Effect of the C/N Ratio on Biodegradation of Ciprofloxacin and Denitrification from Low C/N Wastewater as Assessed by a Novel 3D-BER System

Abstract

Emerging pollutants in the form of pharmaceuticals have drawn international attention during the past few decades. Ciprofloxacin (CIP) is a common drug widely found in effluents from hospitals, industrial and different wastewater treatment plants, as well as rivers. In this work, the lab-scale 3D-BER system was established, and more than 90% of the antibiotic CIP was removed from Low C/N wastewater. The best results were obtained with the current intensity being taken into account, and a different C/N ratio significantly improved the removal of CIP and nitrates when the ideal conditions were C/N = 1.5–3.5, pH = 7.0–7.5 and I = 60 mA. The highest removal efficiency occurred when CIP = 94.2%, NO₃⁻-N = 95.5% and total nitrogen (TN) = 84.3%, respectively. In this novel system, the autotrophic-heterotrophic denitrifying bacteria played a vital role in the removal of CIP and an enhanced denitrification process. Thus, autotrophic denitrifying bacteria uses CO₂ and H₂ as carbon sources to reduce nitrates to N₂. This system has the assortment and prosperous community revealed at the current intensity of 60 mA, and the analysis of bacterial community structure in effluent samples fluctuates under different conditions of C/N ratios. Based on the results of LC-MS/MS analysis, the intermediate products were proposed after efficient biodegradation of CIP. The microbial community on biodegrading was mostly found at phylum, and the class level was dominantly responsible for the NO₃⁻-N and biodegradation of CIP. This work can provide some new insights towards the biodegradation of CIP and the efficient removal of nitrates from low C/N wastewater treatment through the novel 3D-BER system.

1. Introduction

Worldwide, the various kinds of emerging pollutants such as pharmaceuticals compounds, anti-inflammatory drugs, antibiotics, beta-blockers [1] and the massive amount of NO₃⁻ or NO₂⁻ have attracted international attention for the last few decades [2]. Shown in

Table 1. Physical and chemical properties of ciprofloxacin.

Name	Properties	Molecular Formula	Structure
Ciprofloxacin		C17H18FN3O3
CAS number	85721-33-1
* pKa	10.58 ± 0.30
	8.71 ± 0.089
	6.14 ± 0.13
	3.01 ± 0.30
* Log Kowb	0.40
Molecular weight	331.35 g/mol

* The pK_a values of ciprofloxacin were obtained from [30], and the value LogKow^b was taken from [31], respectively.

, the antibiotic ciprofloxacin (CIP) is the 3rd-generation fluoroquinolone group usually used in humans and veterinarians [1,3,4]. The high-proportion of CIP is not fully metabolized in livestock and humans, but instead is evacuated as a parent substance [5]. Hence, CIP can reach the environment through various pharmaceutical industries, sewage treatment plants, livestock activities, landfills, application of sewage sludge [6], manure or treated wastewater to agricultural land [3,7,8]. However, due to the potential development and dissemination of antibiotic resistance, this poses a potential threat to ecosystems and human health [2,3]. Therefore, CIP removal must be considered before being released into the environment [9]. Antibiotic CIP is readily detectable in man-made aquatic-environments (usually found in surface water at ng·L⁻¹ to µg·L⁻¹ levels) [7]; it occurs at high levels in the effluents of Wastewater Treatment plants (WWTPs) (up to 6.55–31 mg·L⁻¹) receiving pharmaceutical wastewater and rivers polluted with industrial waste (up to 14 mg·L⁻¹) [6,10,11]. During wastewater treatment, 80–90% of CIP was removed by adsorption to sludge, which stabilizes the substance [3].

Some scientific reports and data have shown that biodegradation [7] and the adsorption process [9] were the most effective elimination methods for various kinds of antibiotics in WWTPs [4]. It has been shown in some studies that around 50–100% adsorption is the main path for CIP removal through biodegradation in anaerobic sludge system [12,13]. This clearly shows the importance of sludge being released into the environment as a CIP reservoir and developing sludge management strategies [14]. In general, anaerobic digestion was a standard procedure for the stabilization of sludge; it was also meant to extract organic matter without particular regard for the removal of the antibiotics [8]. The biodegradation rate of absorbent CIP durability was 0–40% during sludge treatment [9,15]. A considerable amount of CIP persisted in the digested sludge contained in wastewater treatment plants [10]. Unfortunately, there were no reports on the microbial community structure of CIP degradation [16,17]. The traditional denitrification process depends on four basic denitrifying enzymes [18] including the respiratory of nitrate reductase [19], which is shown as:

$${\mathrm{NO}}_3^{-}\to {\mathrm{NO}}_2^{-}\to \mathrm{NO}\to {\mathrm{N}}_2\mathrm{O}\to {\mathrm{N}}_2$$

However, many processes reported in previous studies were CIP effects on the denitrifying enzymes and the activity of existing denitrifying bacteria [19,20,21]. In recent years, the biofilm electrode reactors (BERs), which combine biological and electrochemical techniques, have been shown to efficiently eliminate nitrogen, heavy metal and antibiotics [10,22]. The autotrophic denitrify bacteria are immobilized on the cathode surface, and the hydrogen produced from the water-electrolysis is also used as an electron donor in the autotrophic denitrification process [21,22]. Therefore, the anode and cathode materials are typically made of carbon and thus provide an inorganic-carbon source that can be used as a pH buffer solution, as shown in Equations (1)–(3):

$$2{\mathrm{H}}_2\mathrm{O}+2{\mathrm{e}}^{-}\to {\mathrm{H}}_2+2{\mathrm{OH}}^{-}$$

(1)

$$\mathrm{C}+2{\mathrm{H}}_2\to {\mathrm{CO}}_2+4\mathrm{H}+4{\mathrm{e}}^{-}$$

(2)

It was shown that the process of denitrification uses hydrogen instead of an organic carbon source [23] to completely covert nitrate to N₂, as per the following reaction [18,24]:

$$2{\mathrm{NO}}_3^{-}+5{\mathrm{H}}_2\to {\mathrm{N}}_2+4{\mathrm{H}}_2\mathrm{O}+2{\mathrm{OH}}^{-}$$

(3)

Since hydrogen is generated from the cathode, it could not cause annoying safety problems, and the dosage was only excessive due to the low solubility of hydrogen [18,24]. The BER only requires an inorganic carbon source [25]. From the literature, the organic-carbon also accumulates in the BERs to enhance the elimination of antibiotics [25,26] and nitrogen through the different C/N ratios [7,27]. The simultaneously heterotrophic-autotrophic bacteria has an improved denitrification process compared to the traditional BER, consumes a large amount of energy and is relatively inefficient [28,29]. We recently introduced a 3D-BER system for the new advanced technique in the potential benefits of low currents, and there is the best capability to treat toxic substances. Hence, effective and economical methods for eliminating CIP antibiotics and nitrate or nitrite were developed. It is also important to improve the methodology to remove the residues of CIP antibiotics that develop when the high manipulative capacity of the anaerobic process occurs in the 3D-BER system.

This study demonstrated that the removal efficiency of antibiotics CIP and nitrogen in the effluent of low C/N wastewater can be improved by using a novel 3D-BER system. The primary objectives of this particular study were: (1) to examine the potential toxicity of ciprofloxacin and nitrogen removal; (2) to comprehensively analyze the effect of pH on antibiotics-ciprofloxacin and denitrification; (3) to appraise the degradation by microbial inhabitants and investigate the intermediate product of CIP. Therefore, this is the first study of the removal of CIP to enhance the denitrification process that can be used for solving many problems in low C/N wastewater. As compared to a traditional low C/N ratio in wastewater treatment systems, this technology is also designed to be a more efficient, useful complement or a cost-effective alternative.

2. Materials and Methods

2.1. Configuration of the 3D-BER System

The schematic view of the three-dimensional bioelectrochemical reactor system (3D-BERs) was used in this study to evaluate the biodegradation of ciprofloxacin and nitrates removal from Low C/N wastewater (Figure 1). The reactor was made of a plexiglass cylinder with a diameter of 10 cm, a height of 20 cm, a total volume of 1.2 L and a working volume of 0.785 L. The cathode consisting of eight graphite rods (with a height of 20 cm and diameter of 0.8 cm) were placed around the periphery of the reactor, while the one graphite rod (with a height of 20 cm and diameter of 1.5 cm) as an anode was fixed in the center of the reactor and was connected using insulated electrical copper wire by DC-power. A DC-controlled power supply: Model No: HQ3003SIII, supplied by Hansheng Puyuan Technology Beijing Co., Ltd. (Beijing, China), (with the ranges 0–3 A and 0–30 V) was used. Electrolysis of water is the decomposition of water into hydrogen (H₂) and oxygen (O₂) due to the passage of an electric current. Meanwhile, from top to bottom, space was filled with granular activated carbon (GAC) functioning as third particle electrodes (size 1.5–3 mm) used for microbial growth that were purchased from Younga Chemical Technology (Nanjing, Jiangsu Co., Ltd., China). The GAC was the best option for the development of bacteria on the cathode and anode on the inner zone and on the bottom zone in the reactor. The pre-treatment of GAC was washed several times with deionized water via a sulfuric acid solution (0.01 M) and then dried for 24–30 h, and was finally heated to 110 ℃.

2.2. Chemicals and Reagents

Ciprofloxacin (Analytical-grade > 99%) was obtained from Sigma Aldrich (Munich, Germany). A stock solution of antibiotic CIP (100–500 µg·L⁻¹) was prepared by using a Milli-Q 18 MΩ system (Millipore, Darmstadt, Germany). The pH was adjusted using 1 M of H₂SO₄ solution prepared with concentrated sulfuric acid (95–97% purity). The high performance liquid chromatography (HPLC) analysis developed by Merck and Fisher was employed with acetonitrile (HPLC quality) and O-phosphate acid (85%). The synthetic wastewater was composed of 30 mg·L⁻¹ (NO₃⁻-N), 3.5 mg·L⁻¹ K₂HPO₄, 4.5 mg·L⁻¹ NaCl, 5.56 mg·L⁻¹ KH₂PO₄, 2.75 mg·L⁻¹ CaCl₂, 500 µg·L⁻¹ C₁₇H₁₈FN₃O₃ (ciprofloxacin) and 0.1 mL L⁻¹ of concentrated trace elements stock solutions, which were prepared by dissolving 1.5 g MgSO₄. 7 H₂O, 2.2 g MnSO₄. H₂O, 2.2g ZnSO₄. H₂O, 0.24 g CoCl₂. 6H₂O, 2 mg NiCl₂. 6 H₂O, 10 mg FeCl₃. 6H₂O, 0.5 mg CuCl₂. 2 H₂O, 0.5 mg Na₂MoO₄. H₂O and 1.5 g CaCl₂ into 1 L of deionized water [32,33]. Nitrates and antibiotics water effluent samples were filtered by using 0.22-micron membrane filters (Millipore, Darmstadt, Germany) prior to testing. The pH of simulated influent wastewater is typically 7.5 ± 0.2, and no further adjustment was required.

2.3. Experimental Procedure

2.3.1. Bacterial Adaptation Phase

All experiments runs were inoculated with a mixed culture of acclimated autotrophic-heterotrophic bacteria (

Table 2. The initial concentration of Low C/N wastewater at different C/N ratios.

Parameters (mg·L−1)	S1	S2	S3	S4	S5	S6	S7
Electric Current (mA)	60	60	60	60	60	60	60
COD	0	15	30	45	75	90	105
NO3−-N	30	30	30	30	30	30	30
C/N	0	0.5	1.0	1.5	2.5	3.0	3.5

). These microbes were enriched from the anaerobic sludge (Mixed Liquor Suspended Sludge: MLSS.100 g), which was taken as a bacterial source of inoculation from the Nanjing municipal wastewater treatment plant (WWTP). Before cultivation, 1.2 L of anaerobic sludge water was placed in a refrigerator with nutritious material at 4 ℃ for seven days. Meanwhile, the synthetic water, which included to NaAc, Nitrates and KH2PO4, was added into the anaerobic sludge with a ratio of C/N/P = 3:1:0.5. During the first three days, sludge water was circulated by a magnetic pump. The concentration of nitrate-nitrogen in the influent was maintained at approximately 30 mg·L−1. After 21 days, anaerobic sludge water was placed in the reactor, and 0.350 L of tap water was added to a total volume of 0.785 L. During the cultivation process, the electrical current was not supplied to the 3D-BER system after two weeks; the direct electrical current was instead gradually adjusted to 10 mA.

2.3.2. Immobilized GAC Biofilm at Various C/N Ratios

The reactor was started for 70 days of batch operations at an electrically applied current of 60 mA. Each cycle of the 3D-BERs (feed, react, settle, decant) was 40 min. The dissolved oxygen (DO) in the reactor was kept in a range from 1.25 mg·L−1 to less than 0.2 mg·L−1, and the pH was maintained from 7.0 ± 0.5 to 8.0 ± 0.3. The pre-treated sludge was mixing with the ratio of GAC/Sludge of 3:1.5, and was introduced into the cathode and anode layers of the 3D-BERs. For the batch experiments, the biofilm in the anode/cathode surface, as well as the third electrode GAC, turned into dark grey at room temperature (27 ± 1.5 ℃). The influent was renewed two times within 24 h due to fast growth microorganism in the reactor. In the 3D-BER system, sustaining an anaerobic condition was needed to expand denitrifying bacteria. By observing the total nitrogen (TN) concentration in the reactor, when TN was exhausted, the residual substrate could be removed, and fresh substrate was fed to the reactor to avoid an endogenous metabolism of the microorganism.

2.3.3. Analytical Methods and Calculations

On a UV-Visible spectrophotometer (Shimadzu UV-1800, Shimadzu, Kyoto, Japan), the nitrate $(\mathrm{N}-{\mathrm{NO}}_3^{-})$, was measured by a UV-spectrophotometric method at wavelength λ = 220–275 × 2 nm. Nitrite $(\mathrm{N}-{\mathrm{NO}}_2^{-}$) was measured by the N-(1-naphthyl) ethylenediamine dihydrochloride spectrophotometric method at λ = 540 nm, and total nitrogen (TN). $\mathrm{N}-{\mathrm{NH}}_4^{+}$ was measured by Nessler’s reagent colorimetric method using a UV-visible spectrophotometer at λ = 420 nm, respectively [32]. The pH and ORP were measured by a pH100 pH meter, delivered by Shanghai Yoke Instrument Co., Ltd. (Shanghai, China). A thermometer was inserted into the reactor to monitor the temperature during the experiments. All samples were stored at 4 ℃ prior to the analysis. At least three samples were taken for each test. Current and voltage were recorded every 60 s to control the pH, bacterial metabolism and hydrolysis process. The removal performance of nitrate and CIP antibiotics was calculated based on the percentages of nitrate reduction and CIP as follows in Equations (4) and (5):

$$\mathrm{The}\mathrm{removal}\mathrm{efficiency}\mathrm{of}\mathrm{Nitrates}\mathrm{reduction}(\%)=\frac{\mathrm{Nitrate}\mathrm{in}-\mathrm{Nitrate}\mathrm{out}}{\mathrm{Nitrate}\mathrm{in}}\times 100$$

(4)

where “in” is the initial concentration of nitrate, and the “out” is the final concentration of nitrate at time t.

$$\mathrm{The}\mathrm{removal}\mathrm{efficiency}\mathrm{of}\mathrm{CIP}\mathrm{reduction}(\%)=\frac{\mathrm{CIP}\mathrm{in}-\mathrm{CIP}\mathrm{out}}{\mathrm{CIP}\mathrm{in}}\times 100$$

(5)

where “in” and “out” are indications of CIP antibiotics concentration in the inlet and the outlet of the 3D-BERS, respectively.

2.4. Ciprofloxacin Concentrations Measurement and By-Products Identification

The CIP detection wavelength was set at 277 nm (fluorescence detection: 240–480 nm), which was determined by using high-performance liquid chromatography (HPLC, Tokyo, Japan) [33]. However, the antibiotics CIP concentrations in the influent were 100 µg·L⁻¹ to 500 µg·L⁻¹, respectively. The CIP separation was carried out using a Phenomenex column of asymmetry column, C18 (3.5 µm × 4.6 mm × 250 mm). The mobile phase was composed of acetonitrile and 0.01 mol L⁻¹ tetra-n-butyl ammonium bromide (C₁₆H₃₆BrN) solution, trimethylamine at pH 3.0 and methanol with a ratio of 88:12 v/v at a flow rate of 1.0 mL min⁻¹. The amount and processing time of the injection were 10 μL and 7.244 min, respectively. In the procedure, the reaction solution was extracted with a syringe needle, which was then directly determined by HPLC [33]. The biodegradation of CIP was monitored with mass-spectrometry (SCIEX Triple Quad^TM 7500 LC-MS/MS, California, USA). The extracts were analyzed by using a gradient method (0.3 mL L⁻¹) and an Acclaim¹²⁰ column C18 (2.1 × 150 mm × 2 μm) under two mobile phases. In phase A: water containing 0.1% formic acid and, in the meantime, in phase B-acetonitrile. To assess biodegradation products, a thorough analysis of positive and negative ion modes was used. However, the molecular ions observed in positive and negative ions were subsequently applied in electrospray ionization techniques to provide a chromatographic shape of these putative biodegradation products. To extract LC-MS/MS spectra of the points determined in the final step, high collision energy was exclusively adjusted, and ion chromatographic response intensity was optimized in the range of 45 eV. From the literature, some researchers confirmed that the BES could improve the mechanism of antibiotic removal from the WWTPs due to adsorption techniques, which is a very economical system. The growth medium of the microbial community utilized the energy and C/N source [34,35]. However, the external calibration curve was used to determine the ciprofloxacin antibiotics concentration with a correlation coefficient (R² = 0.985).

2.5. DNA Extraction, PCR Amplification, and High-Throughput Sequencing

High-throughput sequencing was performed on original inoculants and immobilized biofilm samples taken from the 3D-BERs (anode/cathode) layers (1.0 g) and effluent (100 mL) after long-term acclimation at different C/N ratios. The names of the biofilm samples were identified as S1, S2, S3, S4, S5, S6 and S7. All samples were stored in −85 ℃ for 24 h in a freezer before tests. The bacterial genomic DNA was extracted by the PowerSoil DNA^® Isolation Kit (Waters, Eschborn, Germany: MOBIO) according to the manufacturer’s instructions. The primers 338F (5′-ACTCCTACGGGAGGGAGCAG-3′) and 806R (5′-GGACTACHVGGGTWTCTAAT-3′) was used to amplify the V3–V4 hypervariable regions of the bacteria 16r RNA gene by Macrogen Inc. (Seoul, Korea). The PCR reaction that followed was: Initial denaturation at 98 ℃ for 1.5 min, denaturation at 98 ℃ for 30 s, annealing at 50 ℃ for 30 s and 30 extension cycles at 72 ℃ for 30 s. PCR reactions occurred in a triplicate 20 µL mixture. The PCR products that appeared, as a result, were extracted from a 2% agarose gel and further purified using the Gene JET DNA Gel extraction Kit (Thermo Scientific ^TM, CN: K0691, California, USA). The concentration and purity of DNA were analyzed by the Quanti-Fluor^TM St-Fluorometer (Madison, WI, USA). The resulted productions were pooled in equimolar and paired-end sequenced (300 × 2) on an Illumina MiSeq platform (Illumina, San Diego, CA, USA) according to the protocols developed by Beijing Novogene: Genome Sequencing Company Co., Ltd. (Beijing, China). Using this program, we truncated the adapter with low-quality sequences < Q20, and short sequences < 300-bp were trimmed. Disposing of potential chimeric sequences using the Mothur chimeric UCHIME algorithm. The taxonomic classification of the sequences was conducted using the cutting and filtering readings classifier. The observed operational classifiers’ richness (OTU), diversity index such as ChaoI and abundance coverage estimator (ACE) were determined by the Mothur function [36,37].

2.6. Biofilms GAC Samples Analysis by SEM

The seven biofilms samples of particle electrodes surface morphology for bacteria in GAC-biofilms were examined by using scanning electron microscopy (SEM) analysis. Prior to SEM analysis, a series of batch experimental sample processing techniques (i.e., fixed, washed, dehydrated, dried, coating, etc.) was applied. About 0.8 g of GAC biofilm samples were placed in a centrifugal tube (C.T) of 10 mL, and 6000 revolutions/separation were conducted over 5 min. Then, the supernatant was removed and mixed with a sediment sample at the bottom of the C.T, following which 7 mL of 2.5% glutaraldehyde was added. We placed the C.T in the refrigerator at 4 ℃ for 3 to 4 h, and protected the sample. The tube was centrifuged to centrifuge at 6000 rpm for 5 min, to remove the supernatant after the centrifugation process. We then added about 7 mL of ultrapure water to the GAC biofilm sample left at the bottom of the C.T, then used the centrifuge again for 5 min at 6000 rpm, then removed the top liquid. This clean washing was done three times through the centrifugation process [21,38].

For the next stage, we washed the GAC biofilm, dried it and added 7 mL of 20%, 40%, 50%, 70%, 80%, and 90% ethanol each following 10 min periods. Biofilm samples were dehydrated three times, while 7 mL of ethanol (100%) after 15 min each time. Afterwards, a 7 mL blend of ethanol and acetyl (Ethanol: Isoamyl acetate ratio = 1:1) was discharged into the C.T for 15 min. Finally, we placed the 10 mL samples on filter paper and dried them in a vacuum dryer (Mecha-Tech system Ltd., Bristol, UK) for 24 h at a pre-cooling temperature of 10 ℃ [23,39]. Then the coating (10 nm gold layer) sample was examined with Hitachi S-4300 FEG-SEM (Bavaria, Germany, Europe), and the images were collected digitally. The SEM images are shown in Figure 2, including the morphology of biofilm attached to granular activated carbon (GAC) particle electrodes in the 3D-BERs. The GAC had a high specific area of about 450–900 m² g⁻¹ and a porous structure to facilitate the growth of microbial attachment was reported [20,28,40]. The biofilm samples of the S4, S6 and S7 were abundant compared to the S1, S2, S3 and S5, corresponding to the thicker bacterial community at 60 mA [41]. In the S1 to S3 zones, the microbes were reduced in quantity due to lacking electron donors. It was found that in the 3D-BERs with GAC, biofilm was imbedded due to the abundance of microbial activity [40]. This was clearly observed in rod-shaped graphite anode and cathode, which was authentic and common in the other denitrification system [20,42]. There were also giant microbes found, which may be hydrogen-heterotrophic denitrifying bacteria.

2.7. Statistical Data Analysis

All effluent samples analytical tests and microbiological analyses in this study were performed twice each time, and the batch test results were obtained by averaging the values of repeated measurements. We used Microsoft Excel (2018) (Microsoft, Redmond, WA, USA) and also used Origin Pro 8.5 and IBM SPSS (v.20) (IBM, Armonk, NY, USA) for the statistical analysis. The final results were expressed as mean ± standard error values, and were considered with statistically significant standards of p < 0.05. For all biological studies, seven independent biological tests were performed.

3. Results and Discussion

3.1. Removal Performance and the Effects of C/N Ratios

This analysis shows the average nitrogen concentration (i.e., NO₃⁻-N, NO₂⁻-N, NH₃-N, and TN) in the effluent under different C/N ratios during the batch experiments (i.e., I = 60 mA and pH 6.0–8.0), as shown in Figure 3. In the influent, the NO₃⁻-N concentration remains relatively stable at 30 mg·L⁻¹. When the C/N ratio increases from 0 to 1.5 and 2.5 to 3.5, the average NO₃⁻-N concentration of the influent falls from 26.09 mg·L⁻¹ to 0.9177 mg·L⁻¹ and 11.96 mg·L⁻¹ to 0.199 mg·L⁻¹, respectively (Figure 3a). Moreover, the NO₃⁻-N concentration in the final effluent was high when the ideal condition of the C/N ratio was 1.5–3.5, and the final concentration was 0.916 mg·L^–1–0.199 mg·L⁻¹ as shown in Figure 3c,g. After 30 min, the nitrates concentration dropped rapidly, indicating a significant increase in denitrifying bacteria [19], and that the anoxic condition had been achieved [42]. It was observed that the removal efficiency of nitrogen increased with the amplified C/N ratios.

In particular, the denitrification performance of 3D-BERs with a current intensity of 60 mA was slightly higher in the system as compared with published previous work, as shown in

Table 3. Comparison from the literature of antibiotic Ciprofloxacin removal and denitrification efficiency with previous studies of BER and 3D-BERS.

Reactor Type	Mode	H/A-D	Treatment Source	E.V (L)	Initial (NO3-N) (mg/L)	C/N Source	Power Source	Input Energy	D.R.R (%)	CIP Removal (%)	Country	References
Three-dimensional bioelectrochemical reactor system (3D-BERS)	Batch	H&A	Synthetic municipal wastewater	0.738	30 mg/L	CH3COONa	DC	60 mA	95.53	94.20	China	This Study
Three-dimensional biofilm-electrode reactor (3D-BER)	Continuous	H&A	Synthetic municipal wastewater	3.4	30 mg/L	CH3CooNa	DC	40 mA	98.3	N/A	China	[18]
Three-dimensional biofilm electrode reactors (3D-BERs)	Batch	A	Synthetic wastewater	2	50–100 mg/L	Hydrogen & α-Fe2O3	DC	20–80 mA	95 to 98.9	N/A	China	[20]
A three-dimensional BER	Continuous	H&A	Synthetic groundwater	0.5	20 mg/L	C2H5OH	DC	15mA	100	N/A	China	[22]
Bioelectrochemical System (BES)	Batch	H	Synthetic groundwater	1.0	250 g/m3 day	CH3CooNa	DC	N/A	14.6 ± 0.2 36.2 ± 5.0	N/A	USA	[25]
A combined single-chamberMFC and BER system	Continuous	H&A	Synthetic groundwater	0.480	364 mg/L	CH3OH	MFC	500–700 mV	30	N/A	China	[37]
A combined BES and sulfur autotrophic denitrification system (CBSAD)	Continuous	A	Synthetic groundwater	33.47	20.9–22.0 mg NO3-N/L	Hydrogen	DC	1000 mA	95	N/A	China	[45]
Bioreactor	Continuous	H&A	Synthetic municipal wastewater	0.9	30 mg/L	C6H12O6	--------	--------	N/A	20–22	Singapore	[46]
Fe-C micro-electrolysis reactor and up-flow biological aerated filter (UBAF)	Continuous	H&A	Synthetic municipal wastewater	15	60–100 mg/L	Fe-C micro-electrolysis	N/A	N/A	N/A	90	China	[47]

Noted: H/A-D = Heterotrophic /Autotrophic denitrification; DC = Direct Current; E.V = Effective Volume; C/N = Carbon to Nitrogen ratio; D.R.R= Denitrification Removal Rate; CIP = Ciprofloxacin.

. When the ratio of C/N increased from 0 to 1.5 and 2.5 to 3.5, the average concentration of total nitrogen in wastewater decreased from 29.26 mg·L⁻¹ to 0.955 mg·L⁻¹ and from 17.26 mg·L⁻¹ to 0.651 mg·L⁻¹, respectively. As shown in Figure 3h, TN average removal efficiency of 3D-BERS was 35.60% to 63.90%, 70.59% to 79.58%, 96.81% to 90.83% and 90.83% to 97.82% when the C/N went from 0 to 0.5, 1 to 1.5, 2.5 to 3.0 and 3.0 to 3.5, respectively (p < 0.05) [43,44]. In 3D-BERS, the average removal efficiency was respectively 96.81% and 97.82% when the C/N ratio was 1.5 and 3.5. At the same time, as the ratio of C/N increased from 0 to 1.5 and 2.5 to 3.5, the average NO₂⁻N concentration decreased from 2.52 mg·L⁻¹ to 0.44 mg·L⁻¹ and 1.22 mg·L⁻¹ to 0.398 mg·L⁻¹, respectively (Figure 3b). In addition, the C/N ratio was 1.5 and 3.5; the final effluent concentrations respectively were 0.44 mg·L⁻¹ and 0.39 mg·L⁻¹. However, under these conditions, the initial NO₃⁻-N loading of denitrification was high, so NO₂⁻-N in the wastewater remained at almost zero and was constant [19,43]. Also, the final effluent of NH₃-N in the 3D-BER system decreased from 4.12 mg·L⁻¹ to 2.29 mg·L⁻¹ and from 4.06 mg·L⁻¹ to 1.49 mg·L⁻¹; the different C/N ratios also shifted from 0 to 1.5 and 2.5 to 3.5, and the average NH₃-N concentration in the final effluent was 4.06 mg·L⁻¹ to 1.49 mg·L⁻¹, while the C/N ratio was 1.5 to 3.5, in comparison to other techniques previously studied [27,42,43].

Organic carbon sources are known to be limiting factors for heterotrophic denitrification. The removal of nitrate and total nitrogen increases as the ratio of C/N increases [22]. At the maximum level, the C/N ratio endorses the development of heterotrophic-denitrifying microorganisms, thereby allowing denitrification [21,22,47,48]. If there is a shortage of organic carbon during this period, then the nitrogen removal will usually decrease, and the NO₂⁻-N reductase competes with the electron donor nitrate reductase at a low level of C/N ratio. The reductase of nitrite is more complex and sensitive to the experimental conditions compared to nitrate reductase, resulting in the accumulation of NO₂⁻-N [22].

Autotrophic denitrifiers were dominated by a low concentration of the C/N ratio required to extend the adaptation period with rapid growth, and had a further accumulation of NO₂⁻-N [40]. As the C/N ratio increased between 0.5 and 1.5, the effluent NO₂⁻-N reduced rapidly from 2.52 mg·L⁻¹ to 0.242 mg·L⁻¹. At the ratio C/N = 1.5, the NO₂^—N continues to decline at lower levels. Heterotrophic denitrification may play a pivotal role at an advanced level of the C/N ratio when the rate of denitrification is high. When the C/N ratio was 1.5 and 3.5, the removal efficiency increased by 95.53–85.73% and 80.27–73.85% for NO₃⁻-N and TN, respectively [49]. However, NH₃-N tends to behave in opposite ways to NO₃⁻-N and TN. These results are a good comparison with previous findings [19,21,22]. The NO₃⁻ was commonly used as an acceptor of electrons and could be reduced to NH₄⁺ under anoxic and electron-acceptor conditions [20,45].

The 3D-BER system was an anoxic condition, and the force of the electric field prevents NO₃⁻ from moving to the surface of the anode, so there may be no electron acceptors near the surface of the cathode. In this study, ratio C/N is very close to the complete denitrification in the range of 1.5 to 3.5 and was established by using single heterotrophic denitrification [27,29,50]. Several scientific studies reported on the elimination of nitrogen were conducted by using autotrophic denitrification [48,51]. Under this method, elimination of NO₃⁻-N, NH₃-N and total nitrogen increased while the accumulation of NO₂⁻-N decreased gradually. In this study, the complete denitrification was obtained in C/N = 1.5 and this may be due to the system temperature and denitrification removal rate. However, a similar observation was also reported by other researchers, as shown in Table 3 [40,50].

3.2. Effects of pH on Antibiotics CIP, Nitrate and Total Nitrogen Removal

As shown in Figure 4, the influences of various pH values on the average concentration of nitrogen and ciprofloxacin (Total nitrogen, NO₃⁻-N, and CIP) during the C/N ratio in the batch process from 0 to 3.5 and electrical current intensity was I = 60 mA. A series of batch experiments was conducted in 3D-BERs for 70 days of operation in different C/N ratios to evaluate the NO₃⁻-N and CIP removal. In the novel 3D-BERs, the elimination of NO₃⁻-N decreased from 25.07 mg·L⁻¹ to 9.18 mg·L⁻¹ and from 4.41 mg·L⁻¹ to 0.199 mg·L⁻¹, and for CIP there was a reduction from 377.5 µg·L⁻¹ to 106.36 µg·L⁻¹ and from 115.9 µg·L⁻¹ to 106.3 µg·L⁻¹ while the C/N was between 0 to 1.5 and 2.5 to 3.5 at the pH range of 6.0 to 7.5 [19,20,40,44]. Under those conditions, the maximum removal efficiencies of NO₃⁻-N, TN and CIP were found to be 35.66% to 95.55%, 13.93% to 80.27% and 12.59% to 94.20%, while the C/N ratio was between 0 to 1.5 and 2.5 to 3.5 at a electric current intensity of 60 mA, respectively [19,48,52]. The maximum antibiotic CIP removal efficiency could be reached at 84.61% and 94.20% when the C/N was 3.5 and 1.5 at pH values of 7.0 to 7.5. In this system, the highest removal rate was at pH 7.5 as compared to a pH value of 6.7 to 8.0 [37,40]. However, pH played a pivotal role in ciprofloxacin degradation and gradually NO₃⁻-N removal. It was observed that the average concentration of CIP in influent was decreased from 377 to 106.36 µg·L⁻¹ and 115.9 to 116.99 µg·L⁻¹ when the C/N ratio shifted from 0 to 1.5 and from 2.5 to 3.5. At the C/N = 1.5 and 3.5, the final effluent of CIP concentrations was 106.36 µg·L⁻¹ and 116.99 µg·L⁻¹, respectively, as shown in Figure 5 [50,53,54]. It was shown that denitrification inhibited CIP, and that development was more obvious during the prolonged microbial cultivation period [6,55,56].

Notably, the pH was an essential factor in the denitrification process and removal of antibiotics. For most denitrifying bacteria, the ideal pH was 7.0–8.2, and previous researchers reported that a pH below 6.0 or 9.0 could indicate immobile heterotrophic and autotrophic denitrification [37,48]. However, when the pH = 7.5, antibiotic CIP, NO₃⁻-N and total nitrogen (TN) were 106.36 μg·L⁻¹, 0.917 mg·L⁻¹ and 0.024 mg·L⁻¹, respectively, and the final effluent amount was the lowest level. The CIP was entirely removed within 16 h of batch experiments at the average removal efficiency of 94.20 ± 0.5%, as reported [10] and observed in this study. Furthermore, the highest average removal efficiency of CIP, TN and NO₃⁻-N reached 94.20%, 80.27% and 95.53%, and the enhanced removal efficiency was obtained in 3D-BERS.

3.3. Effect of pH on Biodegradation Mechanism of the Antibiotic Ciprofloxacin

Generally, the ideal pH for Autotrophic–Heterotrophic denitrification systems is chosen to be 7.5, because recent reports have shown that increasing antibiotics CIP levels can significantly reduce the performance of autotrophic denitrification systems [6,56]. For another factor that was selected as a variable, the final result under optimal conditions was pH = 7.5. Therefore, the latest novel 3D-BER system was introduced to enhance the denitrification process and achieved the highest removal efficiency for nitrogen. However, Equations (1) and (3) were initiated to indicate that the pH illustrates the OH⁻ produced during the denitrification process and water electrolysis. In our research work, we found that the pH level of the final effluent was maintained at 7.5 ± 0.2 [51,57]. Moreover, it indicates that the CO₂ produced by the graphite anode not only dissolved in the carbonate but also dissolved in the hydrogen carbonate, as shown in Equation (3), and reacted with OH⁻, so that the pH in the following equations is equal to:

$${\mathrm{CO}}_2+{\mathrm{H}}_2\to {\mathrm{H}}_2{\mathrm{CO}}_3$$

(6)

$${\mathrm{H}}_2{\mathrm{CO}}_3+ {\mathrm{OH}}^{-}\to {\mathrm{H}}_2\mathrm{O}+{\mathrm{HCO}}_3^{-}$$

(7)

$${\mathrm{HCO}}_3^{-}+{\mathrm{OH}}^{-}\to {\mathrm{H}}_2\mathrm{O}+{\mathrm{CO}}_3^{2-}.$$

(8)

The pH of the system was always in a favourable condition for the biological activity of the denitrifying bacteria. Some scientific research suggests that CIP was predominantly eliminated by adsorption during biological wastewater treatment [7,13]. Previous publications demonstrated that CIP is difficult to manage in regular environments and is hardly degraded by biological processes [9]. Sludge adsorption is the main method to remove ciprofloxacin in wastewater treatment [8,26,46]. Novel research by Li and Zhang et al. showed that there was an initial CIP concentration of 100 μg·L⁻¹ after 48 h of incubation with a mixture collected from the aeration tank of a biological wastewater treatment plant treated with saline sewage and with a removal capacity of about 32.2% [26]. However, some work on the biodegradation of CIP by biological wastewater treatment processes is significant. The effect of the degradation mechanism of ciprofloxacin antibiotics is affected by changes in the microbial community [9,46,56]. For this research work, the CIP biodegradation was investigated through a series of batch experiments using nitrified sludge anoxic sludge, and the operating system was 28 h at different C/N ratios. However, the anaerobic sludge is very useful for the biodegradation of CIP, with a concentration of 500 µg·L⁻¹ in the 3D-BER system.

3.4. Identification of Degradation Pathways by LC-MS/MS

After solid-phase extraction (SPE in Germany), biodegradation experiments on ciprofloxacin antibiotics were performed by using LC-MS/MS. Nevertheless, the LC-MS/MS identified a total of four CIP biotransformation products developed by LC-MS/MS: LCQAD-6460, Agilent Technologies, Santa Clara, CA, USA (the molecular structure and intermediates of these products are revealed in Table S1, and the total ion chromatogram (TIC) is demonstrated in Figure S1). These four possible biodegradation pathways of the antibiotic ciprofloxacin are shown in Figure 6. Moreover, in this pathway, C₂H₂ with a piperazinyl substituent of CIP is removed by using demethylations to generate CIP-BBP1 at the pathway point “S” and is transferred to CIP-BBP4 at the conduit point “Q” by the loss of the C₂H₅N fragment of CIP-BBP1 (Table S1). While the biodegradation of CIP intermediates products involved CIP-BBP1 and CIP-BBP4 as previously reported [16,58], similar products were also identified in 3D-BERS. In the first time of this study, these products were investigated and have not been listed in the literature. In the pathway at point “S”, CIP-BBP1 is shaped via hydroxylation (-OH) at the CIP-BBP1-S position, and CIP-BBP4 is also shaped via hydroxylation at the CIP-BBP4-Q position by replacing the piperazinyl group of CIP-BBP1with CIP-BBP4, as presented in Table S1. In the P and Q-points pathways, CIP-BBP2 and CIP-BBP4 were investigated by substituting for both the fluorine at the P-position of CIP [15] and the carboxyl group at the Q-position of CIP with hydroxylation. Meanwhile, the biodegradation of CIP by acetylation of the NH group of piperazine rings caused inactivation of enzyme culture, i.e., it caused the CIP-acetylation process [59]. Therefore, the 4-pathways can be divided into two main reactions: Piperazinyl-substituent decomposition reactions (a), and hydroxylation reactions (b). Figure S1 indicated that the biodegradation mechanism of antibiotic CIP was proposed in the 3D-BER system. Almost all CIP can be converted into the four products mainly through degradation and hydroxylation of the piperazinyl substituent.

Pearson’s bivariate correlation analysis results showed that CIP biodegradation was significantly correlated in SPSS (p < 0.01, R² = 0.996). To better describe the CIP metabolic degradation pathways [4,23] at the gene and predominant enzyme levels, superior isolates of ciprofloxacin in anaerobic sludge can be used, as has been indicated by transcriptomics and metagenomics studies [31,59]. The information obtained from LC-MS/MS was insufficient to fully characterize the chemical structure of these metabolites. As a result, realistic standards will be used in future work to confirm the formation of Ciprofloxacin biotransformation pathways.

3.5. Bacterial Diversity and Community Composition

The evolution of ciprofloxacin degradation and denitrifying bacteria both occur after long-term acclimation to low C/N wastewater.

Table 4. Diversity indices of bacterial communities in different sludge samples of 3D-BERS at different C/N ratios and with ab applied electric current of 60 mA.

Condition (C/N)	Sample ID	No. of Reads	No.of OTUs	ACE	Shannon	Simpson	Chao1	Coverage
0	S1	39,322	182	254.321	2.642	0.321	254.119	0.993
0.5	S2	40,215	189	259.583	2.997	0.224	245.337	0.995
1.0	S3	41,234	254	288.981	3.456	0.423	288.327	0.996
1.5	S4	73,382	382	390.705	6.367	0.838	399.229	0.999
2.5	S5	54,633	267	297.149	3.994	0.455	345.657	0.994
3.0	S6	59,236	263	311.597	4.214	0.754	358.279	0.995
3.5	S7	63,324	353	363.642	5.247	0.936	361.571	0.999

summarizes the microbial species richness and diversity indices for these two samples. Figure 7 illustrates the phylogenetic classification of phylum and class level sequences for the two main samples in different batch operations. Proteobacteria, Bacteroidetes and Firmicutes were identified as predominant phyla in seven samples, but their relative abundance was different in each sample (Figure 7A). However, examination of all of the samples showed that the bacteria which dominated in the samples S4 and S7 were more abundant in those samples than in the other samples. The abundance of Proteobacteria was the highest out of the different phyla in every sludge sample, and accounted for 48.35%, 51.29%, 61.35%, 95.53%, 71.35%, 78.91% and 83.71%, respectively, of total phyla content in each sample, which is why it was considered to be the main phylum of autotrophic-heterotrophic denitrifying bacteria.

It was clearly observed that the richness of Firmicutes increased 7 times (raised from 0.011% to 5.63%) when the NO₃⁻-N concentration increased from 0 to 30 mg/L. Firmicutes are generally accepted to have denitrification potential under anaerobic environments, meaning it could survive under extremely oligotrophic conditions by relying on its ability to produce endospores for resisting environment stress.

According to S1 samples of Firmicutes (0.011%) obtained was less abundant [15] than the number of S7 samples (5.63%). The relative abundance of class levels is shown in Figure 7B using taxonomic classification.

Furthermore, Gammaproteobacteria, Alphaproteobacteria, Bacteroidia, Bacillus, Deltaproteobacteria and Clostridia, were preponderant classes in sludge samples (Figure 7B). The abundance of Gammaproteobacteria accounted for 35.93%, 39.91%, 50.71%, 90.84%, 48.32%, 49.09%, and 74.91% in S1, S2, S3, S4, S5, S6 and S7, respectively, which became the most abundant classes under different C/N ratios.

However, the preponderant classes in the S4 sample were Alphaproteobacteria (29.37%), Gammaproteobacteria (90.84%), Bacteroidia (1.87%), Deltaproteobacteria (4.17%), Bacilli (8.34%) and Clostridia (0.014%) [2,49,59]. It was evident that the distribution of the main classes in these seven samples was significantly different. Nevertheless, the amount of Gammaproteobacteria present in sample S4 was more significant than that of any other samples compared to the distribution of Alphaproteobacteria, Bacteroidia, Deltaproteobacteria, Bacillus, Actinobacteria, and Clostridia [1,3,44]. The equivalent distributions of anaerobic sludge in the S1, S2, S3 and S5 samples were relatively low compared to for the samples S4, S6 and S7 from the 3D-BER system.

It is important to note that the similarities across some variations of S1 to S7 samples shown in Figure 7C indicate the genus levels of more than 65 abundant genera. For any abundant genus with a relative richness of more than 2.5%, the abundance of Pseudomonas was significantly higher in the sample S4 (75.34%) when the C/N ratio was 1.5 [9]. In comparison, the relative abundance of Pseudomonas in the sample S1 was only 11.44%, while the C/N ratio was 0.5 [4,56]. In general, this result has been widely found in previous studies focusing on Pseudomonas and Aeruginosa, especially in terms of the possibility of denitrification [19,60]. Also, Pseudomonas stutzeri [40,61], and the certain type of Pseudomonas.sp bacteria belonging to the genus Pseudomonas.sp C27 has autotrophic and heterotrophic denitrifiers that utilize an assortment of electron donors [39]. As the C/N ratio increases, the impact on the bacteria output involved in autotrophic denitrification and the microbial growth yield of denitrifying autotrophic microbes are smaller than that of heterotrophic bacteria. Despite a gradual increase in the C/N ratio, the denitrifying autohydrogenotrophic bacteria may have been gradually domesticated to become heterotrophic denitrifying bacteria that lead to effective nitrates and to antibiotic CIP elimination [62,63]. Therefore, Pseudomonas abundance in sample S4 indicates that an overall similarity greater than 94% with autotrophic and heterotrophic denitrifying species is related to Pseudomonas in the occurrence of organic substances and applied electrical currents. Moreover, significantly higher amounts of Thauera.sp were detected in sample S4 (6.86%) compared to in sample S1 (0.24%). Bacillus (15.52%), Thiobacillus (28.8%), Flavobacterium (3.81%) [4] and Acinetobacter (9.59%) are the most important genus of S4 samples, but the abundance for the S7 example involved Thiobacillus (23.34%), Bacillus (5.6%), Acinetobacter (4.53%) and Flavobacterium (0.99%) [2,47]. Thiobacillus has been identified as a widespread autotrophic bacterium in recent years, closely related to the oxidation of nitrates to nitrogen [39,51]. In recent years, several species belonging to the genera Thiobacillus, Bacillus and Thauera have been suggested to play an important role in antibiotic CIP and removal of nitrates [56]. Notably, Thiobacillus and Thauera can transfer electrons directly from the carbon electrode and promote nitrate reduction [16,49,59]. The richness and abundance of Pseudomonas (55.04%), Thiobacillus (28.88%) and Thauera (6.68%) [24,64] were enriched in the 3D-BER system. The removal of CIP and nitrates was due to the distinction between the dominant Thauera, Pseudomonas and Thiobacillus, which helped to merge heterotrophic denitrification and autotrophic denitrification processes. In conclusion, the main bacteria enhanced in the third electrode (GAC) continues to be enhanced in both heterotrophic-autotrophic denitrifications, whereas enriched bacteria in S1 are primarily involved in autotrophic denitrification. Therefore, further research is needed to characterize the denitrifying bacteria in the 3D-BER system.

4. Conclusions

Satisfactory antibiotic ciprofloxacin and nitrates removal efficiency was achieved in a novel 3D-BER system from Low C/N wastewater. This system has proved to be more economically and technically attractive than traditional denitrification systems due to the more than 90% reduction of antibiotics and a 55% reduction of sludge production that it causes. Perhaps some resistant bacteria could be used along with antibiotic CIP or dead biomass as an organic substrate for better growth of microbial activities in an anaerobic condition. The microbial community on biodegrading was enriched with a high richness of Pseudomonas (75.44%), making it capable of autotrophic-heterotrophic denitrification, although it also contains Thiobacillus (48.88%) and Bacillus (21.05%) predominant autotrophic bacteria. Biological degradation was the main mechanism for removing CIP and NO₃⁻-N or NO₂⁻-N and TN, with few intermediate products detected. Furthermore, this reactor configuration represents a low-cost form of deployment, operation and maintenance technology, compared to other systems and advanced 3D-bioelectrochemical treatment systems.