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Journal of Water Process Engineering
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Biodegradation of mono azo dye-Reactive Orange 16 by acclimatizing biomass systems under an integrated anoxic-aerobic REACT sequencing batch moving bed biofilm reactor
Chingyeh Ong
a, Khiamin Lee
a,*, Yunfah Chang
baDepartment of Civil Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Jalan Sungai Long, Bandar Sungai Long, Cheras, 43000, Kajang, Selangor, Malaysia
bTaylor’s Business School, Faculty of Business and Law, Taylor’ s University, 1, Jalan Taylors, 47500, Subang Jaya, Selangor, Malaysia
A R T I C L E I N F O Keywords:
Reactive Orange 16 Sequencing batch reactor Moving bed biofilm reactor Biodecolorization Autoxidation
A B S T R A C T
This study dealt with the biodegradation of Reactive Orange 16 (RO16) containing wastewater via a novel integrated anoxic-aerobic REACT sequencing batch moving bed biofilm reactor (SBMBBR). The chemical oxygen demand (COD) removal and biodecolorization from three design schemes pertinent to the effects of (i) dye concentration under stepwise approach,Scheme 1; (ii) dye concentration under shock load approach,Scheme 2 and (iii) hydraulic retention time (HRT) coupled with the increase of biocarrier filling ratio,Scheme 3were evaluated orderly along the evaluation. A complete biodecolorization and over 97 % of COD removal were attained inScheme 1. Although both COD removal and biodecolorization inScheme 2was deteriorated to ap- proximate 40 % by successive RO16 shocks, the microbial strength in removing RO16 molecules (mg RO16/mg biomass) reflected a two-fold amelioration from 0.015 to 0.0304 as the result of actively metabolic system. The maximum co-substrate uptake also reduced from 100 % to 69 %. At higher biocarrier filling ratio (10 % (v/v)) in Scheme 3, the difference in COD removal and biodecolorization rates reflected an improvement of 0.16 and 0.30
% hour-1, respectively, throughout the anoxic-REACT period. However, the performance changed insignificantly due to the reduction of HRT. The presence of the attached-growth biomass system was of great importance, although the suspended-growth biomass dominated more than 76 % of the biomass system in SBMBBR.
1. Introduction
The menace of dye-containing wastewater has reached an all-time high level due to the increasing trend in the world population [1].
Therefore, considerable attention has been paid to the remediation for the discharge of effluents containing synthetic dyes. Synthetic dyes are widely used in industries including textile, paper, leather, cosmetics and pharmaceuticals [2,3]. Amongst these industries, textile industry is notorious for its voracious water consumption and large quantity of wastewater discharge, has been ranked to be the primary culprit in posing such menace [4]. To date, the annual textile production has reached almost 800,000 tons [5]. An estimation of more than 10,000 different synthetic dyes has been commercialized and applied to the textile industry [6]. Azo dye is the most widely adopted class of dyes that constitutes more than 70 % of the textile production [7]. Reactive azo dye, in particular, possesses a broad spectrum of colors, high affi- nity to bind with cellulosic fiber and stable color fastness, is the most
consumable type of azo dyes [8,9]. However, the presence of electro- philic vinyl sulfone groups in their molecular structure incurred them to be hydrolyzed easily, resulting in weak adsorption or fixation towards the substrates [10,11]. Consequently, an approximate 280,000 tons of textile dyes are released annually, leading to a series of environmental and human health issues [12–14].
Albeit the fact that physicochemical treatments are effective for remediating the dye-containing wastewater, their implementation in the big-scale textile industry still being restricted by their steep in- vestment cost, high operation and maintenance costs, intensive energy consumption as well as production of hazardous by-products [15–18].
In contrast, a biological treatment that is more environmentally and economically friendly, is considered to be the most promising approach in remediating dye-containing wastewater [19,20]. Due to the strong electron-withdrawing nature of azo bonds, they are easily be cleaved off under an oxygen-depleted condition. As a result, the decolorization is concordant with the formation of aromatic amines, which are then
https://doi.org/10.1016/j.jwpe.2020.101268
Received 11 November 2019; Received in revised form 17 March 2020; Accepted 22 March 2020
⁎Corresponding author.
E-mail addresses:[email protected](C. Ong),[email protected](K. Lee),[email protected](Y. Chang).
2214-7144/ © 2020 Elsevier Ltd. All rights reserved.
T
being degraded in the aerobic condition [21]. Therefore, the striking arrangement of sequential anaerobic-aerobic/ anoxic-aerobic in com- bating dye-containing wastewater was proposed throughout the years [22–25].
Owing to the ease of operation and the switchable operating layout, the sequencing batch reactor (SBR) has been often reported as a better choice in treating dye-containing wastewater [26,27]. Although the SBR seeded with activated sludge (AS) system could shed light on the improvement of treatment efficiency, the drawbacks of poor settling characteristics, large footprint requirement, limitation in scale-up and maintenance, unstable performance, sensitivity to toxic environment and autoxidation, have hindered the exploitation of this bioremediation approach [28,29]. To circumvent these drawbacks, the incorporation of moving bed biofilm reactor (MBBR) technology into the existing SBR system is paramount importance and imperative.
Moving bed biofilm reactor (MBBR) technology relies decisively on the biomass that attached to specially designed biocarriers to treat a wide variety of wastewater [30–32]. The biocarriers can swim homo- genously inside the reactor by the turbulent energy imparted from the mixer (anoxic condition) and aerator (aerobic condition). The ad- justable filling ratio also safeguards the high biomass concentration and solid retention in the treatment system [33]. Owing to the presence of an extracellular polymeric substance (EPS) on the biofilm, MBBR technology often reflects a stable performance since it is less sensitive to the toxic environment. Several studies have reported that the presence of EPS can reduce the diffusion speed of toxic substances into the deeper biofilm layer due to the increase in mass transfer resistance [34,35]. In the light of biofilm stratification, the deeper zone of the biofilm possesses an anoxic and anaerobic environment simultaneously despite the aerobic condition is executed [30]. Accordingly, the oc- currence of recolorization and autoxidation in the aerobic treatment can be mitigated, leading to the complete or partial decomposition of aerobically refractory products [36]. As such, the features of MBBR technology can be brought to the existing SBR system, offsetting the inadequacy of the SBR system while treating with dye-containing wastewater [37,38].
This study aimed to evaluate the applicability of the sequencing batch moving bed biofilm reactor (SBMBBR) system in treating dye- containing wastewater. Thus, synthetic wastewater contained a typi- cally mono-azo dye, i.e., Reactive Orange 16 (RO16) was treated in a lab-scale integrated anoxic-aerobic REACT SBMBBR. The main objec- tives were as follows: (i) to evaluate the performance of SBMBBR per- taining to COD removal and biodecolorization under three experi- mental design schemes and (ii) to investigate the effects of dye
concentration under stepwise and shock loads approaches i.e.,Schemes 1and2, respectively, as well as the hydraulic retention time coupled with the increase of biocarrier filling ratio,Scheme 3. To the best of our knowledge, this was the first time that MBBR technology was in- corporated into an integrated anoxic-aerobic REACT SBR system to treat dye-containing wastewater.
2. Materials and methods 2.1. Chemicals
Monoazo dye C.I. Reactive Orange 16 with λmax of 494 nm was procured from Sigma-Aldrich (USA) and the COD contributed by 1 g of RO16 per liter was estimated as 788.7 mg L-1. Unless otherwise speci- fied, all chemicals used in this study were of analytical grade and used as received.
2.2. Synthetic dye-containing wastewater
An aqueous stock solution of RO16 was prepared by dissolving an appropriate amount of dye powder into reagent grade water. During the anoxic-aerobic REACT acclimatization step, the synthetic wastewater consisted of organic carbon, organic nitrogen and trace elements solu- tion. The composition of the synthetic wastewater was as follows (concentration in mg L-1): sucrose (563), bactopeptone (188), ammo- nium chloride (172), magnesium sulfate (49), dipotassium hydrogen phosphate (250), iron(III) chloride (11.3) and sodium bicarbonate (14.7). After the acclimatization stage, the bactopeptone was removed but the trace elements solution remained in the SBMBBR. The carbon source, i.e., sucrose, was fixed to 500 mg L−1in this study.
2.3. Setup of SBMBBR
A schematic diagram of the laboratory-scale SBMBBR setup is given inFig. 1. The bioreactor was a 5 L borosilicate glass-made Schott bottle (DURAN GLS 80 Wide Neck Glass Bottle, Germany). The 3.5 L working volume was occupied by 1 L of activated sludge and 2.5 L of synthetic wastewater with a volume exchange ratio (VER) of 71.43 %. For the SBMBBR system, the reactor was subjected to 5 % and 10 % (v/v) biocarrier (K1) based on the process design schemes. The reactor con- tained two reaction systems, i.e., anoxic and aerobic systems. For the anoxic system, a magnetic stirrer (Heidolph MR Hei-Tec, Germany) equipped with a triangular shape of the magnetic stirrer bar provides turbulent energy for the movement of biocarrier and suspension of Fig. 1.Experimental setup of sequencing batch moving bed biofilm reactor (SBMBBR).
sludge. On the other hand, two cylindrical air stones with a base dia- meter of 15 mm and a height of 35 mm, which were connected to an aquarium air pump (SOBO-SB348, China), representing the aerobic system for oxygen supply. The influent and effluent of the SBMBBR system were pulse fed and pulse discharged via two independent peri- staltic pumps (CP/P-07557-10 Drive Masterflex L/S, Cole-Parmer, USA).
2.4. Inoculation and operation of SBMBBR system
Prior to the experimental start-up, the SBMBBR was inoculated with activated sludge (mixed culture) that collected from a municipal was- tewater treatment plant owned by Indah Water Sdn Bhd in Seksyen 7, 40000 Shah Alam, Selangor, Malaysia. Two diverse SBMBBR elapsed time schedules, including Schedules A and B with their relative hy- draulic retention times (HRTs), i.e., 33.6 h and 16.8 h, respectively, were applied in the evaluation as shown in Fig. S1 (Supplementary Material). After two weeks of anoxic-aerobic REACT acclimatization, the COD removal efficiency showed oscillated within a range of 97 % to 98 % and the biocarrier was confirmed covering with biofilm.
Consequently, the dye was added to the influent based on experimental design schemes with different operating conditions. The overall SBMBBR operating conditions and experimental design schemes were tabulated in Table 1. The SBMBBR was operated under an air-condi- tional room with the temperature maintained within 24℃to 26℃for a total operation period of approximately five months.
2.5. Sample collection and analytical methods 2.5.1. Abiotic decolorization
The activated sludge was sterilized at 121℃for 20 min under a pressure of 0.0103 MPa in an autoclave (HIRAYAMA HV-85, Japan).
The autoclaved sludge was added to 250 mL of 100 mg L-1RO16-con- taining synthetic wastewater. The mixture was stirred on a hotplate (Heidolph MR Hei-Tec, Germany) for 22 h (combines the periods of FILL and REACT) with a constant mixing speed of 100 rpm under room temperature. The UV-vis absorption was used as the measurement of abiotic decolorization herein.
2.5.2. SBMBBR treatment performance
The treatment performance for the SBMBBR was evaluated in terms of effluent COD concentration, effluent dye concentration, sludge con- centration and sludge volume index. The samples were collected at the end of both anoxic and aerobic REACT periods. COD assessment was performed according to APHA 5220-C Closed Reflux, Titrimetric Method [39]. The remaining dye concentration from the effluents was analyzed through a double beam UV-vis spectrophotometer (Cary 100, Agilent Technologies, USA). The value was calculated based on the
calibration plots obtained from the integrated area against a series of standard solutions.
The suspended biomass concentration in SBMBBR was measured as mixed liquor suspended solids (MLSS) and mixed liquor volatile sus- pended solids (MLVSS) according to the standard method of APHA 2540-D Total Suspended Solids and Volatile Suspended Solids APHA 2540-E, respectively [39]. The measurement of the attached-growth biomass followed the procedure described by Hosseini Koupaie et al.
[36].
Sludge volume index (SVI) was used to monitor the settle-ability of suspended growth biomass, and it was conducted according to the procedure stated at APHA 2710-D [39].
2.5.3. Time–course profile study
Time–course profile of biodecolorization and COD removal effi- ciencies from phases 2, 4, 6, 7, 8, 9 and 10 under SBMBBR were monitored during the last week of evaluation periods. For this, the hourly samples were collected during REACT mode with a time interval of every 1 h throughout the REACT mode as well as the end of FILL mode. Subsequently, the collected samples were subjected to the ana- lysis of COD and RO16 concentration, as mentioned in Section2.5.2.
2.5.4. Maximum co-substrate uptake
The percentage of co-substrate, i.e., sucrose uptake with respect to its COD value in each phase, was calculated based on the data obtained from the time-course profile, including the efficiencies of biodecolor- ization and COD removal. The one with the highest percentage value throughout 21 h of REACT period at time t was selected as the max- imum co-substrate uptake for that particular phase. Eq.(1)was used to calculate the percentage of sucrose uptake for the biomass in each phase:
= ×
Maximum co substrate uptake COD COD COD
(%) T CODRO t L t 100 %
S
16, ,
,0
(1) where,CODS,0is initial sucrose COD concentration,CODT is the total COD concentration from the influent RO16-containing wastewater comprising the CODs of sucrose and RO16 molecules,CODRO16,t is the removed RO16 molecules at time t calculated based on biodecoloriza- tion at time t and RO16-COD converting value of 0.7887,CODL t, is the leftover COD after the treatment at time t.
2.5.5. The microbial strength in SBMBBR
The microbial strengths in removing COD and RO16 molecules in each phase were interpreted by how much quantity of COD (mg) and RO16 molecules (mg) were removed from the SBMBBR system. The obtained values were then divided by the total biomass (MLVSS) in that particular phase. They can be expressed by Eqs.(2)and(3).
Table 1
The overall SBMBBR operating conditions and experimental design schemes.
Phase Dye Concentration (mg L-1) Biocarriers Filling Ratio
(%) Co-substrate Concentration
(mg L-1) SBMBBR Cycle Elapsed Time
(Schedule) Period involved
(week) Experimental Design
1 10 to 50 5.0 500 A 1 Scheme 1a
2 50 2
3 50 to 100 1
4 100 2
5 100 to 150 1
6 150 2
7 300 2 Scheme 2b
8 600 2
9 1000 2
10 1000 10.0 B 4 Scheme 3c
a The effect of RO16 concentration in a stepwise approach.
b The effect of RO16 concentration in a shock load approach.
c The effect of hydraulic retention time coupled with the increase of biocarrier filling ratio.
=
Microbial strength COD Removed COD mg Total Biomass mg
( ) ( )
( ) (2)
=
Microbial strength RO Removed RO mg Total Biomass mg
( 16) 16 ( )
( ) (3)
Where, the removed COD and removed RO16 were acquired from the average results of the last three days during the last evaluation week of each phase.
2.5.6. The percentage removals of COD and biodecolorization rates The percentage removals of COD and biodecolorization rates throughout the anoxic-REACT period in phases 9 and 10 were calcu- lated by Eq.(4):
=
COD or Biodecolorization rate hour
The percentage removal The percentage removal Anoxic REACT period
(% )
final initial
1
(4) Where,the percentage removalfinalandthe percentage removalinitialis the last and the first hour treatment performance of anoxic-REACT period from the profile studies, respectively. The anoxic-REACT period herein were 16.5 h for phase 9 and 7.5 h for phase 10.
2.5.7. Extracellular polymeric substance (EPS)
The presence of EPS on the biocarriers was confirmed using a Fourier transform-infrared (FTIR) spectrometer (Nicolet iS10, Thermo Fisher Scientific, USA) equipped with an attenuated total reflectance (ATR) sampling accessory. The spectra were collected at the resolution of 4 cm−1and with average scans of 64 from 4000 to 500 cm−1. Prior to this testing, the refrigerant dehumidifier was turned on so as to prevent the excess absorption of moisture from the environment to- wards the dried biofilm.
3. Results and discussion
The experiment pertaining to three discrete schemes including the effects of dye concentration under stepwise and shock loads approaches i.e.,Schemes 1and2, respectively, as well as the hydraulic retention time coupled with the increase of biocarrier filling ratio,Scheme 3was carried out over a period of 132 days (from phase 1–10) but only the details of phases 2, 4, 6, 7, 8, 9 and 10 were further taken into dis- cussion herein since phases 1, 3 and 5 were the transient periods under Scheme 1. Table 2summarises the overall treatment performance re- garding COD removal and RO16 biodecolorization of an integrated anoxic-aerobic REACT SBMBBR, while a time-series graph regarding the end-treatment performance of COD removal and RO16 biodeco- lorization were depicted inFig. 2. The microbial strengths in removing
RO16 and COD, as well as the maximum co-substrate, i.e., sucrose uptake (%) for each phase, were also documented inTable 3. Lastly, the data obtained fromTables 2 and 3were interpreted in conjunction with the time series graphs derived fromFig. 2viz.,Fig. 3(COD removal) andFig. 4(RO16 biodecolorization), as shown in Sections3.2and3.3, respectively, laying down the foundation for the following discussion.
3.1. Mechanism of biodecolorization in SBMBBR
The mechanism of biodecolorization in the SBMBBR system was functioned on accounts of two subsets viz., abiotic adsorption (bio- sorption) and biotic biochemical reaction (biodegradation). In bio- sorption, a very low RO16 biodecolorization efficiency with up to 14 % was attained. The results revealed a truth that biosorption was not the crucial key in mitigating the RO16 concentration herein. In addition, UV-vis spectral scan as shown in Fig. S2 affirmed that the biodecolor- ization in this study was chiefly ascribed to biodegradation, whereby the characteristic peaks of RO16 at the wavelength of 388 nm and 494 nm ( max)were dwindled continuously along with the reaction time.
This observation was concomitant with the pronounced increasing peak at 264 nm, indicating the breakdown of azo bonds was mainly due to enzymatic reaction.
3.2. COD removal
3.2.1. Scheme 1: the effect of dye concentration (stepwise approach) Fig. 3a shows the COD removal efficiency fluctuated within a range of 94 % to 100 % over an evaluation period of 62 days. A dramatic decrease in COD removal efficiency could be seen from phases 1, 3, 5 and the early part (first week) of phases 2, 4 and 6. This observation implied that the SBMBBR system was susceptible to the gradual in- crease of RO16 loadings, which was 10 mg L-1day-1throughout phases 1, 3 and 5. Although the falling trend was prolonged to the early of the phases 2, 4 and 6, it was taken off sharply in the following week. This revealed that the non-acclimatized activated sludge required much time for the microorganisms to recognize RO16 molecules since it is a xe- nobiotic compound [40]. Once the acclimatization of certain RO16 concentrations was accomplished, the COD removal efficiency showed greater than that of transition periods.
Table 2unveils that more than 97 % of COD removal efficiency was attained in phases 2, 4 and 6. The reduction of COD was equivalent to a concentration removal (mg L-1) of 644.18, 668.62 and 709.16 at phases 2, 4 and 6, respectively, as reported inTable 3. Three plausible ex- planations for such excellent COD abatement could be recapitulated herein: (i) the fraction of CODRO16to CODTotalwere tiny underScheme 1, the stepwise approach provided an adequate time frame for the Table 2
The overall treatment performance in SBMBBR system.
Phase RO16 Concentration
(mg L-1) RO16 COD
(mg L-1) COD feed (Base mix + RO16)
(mg L-1) CODRO16/ CODTotal
(%) Reaction Phase Biodecolorizationa(%) COD Removala(%)
2 50 39.44 649.44 6 Anoxic 100.00 98.54
Aerobic 100.00 99.19
4 100 78.87 688.87 11 Anoxic 97.11 98.18
Aerobic 100.00 98.45
6 150 118.31 728.31 16 Anoxic 97.40 97.06
Aerobic 100.00 97.37
7 300 236.61 846.61 28 Anoxic 88.68 47.01
Aerobic 89.44 50.54
8 600 473.22 1083.22 44 Anoxic 57.29 45.78
Aerobic 58.04 48.97
9 1000 788.70 1398.7 56 Anoxic 36.89 39.25
Aerobic 36.33 38.69
10 1000 788.70 1398.7 56 Anoxic 25.60 36.98
Aerobic 26.82 36.27
a The performance was assessed by averaging the results obtained from the last three days during the last evaluation week of each phase.
acclimatization of the SBMBBR system [41]. This was evidenced by the improvement of the microbial strength in removing COD from 0.0677 to 0.0711 mg COD/mg biomass from phase 2 to phase 6 as
demonstrated inTable 3; (ii) the involvement of 500 mg L-1sucrose in the biodegradation, providing the biomass with an accessible substrate for metabolism purpose.Table 3discloses a complete consumption of Fig. 2.Time-series end-treatment performance of SBMBBR system.
Table 3
The microbial strengths in removing RO16 and COD under the SBMBBR system.
Phase Removed CODa
(mg L-1) Removed RO16a
(mg L-1) Removed CODa
(mg) Removed RO16a
(mg) Total Biomass
Weightb(mg) COD (mg)/
Biomass (mg) RO16 (mg)/
Biomass (mg) Maximum Co-substrate Uptake (%)
2 644.18 50 1610.45 125 23,790 0.0677 0.0053 100
4 678.19 100 1695.48 250 21,121 0.0803 0.0118 100
6 709.15 150 1772.88 375 24,923 0.0711 0.0150 100
7 427.88 268.32 1069.70 670.8 25,110 0.0426 0.0267 80
8 530.45 348.24 1326.13 870.6 29,219 0.0454 0.0298 86
9 541.16 363.30 1352.9 908.3 29,906 0.0452 0.0304 69
10 507.31 268.20 1268.28 670.5 35,190 0.0360 0.0191 64
a Both removed COD and RO16 in the expression of concentration and weight based on the aerobic results obtained fromTable 2.
b The biomass weight based on MLVSS.
Fig. 3.Performance of COD removal efficiency in a treatment cycle under SBMBBR system: (a)Scheme 1; (b)Scheme 2and (c)Scheme 3.
Fig. 4.Performance of RO16 biodecolorization in a treatment of cycle under SBMBBR system: (a)Scheme 1; (b)Scheme 2and (c)Scheme 3.
co-substrate under this scheme. This was interrelated to the rise in the aforesaid microbial strength (mg COD/mg biomass). Solís et al. [42]
emphasized that the involvement of co-substrate not only helps in maintaining the vitality of microorganisms but also assisting in biode- colorization. The majority of COD reduction occurred during the FILL period, as shown inFig. 5a. This was ascribed to the reason that sucrose was rapidly reduced in the Fill period by serving as an electron donor and carbon source [43]; (iii) the aromatic intermediate formed from the anoxic-REACT period was degraded in the aerobic-REACT period as illustrated in Fig. 6a. The degradation of an aromatic intermediate during the aerobic reaction period was congruent with the observations from Jonstrup et al. [44] and Ong et al. [45].
3.2.2. Scheme 2: the effect of dye concentration (shock loads approach) Fig. 3b portrays that the COD removal efficiency dropped sharply after the first, second, and third RO16 shocks, i.e., 300 mg L-1, 600 mg L-1and 1000 mg L-1with the CODRO16/ CODTotalratio of 28 %, 44 % and 56 %, respectively, were imposed into SBMBBR system. A sig- nificant decline in the COD removal efficiency was noticed during the early part (first week) of phases 7, 8 and 9. The falling trends could be ascribed to the substrate inhibition exerted by the toxicity of RO16
molecules under the sudden RO16 shocks due to the immaturity of enzymatic system [46–48]. Fortunately, the COD removal efficiency improved progressively and attained an average value of 50.54 %, 48.97 % and 38.69 % in the following week (second week) of phases 7, 8 and 9, respectively, as shown inTable 2. Therefore, the instability on the COD removal efficiency signified that there was a need for an adaptation period so as to the SBMBBR system could contend with the RO16 shocks [49].
In comparison to phase 6, a difference of 46.83 % in COD removal efficiency was heeded in the first RO16 shock. This was equivalent to a disparity of 281.27 mg L-1COD removal concentration, as shown in Table 3.Table 3also reveals that the microbial strength (mg COD/mg biomass) was declined from 0.0711 to 0.0426. Since there was no de- terioration on the RO16 biodecolorization efficiency but amelioration of 10 % was observed as reported in Section3.3.2, the decline in the microbial strength (mg COD/mg biomass) did not pose any negative impact to the SBMBBR system. This was supported by the increment of biomass concentration in phase 7 as compared to phase 6 as discussed in Section3.5. Similarly,Fig. 5b displays that the primary COD removal with more than 60 % was attained at the end of FILL period. Subse- quently, the anoxic-REACT period only attributed approximately 10 % Fig. 5.Time-course profiles of COD removal efficiency during the REACT period of a SBMBBR treatment cycle: (a)Scheme 1; (b)Scheme 2and (c)Scheme 3.
Fig. 6.Time-course profiles for the formation of aromatic intermediate during the REACT period of a SBMBBR treatment cycle: (a)Scheme 1; (b) & (c)Scheme 2and (d)Scheme 3.
of COD removal efficiency. Thereafter, it showed decreasing during the aerobic-REACT period. Furthermore, the aromatic intermediate re- sulted from the biodegradation of RO16 was further degraded in the aerobic-REACT period, as illustrated in Fig. 6b. The observation in phase 7 was likely associated with natural selection. Indeed, the mi- crobial consortia which were unable to survive under 300 mg L−1RO16 containing wastewater were flushed out, while those which were able to produce or secrete the effective working enzymes were screened out and proliferated in the SBMBBR system [50,51]. Consequently, the ef- fective working enzymes could be secreted adequately thereby utilizing RO16 molecules as the primary carbon and energy substrate. This was also supported by an extent of 80 % co-substrate uptake as compared to 100 % in the previous phases, as documented inTable 3. The loser of natural selection underwent disintegration on their structural config- uration, causing the occurrence of cell lysis in the system. The con- sequence of cell lysis could result in the releasing of some macro- molecules, including proteins, DNA, RNA and so on. Those macromolecules possess a backbone of carbohydrate, explaining why there was a built-up of COD herein during the aerobic-REACT period [52].
In spite of the COD removal efficiency dropped to 39.26 % during the first week of phase 8 as illustrated inFig. 3b, the corresponding COD removal concentration, i.e., 425.27 mg L-1was still better than that of 368.11 mg L-1, as obtained from the last day of phase 7. This improvement was stemmed from the fact that the effective microbial consortia have standing out and proliferated due to the natural selec- tion previously. Therefore, the environment containing a high amount of RO16 molecules could favor the growth of the biomass since they exhibited a positive capability in degrading RO16 molecules. However, the enzymatic system in the SBMBBR was still suppressed by the toxi- city from the second RO16 shocks. As a result, the disturbance in mi- crobial syntrophy was speculated, whereby the enzymes for breaking down the co-substrate was induced again [49]. It was evidenced by the increase in the maximum co-substrate uptake from 80 % in phase 7–86
% in phase 8, as shown inTable 3. This remark was also in agreement with the rise in the microbial strength from 0.0426 to 0.0454 mg COD/
mg biomass, as documented in Table 3. Consequently, the effective working enzymes could work well in the SBMBBR system along with the adequate electrons donated from the breakdown of co-substrate, mitigating the toxicity exerted by the RO16 molecules. Eventually, Table 3reveals that the COD removal concentration (mg L-1) was im- proved to 530.45 from 427.88 in phase 7. Likewise,Fig. 5b illustrates that the principal COD reduction occurred in the FILL period. A trifling variation on COD removal efficiency was observed during the anoxic- REACT period. The concentration of resulted aromatic intermediate was abated under the aerobic-REACT period, as evidenced fromFig. 6c.
However, the ability of SBMBBR in degrading the aromatic inter- mediate started to become slower. Pandey et al. [53] and Hosseini Koupaie et al. [36] reported that those aromatic intermediates were more toxic than the parent dyes. Therefore, it could adversely affect the mineralization of aromatic intermediate.
In the last RO16 shock, the SBMBBR system was capable of re- moving a COD concentration of 541.2 mg L-1in phase 9 as compared to 530.45 mg L-1in phase 8 despite the COD removal efficiency of the latter was higher, i.e., 48.97 %. A transformation on the enzymatic system was believed to occur after the acclimatization of 1000 mg L-1 RO16 shock for one week. Indeed, more effective microbial consortia in removing RO16 molecules were formed in phase 9 since the SBMBBR system was able to remove more RO16 molecules, as evidenced in Section 3.3.2, without a high consumption on co-substrate.Table 3 unveils that the maximum co-substrate uptake was only 69 %. In ad- dition, the microbial strength showed almost akin to the value reported in phase 8, i.e., 0.0452 mg COD/mg biomass. Solís et al. [42] stated that microorganisms could acclimatize under extreme conditions, i.e., the presence of RO16 molecules herein by inducing more powerful enzy- matic activity. The authors also added that if the microorganisms have
been acclimatized successively to dye molecules, the cut-down on co- substrate uptake could be noticed.Fig. 5b highlights that only 30 % of COD removal efficiency was attained at the end of the FILL period.
Subsequently, the integrated anoxic-aerobic REACT did not contribute much to the COD reduction. Only approximately 10 % of COD removal efficiency was attained throughout 21 h of reaction time. No im- provement in the performance of COD removal but deterioration was observed under the aerobic-REACT period, as depicted inFig. 6c. There were two plausible reasons which could explain this observation: (i) it was speculated that the absorbed aromatic intermediates in the mi- crobial cell were flushed out by a special efflux pump. Indeed, several studies have reported that aromatic intermediates are more toxic than parent dyes [54,55]. To reduce the concentration of a toxic substance in the cellular membrane, a survival mechanism via the efflux pump was triggered [49]; (ii) it was due to the occurrence of autoxidation. In fact, there is a propensity for the resulted aromatic intermediated to be auto- oxidized under the presence of oxygen [56]. Consequently, this will lead to a formation of more complex and aerobically recalcitrant pro- ducts, leading to the build-up of COD [57]. Da Silva et al. [58] and Mata et al. [26] mentioned that the auto-oxidized aromatic intermediate would form a colored oligomeric structure under an oxygen-rich en- vironment via free radical reactions. The findings in Section3.3.2have supported the statement.
In fact, the actively metabolic system comprising a series of natural selection, disturbance in syntrophy as well as the transformation of the enzymatic system under three RO16 shocks was safeguarded by the attached-growth biomass system in SBMBBR. In addition, the attached- growth biomass possessed a critical role in offsetting the incoming RO16 containing wastewater, especially during the FILL period, since it constituted 12 % to 24 % of the total biomass in this study. Indeed, the pulse-feed of the incoming RO16 containing wastewater was first re- ceived and degraded by the biofilm on the biocarrier since suspended- growth biomass was settled below the biocarriers. Accordingly, it helped to alleviate the toxicity exerted from the shock RO16 molecules, favoring biodegradation on the suspended-growth biomass system.
Bassin et al. [59] reported that suspended-growth biomass is sometimes susceptible to high organic loading. In contrast, attached-growth bio- mass system contained a dense protective layer, i.e., extracellular polymeric substance (EPS), which has been reported to be capable of resisting inhibitory or noxious effects from the toxic compounds such as RO16 molecules herein [35,60]. Venkata Mohan et al. [61] pointed out that the attached-growth biomass system could act as a buffer to reduce the concentration of toxic and recalcitrant chemicals, thereby alle- viating the stress suffered by the suspended-growth biomass system. As a result, the attached-growth biomass system did shed some light on RO16 biodegradation especially underScheme 2.
3.2.3. Scheme 3: the effect of hydraulic retention time coupled with the increase of biocarrier filling ratio
Fig. 3c displays a slight exacerbation on the COD removal efficiency after the HRT was reduced to 16.8 h, although the biocarrier filling ratio was increased to 10 % (v/v). However, it was believed that the rising trend observed in the last week of phase 10 was attributed to the biomass augmentation from the extra 5 % biocarrier. The growth of biofilm required much time, especially in a short HRT system [62].
As shown inTable 2, the COD removal efficiency became worse and showed 36.27 % as compared to 38.69 % in phase 9. Several studies have been reported that a low COD removal efficiency was always ac- companied by a reduction in HRT [44,63,64]. This could be due to the inadequate time for the microorganisms to degrade more organic sub- stances such as both RO16 molecules and co-substrate herein [43].
Subsequently, the microbial strength in removing COD in terms of mg COD/mg biomass in phase 10, dropped to 0.036 as compared to 0.0452 at phase 9 as tabulated inTable 3. In the aspect of time-course profile (especially during the anoxic-REACT period), the COD removal rate (%
hour−1) in phase 10 was higher, i.e., 0.565 than that of 0.408 from
phase 9 as shown inFig. 5c. This was equivalent to a difference of 0.16
% hour-1. A remarkable inefficiency on the COD removal was noticed during the aerobic-REACT period, suggesting the occurrence of cell lysis was reencountered in phase 10. Owing to the reduction in the anoxic- REACT period, most of the microorganisms in the SBMBBR system were speculated could not acclimatize to the shorter HRT, leading to the disintegration of the microbial structure. This explained why there was a built-up of COD.Fig. 6d shows that the aromatic intermediate formed from the breakdown of azo linkages during the anoxic-REACT period was slightly removed in the subsequent aerobic-REACT period, in- dicating partial mineralization was achieved herein.
3.3. RO16 biodecolorization
3.3.1. Scheme 1: the effect of dye concentration (stepwise approach) The performance of RO16 biodecolorization was irresistibly ex- cellent, i.e., 100 % removal efficiency was achieved over 62-days stepwise RO16 loading regardless of any initial concentration values, as shown inFig. 4a andTable 2. Yu et al. [65] mentioned that a start-up period of 3–7 months is necessary for the acclimatization to decolorize azo-containing wastewater completely. Instead, this study showed that a complete biodecolorization with RO16 concentration of 150 mg L-1 was achieved in a shorter acclimatization period, i.e., 50 days. With the biofilm attachment on the biocarrier, a high metabolic potential could be induced [61]. In turn, it shortened the acclimatization period. The complete removal of RO16 molecules was in accordance with the im- provement of the microbial strength in removing RO16 molecules from 0.005 to 0.015 mg RO16/mg biomass, as shown in Table 3. The ex- cellent biodecolorization herein had outcompeted those reported from the current bioreactor configurations [44,66,67]
The outstanding biodecolorization herein could be ascribed to: (i) the mixed culture involved, (ii) the presence of co-substrate, (iii) the experimental design and (iv) the RO16 molecular structure. In com- parison with pure culture, a mixed culture is capable of achieving a higher degree of biodegradation and mineralization as it involves a wide spectrum of microbial consortia. While exposing to a new en- vironment, synergistic metabolic activities speed up the communication between each microbial consortium [68]. The presence of co-substrate, i.e., sucrose, was essential for the reductive cleavage of azo linkages on the RO16 molecules with the assistance of certain enzymes. Indeed, the cleavage of azo linkages is associated with the transfer of four electrons (reducing equivalents), in which two pairs of electrons from the me- tabolism of co-substrate (electron donor) are transferred to azo dye molecule (electron acceptor) each stage, resulting in the breakdown of azo linkages and hence colorless aromatic intermediates were formed [69]. Karim et al. [8] stated the involvement of co-substrate in micro- bial degradation of dyes not only boosts up the ability of dye de- gradation but also shortens the reaction time. Besides, the stepwise increment of RO16 concentration also played a pivotal role in giving excellent biodecolorization. Singh et al. [70] and Dafale et al. [71]
declared that an adequate adaption is needed for the microbial com- munity under the exposure of toxic or recalcitrant compounds. It could provide a sufficient transient period for the improvement of the bio- decolorization rate. Besides, the decolorization efficiency is directly proportional to the number of azo bonds present in the dye molecule [72,73]. A mono azo dye, i.e., RO16, which consists of only one azo bond on its molecular structure, therefore, exhibited high biodegrad- ability. With the presence of sulfonated group (atpara-position of an azo bond) as well as hydroxyl group (at the 2-position of its naphthol ring), the efficiency of RO16 biodecolorization was further reinforced to a higher degree [24].
The time-course profile from Fig. 7a unveiled that a RO16 con- centration of 50 mg L−1could be fully decolorized within an anoxic- REACT period of 7 h in phase 2, while the time required to attain a complete biodecolorization of RO16 was further prolonged to aerobic- REACT period viz., 19th and 18th hour in phases 4 and 6, respectively.
A noticeable extent of RO16 biodecolorization with approximate 10 % occurred under the aerobic phase in both phases 4 and 6. The anoxic and anaerobic zone within the inner biofilm stratification under the aerobic-REACT period could be a putative reason in contributing the biodecolorization herein [67,74].
3.3.2. Scheme 2: the effect of dye concentration (shock loads approach) A complete biodecolorization was undermined after successive RO16 shocks viz., 300 mg L-1in phase 7, 600 mg L-1in phase 8 and 1000 mg L-1in phase 10 were imposed on the SBMBBR system as shown inFig. 4b. The efficiency of RO16 biodecolorization shown inferior to that of reflected from the previous RO16 shock along with the eva- luation ofScheme2. In the first RO16 shock, there was a marginal de- crease in RO16 biodecolorization efficiency, dropping from 100 % (phase 6) to 89.44 % (phase 7) with a difference of approximately 10 % as reported inTable 2. Subsequently, the performance of RO16 biode- colorization was severely affected by the second and last RO16 shocks, resulting in an average biodecolorization efficiency of 58.04 % and 36.33 %, respectively, as documented inTable 2. It was evident that the descending trend in the first evaluation week of phases 7, 8 and 9 was replaced by an ascending trend in the following week, as shown in Fig. 4b. The decline in the first evaluation week was likely caused by the substrate inhibition exerted by the substantial amount of RO16 molecules, which was suddenly introduced into the SBMBBR system without adequate acclimatization (El Bouraie and El Din., 2016; [48]).
Eventually, the RO16 biodecolorization efficiency in phases 7, 8 and 9 corresponded to a RO16 concentration removal of 268.32, 348.24 and 363.30 mg L-1, respectively.
In the first RO16 shock, the microbial strength in removing RO16 molecules was ameliorated to 0.0261 as compared to 0.015 mg RO16/
mg biomass in phase 6, as unveiled inTable 3. Accordingly, an ap- proximate 670.8 mg of RO16 molecules were able to be decolorized in phase 7. Moreover, the maximum co-substrate uptake was diminished to 80 % as compared to 100 % in the previous phases. The findings affirmed that the SBMBBR system was accredited to possess the positive ability in withstanding the first RO16 concentration shock of 300 mg L-
1. Indeed, the prevailing enzymatic system demonstrated that the ef- fective working enzymes in confronting the RO16 molecules had been produced or secreted effectively as a result of the natural selection and proliferation of microbial consortia throughout the acclimatization as described in Section3.2.2. It was evidenced by the time course profile as shown inFig. 7b, in which the SBMBBR system was capable of re- moving half of the inlet RO16 concentration, i.e., approximately 55 % at the end of the FILL period, while it was only 45 % at the end of FILL period under phase 6. While comparing the biodecolorization between anoxic and aerobic periods, not much difference was observed. This finding was in agreement with previous works, whereby no much change on the biodecolorization efficiency in the aerobic period [43,75].
The instability of performance in phase 8 heralded the possible disturbance in syntrophy in the SBMBBR system [49]. There were two groups of microbial consortia living in the system viz., RO16-degrading and co-substrate-degrading microbial consortia. Owing to the inhibition effect exerted from the 600 mg L−1RO16 shock, the microbial con- sortia in degrading the RO16 molecules required many electrons from the electron donor, i.e., sucrose, to break down the azo linkage pos- sessed by RO16 molecules. Therefore, the enzymes for breaking down the co-substrate were induced again herein, as described in Section 3.2.2. In comparison to phase 7, the microbial strength was improved to 0.0298 from 0.0267 mg RO16/mg biomass, which indicated an extra 200 mg of RO16 molecules was able to be decolorized as shown in Table 3. In the aspect of time-course profile, the anoxic and aerobic REACT periods showed an approximate 40 % in biodecolorization, while only around 20 % biodecolorization efficiency was achieved during the Fill period as illustrated inFig. 7b. Again, biodecolorization efficiency between anoxic and aerobic periods shown not much
improvement.
In the course of last RO16 shock, the microbial strength in deco- lorizing RO16 molecules was improved slightly to 0.0304 from 0.0298 mg RO16/mg biomass as compared to phase 8, as shown inTable 3.
This improvement was concomitant with an extra biodecolorization of 10.71 mg RO16 molecules. Moreover, the maximum co-substrate up- take was reduced to 69 % in phase 9. This was inferred that a new enzymatic system had been evolved, i.e., co-substrate independent en- zymatic system during phase 9 as mentioned in Section3.2.2.Fig. 7b illustrates that biodecolorization mainly happened in the FILL and an- oxic-REACT periods. A noticeable decline was observed during the aerobic-REACT period, suggesting a thorny problem, i.e., autoxidation was suspected to be encountered in this phase. Jonstrup et al. [44]
stated that recolorization is a rapid process and it has resulted from the autoxidation. They added that the products spontaneously cross-react with each other, forming polymeric structures which are more re- calcitrant in undergoing degradation aerobically. The decrement on biodecolorization under the aerobic-REACT period could also be ex- plained by the consequence of the efflux pump [49]. These observations verified the occurrence of autoxidation and the surviving mechanism of the efflux pump as described in Section 3.2.2.3.
The RO16 shocks were not found to be exacerbated to the enzymatic system in the purposed SBMBBR system. Henceforth, it could be con- cluded that the purposed SBMBBR system was capable of contending high RO16 shocks. The attached-growth biomass system was believed to be paramount importance and eligible in remediating the refractory pollutants and organic shock loads [52,76,77]. Furthermore, the bio- film attachment on the biocarrier was able to transform an organic or toxic pollutant much faster as compared to suspended flocs [78].
Naresh Kumar et al. [79] compared the biodecolorization between biofilm and activated sludge systems under a successive of Acid Black 10B shocks. They discovered that the biodecolorization in biofilm system was better than that of reflected from the conventional activated sludge system. They also pointed out that such the high stability and biodecolorization in biofilm system was due to the high metabolic po- tential of biofilm operation as a result of high biomass retention cap- ability. Thus, the protruding difference between two distinct biomass systems, providing support for the contribution of the attached-growth biomass system herein. Koçyiğit and Ugurlu [80] mentioned that no adverse effect but an improvement was observed in color removal at varying shock loads despite the COD removal efficiency become lower.
Again, this could be attributed to the presence of the attached-growth biomass system in the SBMBBR. The detailed was also elucidated in Section3.2.2.
3.3.3. Scheme 3: the effect of hydraulic retention time coupled with the increase of biocarrier filling ratio
The performance of RO16 reduction showed fluctuating after the HRT was reduced to 16.8 h, as plotted inFig. 4c. The RO16 biodeco- lorization efficiency reflected a declining trend during the last week of the evaluation period, indicating the inability of the SBMBBR system in degrading a total mass of 5000 mg (two cyclic loadings) of RO16
molecules per day.Table 2reports the RO16 biodecolorization effi- ciency dropped to 26.82 %, accounting for 670.5 mg of RO16 molecules was decolorized at the end of a cyclic treatment. A deterioration of approximately 10 % was accounted in phase 10 in comparison to phase 9 despite the inlet RO16 concentration was the same. In short HRT, the biomass systems had insufficient time to expose to the RO16 molecules and therefore, the probability of decomposition decreased [81]. Ac- cordingly, the microbial strength decreased to 0.0191 from 0.0304 mg RO16/mg biomass at phase 9 as shown inTable 3.
Fig. 7c manifests that the RO16 biodecolorization efficiency in phase 10 showed higher than that of phase 9 within the first 10 h as a result of a faster biodecolorization rate. Indeed, an hourly removal percentage (% hour-1) difference of 0.3 was calculated based on the values obtained from phases 9 and 10, i.e., 1.249 and 1.551, respec- tively during the anoxic-REACT period. This could be concluded that the 10 % biocarrier filling ratio did shed light on the RO16 reduction.
With the incorporation of an extra 5 % (v/v) biocarrier filling ratio, the amount of biofilm attachment was augmented. Since the attached- growth biomass was responsible for the RO16 molecules reduction, therefore the treatment efficiency was promoted [59]. However, the reduction of HRT was the main factor causing the deterioration herein.
Al-Amrani et al. [24] studied the effect of time ratio of operating per- iods in a cycle towards decolorization. They concluded that longer anoxic time is preferable for the biodecolorization whereby the deco- lorization of Acid Orange7 could attain 80 % when an extra 10 h was added to the anoxic reaction time. Koçyiğit and Ugurlu [80] mentioned that a longer anaerobic contact time could improve dye removal effi- ciency.
3.4. UV–vis spectral analysis
The UV–vis spectra from the beginning and the end of each scheme were depicted inFig. 8b–f. In comparison to the control RO16 spectrum inFig. 8a, two of the RO16 characteristic peaks in the visible region, i.e., λmax at 492 nm and 388 nm under Scheme 1 diminished sig- nificantly with respect to time, whilst one of the new peak (inter- mediate peak) kept accumulating nearby the RO16 characteristic peak of 250 nm, i.e., benzene ring in the UV region as portrayed inFig. 8b and c. The intermediate peak in the UV region was further cut down during the aerobic treatment. The disappearance of absorbance peaks in visible region signified a complete cleavage of the azo bond was at- tained [82]. Notably, the RO16 biodecolorization showed a certain extent of improvement under the aerobic-REACT period as evidenced byFig. 8b and c. Besides, the structure of the naphthalene ring, i.e., λmaxat 290 nm, was removed entirely inScheme 1.
InScheme 2, the UV spectrum obtained fromFig. 8d also demon- strated a significant decrease in the absorbance peaks under visible region but not complete removal.Fig. 8d also illustrates that the in- termediate peak in the UV region received destruction during the aerobic treatment. The blunt shape in the UV-region ofFig. 8d was because of non-diluted samples were used. However, the decrement of the absorbance peaks inFig. 8e showed insignificantly. Moreover, a Fig. 7.Time-course profiles of RO16 biodecolorization during the REACT period of a SBMBBR treatment cycle: (a)Scheme 1; (b)Scheme 2and (c)Scheme 3.
Fig. 8.RO16 biodegradation UV–vis absorption spectra (200-700 nm) (a) 100 mg L-1RO16 influent; (b) phase 2; (c) phase 6; (d) phase 7, (e) phase 9 and (f) phase 10.
Fig. 9.Biomass distribution and sludge volume index (SVI) in SBMBBR system.
contrasting trend was observed, in which the aerobic spectrum showed a reappearance of naphthalene ring structure at the peak of 290 nm.
Meanwhile, the peak at 492 nm increased to a certain extent higher than that of the anoxic spectrum. The remarkable phenomenon could be due to the occurrence of autoxidation. The observation herein has been confirmed in Sections3.2.2and3.3.2. InScheme 2, the end pro- ducts shown possessing the characteristic peaks of 260 nm. This peak was found associated with the presence of oxidized aromatic com- pounds, in particular, a phenolic compound, i.e., diketones [83]. In Scheme 3, there is no much difference between the anoxic and aerobic spectra, as shown in Fig. 8e. The problem of autoxidation has been mitigated in this phase.
Henceforth, the initial stage of the RO16 biodegradation was deemed to be the breakdown of azo dye chromophore structures, as evidenced by the diminishment of RO16 maximum absorption in visible wavelength region. This was concomitant with the reflection of a lighter orange or colorless appearance. Subsequently, the resulted in- termediate underwent mineralization in the aerobic treatment. This was tallied with the reduction on the intermediate peak while com- paring anoxic and aerobic spectra. However, partial mineralization was attained throughout this study.
3.5. Biomass distribution, weight and sludge volume index (SVI)
Fig. 9demonstrates that the suspended-growth biomass overtopped the attached-growth biomass utterly in the SBMBBR. Indeed, the sus- pended-growth biomass dominated more than 76 % of the total biomass throughout the evaluation. Conversely, Gu et al. [35] revealed that attached-growth biomass was the dominant biomass system in treating coking wastewater under the application of MBBR technology. The less dominance of attached-growth biomass herein could be attributed to:
(i) 1 L of activated sludge was inoculated into SBMBBR; (ii) the sus- pended biomass possesses more porous floc structure than attached biofilm, which might uptake much of the co-substrate and RO16 mo- lecules along with the reducing agents required for the breakdown of azo bonds [84]; (iii) a more effective microbial consortia in degrading RO16 molecules and intermediates are abundant in suspended-growth system; (iv) the stepwise increment approach in Scheme 1 provided suspended-growth biomass an adequate acclimatization in with- standing the severe dye-containing environment [41].
Fig. 9depicts the trend of the total biomass showed falling in the early ofScheme 1and then steadily ascending to approximate 25 g in phases 6 and 7, 29 g in phases 8 and 9 and eventually 35.19 g in phase 10. A similar trend was observed on the suspended biomass system since it accounted for more than 76 % of the total biomass in the SBMBBR. In addition, it decreased to 18.19 g at phase 4 from 21.7 g at acclimatization after the RO16 molecules were introduced into the SBMBBR. However, this descending trend was replaced by an in- creasing trend since phase 6 onwards till phase 10. This was equivalent to an increment of 5.83 g of suspended-growth biomass. On the other hand, the attached-growth biomass showed slightly fluctuated over the evaluation periods. With the incorporation of 5 % (v/v) biocarrier filling ratio, the attached-growth biomass fromSchemes 1and2showed fluctuating within 2.9 g–4.90 g but it was increased to more than 1.8 folds, i.e., 8.61 g in Scheme 3 after the quantity of extra 5 % (v/v) biocarrier were added into SBMBBR. InScheme 1, the loss of biomass in two biomass systems was probably ascribed to substrate inhibition since the microorganisms were not fully acclimatized to 100 mg L-1RO16 molecules under phase 4 [46,48]. In the light of a stepwise feeding strategy adopted inScheme 1, the suspended-growth biomass system displayed a positive increment throughoutScheme2. This could be at- tributed to the high substrate concentration involved in SBMBBR as well as an adapted enzymatic system had been evolved. Indeed, high substrate concentration could enhance the metabolism of microorgan- isms if enzymatic system reflects a positive ability in degrading RO16 molecules and the relative intermediate. In turn, microorganisms could
proliferate faster, resulting in high total biomass concentration.
In this study, suspended-growth biomass reflected a good settle- ability since the SVI values from 3 experimental design schemes were found within a preferable range of 120 mL g-1, as shown in Fig. 9.
Therefore, high biomass retention in the SBMBBR was ensured. This explained why the suspended growth biomass increased significantly under the SBMBBR [85]. The highest SVI value was attained during the acclimatization, i.e., 107.69 mL g-1. However, the SVI values dropped gradually from 105.8–46.39 mL g−1during the evaluation ofScheme 1 andScheme 2. Subsequently, a slight increment on SVI value was no- ticed inScheme 3. Indeed, the excellent settle-ability herein was in- ferred due to the increase of sludge compactness as a result of microbial acclimatization [86]. However, the presence of pinpoint floc was speculated to be involved inScheme 2andScheme 3since the SVI values were less than 75 mL g-1. Consequently, a cloudy appearance was ob- served in the supernatant above the settled blanket [52,87]. The for- mation of pinpoint floc could be caused by (i) the toxicity exerted by RO16 molecules and their intermediates during the course of shock loads; (ii) the floc disintegration as a result of natural selection and starvation of microorganisms; (iii) the reduction in HRT inScheme 3 [88]. Notably, the sudden rise in SVI value in SBMBBR was caused by the increase in the total biomass concentration as a result of two op- erating cycles applied per day [89]. Apart from this reason, the increase in the shear stress due to the 10 % (v/v) biocarrier filling ratio delayed the formation of compressed and settleable sludge flocs, causing the increase in SVI value [36].
3.6. Extracellular polymeric substance (EPS)
Fig. 10shows the presence of several intense frequency peaks on the five segregated regions ranging from I to V, which were found to be related to the stretching and bending vibrations possessed by EPS macromolecules comprising polysaccharides, proteins, nucleic acids and lipids [90]. Thus, this observation substantiated the existence of EPS under the entire SBMBBR evaluation period.
The spectral region V showed a relatively more intense peak (1021 cm-1-1023 cm-1) in the spectrum, indicating the characteristics of polysaccharides (C-H Trans), which is the carbohydrate-like backbone for the EPS [91]. The observation was in line with the composition of polysaccharides reported in the EPS constitution. Interrelatedly, the broader peak appeared at 3269 cm-1to 3272 cm-1on the spectral region I was ascribed to the stretching vibration of the O–H bond. This evi- denced that the abundance of hydroxyl groups presented in the car- bohydrate backbone of EPS as well as the possibility of residual moisture content in the sample [92,93]. Furthermore, the two peaks displayed in spectral region III, which were ascribed to C=O stretching vibration of amine I (1628 cm-1to 1636 cm-1) and N–H bending vi- bration of amine II (1534 cm-1to 1541 cm-1), respectively. These two peaks are unique and they are an essential testimony of protein sec- ondary structure in EPS [94,95]. The bands of 2852 cm-1 and 2953 cm−1in spectral region II correspond to C–H stretching vibrations of lipids in the EPS as it was allocated within the domain 2850 cm-1to 2970 cm-1[92,96]. Lastly, the less intense peak of 1238 cm-1to 1254 cm-1in spectral region IV was due to the presence of nucleic acid in the EPS [91].
4. Conclusions
MBBR technology was hitherto incorporated into an integrated an- oxic-aerobic REACT SBR system, which had paved a new way for the treatment of dye-containing wastewater. The primary outcome derived from three experimental design schemes can be summarized into:
•
Scheme 1In the early acclimatization, the dependence of co-substrate was
inevitable for RO16 biodegradation since it functioned as an electron donor for the whole process. More than 97 % COD removal efficiency and a complete biodecolorization were attained under this scheme. A threefold improvement of microbial strength in removing RO16 mole- cules was noticed, i.e., from 0.005 to 0.015 mg RO16/mg biomass. This was congruent with the improvement of microbial strength in terms of mg COD/mg biomass, i.e., from 0.0677 to 0.0711. No autoxidation was encountered.
•
Scheme 2A discernible decline in COD and RO16 removal efficiencies, i.e., both COD and RO16 removals were deteriorated to approximate 40 % at the end of this scheme. The decrease in COD removal was in agreement with the decline in microbial strength (mg COD/mg bio- mass) from 0.071 in phase 6 to 0.0452 in phase 9 as a result of the reduction on maximum co-substrate uptake from 100 % to 69 %.
However, the microbial strength in removing RO16 molecules was doubled up to 0.0304 mg RO16/mg biomass as compared to 0.015 in phase 6. The presence of the EPS on the biofilm shielded the attached- growth biomass system from the toxicity exerted by RO16 shocks.
Henceforth, the attached-growth biomass system provided advantages for the suspended-growth biomass systems and safeguarded the actively metabolic potential under this scheme. A labile extent of autoxidation was encountered in phase 9.
•
Scheme 3The treatment performance on COD and RO16 removals were fur- ther undermined to approximate 36 % and 27 %, respectively, under this scheme. The falls on both COD and RO16 removal efficiencies were concomitant with the impairment of microbial strength in removing COD and RO16 molecules, in which it dropped from 0.0452 to 0.036 mg COD/mg biomass and 0.0304 to 0.0191 mg RO16/mg biomass, respectively. The short HRT was the main reason for the deterioration herein despite the 10 % (v/v) biocarrier filling ratio did show ameli- oration on treatment efficiencies. Nonetheless, the previous autoxida- tion was eliminated in phase 10.
The results of the study demonstrate that the integrated anoxic- aerobic REACT SBMBBR system is suitable for the treatment of RO16- containing wastewater with up to 300 mg L-1. The SBMBBR system was also proven to be capable of contending RO16 shocks. However, a higher biocarrier filling ratio, longer cyclic reaction time and post- treatment are required in order to achieve better performance.
Declaration of Competing Interest
The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influ- ence the work reported in this paper.
Acknowledgment
This work was supported by the Universiti Tunku Abdul Rahman under the UTARRF Project No. IPSR/RMC/UTARRF/2016-C1/L2.
Appendix A. Supplementary data
Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jwpe.2020.101268.
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