Abstract

To develop a better understanding for fixed biomass processes, the development of a nitrifying bacterial biofilm, as well as the performance of treatment during modifications to operational conditions of a full-scale submerged biological filter were examined. The development of the nitrifying biofilm was investigated at four depth levels (1, 2, 4 and 5 feet). The result of bacterial subpopulations analyzed by qPCR relative to the physico-chemical parameters of the wastewater during the various tests (sustained aeration, modified backwash parameters and inflow restriction) revealed an increase of the relative presence of nitrifying microorganisms throughout the biofilm (especially for nitrite oxidizing bacteria (NOB)), but this was not necessarily accompanied by a better nitrification rate. The highest observed nitrification rate was 49% of removal in the test cell during backwashing conditions, whereas the relative ammonia oxidizing bacteria (AOB) population was 0.032% and NOB was 0.008% of the total biomass collected. The highest percentage of nitrifying bacteria observed (0.034% AOB and 0.18% NOB) resulted in a nitrification rate of 21%. The treatment of organic matter determined by measuring the chemical and biochemical oxygen demand (COD, CBOD5) was improved.

INTRODUCTION

The difficulties in regulating nitrification in wastewater effluent are well known. The common installation is two biological treatment units in series: the first for removing carbon nutrients (secondary treatment) and the second for nitrification (tertiary treatment). Single-sludge secondary treatment process is also common but harder to operate. However, it has been shown that biological aerated filters (BAFs), under optimal conditions, can sustain both treatments in a single step (Mendoza-Espinosa & Stephenson 1999). This offers an economical avenue for limiting levels of ammonia in the effluent. Wastewater effluent nitrification involves a two-step process in which ammonia-oxidizing organisms (ammonia-oxidizing bacteria (AOB)) and ammonia-oxidizing archaea (AOA) (Koops et al. 2006; Stahl & Torre 2012) oxidize ammonia to form nitrite, and nitrite-oxidizing bacteria (NOB) (Teske et al. 1994) subsequently oxidize nitrites to nitrates. Among the NOB, some species such the recently discovered Nitrospira bacteria (comammox) can completely oxidize ammonia (Daims et al. 2015). The biochemical relationship between AOB and NOB is well known and has been used as a tertiary treatment with activated sludge (Blackall & Burrell 1999). Attempts to use this biochemical relationship for a one step nutrient removal in a fixed biomass process was revealed to be complicated and not easily applicable (Bovendeur et al. 1990).

Several studies at pilot scale (Mendoza-Espinosa & Stephenson 1999; Gullicks et al. 2011; Pramanik et al. 2012) have reported that the performance of BAFs in reducing the nutrient load of primary influent may be influenced by the filtration rate (hydraulic loading), organic loading, aeration, efficiency of oxygen use and backwashing of the filters. How these influence the distribution of microbial populations on the biofilm remains largely unexplored.

Pramanik et al. (2012) reported that the use of BAFs was more efficient in removing ammonia from wastewater than with activated sludge. However, information at full scale on the efficiency of BAFs under different operating and environmental conditions is still lacking. Furthermore, correlations between bacterial populations and the efficiency of the treatment have, until now, only been studied under small-scale laboratory conditions (Li et al. 2016) and are yet to be tested in a large-scale setting.

The objective of this study was to correlate the development of nitrifying bacteria with water quality following operational assays and validate if nitrification is directly favored within a fixed growth biofilm in a large-scale wastewater treatment facility.

MATERIAL AND METHODS

WWTP description and wastewater characteristics

A municipal wastewater treatment plant (WWTP; Auteuil Laval, Quebec, Canada) using a biological filtration process was used for these studies. The mean inflow of wastewater (33,000 m3/d), after a preliminary decantation, was distributed equally on two separated down-flow aerobic biofilters using a Biocarbon® (OTV) submerged biofilter (attached growth). As shown in Figure 1(a) and 1(b), each biofilter is composed of 10 process cells filled with 1.8 m of ‘Biodagen’ (grain-sized expanded schist with a nominal diameter of 3–6 mm).

Figure 1

Detailed plan of (a) biofilters with treatment cell (gray) and selected test cell (green) and water distribution and (b) sectional view of a media bed with each depth, as used in the current work: 0.3 m (1′), 0.6 m (2′) presented as upper part of the cell, 1.2 m (4′) and 1.5 m (5′) deep presented as lower part of the cell. Media bed of a treatment cell has a total height of 1.8 m (6′); plan view with site location (cross) of sampling devices (grain sized schist and wastewater). Setup is the same for all four selected treatment cells. Please refer to the online version of this paper to see this figure in colour: http://dx.doi.org/10.2166/wst.2018.447.

Figure 1

Detailed plan of (a) biofilters with treatment cell (gray) and selected test cell (green) and water distribution and (b) sectional view of a media bed with each depth, as used in the current work: 0.3 m (1′), 0.6 m (2′) presented as upper part of the cell, 1.2 m (4′) and 1.5 m (5′) deep presented as lower part of the cell. Media bed of a treatment cell has a total height of 1.8 m (6′); plan view with site location (cross) of sampling devices (grain sized schist and wastewater). Setup is the same for all four selected treatment cells. Please refer to the online version of this paper to see this figure in colour: http://dx.doi.org/10.2166/wst.2018.447.

Each process cell opens (filtration mode) or closes (waiting mode) depending on the inflow level and is operated in rotation within the biofilter. In filtration mode, the designed hydraulic surface load on each media bed (treatment cell) is assumed to vary on average from 4.8 to 9.4 m3/m2/h (filtration time of 22.5 and 11.5 min) with a maximum inflow of between 33,000 and 65,000 m3/d. Experimental data suggest an average inflow hydraulic surface load of 7 m3/m2/h (filtration time of 15.4 min) throughout the study. Two side-by-side cells (shown in green in Figure 1) were selected from each biofilter. The wastewater to the treatment plant consisted mainly of domestic sewage. The typical influent characteristic load (and concentration) for a 5-day carbonaceous biochemical oxygen demand (CBOD5), chemical oxygen demand (COD), total suspended solids (TSS), ammonium-nitrogen (NH4+-N) and total phosphorus (Ptot) levels were 3,033 kg-O2/d (93 g-O2/m3), 9,402 kg-O2/d (286 g-O2/m3), 6,400 kg/d (197 g/m3), 369 kg-N/d (11.3 g-N/m3) and 91 kg-P/d (2.9 g-P/m3), respectively. The average removal efficiency of the plant prior to this study was 87% for CBOD5, 84% for COD, 95% for TSS, and 75% for Ptot, respectively. This WWTP was not able to sustain nitrification prior to the beginning of the study. The bioreactors had mean temperatures of 18.5 ± 1.5 °C and were operated at a pH of 6.8 ± 0.3 during the sampling times.

Testing conditions

Three different operational conditions were assessed. 1. Aerobic conditions in which 3,000 m3/h of processed air was injected through the test biofilter, whereas in the control biofilter, no processed air was injected. 2. Modified backwash parameters. Modified backwash parameters for the cells in the test biofilters were as follows: unclogged procedure for 2 min with an air injection alone, followed by backwash and water rinse for 6 min followed by water alone for 1 min (9 min). Filtration time between backwashes was reduced from 16 to 12 h. Control biofilters followed the same procedure except that the backwash lasted 11 min, instead of 6 min, and filtration time between backwashes was 16 hrs. The rotation of waiting cells of a Biocarbon® filter was set at 30 min in both the test and control cells. 3. Inflow adjustment. The inflow was restricted to 2.3 m2 (test) instead of 10.1 m2 (control). This restriction caused an average reduction of 36% of the normal wastewater inflow, a reduction of the nutrient load and an average filtration time of 21 min with a flow rate of 5 m3/m2/h (controls had no restriction and displayed an average filtration time of 15.4 min with a flow rate of 7 m3/m2/h). Processed air was maintained at 3,000 m3/h in the test and the control cell for the modified backwash and inflow tests.

Aerobic conditions and modified backwash parameters were used to compare results from one biofilter (test) to the other (control) while the inflow adjustment compared a process cell with an adjacent control cell. For each test condition, a minimum conditioning period of 5 weeks was followed before schist grain sampling.

Samples

Expended schist grain (Biodagen)

Approximately 100 mL of samples of Biodagen were taken by core sampling in the media bed of the test and control cell (Figure S1, available with the online version of this paper) between two backwashes at approximately middle cycle of filtration for each testing conditions. Two side by side cells were selected per biofilter. Selected cells were symmetrically opposed in each biofilter. Each sampling was done at four different depths over the schist bed (Figure 1(b)): 0.3, 0.6, 1.2 and 1.5 m (1, 2, 4, and 5 feet). Samples were taken at the same times of water sampling. Each schist sample was preserved at −20 °C in a sterile 50 mL tube containing glycerol for microbiological analysis.

Wastewater sampling and physico-chemical analysis

Wastewater was collected with an Easy-Load® peristaltic pump (Masterflex®, model 7518-60) with LS-15 Masterflex® tubing, in accordance with the Ministère de l'Environnement du Québec's Centre d'Expertise en Analyse Environnementale (CEAEQ-MDDEFP) methodologies (Figure 1(b)). The protocol involved sampling proportional to the time with uptake rate of 1 L per hour to compose a single sample of 14 h. The sample was kept at 4 °C in a polyethylene bottle. The wastewater was sampled twice a week at each sampling depth (including the influent biofilter) and each produced value comes from a triplicate analysis. A mean for each parameter has then been produced with raw data (Table S1, available online) collected and presented in Figure 2. Table 1 shows the physico-chemical parameters monitored during the study and reference methods (APHA et al. 2012). Data were obtained using a number of samples (n = 5–10) for each parameter analyzed. The mean value of the inflow for each biofilter was calculated as this will influence the nutrient load and filtration rate. The data were then correlated to the development of the nitrifying bacterial population.

Table 1

Physico-chemical parameter monitored with limits of detection and references

ParameterUnitLODStandard Methods reference
COD mg-O2/L 16 SM 5220 
CBOD5 mg-O2/L 1.0 SM 5210 
N-NH4+ mg-N/L 1.0 SM 4500-NH3 
Total alkalinity mg-CaCO3/L 4.0 SM 2320 
ParameterUnitLODStandard Methods reference
COD mg-O2/L 16 SM 5220 
CBOD5 mg-O2/L 1.0 SM 5210 
N-NH4+ mg-N/L 1.0 SM 4500-NH3 
Total alkalinity mg-CaCO3/L 4.0 SM 2320 

SM: Standard Methods for the Examination of Water and Wastewater (APHA et al. 2012).

LOD: limit of detection.

CBOD5: carbonaceous biochemical oxygen demand incubates for 5 days at 20 °C ± 1 °C.

Figure 2

COD, CBOD5, NH4+-N) during (a) aerobic testing condition (mean on 42 days), (b) backwashing testing condition (mean on 31 days) and (c) during inflow testing condition (mean on 43 days). Each graph is separated in test and control section. COD (chemical oxygen demand), CBOD5 (5-day carbonaceous biochemical oxygen demand) and NH4+-N (ammonia nitrogen). Statistical analysis was performed with ρ = 0.05. Error bars represent the standard deviation on raw data for each parameter.

Figure 2

COD, CBOD5, NH4+-N) during (a) aerobic testing condition (mean on 42 days), (b) backwashing testing condition (mean on 31 days) and (c) during inflow testing condition (mean on 43 days). Each graph is separated in test and control section. COD (chemical oxygen demand), CBOD5 (5-day carbonaceous biochemical oxygen demand) and NH4+-N (ammonia nitrogen). Statistical analysis was performed with ρ = 0.05. Error bars represent the standard deviation on raw data for each parameter.

Biofilm analysis

DNA extraction

Biological material was extracted from expanded schist grains of the biofilm by vigorously shaking 6 grains of similar size in the PBS buffer (10 mL) for 3 min followed by washing in PBS buffer (5–10 mL; three times). Genomic DNA was then extracted from 1.7 mL of biofilm mixture using a nucleic acid extraction kit (Wizard Genomic DNA purification kit, Promega, Madison, WI, USA), according to the manufacturer's instructions. Each sample was eluted in a DNA rehydration solution prior to being diluted (100–1,000 fold) with water and stored at −20 °C. DNA levels were quantified using the Quant-iT™, PicoGreen®, dsDNA test kit (ThermoScientific, Burlington, ON, Canada) (Ahn et al. 1996).

qPCR amplification of amoA and nxrB

To estimate the relative abundance of AOB and NOB populations in the extracted biofilm, the partial sequence of either the amoA (AOB) or nxrB (NOB) genes from a selection of the principal AOB and NOB species found in nitrifying WWTPs (Table S2, available online) were aligned using the ClustalW2 software (European Bioinformatics Institute, Hinxton, Cambridgeshire, UK). As in previous studies, we considered that AOA populations were negligible (Graham et al. 2007) and that Nitrosomonas spp. and Nitrospira spp. were considered the principal AOB and NOB species in the wastewater samples (Liu & Jansson 2010). Degenerate primers (Table 2) were designed in order to obtain a broad spectrum of AOB and NOB populations. qPCR analyses were performed using 96-well plates (Eppendorf, Mississauga, ON, Canada) and the Perfecta® SYBR®Green FastMix® kit (Quanta, Beverly, MA, USA) according to the manufacturer's instructions. DNA amplifications by qPCR for amoA were done using 40 cycles of initial denaturation at 95 °C for 5 min; denaturation at 95 °C for 10 s, annealing at 46.6 °C for 30 s and elongation at 72 °C for 10 s. Amplifications of nxrB were done using 40 cycles of initial denaturation at 95 °C for 5 min; denaturation at 95 °C for 10 s, annealing at 59.3 °C for 30 s and elongation at 72 °C for 10 s. An enriched nitrifying biofilm from pilot-scale nitrifying process (STEPPE laboratory, École de technologie supérieure) was used as positive control.

Table 2

Newly designed oligonucleotides primers for qPCR analysis

PrimersSequence (5′–3′)Target sizeAmplicon size (bp)Specificityd
16SrDNA790f GAT ACC CTG GTA GTC CAC GC 790–810a 160 Total bacteria 
16SrDNA950r CAT RMT CCA CCG CTT GTG CGG G 922–950a 
amoA930f GTA TCM ATG YTG ATG TTC 930–947b 141 AOB 
amoA1069r CCC TCK GSA AAG CCT TCT TCA 1,049–1,070b 
nxrB1061f AGC CAR CAG ATC ATC TTC CGG TA 1,061–1,083c 124 NOB 
nxrB1184r G GST TCA ACA AYT CGG GCA AGG 1,163–1,184c 
PrimersSequence (5′–3′)Target sizeAmplicon size (bp)Specificityd
16SrDNA790f GAT ACC CTG GTA GTC CAC GC 790–810a 160 Total bacteria 
16SrDNA950r CAT RMT CCA CCG CTT GTG CGG G 922–950a 
amoA930f GTA TCM ATG YTG ATG TTC 930–947b 141 AOB 
amoA1069r CCC TCK GSA AAG CCT TCT TCA 1,049–1,070b 
nxrB1061f AGC CAR CAG ATC ATC TTC CGG TA 1,061–1,083c 124 NOB 
nxrB1184r G GST TCA ACA AYT CGG GCA AGG 1,163–1,184c 

aPosition relative to the 16S rRNA gene of Escherichia coli (J01859).

bPosition relative to the amoA gene of Nitrosomonas europaea (L08050.1).

cPosition relative to the nxrB gene of Nitrospira marina (KC884909.1).

dSpecificity is related to selected species present in wastewater.

The amoA and nxrB genes were amplified from the same genomic DNA as that used for the amplification of the 16S rRNA which was used as an indicator of the total bacterial population (supplemental information, Figure S2). qPCR amplification produced amplicons of 141-and 150-bp for the amoA and nxrB genes, respectively. The data were quantified using a standard curve method (Morrison et al. 1998; Schefe et al. 2006) and the size of the amplicons confirmed by electrophoresis.

Data analysis

Relative populations of AOB (amoA) and NOB (nxrB) detected in the samples were calculated in relation to the total genomic DNA collected within the biofilm sample as previously reported (Dionisi et al. 2002; Dryburgh 2011; Pester et al. 2014; Gruber-Dorninger et al. 2015). Estimation of the relative abundance of total bacteria within the biofilm, as determined using 16S rRNA levels, were calculated and presented in supplemental materials.

Statistical analysis

To determine if there were significant differences between the physico-chemical parameters and bacterial populations in test and control cells, statistical analyses were performed using an F-test and a one-way Student t-test. Statistical significance was set at P ≤ 0.05.

RESULTS

Physico-chemical analysis

Aerobic conditions

At each depth analyzed (1, 2, 4 and 5 feet), data on aerobic conditions (Figure 2(a)) show a reduction of 57 and 74% in COD and CBOD5 from the wastewater inlet to the lower part of the aerated test cell. In contrast, a negligible reduction of COD (0.4%) and a 16% increase in CBOD5 was observed from the inflow throughout the control cell. In the aerated cell, a 17% reduction in NH4+-N was observed. There was no removal of NH4+-N in the control cell but rather a significant increase of 11%. Statistical analyses indicate a significant difference for COD, CBOD5 and NH4+-N reduction between the test and control cell. Data on alkalinity (data not shown) and NH4+-N indicated a reduction of 33 ± 11 kg/d of NH4+-N and an alkalinity consumption of 210 ± 22 kg/d (ratio alkCons/NReduc = 6.4 ± 2.8). This decrease further corroborates the reduction in NH4+-N content, since the consumption of alkalinity is theoretically 7.14 fold higher than the amount of nitrified NH4+-N (Metcalf & Eddy 2014). The average influent flows on the biofilters were similar throughout the experiments (16,895 m3/d for biofilter 1 and 16,930 m3/d for biofilter 2). Data of dissolved oxygen (Figure 3(a)) show an increase in the lower part of the testing cell with values between 4 and 6 mg-O2/L. Data of the control cell were in the limits of detection with average values of 1 mg-O2/L.

Figure 3

Results of dissolved oxygen throughout the filter test and control cell during (a) the aerobic condition (mean on 42 days), (b) backwash testing condition (mean on 31 days) and (c) during inflow testing conditions (mean on 43 days). Data were collected twice a week to produce the mean.

Figure 3

Results of dissolved oxygen throughout the filter test and control cell during (a) the aerobic condition (mean on 42 days), (b) backwash testing condition (mean on 31 days) and (c) during inflow testing conditions (mean on 43 days). Data were collected twice a week to produce the mean.

Modified backwash parameters

Data on backwash conditions (Figure 2(b)) indicate a reduction of COD and CBOD5 of 49 and 83%, respectively, in the test cell, while a reduction of 45 and 77% was observed in the control cell. The observed efficiency for the treatment of carbonaceous nutrients was not significantly different between test and control cells. However, the percent of NH4+-N removed in the test cells (49%) was significantly higher than in control cells (16%). Data on alkalinity and NH4+-N indicate a reduction of 102 ± 19 kg/d of NH4+-N and an alkalinity consumption of 737 ± 158 kg/d for the test cell (ratio alkCons/NReduc = 7.2 ± 2.9). A reduction of 35 ± 19 kg/d of NH4+-N and an alkalinity consumption of 246 ± 100 kg/d was observed for control cells (ratio alkCons/NReduc = 7.0 ± 4.8), thereby confirming the increased nitrification. The average influent flow on the biofilters was 16,528 m3/d for biofilter 1 and 18,070 m3/d for biofilter 2. Interestingly, the treatment appears to peak at a depth of 4 feet (Figure 2(b)). Data of dissolved oxygen (Figure 3(b)) show an increase at 4′ depth in both test and control cell with an average values between 5 and 6 mg-O2/L. There is no significant differences of dissolved oxygen between the test and control cell.

Inflow adjustment

Data on inflow restriction conditions (Figure 2(c)) show a significant reduction of 69 and 87% of COD and CBOD5 in the test cell while a reduction of 53 and 62% was observed for the control cell, in particular for the lower portion of cell. Removal of NH4+-N was observed only in the test cell, and reached 21%. Significant differences in NH4+-N removal between the test and control cell was also confirmed for the lower depths of the test cells. Data on alkalinity and NH4+-N show a reduction of 39 ± 14 kg/d of NH4+-N and an alkalinity consumption of 292 ± 95 kg/d for the test cell (ratio alkCons/NReduc = 7.5 ± 5.1), confirming the nitrification process. The average influent flow on biofilters was 15,670 m3/d for biofilter 1 and 17,070 m3/d for biofilter 2. Data of dissolved oxygen (Figure 3(c)) show an increase in the lower part of the testing cell with values between 5 and 6 mg-O2/L and significant differences of the 1′ and 5′ feet depth between the test and control cell. Data of the control cell show dissolved oxygen at the 2′ and 4′ depth while at 1′ and 5′ values were at an average of 1 mg-O2/L.

Nitrifying and total bacteria analysis

Data on aerobic conditions for the test cell (Figure 4(a)), indicate that AOB populations within the biofilm were significantly more enriched at a depth of 4 and 5 feet than in the control cell, with values between 0.0095 to 0.013% for the test cell and 0.004 to 0.008% for the control cell. Data on the relative NOB population showed that they are generally present at significantly higher levels in the test cell as compared to the control cell, in particular at 2, 4 and 5 feet with values ranging from 0.0009 to 0.008% in the test cell and 0.0004 to 0.002% in the control cell. The relatively low levels of NOB suggest a very low nitrification activity in the biofilters. However, as indicated above, 17% of NH4+-N was removed in the lower portion of the test cell. Each test and control cell was operated continuously under test conditions for 5 weeks prior to sampling. Despite this, the biofilm in the control cell was able to maintain AOB and NOB populations, even though the conditions were theoretically not suitable for these bacteria due to the lack of oxygen. These results support the fact that the fixed biomass is an effective and resilient biofilm structure. Indeed, BAFs display an advantage in the attachment of nitrifiers which allowed them to survive until better growth conditions (Bovendeur et al. 1990). However, an activity test would be required to confirm the viability of the AOB/NOB presence detected.

Figure 4

Results of relative overall presence of nitrifying bacteria (% AOB and % NOB) in the biofilm under testing conditions. Each graph is separated in test and control section. AOB (ammonia oxidizing bacteria) and NOB (nitrite oxidizing bacteria) analysis during (a) aerobic testing condition (mean on 42 days), (b) under backwashing testing condition (mean on 31 days) and (c) under inflow testing condition (mean on 43 days). Statistical analysis was performed with ρ = 0.05. Error bars represent the standard deviation on raw data for both AOB and NOB analysis.

Figure 4

Results of relative overall presence of nitrifying bacteria (% AOB and % NOB) in the biofilm under testing conditions. Each graph is separated in test and control section. AOB (ammonia oxidizing bacteria) and NOB (nitrite oxidizing bacteria) analysis during (a) aerobic testing condition (mean on 42 days), (b) under backwashing testing condition (mean on 31 days) and (c) under inflow testing condition (mean on 43 days). Statistical analysis was performed with ρ = 0.05. Error bars represent the standard deviation on raw data for both AOB and NOB analysis.

Data on modified backwash parameters (Figure 4(b)) show a significant increase in the AOB relative population throughout the lower part of the media bed between the test (maximum value of 0.032%) and control cell (maximum value of 0.019%). The highest values were observed at a depth of 4 and 5 feet in the test cell. Data on the relative NOB population showed a significant difference at a depth of 4 and 5 feet between the test and control cell. The highest value in the test cell was reached at 4 feet with NOB making up 0.008% of the bacterial population and 0.004% of the control cell. In contrast, the progression of the relative NOB population in the control cell was almost negligible, except at 4 feet which represents the point where there is the most intense degree of aeration in the filter bed. Maximum NH4+-N removal was observed at a depth of 4 feet, with values of 49 and 16% in test and control cell, respectively. These results indicate a positive response to the modified backwash conditions for the development of nitrifying bacterial populations and, consequently, the removal of ammonia.

Data on inflow adjustment conditions resulted in unexpected observations (Figure 4(c)). The relative AOB populations in both the test and control cells appeared to stabilized between 0.020 and 0.034%. The relative abundance of AOB at 4 feet depth of the test cells was 0.01% for the aeration assay, 0.032% for the modified backwash parameter and 0.034% for the inflow restriction assay. Data on relative NOB population show a significant difference between the lower parts (4 and 5 feet) of the test cell as compared to the control cell. Maximum relative NOB abundance for the test cell at 4 and 5 feet was 0.18 and 0.16%, compared to 0.042 and 0.047% for the control cell. The maximum values were reached at a depth of 1 foot with the NOB population making up 0.10% of the bacterial population. These surprising data with regard to the NOB population under inflow adjustment condition, as compared to the other testing conditions, may be due to a slight increase in the wastewater temperature. The temperature monitored over the biofilters was at 17 °C, 15.8 °C and 18.3 °C for aeration, modified backwash and inflow adjustment conditions. The data indicated a statistical difference (P ≤ 0.1) between the temperature of the inflow adjustment and the modified backwash assays.

DISCUSSION

Results from the three conditions tested indicate an improvement in the treatment under each situation with regard to organic load and nitrification. There were few notable differences in the relative presence of total bacterial population within the biofilm sample between the upper and lower regions of the test cells in the different assays. This represented between 15 and 20% of the entire biofilm (Figure S2, available with the online version of this paper). However, the relative nitrifying subpopulation of bacteria within the overall bacterial population was increased.

Relation between relative % of nitrifying bacteria and process yield

Results shown in Figures 2 and 4 indicated that the relatively small population of nitrifying bacteria within the fixed biofilm was sufficient to produce quantifiable changes in NH4+ concentration. As shown in Figure 4, the lowest and highest relative AOB population observed in the test cells was 0.010% (Figure 4(a), aerobic conditions at 1 foot) and 0.034% (Figure 4(c), inflow adjustment at 4 feet). Maximum nitrification rate was noted when the AOB population reached 0.032% of the total bacterial population (Figure 4(b); testing cell; modified backwash parameters at 4 feet). When this situation occurred, the relative bacteria population was composed of 0.008% NOB within the biofilm and had an increased nitrification rate of 49% removal between the top of the cell and at a depth of 4 feet. At the maximum NOB peak of growth (Figure 4(c); testing cell; inflow adjustment at 4 feet), a relative NOB population of 0.18% was observed. This represents approximately 23 times more NOB as compared to the population levels observed with the modified backwash parameters. However, the AOB population was estimated at 0.034% of the total population within the biofilm and the nitrification rate evaluated at 21% as compared to 49% nitrification in the backwash test. These results suggest that the NOB population is not strictly dependent on nitrification for growth and that the higher percent population is not directly associated with better nitrification. In accordance with a previous study (Chandran & Smets 2000), these results suggest that the nitrification process is mainly controlled by the AOB rather than the NOB.

Effect of aeration on overall treatment and nitrifying community

Aeration is critical for the proper operation of BAFs. Too much air produces overgrowth on the biofilm and therefore a premature clogging of the filter; not enough air leads to the development of anoxic zones within the filter, thereby significantly reducing the treatment performance (Mendoza-Espinosa & Stephenson 1999; Gerardi 2006). Tests on aeration (3,000 m3/h) in one biofilter showed a significant improvement of wastewater treatment in the test cells as compared to control cells (Figure 2(a)) as test cells displayed a better reduction rate of COD and CBOD5.

The biofilm assay results indicate the presence of nitrifying bacteria at all depths sampled. From an engineering perspective, the presence of a particular functional group of bacteria in the treatment reactor may be sufficient for nitrification (de Beer & Muyzer 1995). However, our results indicate that the increase in nitrifying bacterial population within the whole biofilm in the aerated cell is correlated with NH4+-N treatment efficiency. The relative nitrifying population present in the biofilter is different from those reported by other studies on wastewater treatment systems using a mixed liquor solids suspension process (Dionisi et al. 2002; Harms et al. 2003). Dionisi et al. (2002) reported that AOB made up 0.0033% and NOB 0.39% of the total bacteria in their WWTP study while Harms et al. (2003) reported 1.7% AOB and 8.6% NOB. In the present study, the highest levels of AOB and NOB were 0.013% and 0.008%, respectively (Figure 4(a)). Even though these levels were relatively low, they were able to decrease NH4+-N by 17% in the test cell. Even if nitrifying bacteria were present in the control cell, no nitrification was observed. Instead, increases of N-NH4+ observed can be explain by possible amonification of urea with the residual dissolved oxygen or the action of denitrifying bacteria. However, the testing conditions were not suitable to gain an efficient treatment in the control cell due to the lack of O2 (Figure 3(a)). Indeed, COD and CBOD5 removal in the control cell was very limited, with a COD reduction of 0.4% and an increased CBOD5 of 16%. Graham et al. (2007) reported that NOB bacteria are more susceptible to unsuitable conditions such as low O2 concentrations, pH under 6.5 (Jiménez et al. 2011; Claros et al. 2013), inhibitory compounds and temperature, which may explain the low NOB abundance obtained in our aeration experiments. Several studies have reported that the NOB population is dominant relative to the AOB population for functional nitrification (Daims et al. 2000; Dionisi et al. 2002; Gieseke et al. 2003; Harms et al. 2003). Our results, however, indicate a higher AOB population as compared to the NOB population when the nitrification process is not fully functional. However, the fact that nitrifying bacteria were still present in the control cell after more than 5 weeks without oxygen input was unexpected.

Effect of shorter and more frequent backwash on overall treatment and nitrifying community

Feng et al. (2017) have recently reported the effect of backwashing on microbial community structure at a genus level in a BAF. However, there remains a lack of information regarding these effects on the development of specific microbial populations and their importance for treatment efficiency. Modifying the backwash experimental conditions to shorter cycles appears to have a beneficial impact on the NH4+-N treatment. A recent study reported improved biofouling control of backwashing to increase the filtration time, regardless of the wastewater treatment recovery and the effect on microbial community structure (Hasnain et al. 2017). However, our results showed that a less intense backwash will, over time, improve the treatment quality with an increased removal of NH4+-N, particularly in the deeper portion of the cell. Ammonia removal was increased at both the 4 and 5 foot mark in the test cell, although the higher efficiency was observed at 4 feet. The decreased efficiency observed of the NH4+-N treatment from the 4 to the 5 foot depth may result from the lack of uniformity or preferential current of the water down-flow which can induce slight changes in water parameters such as nutrient loading rate of N-nitrogen or faster inflow rate (Metcalf et al. 2004).

The relative presence of nitrifiers within the test cells increased throughout the treatment. This supports the notion that a less intense backwash maintains the bacterial population and allows the relative population to expand. This happens primarily in relation to the depth of the media bed as we have noted no significant changes in the % AOB and NOB between the test and the control cell in the shallow portion of the media bed. A gradual increase in AOB and NOB populations in the test cell from the upper part to the lower part of the media bed was observed. This contrasts with the control cell where no significant increases were observed. Likewise, Feng et al. (2017) reported a decrease in the relative presence of the majority of microbial genus in the lower part of the cell.

Effect of reduced inflow rate on overall treatment and nitrifying community

Decreasing the flow rate by approximately 36% through the test cell allowed the percent NOB population to significantly increase. Furthermore, the increase in the relative population of NOB was much higher in the reduced inflow rate assay than the results obtained in either the aeration or modified backwash parameter assays. This supports the notion that NOB appear to be the dominant nitrifying bacterial population, in particular when environmental conditions are favorable for their development (Daims et al. 2000; Dionisi et al. 2002; Winkler et al. 2012), such as a long contact time between the wastewater and the biofilm. However, the presence of Nitrotoga that can achieve nitrification at low dissolved oxygen could partly explain the increases of NOB over AOB (Keene et al. 2017). It should be noted that the relative AOB population did not significantly change between the test and the control cell in this assay. Changes in the relative populations of AOB and NOB did not result in altered rates of nitrification, which was unexpected (Figure 2(b) and 2(c)).

The test conditions show relatively little variation between control and test cells on the overall bacterial population within the biofilm (Figure S2). This means that observed differences in NOB population was not associated with the overall bacterial population. This suggests that the presence of NOB does not rely exclusively on nitrites for growth. It has been shown by Koch et al. (2015) and by Daims et al. (2015) that NOB specific strains of Nitrospira (moscoviensis and inopinata) can metabolize urea into ammonium and sustain complete nitrification in the presence of O2. This may have a direct influence on the rates of nitrification when the NOB population is dominant in a nitrifying biofilm because of the additional NH4+-N that may supplied by Nitrospira spp. Winkler et al. (2012) also tried to explain the disproportion of the NOB population relative to the AOB population in aerobic granular sludge by a nitrite loop hypothesis where nitrite oxidation/nitrate reduction by denitrifiers supply the NOB in nitrite independent of AOB. With lower loading rate (inflow restriction assay), this could mean lower N-NH4+ to be oxidized for NOB. In this case, NOB can be produced by themselves the ammonification of urea. This N-NH4+ can be oxidized by NOB (comammox) reducing by the same way the N-NH4+ available for the AOB and producing the nitrite independent from AOB. If denitrifiers are present, then the reciprocal feeding may have happened. This may be occurring on a fixed biofilm and may partly explain our present results. It is difficult to estimate the nitrification rate based on the population of nitrifying bacteria, as the NH4+-N is consumed/oxidized by the AOB at the same time as it is supplied by the reverse metabolic pathway by certain NOB species. This may result in a misinterpretation of the nitrification rate. Furthermore, the decreased flow rate increased the removal of both COD and CBOD5 with lowest values observed in this study for these parameters, but did not reach the expected levels of nitrification. Our observations lead us to hypothesize that in order to sustain a good nitrification rate in this type of process, and if the presence of AOB is to be improved at the same time as that of NOB, the washing conditions must be modified, the flow rate reduced and the process subjected to sustained aeration.

CONCLUSION

The study shows that modifications in operational conditions related to sustained aeration, modified backwash parameters and decreased flow rate on a treatment cell can increase the efficiency of wastewater treatment. This improved efficiency was not directly correlated with an increased abundance of nitrifying bacteria and, in particular, the relative NOB population, in the biofilter. We suggest the possibility of maximizing the nitrification process while reducing biomass by limiting the metabolic pathway for NOBs. This modification to the operating procedure of the WWTP could upgrade the efficiency of the biofilter to increase ammonia removal as well as decreasing COD and CBOD5 effluent concentrations.

ACKNOWLEDGEMENTS

Financial support from the National Sciences and Engineering Council of Canada and from the Fonds de recherche du Québec – Nature et technologies and the City of Laval in the form of Industrial Innovation Studentship to Bourgeois F-R. M. Gagné, J. Charbonneau, S. Dubé (City of Laval), are thanked for their assistance and for providing access to the Laval WWTP. J. Dufresnes (INRS-IAF) is thanked for her technical assistance. We also thank Louisa Sades (STEPPE-ÉTS) for her assistance with the analyses.

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Supplementary data