In this study, the denitrification of nitrate-contaminated groundwater by the heterotrophic denitrification (HD) method was studied to produce drinking water. Preliminary tests were carried out in a denitrification reactor, consisting of an opaque PVC column filled with a plastic packing, and fed with a synthetic solution based on glycerol, in which activated sludge bacteria were added. The performance of the reactor was monitored by measuring physicochemical parameters such as pH, turbidity, nitrates, and nitrites. This monitoring was carried out for different tests within the same reactor to evaluate the adaptation possibilities of the denitrifying bacteria. At the end of each test when the substrate was exhausted, a new synthetic solution was added to the reactor under discontinuous aeration (aeration period = 1 h). The results obtained showed that the nitrate removal efficiency reached a value of 99.42% after 8 h of treatment depending on the adaptation of the denitrifying bacteria. This experiment also showed that the concentration of produced nitrite depends on the aeration mode and it reached a value below the detection limit in the sequential aeration mode after 12 h of treatment under discontinuous aeration (aeration period = 1 h).

  • The biological denitrification using activated sludge is used to remove nitrate from groundwater.

  • The residual nitrate concentration depended on the bacteria adaptation.

  • The nitrate removal efficiency increased to 99.42% after 8 h under discontinuous aeration.

  • The mode of aeration influenced the nitrate removal and also the removal of accumulated nitrite.

  • The produced nitrite was removed after 12 h in discontinuous aeration.

Graphical Abstract

Graphical Abstract
Graphical Abstract

Life on Earth relies on the existence of water (Saleh & Gupta 2012; Saravanan et al. 2014). Water pollution by organic and inorganic compounds is a matter of great public concern (Gupta et al. 2002; Gupta & Saleh 2013; Jerroumi et al. 2020). Recent anthropogenic activities are greatly increasing the amount of nitrogen cycling between the living world and the soil, and the water and the atmosphere (Galloway et al. 2003). The main sources of nitrate pollution are agricultural activities (especially fertilization), wastewater, leakage from landfills, and nitrogen solubility in the atmosphere (Abdel-Aziz et al. 2019). The nitrate contamination of groundwater has become an environmental concern and a public health problem. In fact, nitrate is a highly mobile form of nitrogen in the soil, making it a dangerous pollutant in water, and can cause adverse effects on human health, including methemoglobinemia in infants (Chen et al. 2017; Zhai et al. 2017). For those reasons, the World Health Organization and Morocco stipulated that the maximum nitrate concentration in drinking water is 50 mg NO3/L () (WHO 2011; NM03.7.001).

There exist several methods with different levels of performance and cost for water treatment (Jerroumi et al. 2019; Amarine et al. 2020a, 2020b). The most commonly used methods are ion exchange, reverse osmosis, and biological denitrification. Biological denitrification for water treatment, among others, was proven to be more economical, practical, and the most promising and versatile approach (Pu et al. 2014). In the geochemical carbon–nitrogen cycle processes, nitrates and additional carbon sources act as terminal electron acceptors and donors, respectively, to reduce nitrates to nontoxic nitrogen gas (Jiang et al. 2018). This action was carried out by denitrifying bacteria. Among these bacteria in the literature, we find the following: Thiosphaera pantotropha (Robertson et al. 1988), Bacillus (Gokce et al. 1989), Alcaligenes faecalis (Papen et al. 1989), Pseudomonas putida (Kim et al. 2008), Agrobacterium (Chen & Ni 2012), and Chryseobacterium (Kundu et al. 2014). Under anoxic conditions, four genera (Pseudomonas, Arcobacter, Comamonas, and Plaudibacter) were detected in the granular reactors that were fed nitrite or nitrate, and the genera Fusibacter, Cloacibacterium, and Erysipelothrix were uniquely identified in nitrate media (Pishgar et al. 2019). In the denitrification process, the reactions for nitrate reduction are (Zhao et al. 2011):
formula
The theoretical stoichiometric equation for the denitrification with glycerol as a carbon source (C3H8O3) may be given by the following reaction:
formula

In fact, bacterial cells take longer to grow and adapt to the new environment (Shimp & Pfaender 1987). In the same way, Dey & Mukherjee (2010) and Cozma et al. (2012) tested the biodegradation of phenol in a batch reactor. They found that the duration of adaptation phase increased with the substrate concentration. Indeed, the evolution of the adaptation phase was mentioned by some authors (Lekhlif et al. 2015b). They treated a synthetic petrochemical wastewater containing benzoic acid in a laboratory scale using a sequential batch reactor (SBR). They tested concentrations of synthetic solutions of 50, 100, 150, and 200 mg/L of benzoic acid, which were inoculated by an activated sludge collected from the municipal sewage treatment plant and aerated for acclimatization in synthetic solution of 200 mg/L of benzoic acid.

This paper is dedicated to the investigation of the nitrate removal efficiency of nitrate-rich groundwater as a function of several parameters such as the C/N/P ratio required for bacterial multiplication, the bacteria adaptation, and the aeration period.

In this study, we treated the nitrate-rich groundwater with heterotrophic denitrification (HD) using activated sludge as a source of bacteria and glycerol as a source of carbon. In addition, a comparison of process performance using continuous and batch aeration is provided.

Materials

The nitrate-contaminated synthetic water was prepared from sodium nitrate, phosphorus, and glycerol in a Chemical Oxygen Demand (COD)/N/P ratio of 50/4/1. Pure glycerol was used as an organic substrate. The salts of sodium nitrate and sodium phosphate were used as the source of nitrogen and the source of phosphorus, respectively. The seeding of the reactor was carried out using bacteria extracted from the activated sludge. The composition of the synthetic water for 5 L of distilled water is presented in Table 1.

Table 1

Composition of the synthetic solution according to the COD/N/P ratio of 50/4/1 with the quantity of sludge used

Constituents of the synthetic solutionQuantities
NaNO3 (100 mg/L of nitrates) 0.6855 g 
NaH2PO4, 2H20.1420 g 
Glycerol (P = 99.5%; d = 1.26; M = 92.10 g/mol) 0.92 mL 
Activated sludge 10 g 
Constituents of the synthetic solutionQuantities
NaNO3 (100 mg/L of nitrates) 0.6855 g 
NaH2PO4, 2H20.1420 g 
Glycerol (P = 99.5%; d = 1.26; M = 92.10 g/mol) 0.92 mL 
Activated sludge 10 g 

Experimental equipment and procedures

The pH was determined using a consort C6010 pH meter, whose probe also measures the temperature. Turbidity was measured using a turbidimeter type HACH 2100P. The nitrate and nitrite concentrations were measured using an ion chromatograph device of the type ΩMetrohm 881 Compact IC pro according to the standard NF EN ISO 10304-1. A Metrosep A Supp 5-250/4.0 column, composed of an anion exchange resin, was used for anion separation. The mobile phase consisted of a mixed solution of 3.2 mM Na2CO3 and 1 mM NaHCO3 with a flow rate of 0.7 mL/min. The peaks of the different ions analyzed were identified and quantified by MagIC Net software, version 2.3.

The experiments were carried out in a reactor operating in a batch mode. It was constituted by a cylindrical column, with a height of 1 m and a diameter of 12.5 cm, which is made of opaque PVC. Air was introduced at the bottom of the column with an air flow equal to 10 L/h. A grid was placed above the air sparger, at the bottom of the column, to support a bed of plastic packing medium. Another grid at the top of the column was used to prevent this packing from floating due to the effect of aeration (Figure 1). Tests were carried out on synthetic solutions of nitrate (100 mg/L), which were prepared in the laboratory (Table 1). The denitrification reactor, filled by plastic packing media, was fed the synthetic solution with the addition of activated sludge bacteria, and it was also aerated with a pump serving a diffuser placed at the bottom of the reactor in a sequential manner (15 min/60 min). The monitoring of parameters was carried out on samples taken over time (Figure 1).
Figure 1

Experimental setup.

Figure 1

Experimental setup.

Close modal

Design of experiments

In this study, four tests were performed:

  • The first test studied the effect of the COD/N/P ratio on nitrate removal efficiency. For this test, three ratios were used: 100/4/1, 50/4/1, and 25/4/1, respectively.

  • The second test studied the effect of adaptation on nitrate removal efficiency for synthetic waters with an initial concentration of nitrate (100 mg-nitrate/L), which was prepared according to the ratio of 50/4/1.

  • The third test studied the effect of aeration on nitrate removal efficiency by comparing batch and continuous aeration using synthetic water containing 100 mg/L of nitrate, which was prepared according to the 50/4/1 ratio.

  • The fourth test consisted of nitrate removal by avoiding the accumulation of nitrites, which are produced by the reduction of nitrates by changing the aeration cycle.

In each of these tests, the first synthetic solution was added to the submerged aerobic fixed-film reactor, which was filled with a clean packing. After the total nitrate removal or the stability of concentration in the case of the first adaptations, the bioreactor was evacuated and drained out for 15 min. Furthermore, the second synthetic solution was added without washing the bioreactor. This procedure was used for all the tests cited above. Samples were taken and filtered through syringe filters with a porosity of 0.45 μm to determine nitrate and nitrite concentrations. All the experiments were repeated three times and average values were reported.

Effect of COD/N/P ratio on nitrate removal

To investigate the effect of COD/N/P ratio on nitrate removal, we varied the COD/N/P ratio to reveal its effect on the nitrate removal efficiency for the first adaptation phase of denitrifying bacteria. In this study, we used three ratios: 100/4/1, 50/4/1, and 25/4/1.

The carbon/nitrogen ratio (C/N) is a measure of the electron donor/acceptor ratio in the denitrification process. For effective denitrification to take place, an appropriate dose of a carbon source must be used (Obaja et al. 2005). Figure 2 shows a COD/N/P ratio of 100/4/1 and a nitrate removal efficiency of 96.36% after 28 h.
Figure 2

Effect of COD/N/P ratio on nitrate removal by biological denitrification ((a) residual nitrate content and (b) nitrate removal efficiency) [10 g of activated sludge; continuous aeration; 100 mg/L of nitrate].

Figure 2

Effect of COD/N/P ratio on nitrate removal by biological denitrification ((a) residual nitrate content and (b) nitrate removal efficiency) [10 g of activated sludge; continuous aeration; 100 mg/L of nitrate].

Close modal

For the other two ratios (COD/N/P = 50/4/1, 25/4/1), limiting conditions after 12 h were noticed. In fact, the removal rates were 60.6 and 27.19%, respectively. The COD/N/P ratio is a limiting factor influencing the HD rate, since organic carbon is essential for the survival of heterotrophic bacteria. Indeed, a high C/N ratio could accelerate the growth of heterotrophic denitrifying bacteria in the biofilm and thus favor the total denitrification rate (Zhou et al. 2007; Warneke et al. 2011).

Study of denitrification parameters with the ratio of 50/4/1 and with continuous aeration

pH evolution

The evolution of pH as a function of time is reported in Figure 3.
Figure 3

pH evolution for three adaptations [10 g of activated sludge; continuous aeration; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Figure 3

pH evolution for three adaptations [10 g of activated sludge; continuous aeration; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Close modal

In this figure, an increase in pH is observed for the different tests. This is probably due to the process of anoxic HD that takes place within the biomass attached as a biofilm to the packing supports. This results in the degradation of biofilm proteins (release of hydroxide ions). Indeed, many authors (Vives Fàbregas 2004; Akın & Ugurlu 2005) have observed this phenomenon. This behavior may also be due to the release of carbon dioxide. Lekhlif et al. (2015a) observed the same evolution in tests carried out on glycerol. Other authors focused on the causes of the increase in pH. The detachment of the biofilm can trigger a phenomenon of degradation of proteins in the biomass. Et-taleb et al. (2014) showed that the increase in pH can be explained by the consumption of nitrogen compounds during reactor aeration. They showed identical pH behavior after almost 2 h of aeration in an SBR system.

Turbidity evolution

Turbidity results are displayed in Figure 4, showing an increase in turbidity at the beginning of the first adaptation test. This is probably due to the formation of bacteria flocs, which are increasing. They did not sufficiently adapt to settle on the packing. Thus, the turbidity decreased at the end. This phenomenon probably corresponds to the beginning of adaptation. In the other tests, it was found that the turbidity decreased from the beginning. Bacteria are formed in the form of biofilm, which ensures cohesion by flocculation of the bacteria due to its polysaccharides (Sheng et al. 2010; Liu & Guo 2013; Ding et al. 2015; Lekhlif et al. 2015b; Sehar & Naz 2016).
Figure 4

Turbidity evolution for three adaptations [10 g of activated sludge; continuous aeration; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Figure 4

Turbidity evolution for three adaptations [10 g of activated sludge; continuous aeration; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Close modal

The third adaptation test, however, shows a slight increase in turbidity. This could be the result of the endogenous respiration process that occurs in the biofilm after the organic substrate has been almost exhausted (Zhao et al. 2015; Kherbeche et al. 2017), causing the biofilm to slough off the packing material.

Evolution of nitrates and nitrites

The monitoring of nitrate and nitrite concentrations for the three adaptations is shown in Figure 5.
Figure 5

Evolution of (a) nitrate and (b) nitrite concentrations during the three adaptations [10 g of activated sludge; continuous aeration; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Figure 5

Evolution of (a) nitrate and (b) nitrite concentrations during the three adaptations [10 g of activated sludge; continuous aeration; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Close modal

Figure 5 shows that nitrites are produced during the aerobic treatment process. These nitrites are probably formed in areas where aeration is inaccessible when oxygen diffusion becomes limiting, especially in the deep layers of the biofilm; the transformation of nitrates into nitrites takes place in an anoxic environment. The nitrite production increases as the biofilm matures and becomes more compact. Their concentration then decreases slightly over time.

Many researchers noted that this denitrification occurs in an aerobic environment (Robertson et al. 1988; Ahn 2006; Joshi et al. 2014). Cheikh et al. (2013) studied the novel water denitrification technique based on immobilized bacterial biomass using various plastic wastes as supports in packed columns. They showed that the denitrifying bacteria, obtained from activated sludge present in a local municipal wastewater treatment plant, were capable of treating the water at an inlet nitrate concentration of 600 mg/L with an elimination rate close to 100%. Many researchers used activated sludge, collected from wastewater treatment plants, as a source of denitrifying bacteria (Zhao et al. 2012; Maintinguer et al. 2013).

As for nitrates, we noted a decrease in their concentration, which improved as the bacteria became adapted. Nitrate removal or reduction results from the heterotrophic degradation process, in which bacteria consume nitrogen as a trace element, i.e., nitrogen assimilation into the biomass. It also occurs through the process of anoxic denitrification within the biofilm, as demonstrated by the denitrification process (reduction of nitrates to nitrites) outlined above. Many authors have observed this behavior of nitrates in the first stage of the denitrification reaction. De Filippis et al. (2013) observed that during the denitrification of highly loaded wastewater in an SBR reactor, nitrates were reduced to nitrite ions and then totally reduced to nitrogen gas. Aouati et al. (2017) conducted a biological denitrification study on synthetic waters using a mixed culture taken from a treatment plant. They chose sodium succinate, the buffer mixture of acetic acid, and sodium acetate and date flour as the carbon source. They showed that nitrate degradation using all three carbon sources is complete and does not stop at the nitrite stage after 80 h of treatment. The occurrence of nitrite is probably related to the removal of nitrate according to the denitratation reaction that takes place within the biofilm in the absence of oxygen.

Figure 5 also shows that the nitrite concentration ultimately stabilizes for all three tests. It reaches values of 0.86, 5.79, and 38.85 mg-nitrite/L for the first, second, and third adaptation, respectively. This stabilization can probably be explained by the absence of nitrite-reducing enzymes (nitrite reductase).

Effect of aeration mode on denitrification kinetics

Figure 6 shows the evolution of nitrate and nitrite concentrations as a function of time for the three adaptation phases. The same remark, indicated in all the tests previously described, was also noted in this section, concerning the behavior of nitrates during the denitrification reaction with nitrite (NO2) accumulation for continuous and discontinuous aeration (Figure 6).
Figure 6

Nitrate and nitrite concentrations versus aeration time ((a) and (b) continuous aeration, (c) and (d) discontinuous aeration with aeration period = 2 h, (e) and (f) discontinuous aeration with aeration period = 1 h) [10 g of activated sludge; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Figure 6

Nitrate and nitrite concentrations versus aeration time ((a) and (b) continuous aeration, (c) and (d) discontinuous aeration with aeration period = 2 h, (e) and (f) discontinuous aeration with aeration period = 1 h) [10 g of activated sludge; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Close modal

To remove accumulated nitrite, we changed the aeration mode. Subsequently, we noticed a decrease in the concentration of nitrite until it disappeared completely. The complete elimination of the produced nitrites was obtained after 12 h of treatment under discontinuous aeration (aeration period = 1 h). This elimination can be explained by the production of enzymes (nitrite reductase) capable of transforming nitrites into nitrogen gas.

The nitrate removal efficiency was also affected by the change in aeration mode (Figure 7). Table 2 shows the nitrate removal efficiency obtained in the first 8 h of treatment as a function of aeration period for the third phase of bacterial adaptation. A slight increase in the yield was observed in the case of 2 h and 1 h of aeration period.
Table 2

Evolution of nitrate removal efficiency as a function of the aeration period tested after 8 h of treatment

Aeration period (h) 
Nitrate removal efficiency (%) 25.26 99.12 99.42 
Aeration period (h) 
Nitrate removal efficiency (%) 25.26 99.12 99.42 
Figure 7

Nitrate removal efficiency versus aeration time ((a) continuous aeration, (b) discontinuous aeration with aeration period = 2 h, (c) discontinuous aeration with aeration period = 1 h) [10 g of activated sludge; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Figure 7

Nitrate removal efficiency versus aeration time ((a) continuous aeration, (b) discontinuous aeration with aeration period = 2 h, (c) discontinuous aeration with aeration period = 1 h) [10 g of activated sludge; 100 mg/L of nitrate; COD/N/P = 50/4/1].

Close modal

The tests carried out in this study on the heterotrophic biological denitrification reaction under alternative aeration, using activated sludge as a source of bacteria, of synthetic water contaminated by nitrates. The results obtained draw the following conclusions:

In all the tests, the pH increased due to the anoxic HD process that takes place in the biomass attached as a biofilm to the plastic packing supports. Turbidity varied initially for the first addition and after each replenishment of the bioreactor with the synthetic solution. It was affected by the solid suspension of activated sludge used to inoculate the synthetic solution as well as by loose biofilm particles that were not drained. Nitrates are removed over time, especially when the bacteria are adapted. Nitrate removal could be explained either by a heterotrophic assimilation process, in which the bacteria consume nitrogen as a trace element or by an anoxic process that settles within the biofilm when the diffusion of oxygen becomes limiting. Furthermore, the accumulation of nitrites can be explained by the absence of the enzyme nitrite reductase. The complete disappearance of nitrites is mainly due to the change in the aeration mode. The efficiency of the HD reaction increased to 99.42% after 8 h of treatment with the total disappearance of nitrite after 12 h of treatment under discontinuous aeration (aeration period = 1 h).

This research did not receive any specific grant from funding agencies in the public, commercial, or not-for-profit sectors.

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.

Abdel-Aziz
M.
,
El-Ashtoukhy
E. Z.
,
Zoromba
M. S.
,
Bassyouni
M.
&
Sedahmed
G.
2019
Removal of nitrates from water by electrocoagulation using a cell with horizontally oriented Al serpentine tube anode
.
Journal of Industrial and Engineering Chemistry
82
,
105
112
.
Amarine
M.
,
Lekhlif
B.
,
Mliji
E.
&
Echaabi
J.
2020a
Nitrate removal from groundwater in Casablanca region (Morocco) by electrocoagulation
.
Groundwater for Sustainable Development
11
,
100452
.
Amarine
M.
,
Lekhlif
B.
,
Sinan
M.
,
El Rharras
A.
&
Echaabi
J.
2020b
Treatment of nitrate-rich groundwater using electrocoagulation with aluminum anodes
.
Groundwater for Sustainable Development
11, 100371.
Aouati
M. K.
,
Bougherara
H.
,
Zeroual
S.
&
Kebabi
B.
2017
Effet de la nature de la source de carbone sur la dénitrification
.
Sciences & Technologie A Sciences exactes
45,
9
15
.
Cheikh
A.
,
Yala
A.
,
Drouiche
N.
,
Abdi
N.
,
Lounici
H.
&
Mameri
N.
2013
Denitrification of water in packed beds using bacterial biomass immobilized on waste plastics as supports
.
Ecological Engineering
53
,
329
334
.
Cozma
P.
,
Hlihor
R.-M.
,
Apostol
L. C.
,
Diaconu
M.
,
Pogăcean
M. O.
&
Gavrilescu
M.
2012
Aerobic biodegradation of phenol by activated sludge in a batch reactor
.
Environmental Engineering & Management Journal (EEMJ)
11
,
2053
2058
.
De Filippis
P.
,
Di Palma
L.
,
Scarsella
M.
&
Verdone
N.
2013
Biological denitrification of high-nitrate wastewaters: a comparison between three electron donors
.
Chemical Engineering Transactions
32
,
319
324
.
Dey
S.
&
Mukherjee
S.
2010
Performance and kinetic evaluation of phenol biodegradation by mixed microbial culture in a batch reactor
.
International Journal of Water Resources and Environmental Engineering
3
,
40
49
.
Ding
Z.
,
Bourven
I.
,
Guibaud
G.
,
Van Hullebusch
E. D.
,
Panico
A.
,
Pirozzi
F.
&
Esposito
G.
2015
Role of extracellular polymeric substances (EPS) production in bioaggregation: application to wastewater treatment
.
Applied Microbiology and Biotechnology
99
,
9883
9905
.
Et-Taleb
S.
,
Elhaouti
R.
,
Abbaz
M.
,
Lhanafi
S.
,
Ez-Zahery
M.
,
Aba-Aaki
R.
&
El Alem
N.
2014
Comparaison du phénomène de colmatage des eaux usées à travers deux types de sables: l'un vierge et l'autre utilisé lors de traitement par infiltration (Comparison of the clogging phenomenon wastewater through two types of sand: one virgin and one used during treatment with infiltration)
.
Journal of Materials and Environmental Science
5
,
1906
1912
.
Galloway
J. N.
,
Aber
J. D.
,
Erisman
J. W.
,
Seitzinger
S. P.
,
Howarth
R. W.
,
Cowling
E. B.
&
Cosby
B. J.
2003
The nitrogen cascade
.
Bioscience
53
,
341
356
.
Gokce
N.
,
Hollocher
T.
,
Bazylinski
D.
&
Jannasch
H.
1989
Thermophilic Bacillus sp. that shows the denitrification phenotype of Pseudomonas aeruginosa
.
Applied and Environmental Microbiology
55
,
1023
1025
.
Gupta
V. K.
&
Saleh
T. A.
2013
Sorption of pollutants by porous carbon, carbon nanotubes and fullerene-an overview
.
Environmental Science and Pollution Research
20
,
2828
2843
.
Gupta
V. K.
,
Jain
C.
,
Ali
I.
,
Chandra
S.
&
Agarwal
S.
2002
Removal of lindane and malathion from wastewater using bagasse fly ash – a sugar industry waste
.
Water Research
36
,
2483
2490
.
Jerroumi
S.
,
Nagala
S.
,
Lekhlif
B.
,
Jamal
J.
&
Afrine
L.
2019
Investigation of electrocoagulation on the removal of nickel in waste water from an electroplating bath using aluminium and iron electrodes
.
Morocccan Journal of Chemistry
4
,
727
738
.
Jerroumi
S.
,
Amarine
M.
,
Nour
H.
,
Lekhlif
B.
&
Jamal
J. E.
2020
Removal of nickel through sulfide precipitation and characterization of electroplating wastewater sludge
.
Water Quality Research Journal
55
,
345
357
.
Joshi
K.
,
Joseph
J.
,
Srinikethan
G.
&
Saidutta
M.
2014
Isolation and characterization of Psedomonas syringae for nitrate removal under aerobic conditions
.
Journal of Biochemical Technology
5
,
693
697
.
Kherbeche
A.
,
Nsakou
S. N.
,
Lekhlif
B.
,
Hébrard
G.
&
Dietrich
N.
2017
Study of the initial glycerol concentration effects upon bacterial cells adaptation and biodegradation kinetics on a submerged aerated fixed bed reactor using Biocell® packing
.
Journal of Materials and Environmental Science
8
,
3280
3289
.
Kim
M.
,
Jeong
S.-Y.
,
Yoon
S. J.
,
Cho
S. J.
,
Kim
Y. H.
,
Kim
M. J.
,
Ryu
E. Y.
&
Lee
S.-J.
2008
Aerobic denitrification of Pseudomonas putida AD-21 at different C/N ratios
.
Journal of Bioscience and Bioengineering
106
,
498
502
.
Kundu
P.
,
Pramanik
A.
,
Dasgupta
A.
,
Mukherjee
S.
&
Mukherjee
J.
2014
Simultaneous heterotrophic nitrification and aerobic denitrification by Chryseobacterium sp. R31 isolated from abattoir wastewater
.
BioMed Research International
2014
,
1
12
.
Lekhlif
B.
,
Dietrich
N.
,
Kherbeche
A.
&
Echaabi
J.
2015a
Influence of initial glycerol concentration upon bacterial cells adaptability and biodegradation kinetics on a submerged aerated fixed bed reactor using Biocell (PE05) packing
.
African Journal of Environmental Science and Technology
9
,
71
79
.
Lekhlif
B.
,
Kherbeche
A.
,
Hebrand
G.
,
Dietrich
N.
&
Echaabi
J.
2015b
Influence of initial glycerol concentration upon bacterial cells adaptability and biodegradation kinetics on a submerged aerated fixed bed reactor using Biocell®(PE05) packing
.
African Journal of Environmental Science and Technology
9
,
71
79
.
Maintinguer
S. I.
,
Sakamoto
I.
,
Adorno
M.
&
Varesche
M.
2013
Evaluation of the microbial diversity of denitrifying bacteria in batch reactor
.
Brazilian Journal of Chemical Engineering
30
,
457
465
.
NM03.7.001 03.7. 001
.
2006
Par arrêté conjoint du ministre de l'Équipement et du Transport du ministre de la Santé, du ministre de l'Aménagement du territoire, de l'Eau et de l'Environnement et du ministre de l'Industrie, du Commerce et des Télécommunications
.
Papen
H.
,
Von Berg
R.
,
Hinkel
I.
,
Thoene
B.
&
Rennenberg
H.
1989
Heterotrophic nitrification by Alcaligenes faecalis: NO2−, NO3−, N2O, and NO production in exponentially growing cultures
.
Applied and Environmental Microbiology
55
,
2068
2072
.
Pishgar
R.
,
Dominic
J. A.
,
Sheng
Z.
&
Tay
J. H.
2019
Denitrification performance and microbial versatility in response to different selection pressures
.
Bioresource Technology
281
,
72
83
.
Pu
J.
,
Feng
C.
,
Liu
Y.
,
Li
R.
,
Kong
Z.
,
Chen
N.
,
Tong
S.
,
Hao
C.
&
Liu
Y.
2014
Pyrite-based autotrophic denitrification for remediation of nitrate contaminated groundwater
.
Bioresource Technology
173
,
117
123
.
Robertson
L. A.
,
Van Niel
E. W.
,
Torremans
R. A.
&
Kuenen
J. G.
1988
Simultaneous nitrification and denitrification in aerobic chemostat cultures of Thiosphaera pantotropha
.
Applied and Environmental Microbiology
54
,
2812
2818
.
Saravanan
R.
,
Gupta
V.
,
Narayanan
V.
&
Stephen
A.
2014
Visible light degradation of textile effluent using novel catalyst ZnO/γ-Mn2O3
.
Journal of the Taiwan Institute of Chemical Engineers
45
,
1910
1917
.
Sehar
S.
&
Naz
I.
2016
Role of the biofilms in wastewater treatment
. In
Microbial Biofilms – Importance and Applications
(Rodriguez-Couto, S. & Shah, M.P. eds.). Elsevier, Amsterdam, pp.
121
144
.
Vives Fàbregas
M. T.
2004
SBR Technology for Wastewater Treatment: Suitable Operational Conditions for a Nutrient Removal
.
Universitat de Girona, Girona, Spain
.
Warneke
S.
,
Schipper
L. A.
,
Bruesewitz
D. A.
,
Mcdonald
I.
&
Cameron
S.
2011
Rates, controls and potential adverse effects of nitrate removal in a denitrification bed
.
Ecological Engineering
37
,
511
522
.
WHO
2011
Guidelines for drinking-water quality
.
WHO Chronicle
38
,
104
108
.
Zhai
Y.
,
Zhao
X.
,
Teng
Y.
,
Li
X.
,
Zhang
J.
,
Wu
J.
&
Zuo
R.
2017
Groundwater nitrate pollution and human health risk assessment by using HHRA model in an agricultural area, NE China
.
Ecotoxicology and Environmental Safety
137
,
130
142
.
Zhao
Y.
,
Feng
C.
,
Wang
Q.
,
Yang
Y.
,
Zhang
Z.
&
Sugiura
N.
2011
Nitrate removal from groundwater by cooperating heterotrophic with autotrophic denitrification in a biofilm–electrode reactor
.
Journal of Hazardous Materials
192
,
1033
1039
.
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