Long-term post-storage reactivation of a nitrifying sludge in a sequential batch reactor: physiological and kinetic evaluation

Abstract

Production, preservation and recovery of sludge with stabilized nitrifying activity over long time can be difficult. Information on the ability of nitrifying sludge to regain its nitrifying activity after long-term storage is still scarce. In this work, the physiological and kinetic changes during the reactivation and stabilization of a nitrifying sludge previously exposed to ampicillin (AMP) were evaluated in a sequential batch reactor (SBR) after its long-term storage (1 year) at 4 °C. After storage, both ammonium and nitrite oxidizing processes were slow, being nitrite oxidation the most affected step. During the reactivation stage (cycles 1–6), physiological and kinetic activity of the nitrifying sludge improved through the operating cycles, in both its ammonium oxidizing and nitrite oxidizing processes. At the end of the reactivation stage, complete nitrifying activity was achieved in 10 h, reaching ammonium consumption efficiencies (ENH₄⁺) close to 100% and nitrate yields (YNO₃⁻) of 0.98 mg NO₃⁻-N/mg NH₄⁺-N consumed without nitrite accumulation. During the stabilization stage (cycles 7–17), results indicated that the sludge could maintain a steady-state respiratory process with restoration percentages of 100% for nitrifying specific rates (qNH₄⁺ and qNO₃⁻) with respect to their values obtained before storage. Furthermore, during the addition of 15 mg AMP/L (cycles 18–21), the sludge preserved its metabolic capacity to biodegrade 90% of AMP in 2 h. Therefore, long-term storage of nitrifying sludge could be used to preserve nitrifying inocula as bioseeds for bioremediation and bioaugmentation strategies.

Introduction

One of the biggest problems worldwide is water pollution, largely due to anthropogenic activities without care for the environment (Zhang et al. 2021). Ammonium is a nitrogenous compound which is released through the discharge of wastewater from the chemical, mining, agricultural and food industry, among others, causing environmental damages and even the death of aquatic organisms (Beristain-Cardoso et al. 2009; Kouba et al. 2019; Stevens 2019). Total ammonium concentrations released by these industries have been reported to range between 100 and 600 mg/L (Qasim and Mane 2013). Therefore, the elimination of ammonium from wastewater is necessary for preserving the environment (Lu et al. 2020). Nitrification is an important biological process in the removal of ammonium nitrogen in wastewater treatment plants. This process is mainly carried out by chemolithotrophic bacteria in two aerobic stages: the oxidation of NH₄⁺ to NO₂⁻ by the respiratory activity of ammonium oxidizing bacteria (AOB) and subsequently the oxidation of NO₂⁻ to NO₃⁻ by nitrite oxidizing bacteria (NOB) (Tang and Chen 2015). However, this process is sensitive to changes in dissolved oxygen (DO) and substrates concentration, temperature, pH and the presence of organic compounds. Variations of these parameters can cause the loss of nitrifying activity (Sahaya Sukeetha et al. 2020). Additionally, compared to heterotrophs, nitrifying bacteria have a slow growth (µ = 0.2 ± 0.1 day⁻¹) which makes difficult the recovery of the nitrification process in wastewater treatment plants (Pollard 2006). Nitrification performance of microorganisms in different types of wastewater and treatment systems must be evaluated because during wastewater treatment, the continuous variation of wastewater composition can lead to incomplete, slow and unstable nitrifying activity with low ammonium removal efficiencies and rates in addition to low ammonium conversion yields into nitrate (Texier et al. 2012). Bioaugmentation is one of the proposed strategies to recover or enhance biological processes involved in the removal of contaminants in wastewater plants, including nitrification (Boon et al. 2000; Wang et al. 2009; Tang and Chen 2015).

Studies at laboratory-scale under controlled conditions on nitrifying consortia usage are extensive. Recently, the metabolic and cometabolic abilities of nitrifying sludge and AOB for the simultaneous biotransformation of ammonium and diverse contaminants such as carbonaceous, aromatic, emerging organic and sulfurous compounds have been reported (Tran et al. 2013; Margot et al. 2016; Salas-Cortés et al. 2017; Huiliñir et al. 2018; Martínez-Jardines et al. 2018; Suárez-García et al. 2019; Trejo-Castillo et al. 2021). In recent years, an evident need for producing stabilized nitrifying sludge to investigate their physiological contribution in the removal of different contaminants has been recognized as it can represent a good opportunity for bioremediation strategies (Bejarano Ortiz et al. 2020; Ramírez Muñoz et al. 2020). This could be exploited for the enhanced biodegradation of pollutants in wastewater treatment by using enriched nitrifying sludge (Mai et al. 2016). Therefore, the production, preservation and recovery of sludge with stabilized nitrifying activity over long time, which in high efficiencies, yields and specific rates of the process can be maintained, have become important tasks for studies at the laboratory level and practical purposes such as bioaugmentation. However, the start-up and maintenance of nitrifying reactors in steady-state nitrification as source of inoculum can be long and difficult. In this sense, it is relevant to evaluate the capacity of nitrifying sludge which has been remained in long-term storage for recovering its physiological and kinetic abilities and subsequently being used as nitrifier source in nitrifying processes. Therefore, more studies on the possibility to store stabilized nitrifying sludge are essential for basic and applied research.

It is known that during short- and long-term storage, diverse effects on sludge from different biological systems used in wastewater treatment can be observed, including changes in color, structure, sedimentability and substrates consumption capacity (Yan et al. 2017). Regarding this, in the study of Zeng et al. (2007), phthalic acid (PA)-degrading aerobic granules were cultivated and stored at 4 °C for 8 weeks. The authors observed a decrease of the heterotrophic activity for PA biodegradation as a diminishment of 96% was determined in the specific oxygen utilization rate (SOUR). After 7 operation days in sequential batch reactors (SBR), granules restored their heterotrophic activity registered before the storage. Wang et al. (2008) evaluated through the SOUR, the heterotrophic and autotrophic activity of an aerobic granular sludge after its storage at 4 °C for 7 months. Their results indicated that after this storage time, the heterotrophic and nitrifying activity decreased by 80%. However, activities of heterotrophs and nitrifiers could be completely recovered in a SBR within 16 and 11 days of reactivation, respectively. Zhang et al. (2020) reported that the storage of an autotrophic nitrifying granular sludge at room temperature for 40 days caused the breakdown of the granules since the first day of storage and a 24% decrease of SOUR. The authors reported that a rapid recovery of the nitrifying sludge activity was achieved after 27 operation days in a SBR by adding a concentrate of exogenous nitrifying bacteria. Even though nowadays studies on the storage of aerobic granules are available, physiological and kinetic information on the effects of long-term storage of nitrifying consortia is still scarce. The nitrifying oxidation products (nitrite and nitrate) are rarely measured as well as the nitrifying kinetic rates. Moreover, some authors mentioned that it is necessary to investigate more the reactivation capacity of sludge cultivated on toxic contaminants and the preservation of their metabolic abilities for degrading these compounds after storage under non-selective conditions (Zeng et al. 2007). This is especially needed for nitrifying sludge with abilities to degrade organic compounds.

The main objective of this study was to evaluate the reactivation and stabilization capacity of a nitrifying sludge for carrying out the ammonium and nitrite oxidizing respiratory processes after its long-term storage (1 year). The reactivation and stabilization processes were performed in a SBR. In the present study, ampicillin (AMP)-degrading nitrifying sludge was used as a model to evaluate the storage effects on its ability to recover and stabilize, along the operation cycles both its nitrifying activity and antibiotic consumption capacity. Throughout the operating cycles of the SBR, the physiological evaluation of nitrification was performed by determining ammonium consumption efficiencies (ENH₄⁺) and ammonium oxidation to nitrite and nitrate expressed as yields (YNO₂⁻, YNO₃⁻), while kinetic changes were assessed through lag phases (ʎ) and specific rates of ammonium consumption (qNH₄⁺) and nitrite (qNO₂⁻) and nitrate (qNO₃⁻) formation.

Materials and methods

Nitrifying sludge activity background

Prior to its storage, the sludge used in this study was obtained at the end of operation of a SBR system fed with ammonium (100 mg N/L) and AMP (14.4 mg/L) under nitrifying conditions. Each operating cycle was of 24 h. From the first cycle of operation until cycle 109 in the absence of AMP, the sludge maintained a complete nitrifying activity and in a steady state. With AMP addition (cycle 110), the sludge presented a complete and stable nitrifying activity only during the last 15 cycles of operation (cycles 150–165), reaching an ENH₄⁺ of 99.8% ± 0.2, YNO₂⁻ of zero and YNO₃⁻ of 0.87 ± 0.10 mg NO₃⁻-N/mg NH₄⁺-N consumed. Specific rates of nitrification were the following: qNH₄⁺ of 1.2 ± 0.2 mg NH₄⁺-N/mg microbial protein d and qNO₃⁻ of 1.2 ± 0.2 mg NO₃⁻-N/mg microbial protein d. The nitrifying sludge presented the capacity to remove the totality of AMP in less than 30 min at a specific rate of 1.92 mg AMP/mg microbial protein d. The sludge maintained good settling properties with a volumetric index of 17.9 ± 1.4 mL/g.

Storage conditions of the nitrifying sludge

Before storage, the microbial biomass obtained from the SBR (500 mL) was twice washed with an isotonic saline solution (NaCl, 9 g/L). The sludge was suspended in a liquid phase (0.8 L) consisting of mineral culture medium (g/L): (NH₄)₂SO₄ 0.26, NH₄Cl 0.34, KH₂PO₄ 0.33, MgSO₄ 0.23, NaCl 0.23, NaHCO₃ 2.0 and CaCl₂ 0.01. The nitrifying sludge was placed in a 1-L plastic container with a liquid volume of 0.8 L and a headspace of 0.2 L. The container was sealed to prevent contamination and stored for one year under refrigeration at a temperature of 4.0 ± 0.2 °C.

Reactivation and stabilization of nitrification: Sequential Batch Reactor setup

A 2-L laboratory-scale SBR with a working volume of 1.6 L was employed (Fig. 1). After one year of storage, the nitrifying sludge (400 mL) was washed three times with an isotonic saline solution. After inoculation of the SBR, the biomass concentration as microbial protein was 400 ± 6 mg/L. The culture medium fed into the reactor consisted of nitrogen (medium N) and carbon (medium C) sources. The chemical composition of medium N was (g/L): (NH₄)₂SO₄ 1.73, NH₄Cl 1.39, KH₂PO₄ 2.73, MgSO₄ 1.26 and NaCl 1.0. Medium C was composed of the following (g/L): NaHCO₃ 12.83 and CaCl₂ 0.005. Bicarbonate was used as carbon source and pH buffer. At the beginning of each operation cycle, the initial ammonium concentration was 101 ± 2 mg NH₄⁺-N/L, giving a C/N ratio of 2.5. The SBR was operated under the following conditions: temperature of 25 ± 5 °C, initial average value of pH of 7.4 ± 0.2, stirring rate of 200 rpm and continuous aeration giving a DO concentration of 5.8 ± 0.1 mg/L. This remanent DO concentration indicated that oxygen availability was not limited, satisfying the stoichiometric oxygen requirements of nitrification for an initial ammonium concentration of 101 ± 2 mg NH₄⁺-N/L.

Fig. 1.

Sequential batch reactor setup

Each operation cycle of the SBR was divided into four phases: feeding (5 min), reaction (from 30 d to 10 h), sedimentation (45 min) and drainage (10 min) (Fig. 1). Operation phases were electronically controlled by using timers and peristaltic pumps. Reactivation stage was the first operation step of the SBR, wherein in order to reach the total consumption of ammonium, the operating cycles (1–6) had a variable reaction time duration: 30, 26, 22, 19, 16 and 7 days, respectively. Afterward, the second step corresponded to the stabilization stage, which was conducted during cycles 7–17. In the second stage, cycles consisted of 11 h. Finally, in a third operation step, nitrifying sludge activity in the presence of 15 mg AMP/L was evaluated along cycles 18–21. In the third stage, cycles duration was of 11 h. In order to assess the abiotic removal of the antibiotic in the SBR system, abiotic control assays in the absence of biomass were also performed under the same culture conditions.

Samples were taken at different times along the reaction time of each cycle for establishing the respective kinetic profiles of ammonium consumption, nitrite and nitrate formation as well as AMP consumption. Samples were centrifuged (4000 rpm) and filtered (0.2 µm) prior to their analysis.

Analytical methods

The modified Lowry colorimetric method was used to determine the microbial protein concentration in the reactor (Lowry et al. 1951; Martinez et al. 2000). DO and pH were quantified with selective electrodes (DO HI 98186 Hanna, Romania and Digi-Sense digital pH, USA, respectively). Ammonium quantification was performed using an ammonia selective electrode (Phoenix Electrode Company, USA). For the analysis of nitrite and nitrate, a liquid chromatograph (HPLC, PerkinElmer Series 200, USA) was used with an ion exchange column (IC-Pak Anion HC, 4.6 × 150 mm, Waters, USA) and a UV–visible detector. The wavelength was 214 nm, and the flow of the mobile phase was 2 mL/min (Trejo-Castillo et al. 2021). For the quantification of AMP (Sigma-Aldrich, product No. A6140), a liquid chromatograph (HPLC, PerkinElmer Series 200, USA) was used with a C18 column (stationary phase Microsorb 100, 4.6 × 150 mm, Varian, USA) at a wavelength of 230 nm and the flow of mobile phase was 1 mL/min (Ramírez Muñoz et al. 2020). For all analytical methods, standard curves were made in triplicate and analyzed as linear regressions, obtaining coefficients of variation for slopes lower than 7% and coefficients of determination (r²) greater than 0.99.

Evaluation of the nitrifying sludge activity in the SBR reactor

The nitrifying sludge activity in the SBR reactor was physiologically evaluated in terms of ammonium consumption efficiency (ENH₄⁺, [mg NH₄⁺-N consumed/mg NH₄⁺-N fed] × 100) and products formation yields (YNO₂⁻, [mg NO₂⁻-N produced/mg NH₄⁺-N consumed] and YNO₃⁻, [mg NO₃⁻-N produced/mg NH₄⁺-N consumed]), while it was kinetically assessed using the lag phases (λ) and the specific consumption rates of NH₄⁺-N (qNH₄⁺, [mg NH₄⁺-N/mg protein d]) and ampicillin (qamp, [mg AMP/mg protein d]) and the specific formation rates of NO₂⁻-N (qNO₂⁻, [mg NO₂⁻-N/mg protein d]) and NO₃⁻-N (qNO₃⁻, [mg NO₃⁻-N/mg protein d]).

Lag phases and specific consumption and formation rates were calculated by fitting experimental data to the adjusted Gompertz model by using the package NCSS 2021 software. Concentration of NH₄⁺-N or AMP consumed, NO₂⁻-N or NO₃⁻-N produced Y(t) (mg/L), at time t (days or hours), can be expressed as a function of time according to Eq. (1) where the parameters A, B and C represent the maximum concentration of substrate consumed or product formed (mg/L), the volumetric rate of substrate consumed or product formed (mg/L d or mg/L h) and the inflection point of the curve (d or h), respectively. Increments in the amount of NH₄⁺-N and AMP consumed (mg/L) were calculated along time regarding the, respectively, initial NH₄⁺-N or AMP concentrations. Consequently, Y(t) represents the increment of NH₄⁺-N or AMP consumed. In the case of nitrite and nitrate formation through time, Y(t) represents the concentration of NO₂⁻-N or NO₃⁻-N formed (mg/L). Once the parameters A, B and C were computed, lag phases (λ, d or h) (Eq. 2), maximum volumetric consumption or formation rates (V_max, mg/L d or mg/L h) (Eq. 3) and specific rates (q, mg/mg protein d or mg/mg protein h) (Eq. 4) were calculated. The r² values (r² > 0.98) and the model significance were used as adjusting criterion.

$$Y\left(t\right)=A\exp (-\exp (-B\left(\left(t\right),-,C\right)))$$ 1

$$\mathrm{\lambda }=\frac{BC-1}{B}$$ 2

$$V\max =\frac{\mathrm{AB}}{\exp }$$ 3

$$q=\frac{V\max }{\mathrm{Protein}\mathrm{concentration}}$$ 4

Results and discussion

Nitrifying sludge reactivation

Kinetic profiles during the reactivation stage of nitrifying activity of the sludge after its long-term storage are presented in Fig. 2. For the first cycle of operation (cycle 1), only the increase in the ammonium consumption was evaluated (Fig. 2a). The results obtained in this cycle showed that 28 d were required for the total consumption of 100 mg NH₄⁺-N/L. For the following cycles from cycle 2 to cycle 6, the time necessary to consume the totality of initial ammonium was decreasing gradually from 24 to 6 days (Table 1). This indicated that after storage and since the first cycle, the sludge maintained the metabolic ability to consume ammonium and this was getting better throughout the SBR cycles. Products resulting from the oxidation of ammonium were identified and quantified as nitrite and nitrate (Fig. 2b, c, Table 1), evidencing that ammonium consumption and its oxidation were associated with the nitrifying respiratory activity of the sludge. The maximum nitrite accumulation was observed in cycle 2 of operation with a maximum nitrite concentration of 86 mg N/L after 26 days (Fig. 2b), corresponding to a nitrite yield of 0.84 (Table 1). Concomitantly, the formation of nitrate in the same culture at cycle 2 was low, reaching only a value of 15 mg N/L after 26 days (Fig. 2c) and resulting in a nitrate yield of 0.14 (Table 1). These results showed that storage of the sludge affected mainly the nitrite oxidizing respiratory process. Along the cycles, nitrite accumulation decreased, whereas nitrate formation increased (Fig. 2b, c). These trends were accompanied by a decrease of final YNO₂⁻ values and an increase of final YNO₃⁻ values along cycles (Table 1). Since cycle 4, there was no significant accumulation of nitrite at the end of culture (YNO₂⁻ = 0.05 mg NO₂⁻-N/mg NH₄⁺-N consumed) and nitrate was obtained as the main product of ammonium oxidation (YNO₃⁻ = 0.94 mg NO₃⁻-N/mg NH₄⁺-N consumed). Finally, during cycles 5 and 6, NO₂⁻-N was not detected and 100 mg NO₃⁻-N/L were produced in 16 and 7 d, respectively (Fig. 2b, c). Therefore, the decrease in the final and transitory accumulation of NO₂⁻-N as well as its oxidation to NO₃⁻-N in a shorter time through the cycles showed the recovery and increment of the nitrite oxidizing activity. These results showed that after long-term storage and during the reactivation stage in the SBR, the sludge nitrifying activity physiologically improved through the operating cycles, in both its ammonium oxidizing and nitrite oxidizing activity after long-term storage, reaching ENH₄⁺ close to 100% and YNO₃⁻ of 0.99 mg NO₃⁻-N/mg NH₄⁺-N consumed (Table 1).

Fig. 2.

Kinetic profiles during reactivation stage of the sludge nitrifying activity after long-term storage: a increase in ammonium consumption, b nitrite accumulation and c nitrate formation. Cycle 1 (○), Cycle 2 (□), Cycle 3 (◊), Cycle 4 (△), Cycle 5 (●) and Cycle 6 (▲). Experimental values (symbols) and data adjusted by the Gompertz model (continuous line)

Table 1.

Physiological recovery stages of the sludge nitrifying activity after long-term storage

Stage	Cycle	Time required for total ammonium consumption (days)	ENH4+ (%)a	YNO2−b	YNO3−c
Reactivation	1	28	99.9 ± 0.2	NMd	NMd
	2	24		0.84 ± 0.02	0.14 ± 0.01
	3	21		0.50 ± 0.02	0.49 ± 0.02
	4	17		0.05 ± 0.02	0.94 ± 0.01
	5	15		0	0.99 ± 0.01
	6	6		0	0.99 ± 0.01
Stabilization	7–17	0.4	99.7 ± 0.1	0	0.98 ± 0.01
With AMP added	18	0.4	99.9 ± 0.1	0.28 ± 0.01	0.72 ± 0.02
	21	0.4	99.9 ± 0.1	0.28 ± 0.01	0.72 ± 0.02

^aENH₄⁺: Ammonium consumption efficiency (mg NH₄⁺-N consumed/mg NH₄⁺-N fed) × 100

^bYNO₂⁻: Nitrite production yield (mg NO₂⁻-N/mg NH₄⁺-N consumed)

^cYNO₃⁻: Nitrate production yield (mg NO₃⁻-N/mg NH₄⁺-N consumed)

^dNM: not measured

Kinetic data of the sludge reactivation are presented in Table 2. From cycles 1 to 6, the lag phase estimated by the Gompertz model for initiating ammonium consumption decreased gradually from 14.2 to 1.0 days. In terms of specific rate, the oxidation of ammonium started to be significantly faster along the cycles since cycle 4. From cycle 4 to 5, its specific rate was 1.5-fold higher while from cycle 5 to 6, it was up to 8.4-fold higher. Similarly, the lag phase for nitrate formation tended to decrease along the cycles while the specific rate increased. The results obtained during this stage indicate that after long-term storage, the nitrifying consortium was able to reactivate and physiologically and kinetically improve its metabolic capacity through the operating cycles, in both its ammonium oxidizing and nitrite oxidizing processes.

Table 2.

Kinetic recovery stages of the sludge nitrifying activity after long-term storage

Stage	Cycle	Ammonium consumption		Nitrite formation		Nitrate formation
		ʎa	qNH4+b	ʎa	qNO2−c	ʎa	qNO3−c
Reactivation	1	14.2 ± 0.1	0.014 ± 0.001	NMd	NMd	NMd	NMd
	2	8.3 ± 0.1	0.012 ± 0.001	6.2 ± 0.1	0.009 ± 0.001	21.2 ± 0.1	0.007 ± 0.001
	3	6.0 ± 0.1	0.014 ± 0.001	16.0 ± 0.1	0.014 ± 0.002	6.4 ± 0.1	0.012 ± 0.002
	4	4.1 ± 0.1	0.016 ± 0.001	2.8 ± 0.1	0.009 ± 0.001	9.3 ± 0.1	0.024 ± 0.004
	5	3.7 ± 0.2	0.024 ± 0.002	- -	0	5.7 ± 0.1	0.021 ± 0.005
	6	1.0 ± 0.2	0.201 ± 0.002		0	1.0 ± 0.1	0.165 ± 0.001
Stabilization	7–17	0.04 ± 0.01	1.2 ± 0.1	-	0	0.04 ± 0.01	1.2 ± 0.1
With AMP added	18	0.04 ± 0.01 0.04 ± 0.01	1.2 ± 0.1 1.2 ± 0.1	0.02 ± 0.001 0.02 ± 0.001	0.019 ± 0.001 0.019 ± 0.001	0.05 ± 0.01 0.05 ± 0.01	0.96 ± 0.01 0.96 ± 0.01
	21

^aʎ: Lag phase (days)

^bqNH₄⁺: Specific consumption rate (mg NH₄⁺-N consumed/mg protein d)

^cqNO₂⁻ and qNO₃⁻: Specific production rates (mg NO₂⁻-N or NO₃⁻-N formed/mg protein d)

^dNM: not measured

The effects of long-term storage of aerobic sludge under different conditions have been mainly reported through chemical oxygen demand (COD) removal efficiency and SOUR, used as bioactivity measurement factors (Vollertsen et al. 2001; Wan et al. 2014). Gao et al. (2012) showed that regardless of long-term storage conditions (8 months) such as temperature (− 25 °C, 4 °C and room temperature) and storage substrate (distilled water and 400 mg/L glucose solution) in an aerobic granular sludge, the heterotrophic activity was recovered in 90 and 100% for COD removal efficiency and SOUR, respectively, after 10 days of reactivation in a SBR. The recovery in ammonium removal efficiency was 80% after 30 days of reactivation; however, this consumption could be more associated with microbial growth than ammonium oxidation into nitrate via nitrification. In a similar way, Yuan et al. (2012) achieved the recovery of 90–100% of SOUR, 90–99% of COD removal and 38–100% of ammonium removal efficiencies in an aerobic granular sludge after a long-term storage (6 months) at room temperature and after 7 days of reactivation under different reactivation conditions (different organic loading rates, ammonium concentrations and shear force) in a SBR system. In the study, the authors suggested that ammonium was used mainly for heterotrophic biosynthesis because there was no accumulation of oxidized products, such as nitrite and nitrate, during cultivation and reactivation processes. Aerobic granules can simultaneously perform organic matter removal and nitrification processes, but, in the literature, more information is available on the recovery performance of the heterotrophic activity after storage than on nitrifying activity. Despite the reported works, physiological or kinetic information reported on nitrifying consortia is still limited. The present study provides a detailed evaluation of both physiological and kinetic activity during the recovery of the nitrifying process through the operating cycles of a SBR, showing that it is feasible to store nitrifying sludge for long time and use it later as bioseed for nitrifying processes. These results show that nitrifying bacteria can survive after long-term storage despite being reported as sensitive bacteria under adverse conditions (Sahaya Sukeetha et al. 2020). It has been reported that flocs are usually more sensitive to ambiental alterations than granules, so the physiological and kinetic information obtained in the present study with nitrifying flocs is also useful for granules.

Stabilization of the nitrifying process in the SBR reactor

The stabilization stage comprised eleven cycles, from cycles 7–17 (Fig. 3). During these cycles, the total ammonium consumption occurred in a time of 10 h (Fig. 3a and Table 1). The kinetic profiles for the increase in ammonium consumption and nitrate formation during the stabilization period maintained a similar behavior from cycles 7–17 (Fig. 3a, b). Physiologically, high ammonium consumption efficiencies were achieved with an average value of 99.7%, without accumulation of nitrite and with high nitrate formation yields of 0.98 mg NO₃⁻-N/mg NH₄⁺-N consumed (Fig. 3c and Table 1). Lag phases and specific rates of ammonium consumption and nitrate formation did not have significant differences through the operating cycles of this stage (Table 2). The lag phase was only of 1 h for both processes, 24-fold lower than its value obtained in cycle 6. The average specific rate in the consumption of ammonium and formation of nitrate had an increase of 6 and 7 times, respectively, compared to its value in the last operating cycle of the reactivation stage. During the stabilization stage, the results obtained have a coefficient of variation of less than 10%, indicating that the sludge had a complete and stable nitrifying activity; therefore, during this stage the sludge could maintain a steady-state regime. The recovered nitrifying performance of the sludge was good, presenting high ENH₄⁺ and YNO₃⁻ values and restoration percentages of 100% for qNH₄⁺ and qNO₃⁻ with respect to their values obtained before storage. The time required for the total recovering of nitrification performance was 120 days which is longer than periods reported for recovering heterotrophic activity of aerobic sludge (Gao et al. 2012; Yuan et al. 2012). This was probably due to the lower specific rates of growth of the AOB and NOB, the previous exposition of the sludge to AMP and the higher time of storage, among other factors.

Fig. 3.

Nitrifying sludge performance during stabilization stage: a profiles of the increase in consumed ammonium; b nitrate formation profiles; c efficiencies and yields, ENH₄⁺ (), YNO₂⁻(♦), YNO₃.⁻ (■). Cycle 7 (◊), Cycle 9 (○), Cycle 13 (△) and Cycle 17 (□). Figure 2a, b: experimental values (symbols) and data adjusted by the Gompertz model (continuous line)

Consumption of ampicillin by the reactivated and stabilized nitrifying sludge

From cycle 18–21, the nitrifying sludge was exposed to a concentration of 15 mg AMP/L. The profiles obtained for the increase in ammonium consumption with AMP did not present significant differences with respect to those observed in the stabilization stage (cycle 17) (Fig. 4a). However, the addition of the antibiotic caused an accumulation of NO₂⁻-N until the end of the cycles (Fig. 4b). On the other hand, a decrease in the NO₃⁻-N formation was observed with respect to the stabilization period (cycle 17) (Fig. 4c). Physiologically, at this stage, the ammonium consumption efficiencies were close to 100%, there was an increase in YNO₂⁻ at 0.28 mg NO₂⁻-N/mg NH₄⁺-N consumed, leading to a 26.5% decrease in YNO₃⁻ (Table 1). Kinetically, the nitrifying activity did not have significant changes in the lag phase or the specific rate of ammonium consumption due to AMP addition, but a decrease of 20% in qNO₃⁻ was observed (Table 2). These results indicate that the addition of 15 mg AMP/L negatively affected the nitrite oxidizing process at the physiological and kinetic levels. Contrarily, Ramírez Muñoz et al. (2020) have reported no physiological or kinetic effects of AMP (10–50 mg/L) on nitrification, probably because a sludge without prior AMP exposure was used and the evaluation was made during a unique 24 h batch culture. This suggests that sludge previously and repeatedly exposed to AMP would be more sensitive to the antibiotic effects and might provoke nitrite accumulation in dynamic nitrifying systems.

Fig. 4.

Behavior of the nitrifying sludge with 15 mg AMP/L: a increase in ammonium consumption, b nitrite formation, c nitrate formation and d increase in AMP consumption. Cycle 17, stabilization stage (♦), Cycle 18 (○), Cycle 21 (△), abiotic assay (◊). Experimental values (symbols) and data adjusted by the Gompertz model (continuous line)

Finally, the capacity of the nitrifying sludge was evaluated for the biological elimination of AMP. Biological tests with biomass and abiotic assays without biomass were carried out in the same SBR reactor, under the same experimental conditions used in the reactivation and stabilization stages (Fig. 4d). The abiotic results showed that AMP concentration was stable for 2 h under the conditions used. After this time, there was an increase in the AMP eliminated, probably due to physicochemical processes. In the presence of sludge and since the first cycle of AMP addition, 13.5 mg/L of AMP was biologically consumed in 2 h of culture, corresponding to 90% of the initial concentration. Therefore, after storage, the sludge conserved its metabolic ability to consume the antibiotic, probably due to the permanence of the microorganisms responsible for AMP biodegradation during the storage process. This could be possible through the liberation of cell lysis products that might be used as substrates and/or the endogenous respiration of the aerobic sludge (Yuan et al. 2012).

Conclusions

The results indicated that after long-term storage (1 year) and since the first cycle of operation, the sludge maintained the metabolic capacity to carry out nitrification; however, the nitrifying activity was slow, being the nitrite oxidizing respiratory process the most altered. During the reactivation stage in the SBR, the nitrifying activity improved physiologically and kinetically through the operating cycles, in both its ammonium oxidizing and nitrite oxidizing respiratory processes. After 120 days of reactivation, ENH₄⁺ close to 100% and high YNO₃⁻ values were reached after only 10 h of culture. After the reactivation period, it was also evidenced that the nitrifying consortium maintained a stable and steady-state process at a physiological and kinetic level. The present study shows that long-term storage of nitrifying sludge can be used as a strategy to preserve a sludge reservoir, allowing its later use as bioseed for nitrifying processes or inoculum for basic studies.

On the other hand, after storage, the sludge maintained its metabolic ability to consume 90% of AMP (13.5 mg/L) in 2 h and 100% of ammonium (100 mg N/L) in 10 h. However, the nitrite oxidizing process was more sensitive to the antibiotic exposure than the ammonium oxidizing process, causing nitrite accumulation. This should be considered if the storage purpose would be to recover the simultaneous capacity of the sludge to perform complete nitrification and biodegradation of the organic molecule.

Author contributions

DMC took part in methodology, investigation, writing the original draft and visualization. OOL was responsible for formal analysis, investigation, and writing, reviewing and editing. FMCL was involved in formal analysis, resources, writing, reviewing and editing, and supervision. ACT participated in conceptualization, formal analysis, resources, writing, reviewing and editing, supervision, project administration, and funding acquisition. All authors have read and approved the manuscript.

Data availability

All data generated or analyzed during this study are included in this article.

Declarations

Conflict of interest

The authors declare that they have no conflict of interest in the publication.