significantly reduced in the remaining days of the biodegradation process, while Stenotrophomonas and Dysgonomonas growth was inhibited up to the point of extinction after day 1 of the biodegradation process for all the batch reactions inoculated with the water and sediments microorganisms. The evident shifts in microbial composition and biodiversity during the biodegradation of the n-alkane derivatives are brought about by the susceptibility of these hydrocarbons to the different strains found in the microbial consortia of the water and the sediments. These shifts can also be attributed to the produced metabolites that can be easily degraded by less tolerant microorganisms that utilize the produced metabolites for energy production and growth. During the biodegradation process, there are various shifts in the chemical properties of the batch culture media, brought about by the produced metabolites and the electron transfer mediated microbial activities that result in the changes in the chemical and physicochemical characteristics of the reaction culture media. These changes in the chemical and physicochemical properties of the culture media are accompanied by changes in the population dynamics of the biodegrading consortia. Mori et al. (2020) [61] demonstrated how microbial interactions influence the development of specialized niches that support and maintain microbial structural changes pertinent to the systematic breakdown of petroleum hydrocarbon pollutants. More hydrocarbon-tolerant microbial species were capable of biodegrading more complex and less soluble parent hydrocarbons into simpler metabolites that could easily be utilized by other less tolerant microorganisms through metabolic dependencies and creation of specialized niches suitable for the catalytic activities of the coexisting genera within the soil bacterial consortium.

3.3.4. Correlation Analysis of the Physicochemical Parameters and Microbial Communities in the Batch Cultures for Four Days Biodegradation Process

Figure 9 shows the Kendall correlation plots of the physicochemical parameters and the microbial communities in the batch cultures for the four days of the biodegradation process. Enterococcus genera, which was found to be dominant during the four-day biodegradation process in all the batch culture media, displayed a positive correlation with the EC on the fourth day of biodegradation for microorganisms isolated from microorganisms isolated from the sediments. Positive correlations between the Enterococcus genera and ORP, pH, and TDS were also observed for batch cultures inoculated with microorganisms isolated from water and sediments on the fourth day of biodegradation. These positive correlations are in agreement with the increased relative abundance of Enterococcus genera that is observed in Figure 8. Escherichia genera also showed a positive correlation with electrical conductivity (EC), ORP, pH, and TDS on the first and second day of the biodegradation process for microorganisms isolated from the water and sediments samples collected from the Blesbolspruit wetland. These positive correlations, therefore, confirmed that these two genera were the most active in the biodegradation of the n-alkane derivatives batch cultures. These two species are facultative anaerobic microorganisms that maintain their metabolic activity in both aerobic and anaerobic conditions. This explains why their dominance in both the aerobic and anaerobic conditions that was observed during the biodegradation process. These genera are tolerant to varying environmental conditions.

3.3.5. Predicted Metabolic Function Variations

The various gene functions associated with the biodegradation of n-alkane derivates inoculated with microorganisms from the water and sediment samples collected from Site 1 of the Blesbokspruit wetland. There was a moderate presence of microorganisms responsible for mycothiol (MSH) biosynthesis, which plays a crucial role in maintaining the redox balance within the microbial cells and also protects the cells from oxidative stress brought about by reactive oxidative species generated during the aerobic biodegradation of hydrocarbons such as the n-alkane derivatives [62,63]. Trace amounts of mycothiol are often produced by Actinobacteria genera as a response mechanism against harsh external environmental conditions and, therefore, could have been produced during the initial stage of aerobic biodegradation, where strongly oxidizing molecular oxygen radicals are produced [64]. This can be confirmed by the presence of Actinobacteria genera on the first day of biodegradation for microorganisms isolated from the water and sediment samples. The tricarboxylic acid V (TCA) cycle metabolic activity was more dominant on day 1 of the biodegradation process for both water and sediment microorganisms. The TCA cycle is the second stage of microbial cellular respiration in which fuel molecules such as hydrocarbons are broken down under aerobic respiration (in the presence of oxygen) to produce energy. This is the last step in the aerobic biodegradation of n-alkanes and their derivatives which results in the mineralization of hydrocarbons to produce energy and carbon dioxide (CO₂). This, therefore, infers that some of the n-alkane derivatives in the batch cultures could have started to be broken down to complete mineralization within the first day of the biodegradation process [65]. L-1,2 propanediol degradation was observed on day 2 of the biodegradation process (Figure 10b). This metabolic activity indicates that propanediol formed through the subterminal oxidation of 1-propanol during the initial aerobic biodegradation process. The produced propanediol would then be oxidized to form a ketone and subsequently an acetyl ester and fatty acids that would then be converted to acetyl CoA to be oxidized in the TCA to produce energy and carbon dioxide. Fatty acid salvage and fatty acid and beta-oxidation I were the predominant metabolic functions observed on day 4 (Figure 10c) of the biodegradation process, with relatively higher significance observed in microorganisms isolated from water. Fatty acid salvage is the enzyme-driven process where short-chain fatty acids are elongated into long-chain fatty acids through condensation of acetyl CoA. Butyric acid, valeric acid, and the downstream fatty acids metabolites could have possibly been involved in the fatty acid biosynthesis process to form longer chain fatty acids, which were then broken down through β-oxidation to form the acetyl CoA that is further oxidized through the TCA metabolic pathway to form adenosine triphosphate (ATP) for cellular energy use and storage. The protocatechuate degradation II pathway, which is a common route in biodegradation of aromatic hydrocarbons, was also observed from microorganisms isolated from both the water and sediments samples. This metabolic pathway is catalyzed by dioxygenase, which is an enzyme that oxidizes n-alkanes and their derivatives. The sulfate reduction pathway was also observed on day 4 of the biodegradation process for microorganisms isolated from the water and sediment samples. Sulfate reduction is a microbial redox process where sulfates are reduced into insoluble sulfides in the presence of organic substances as electron donors under anoxic conditions. These electron donors include hydrocarbons that get oxidized during the electron transfer process, which results in the reduction of the soluble sulfates to form sulfide precipitates. This dual process has been used in the simultaneous precipitation of toxic metals and biodegradation of hydrocarbons such as n-alkanes in the bioremediation of polluted water. The Phylogenetic Investigation of Communities by Reconstruction of Unobserved States (PICRUSt) results also revealed that there was sulfate assimilation to form cysteine.

4. Implications of the Findings

Short-chain alkane derivatives, such as 1-propanol, butyric acid, and valeric acid, are dominant and less complex hydrocarbons found in FTWW. Their high susceptibility to microbial biodegradation, as opposed to high molecular weight aromatic and polyaromatic hydrocarbons that are also found in FTWW, allow them to be used in theoretical baseline studies in an attempt to investigate and elucidate the biochemical processes and operational dynamics in the biodegradation of FTWW hydrocarbons. Numerous studies have been conducted to demonstrate the capabilities of various microorganisms in biodegrading the hydrocarbons from petroleum, oil, and FTWW. However, the common challenge in most of these studies is to determine the most ultrafunctional site within the sampling area that harbors the most active hydrocarbon-utilizing microorganisms. In this study, we presented an unbiased approach where microorganisms isolated from the water and sediment samples collected at different sampling sites with varying pollution gradients are tested for their COD reduction efficiencies in biodegradation of the short-chain n-alkane derivatives. Microorganisms with the highest COD reduction efficiency indicated high biodegradation potential of hydrocarbons in FTWW and were therefore used for downstream studies to investigate the applicability of physicochemical and microbiological system performance indicators in the biological treatment of FTWW. Ultrafunctional microorganisms isolated from the water and sediment samples from site 1 exhibited the highest biodegradation capabilities, as shown in Figure S1. As shown on the GIS Map and site description in the previous study, site 1 had extensive agricultural activity, which could be the major source of pollution. Sheep skins were also dumped openly into the water at this wetland segment. Hydrocarbon-degrading microorganisms have been isolated from abattoir waste and have shown remarkable biodegrading capabilities. Ogbonna et al. (2012) [65] isolated hydrocarbon-utilizing microorganisms from abattoir wastewater and soil for biodegradation of polyaromatic hydrocarbons in mineral salt broth, and the results obtained showed that these microorganisms could degrade low molecular weight polyaromatic hydrocarbons as opposed to the high molecular weight components. Ariyo et al. (2016) [66] also isolated heterotrophic microorganisms from soil samples of abattoir waste, which showed great potential for hydrocarbon biodegradation. There is a high possibility that microorganisms found on the sheep skin will overlap with the microbial communities found in abattoir waste because the sheep skin is a component of abattoir waste. This, therefore, could account for the high tolerance and biodegradation capacity of the microorganisms isolated from the water and sediments from site 1 in this study.

The COD reduction results indicated a high microbial activity on the first day of biodegradation in the propanol, butyric acid, valeric acid, and the PBV mixed batch cultures, and this was attributed to the high cell proliferation and metabolic activity of the microbial cells due to enough essential nutrients, such as oxygen, nitrates, phosphates and cations. Figure 10a showed relatively high metabolic activity for the TCA cycle, which, therefore, confirms the accelerated oxidation of some of the short-chain n-alkane derivatives in the presence of sufficient oxygen and nutrients in the initial stages of the biodegradation process. The decline in pH of the batch cultures could, therefore, be ascribed to the trapped carbon dioxide during the mineralization of the hydrocarbon substrates. The high biodegradation rates were observed at the pH of about 7 and steadily declined with the decrease in pH, mainly because most hydrocarbon-utilizing microorganisms attain their optimal catalytic activities at near neutral pH. The decline in biodegradation rate in the remaining 3 days was also ascribed to the insufficient oxygen availability as the biodegradation reaction progressed, and this was evidenced by the sharp decline in the ORP values. This decline in ORP on day 1 indicated a shift in the system chemistry of the batch cultures from an aerobic to an anaerobic condition. These shifts, which could be brought about by the inconsistent oxygen supply in batch culture media, were confirmed by the sulfate-reducing metabolic activity in day 4 of the biodegradation process. The sulfate reduction pathway is an anaerobic process where sulfates are reduced to sulfides with concomitant oxidation of the electron donor species, such as hydrocarbons. Based on these findings, it is advisable to use mechanical aeration methods to maintain an aerobic niche within the system bioreactor to avoid these shifts if aerobic biodegradation is desired. The decline in biodegradation efficiencies of the microorganisms collected from both the water and sediments on day 1 was also ascribed to the limitations of nutrients, such as nitrates, phosphates, and cations, and this was confirmed by the decrease in EC and TDS readings. Therefore, EC and TDS can be used to indirectly monitor the nutrient levels in a biological treatment system, which could inform on the nutrient dosing time and rates. The metabolic functional responses in Figure 10 were able to reveal the possible biodegradation pathways of the n-alkane derivatives in the batch cultures, which could also be explained by the physicochemical variations observed during the biodegradation process.

5. Conclusions

Microorganisms isolated from the water and sediments collected from site 1 of the Blesbokspruit wetland segment exhibited a very high COD reduction of the n-alkane derivatives in a short reaction time (96 h) and elevated initial concentration of 300 ppm. This indicated that the microbiome at this site of the wetland was more ultrafunctional and tolerant to the short-chain n-alkane derivatives found in FTWW and, therefore, had a greater potential for biodegradation of FTWW. The batch culture systems shifted from aerobic to anaerobic conditions over time, and this was confirmed by the drastic decline in ORP values and the detection of aerobic metabolic activities at the early stage of biodegradation and a shift into anaerobic metabolic activities at the later stage of biodegradation. Enterococcus and Escherichia genera were more dominant on most days of biodegradation, therefore, indicating that these genera were actively involved in the biodegradation process of the n-alkane derivatives. These genera displayed positive correlations with EC, ORP, pH, and TDS during the four-day biodegradation process. This study demonstrated that physicochemical and microbiological indicators can be used to infer the biodegradation patterns and dynamics of n-alkane derivatives in Fischer–Tropsch wastewater. These findings can be used in the eminent system diagnosis and improvement of FTWW biological treatment plants. Although it is crucial to confirm the biodegradation pathways of the n-alkane derivatives in the batch cultures in order to trace the produced metabolites during the biodegradation process, this simple and fast approach can be adopted by most wastewater treatment works in developing countries where high cost and rigorous analytical methods, such as Gas Chromatography–Mass Spectroscopy (GC-MS), are not feasible. This study provides new knowledge on the system chemistry and microbial community and function variations in biodegradation of n-alkane derivatives from Fischer–Tropsch wastewater, which would allow for timely system control restoration and therefore avoid ultimate system failure.