28 In this study, a municipal lagoon with high wintertime effluent total ammonia 29 nitrogen (TAN) concentrations was upgraded with a pilot -scale nitrifying -nitrifying - 30 denitrifying ( NIT-NIT-DENIT ) moving bed biofilm reactor (MBBR) treatment train to 31 characterize its effluent over wintertime operation , investigate the feasibility of upgrading 32 lagoons to achieve substantial biological total nitrogen r emoval across ultra-low 33 temperatures (0.6 – 3.0°C) and investigate nitrification inhibition pathways in facultative 34 lagoon systems at ultra -low temperatures . Throughout the study, it was observed that the 35 system substantially reduced total nitrogen (TN) and total phosphorus (TP) effluent 36 concentrations by an average of 69.8 ± 24.5% and 74.7 ± 20.1% , respectively . 37 Furthermore, it was observed that sulfide toxicity may play an important role in the 38 inhibi tion of nitrifying organisms in lagoons. Finally, t he MBBR treatment technology has 39 emerged as a suitable and sustainable upgrade technology for existing lagoon and waste 40 stabilization pond facilities operating in temperate, northern, and cold climate countries. 41 42 Keywords: wastewater treatment, water quality, pollution, seasonal, environmental 43 management 44 3 1. Introduction 45 The impoverishment of surface waters is a global problem that impact many 46 communities around the world . Anthro pogenic release of nutrients to natural waters is 47 widely recognized as a leading cause of eutrophication and water toxicity (Smith et al., 48 1999; Lewis et al., 2011) . Furthermore, i t has been recognized that the effluent discharge 49 of water resource recovery facilities (WRRF) is a major source of these anthropogenic 50 nutrients (nitrogen & phosphorus) that are released to the natural environment (Preston 51 et al., 2011) . As a result of the impact of wastewater discharge, several countries and 52 states have imposed stringent effluent discharge regulations for total nitrogen (TN) and 53 total phosphorus (TP) (Directive Council of the European Union, 1991, 1996; Ministry of 54 Ecology and Environment of the People’s Republic of China, 1996; USEPA, 2002; MOE 55 Ontario, 2008) . 56 In many countries , waste stabilization ponds, also named lagoons , remain one of 57 the most common type s of WRRF s. Currently, over 8,000 , 3,000, 2,500 and 1,200 58 lagoons are in operation in the U .S., Germany, France and Canada, respectively. (Mara, 59 2009; USEPA, 2011; Statistics Canada, 2016) . Lagoons typically consist of single or 60 multi -celled basins that hold wastewater for extended periods (7 to 30+ days) (USEPA, 61 2000; Delatolla and Babarutsi, 2005; Bruce, 2 009). They can be operated in non -aerated 62 (facultative) or aerated configurations, and lagoon -type WRRF installation s often combine 63 several types of cells to achieve the desired treatment level. Lagoons are designed to 64 reduce the concentration of total suspended solids (TSS) , carbonaceous biochemical 65 oxygen demand (CBOD) and, at some installations, total ammonia nitrogen (TAN) , prior 66 4 to their discha rge to receiving waters (Asano et al., 2007) . Due to their simplicity and low 67 operational intensity, lagoons remain one of the preferred methods of treating municipal 68 wastewaters in smaller communities where general land availability allows for their 69 construction (Mittal, 2006; Muga and Mihelcic, 2008) . 70 Despite th e extensive installation and use of lagoons in many countri es, these 71 systems are subject to several drawbacks and limitation s. The n utrient treatment removal 72 capabilities of lagoons situated in northern climates often become s significantly limited 73 due to temperature -induced decrease s in the metabolic activity of the microbial 74 population s (Andreottola et al., 2000; Wessman and Johnson, 2006; Delatolla et al., 2009, 75 2010, 2012; Hoang, 201 3; Hoang et al., 2014; Ragush et al., 2015; Young et al., 2016a) . 76 During fall and winter in temperate and northern climates, waste water temperature s in 77 lagoons can often fall below 1.0°C (Heaven et al., 2003; Krkosek et al., 2012; Ahmed et 78 al., 2019) . Cold-temperature driven loss of bacterial ly-mediated nitrification in lagoons 79 located in northern climates has often been shown to cause effluent TAN concentrations 80 to substantially increase (Painter and Loveless, 1983) , leading to the harmful release of 81 ammonia to receiving waters (Painter and Loveless, 1983) . Additionally, r emoval of 82 phosphorus in lagoons is typically performed via the chemical addition of coagulants or 83 lime (Narasiah et al., 1994) , as few naturally occurring biological pathways exist within 84 these systems that are capable of effectively reduc ing the release of phosphorus. Limited 85 overall performance and specifically limited nutrient removal during low temperature 86 operation has created a need to upgrade these sy stems to maintain compliance with 87 increasingly stringent nutrient regulations , and specifically to remove nitrogenous and 88 phosphor ous compounds prior to discharge . 89 5 Treatment technologies designed to upgrade lagoons should be appropriate in 90 their design to require similar , low level operational intensity ; as lagoon facilities often do 91 not have full -time operators always present. Several technologies have been employed 92 to upgrade lagoons ; such as extended aeration of aerated lagoons (Melcer et al., 1995) , 93 trickling filters (Archer and O’Brien, 2005; Avsar et al., 2008) , rotating biological 94 contactors ( RBC ) (Hassard et al., 2015) , and constructed wetlands (Cameron et al., 2003; 95 Butterworth et al., 2016) . In recent years, studies have outlined the use of rock or 96 aggregate -based attached growth systems (Swanson and Williamson, 1980; Mara and 97 Johnson, 2006; Mattson et al., 2018) , stationary in -lagoon fixed film media (Shin and 98 Polprasert, 1988; Srinivas, 2007; Gan and Champagne, 2015) and moving bed biofilm 99 reactor ( MBBR ) systems (Wessman and Johnson, 2006; Delatolla et al., 2010; Hoang, 100 2013) to upgrade lagoons either by enhancing microbially -mediated nitrification or by 101 increasing the lagoon system’s volumetric or loading capacity . Numerous studies have 102 recently demonstrated the effectiveness of nitrifying attached growth technologies such 103 as the MBBR to enhance TAN-removal performance at the end of lagoons to meet 104 stringent ammonia effluent guidelines at ultra -low (0.6 – 3.0°C) temperatures (Delatolla 105 et al., 2010; Hoang , 2013; Almomani et al., 2014; Young et al., 2016b; Ahmed et al., 2019; 106 Patry et al., 2019) . Post-carbon nitrifying MBBR systems utilize plastic media with a high 107 specific surface area to encourage the adhesion and attachment of nitrifying bacterial 108 communities . The microbial populations in nitrifying MBBR systems have been shown to 109 be highly diversified (Otawa et al., 2006) , allowing nitrifying MBBR system s to operate in 110 a wide range of temperatures (1°C – 20+°C). Biologically active organisms in nitrifying 111 biofilms are ammonia oxidizing bacteria (AOB) and nitrite oxidizi ng bacteria (NOB) 112 6 (Gieseke et al., 2003) . In locales where total nitrogen is regulated, nitrifying MBBR 113 systems can be supplemented with a fol low-up denitrifying MBBR system to convert 114 produced nitrate to nitrogen gas, effectively removing nitrogen from the aqueous 115 environment. The denitrifying MBBR technology has been utilized for over 20 years 116 (Aspegren et al., 1998) however it is almost exclusively applied in larger, mechanical 117 plants, wher e mesophilic con ditions exist (10°C+). In the context of lagoon facility 118 upgrades and retrofits with denitrifying MBBR technology , current available literature do es 119 not address operation below 5°C (Dale et al., 2015) . As a result , a fundamental lack of 120 knowledge remains regarding the capacity of utilizing the MBBR technology to perform 121 total nitrogen removal (nitrification and denitrification) as an add -on technology at a 122 municipal lagoon in temperate and cold climates, where water temperatures below 1°C 123 may routinely be encountered during wintertime . In addition, there is a gap of knowl edge 124 with respect to biologically mediated upgrade systems for lagoons to enhance total 125 phosphorus removal . Ensuring effective total nitrogen and total phosphorus removal in 126 lagoon treatment systems is necessary to ensure the perenniality of water resources in 127 various countries (Brettle et al., 2016; Statistics Canada, 2018) and as such feasible and 128 appropriate nutrient removal upgrade technologies are urgently needed as low 129 operational intensity add -on technologies for lagoon WRRFs . 130 The aim of this study is to investigate the performance of a nitrifying and denitrifying 131 multi -reactor MBBR system installed as an upgrade system to facultative lagoons 132 operating at ultra -low temperatures ( 0.6 - 3.0°C), quantify the removal efficiency of TAN 133 and NO x within the multi -reactor MBBR treatment system at ultra -low temperatures, and 134 to understand the inhibition of nitrification within facultative lagoon systems during ultra - 135 7 low temperature operation . Furthermore, the study also investigate s phosphorus 136 assimilation in a nitrifying -denitrifying upgrade MBBR system as an additional benefit to 137 the operation and as a potential ly significant pathway for total phosphorus removal from 138 lagoon effluent. 139 8 2. Materia ls and methods 140 Study site 141 A pilot-scale wastewater treatment plant was installed at the Casselman, Ontario, 142 Canada ( 45°19'26.3"N, 75°04'48.6"W ) municipal facultative lagoon WRRF from Dec. 143 2017 until Jun . 2018 . The facultative lagoon is characterized by high effluent TAN 144 concentrations in wintertime due to seasonal temperature -driven loss of bacterial 145 nitrification. The lagoon system consists of three cells operated in -series . Cells #1 (7.4 146 ha) and #2 (7.2 ha) are operated as facultative cells, but cell #3 (4.1 ha) is equipped with 147 bottom -mounted aerators and is aerated prior to planned discharge to alleviate sulfides 148 and enhance effluent polishing (Figure 1) . Cell #3 was not aerated from day 1 -76. Pre- 149 planned aeration in cell #3 began on day 77 of the study, lasting until the end of April (day 150 125) due to required lagoon discharge . Spring discharge of the WRRF commenced on 151 day 116 of the study and lasted approximately 30 days . The lagoon is typically discharged 152 twice a year, in th e fall and in the spring. The study was separated in three distinct periods 153 of comparable lengths: i) ice -forming period (days 1 -50), ii) full ice -cover period (days 51 - 154 115) and iii) spring thaw period (days 116 -165). 155 Pilot plant configuration 156 The pilot pl ant house d an influent storage vessel and three MBBR reactor s in 157 series . The pilot -scale treatment plant was fed with the effluent of the last cell of the three - 158 cell lagoon treatment system in Casselman, Ontario. The lagoon effluent was pumped 159 from the lagoon and into an influent storage vessel in the pilot via a 0.75 kW submersible 160 9 pump suspended by a b uoy approximately 40 cm below the water surface (near to the 161 surface) provid ing flow to the MBBR treatment train . As such, the treatment train 162 consisted of an influent storage vessel (retention time of less than 10 minutes), followed 163 by three cylindrical 223 L MBBR reactors (total dimensions: 61 cm Ø x 122 cm H) in- 164 series; a nitrif ying MBBR reactor followed by a second nitrif ying MBBR reactor followed 165 by a denitrifying MBBR reactor , referred to in this manuscript as a NIT -NIT-DENIT 166 configuration (Figure 1). This choice of configuration (NIT-NIT-DENIT) was selected due 167 to concerns of potential sulfide toxicity to the nitrifying microbial communities, and due to 168 intrinsic improvements to nitrifying removal efficiency due to the process being rate - 169 limited at cold temperatures . Sizing of the reactors in the pilot unit was based on the 170 kinetics calculated at projected minimum temperatures, the total flowrate and the TAN 171 loading. Carrier movement in the first and second nitrifying MBBR reactors (R1 & R2) 172 were maintained by supplied aeration. The carriers in the denitrifying reactor (R3) were 173 maintained in motion with a mechanical mixer. The airflow in the reactor was adjusted 174 periodically to maintain adequate carrier motion throughout the experiment. The 175 mechanical mixer speed was also adjusted periodically to maintain adequate carrier 176 motion. The effluent of the pilot -scale treatment system was then recirculated to the first 177 cell of the lagoon treatment system. The ca rbon source dosed in the third, denitrifying 178 reactors was HydrexTM 6860 supplied by Veolia Water Technologies (Montreal, Canada). 179 All MBBR carrier media utilized in this study were AnoxTM K5 carrier s (surface area to 180 volume ratio of 800m2/m3). All reactors in this study had a carrier fill percentage of 50% 181 and had an operational HRT of 4.0 hours, based on the design criteria of low temperature 182 10 nitrifying MBBR systems treating lagoon effluent (Young et al., 2017a; Ahmed et al., 183 2019) . 184 Pilot plant start -up 185 Prior to the beginning of the study, all three MBBR reactors in the pilot plant were 186 seeded with carriers harvested from an integrated fixed -film activated sludge (IFAS) 187 facility in Hawkesbury, ON, Canada , to promote rapid inoculation and growth of the 188 nitrifying community (Young et al., 2017b) . The carriers were collected with 20 L HDPE 189 strainers and transported in drained 200 L barrels to the Casselman lagoon facility, 190 approximately 60 kilometres away , and loaded into the MBBR reactors . The three MBBR 191 reactors in the pilot treatment unit were o perated in a continuous flow configuration for a 192 period of one month prior to the start of the study to acclimatize the seeded carriers and 193 were fed with TAN -augmented lagoon effluent with the help of a peristaltic pump injecting 194 Figure 1: Schematic of the Casselman municipal facultative lagoon WRRF and MBBR nitrogen removal MBBR pilot systems. Cell#1 Cell#2 Cell#3Lagoon R3 (Denitrifying )R1 (Nitrifying ) R2 (Nitrifying ) Pilot PlantEffluentInfluent 11 ammonium chloride solution into the first reactor in -series to stimulate the growth of 195 ammonia -oxidizing bacteria . The water quality parameters of the first reactor in the 196 treatment train during the acclimatization period were as follows: TAN ≥ 25.0 mg-N/L, 197 BOD ≤ 30.0 mg/L, DO ≥ 5.0 mg/L . All three reactors demonstrated steady state operati on 198 after one month of operation; with steady state defined as TAN and NO 3- removal rate 199 fluctuations within a ±10% variation of the geometric mean . 200 Water quality analyses 201 Water samples w ere collected from the influent holding tank and each of the 202 reactors, and analyzed at the pilot facility every 2 -5 days over a period of 165 days . 203 Samples which could not immediately be analysed were stored at 4°C for a maximum of 204 24 hours prior to analysis in the laboratory . The following constituents were analyzed : 205 biological oxygen demand (BOD) (5210 B) (APHA, WEF, 2012) , chemical oxygen 206 demand (COD) (SM 5220 D) (APHA, WEF, 2012) , total suspended solids (TSS) (SM 207 2540 D) (APHA, WEF, 2012) , total ammonia nitrogen (HACH Method 8155) (Reardon et 208 al., 1966; Hach Company, 2015) , nitrate (SM 4500 -NO 3- B) (APHA, WEF, 2012) , nitrite 209 (SM 4500 -NO 2- B) (APHA, WEF, 2012) , alkalinity (SM 2320 B) (APHA, WEF, 2012) , total 210 sulfide (SM 4500 -S2-D) (APHA, WEF, 2012) , sulfate (US EPA 375.4) (U.S. Environmental 211 Protection Agency, 1978) and total phosphorus (SM 4500 -P E) (APHA, WEF, 2012) . 212 Dissolved oxygen (DO) and pH were measured using HACH LDO101 and PHC201 213 probes paired with a HACH HQ40d m eter (Loveland, CO). 214 12 Statistical analyses 215 Statistical analyses were performed in order to establish statistical significance (p- 216 value and Pearson’s R) between water quality parameters throughout the study, linear 217 regressions with a 95% confidence interval were performed between parameters of 218 interest . The assumptions associated with a linear regression model were validated. 219 3. Results and discussion 220 Lagoon effluent character istics 221 Domestic wastewater entering the Casselman municipal lagoon flow s through 222 cells #1 and #2, which are facultative, followed by cell #3, which is intermittently aerated 223 prior to seasonal discharge . The effluent of cell #3 fed the pilot treatment plant and was 224 monitored over a period of 165 days, which encompassed the seas onal spring discharge 225 period from the end of December to early June. The following w ater quality parameters 226 (COD, TSS, nitrite, alkalinity, sulfate, and pH) were separated in three different lagoon 227 operation periods : the ice -forming period (days 1 -50), and the full ice-cover period (days 228 51-115) and the spring thaw period (days 116 -165) (Table 1). BOD (data not shown) and 229 TSS (Table 1) concentrations met or exceeded regulatory standards (<25 mg/L cBOD 5 230 and <25 mg/L TSS) (Canadian Cou ncil of Ministers of the Environment, 2009) throughout 231 almost the entire study. During the snowmelt period , COD and alkalinity decreased 232 slightly (Table 1). It is possible that COD and alkalinity concentrations in the third cell of 233 the lagoon may have decreased slightly during the spring thaw period either due to a 234 restart of aerobic carbon removal and nitrifying activity, or potentially due to dilution. Due 235 to the location of the pump used to coll ect water from the lagoon (near the surface), some 236 13 biological processes and concentration gradients of different water quality constituents 237 may not have been fully observed , particularly during the full ice -cover period. 238 Table 1. Average lagoon effluent concentrations of water quality parameters throughout the 239 study period. 240 Parameter Ice-forming period, Average ± SD, (n) Full ice -cover period, Average ± SD, (n) Spring thaw period, Average ± SD, (n) COD (mg DO/L) 59.2 ± 5.6 , (13) 51.1 ± 9.1 , (16) 32.6 ± 5.0, (11) TSS (mg/L) 13.6 ± 3.2 , (5) 11.6 ± 10.7 , (18) 12.4 ± 11.7 , (10) NO 2- (mg-N/L) 0.01 ± 0.01 , (13) 0.00 ± 0.01 , (17) 0.46 ± 0.73 , (11) Alk. (mg CaCO 3/L) 315.3 ± 16.3 , (7) 299.0 ± 41.6 , (16) 203.1 ± 60.0 , (11) SO 42- (mg-S/L) 40.7 ± 5.6 , (5) 42.6 ± 6.2 , (18) 36.9 ± 2.9 , (8) pH 7.9 ± 0.6 , (12) 7.6 ± 0.2 , (19) 7.6 ± 0.4 , (10) 241 Temperature, TAN, nitrate, total sulfides (TS), TP and DO values in the lagoon 242 system demonstrated distinct concentrations during three distinct periods of operation 243 (Table 2): i) ice -forming period (days 1 -50), ii) full ice -cover period (days 51 -115) and iii) 244 spring thaw period (days 116 -165). During the ice -forming period (Days 1 -50), water 245 temperatures in cell #3 were sufficiently low (1.8 ± 0.5°C) to limi t benthic nitrification, 246 leading to lagoon -wide increases in TAN concentrations (peaking at 18.41 mg -N/L) and 247 low nitrate concentrations (max. of 2.09 mg -N/L). TAN concentrations in cell #3 during 248 this first period increased from an initial concentration o f 14.59 mg -N/L at day 1 of the 249 study to 18.41 mg -N/L around day 50 of the study . Furthermore, the lack of observable 250 nitrite build -up throughout the period is indicative that nitratation was likely not suppressed 251 due to toxic effects . 252 253 14 Table 2. Average ± SD of TAN, nitrate, TS and TP concentrations in cell #3 throughout the 254 study. 255 256 As a result the observed data suggests limited nitrifying and denitrifying microbial 257 activity in the lagoon system in wintertime may o ccur due to a temperature -induced 258 (<4.0°C) reduction in microbial activity and low-level sulfide toxicity in the benthos , which 259 is known to be inhibit ory to the nitrifying and denitrifying microbial communities (Joye and 260 Hollibaugh, 1995; Senga et al., 2006; Bejarano Ortiz et al., 2013) . Hypoxic conditions 261 prevailed throughout cell #3 across days 1 to 50. The low DO conditions (0.8 ± 0.6 mg 262 O2/L) were also potentially exacerbated by the production of hydrogen sulfide in the 263 benth ic zone of the lagoon, causing a high sediment oxygen demand (Wang, 1980; Chen 264 et al., 2017; D’Aoust et al., 2018 , Alqaralleh et al., 2019 ). However, TS concentrations in 265 the water column remained low during this period (0.1 4 ± 0.14 mg -S/L). Meanwhile, TP 266 concentrations were stable and saw little fluctuations , with concentrations of 0.38 ± 0.05 267 mg-P/L being observed during this period . 268 During th e ice-cove r period ( days 51 -115), ice and snow covered the entirety of 269 the lagoon and water temperatures averaged 2.1 ± 0.8 °C. TAN concentrations further 270 In-lagoon (Cell #3) water quality parameters Ice-forming period, Average ± SD, (n) Full ice -cover period, Average ± SD, (n) Spring thaw period, Average ± SD, (n) Temperature (°C) 1.8 ± 0.5, (12) 2.1 ± 0.8, (20) 13.4 ± 6.2, (11) TAN (mg -N/L) 17.01 ± 1.62, (14) 19.23 ± 3.52, (18) 12.58 ± 5.60, (11) NO 3- (mg-N/L) 1.32 ± 0.62, (13) 0.33 ± 0.25, (18) 0.42 ± 0.42, (11) TS (mg-S/L) 0.14 ± 0.14, (4) 6.26 ± 6.17α, (10) / 0.07 ± 0.08β, (9) 0.04 ± 0.04, (11) TP (mg -P/L) 0.38 ± 0.05, (4) 0.38 ± 0.05, (4) 0.08 ± 0.03, (11) Dissolved Oxygen (mg O 2/L) 0.8 ± 0.6, (13) 0.3 ± 0.1α, (7)/ 9.0 ± 5.5β, (13) 8.9 ± 4.4, (11) α: concentration prior to start of aeration in cell #3; β: concentration after start of aeration in cell #3 15 increased during this period , reaching an average concentration of 19.23 ± 3.52 mg -N/L 271 and ultimately plateauing at 20.27 mg -N/L between days 94 to 102 (Supplemental Figure 272 1). Seasonal increases in TAN concentrations show strong correla tion to decrease s in 273 water temperature (R=0.651, p<0.02), highlight ing the sensitivity of lagoon nitrifying 274 communities to the wastewater temperature (Van Dyke et al., 2003) . Ice cover completely 275 inhibited aerobic conditions in the top portion of the lagoon water column (Macdonald et 276 al., 1991; Chen et al., 2017; D’Aoust et al., 2017) , 277 The low DO concentrations of the lagoon during ice -cover and the lack of 278 reaeration processes in cell #3 during wintertime caused widespread anaerobic 279 processes to take hold. S pecifically , sulfate -reduction in the lagoon led to a n average TS 280 concentration of 6.94 ± 6.15 mg-S/L across the ice -cover period , peak ing at 17.11 ± 5.66 281 mg-S/L (Supplemental Figure 1) . As ice -cover forms and becomes covered by snow , light 282 and reaeration processes cease d, causing the death and degradation of most lagoon 283 macrophytes , which could potentially provide carbon to sustain sulfate -reduction. With 284 this onset of TS production, a rapid decrease in residual nitrate concentration (1.98 to 285 0.09 mg -N/L) was observed to occur simultaneously with increases in TS concentration s 286 (Supplemental Figure 1) . This correlation between TS and nitrate concentration is 287 believed to be caused by sulfide toxicity to nitrifying microorganisms (Joye and 288 Hollibaugh, 1995; Senga et al., 2006; Bejarano Ortiz et al., 2013) , and may be indicative 289 that sulfide toxicity in the benthic zone further limits nitrification in la goons at low 290 temperatures , beyond effects caused by low temperatures alone . The identifica tion of 291 sulfide benth ic toxicity of nitrifiers in lagoons during low temperatures as a contributing 292 factor to loss of nitrification modifies conventional belief that nitrification in lagoons located 293 16 in northern or temperature climates ceases solely due to temperature inhibition of 294 nitrifying organisms at temperatures lower than 4°C (Sharma and Ahlert, 1977; Painter 295 and Loveless, 1983) . As both nitrification and sulfate -reduction processes occur in the 296 benthos of lagoons, it is plausible that in-sediment sulfide exposure may be a secondary, 297 significant inhibitor of nitrification (Caffrey et al., 2019) in wastewater lagoons. Another 298 possible cause for decreases in nitrate could be due to potential autotrophic denitrification 299 performed by sulfate -reducing bacteria (Shao et al., 2010) , but the authors believe this is 300 less likely due to rela tively low nitrate concentrations . During this period, TP 301 concentrations did not fluctuate significantly, averaging a concentration of 0.32 ± 0.10 302 mg-P/L. 303 Hypoxic conditions in the lagoon continued from day 51 until day 77 (0.3 ± 0.0 mg 304 O2/L), at which poi nt DO concentrations increased significantly due to the planned lagoon 305 cell mechanical aeration. On day 77 of the study, aeration in cell #3 was activated to 306 polish the lagoon effluent in preparation for seasonal discharge to receiving waters during 307 the th aw period . The start of the aeration was governed by the partner municipality and 308 was required due to the limited allowed discharge period outlined in their Environmental 309 Compliance Approval (ECA). The mechanical aeration had the effect of stripping sulfides 310 out of the water column of cell #3. The average DO concentrations in the cell after aeration 311 commenced increasing slowly to 1.6 mg O 2/L by day 92. Average DO concentrations from 312 day 92 to day 115 were of 12.6 ± 0.6 mg O2/L, outlining a lag period of approximately 15 313 days between the beginning of aeration and the cessation of widespread pond hypoxia, 314 possibly due to sulfidic oxygen demand in the lagoon masking immediate effects of 315 17 aeration on DO concentrations. Meanwhile , after stripping sulfides, TS concentrations 316 decreased significantly to 0.0 9 ± 0.09 mg/L. 317 Finally, during the spring thaw period ( days 116 -165), waste water temperatures 318 increased rapidly from 1.3°C on day 116 to 10.3°C on day 123 and 15.1°C on day 130. 319 This rapid increase in waste water temp erature led to rapid dilution effects, causing TAN 320 concentrations to decrease to 9.00 mg -N/L. Nitrate concentration s during this period were 321 low (<2% of TN as NO 3-) until temperatures reached 15°C and above , and subsequently 322 increased to 2.51 mg -N/L (31% of TN as NO 3-). Meanwhile, TS concentrations (0.04 ± 323 0.04 mg -S/L) remained suppressed due to sustained aeration in cell #3 of the lagoon . At 324 the same time, TP concen trations significantly decreased, with an average of 0.08 ± 0.03 325 mg-P/L in the water column of cell #3. The decrease in TP concentration is believed to 326 have occurred due to spring thaw period induced dilution and biochemical effects. 327 Spring thaw caused a reaeration of the lagoon water column due to a significant 328 increase in influent flow rates while simultaneously reallowing wind rea eration pathways 329 (surface oxygen transfer) to increase water column DO. This s ubsequently caused the 330 cessation of widespread hypoxic conditions in cell #3 of the lagoon . In the presence of 331 oxygen, benthic sulfate -reduction was demonstrated to cease (Atkinson et al., 1995) , 332 potentially prevent ing sulfide -induced dissolution of naturally occurring benthic Fe(III) - 333 (oxyhydr)oxides and the reduction of Fe(III) to Fe(II) (Jansson, 1987; Ding et al., 2012; 334 Kumar et al., 2018) , lower ing total phosphorus (and specifically orthophosphate) 335 concentrations in the water column . The presence/absence of sulfides has been 336 recognized in many studies to be a major cause of release/adsorption of phosphorus to 337 18 the water column, due to the interactions between sulfide and other phosphate -containing 338 minerals and compounds (Nürnberg, 1984; Morse et al., 1987; Boström et al., 1988) . It is 339 therefore believed that upon reaeration of the lagoon, surface adsorption of phosphate to 340 newly available ferric reaction sites in the bent hic zone could potentially be a second, 341 minor pathway for phospho rus removal in the lagoon cells during spring -melt. 342 Upgrade MBBR treatment train performance 343 3.2.1. Nitrif ying MBBR reactors 344 The pilot treatment plant installed at the outlet of a multi -cell facultative wastewater 345 lagoon was operated across a 165-day period spanning winter and spring. During the 346 study, the temperature from the pilot plant ’s influent wastewater remained below 5°C for 347 a period of 119 days . The performance of the upgra de MBBR pilot system is separated 348 and presented as three distinct operational periods : i) ice -forming period (days 1 -50), ii) 349 full ice cover period (days 51 -115) and iii) spring thaw period (days 116 -165). These 350 periods were selected due to an interest in delineating treatment performance across 351 influent wastewater character istics entering the pilot system and associating performance 352 of the upgrade pilot treatment train to the lagoon operation . The water quality 353 characteristic results are shown below in Figure 2 across these three periods . 354 19 Ice-forming period (days 1 -50) 355 The temperatures in the two ni trifying reactors in series, R1 and R2 , decreased to 356 an average value of 3.9 ± 1.9°C and 3.4 ± 2.2°C during the ice -forming period of lagoon 357 operation , respectively. The representative period (days 1-50) removal performance in 358 terms of removal percentage and removal rate in R1 and R2 was of 55.6 ± 26.6% (85.0 359 ± 43.0 g-N/m3·d-1) and 74.8 ± 29.8% (49.1 ± 44.1 g-N/m3·d-1). The total treatment train 360 Cell #3 R1 R2 R30510152025TAN (mg/L) Cell #3 R1 R2 R30510152025 Cell #3 R1 R2 R30510152025Days 1-50 Days 51-115 Days 116-165 Cell #3 R1 R2 R30510152025NOx (mg-N/L) Cell #3 R1 R2 R30510152025 Cell #3 R1 R2 R30510152025Days 1-50 Days 51-115 Days 116-165 Cell #3 R1 R2 R3050100150200250300350Alkalinity (mg-CaCO3/L) Cell #3 R1 R2 R3050100150200250300350 Cell #3 R1 R2 R3050100150200250300350Days 1-50 Days 51-115 Days 116-165 Cell #3 R1 R2 R30510152025TN (mg/L) Cell #3 R1 R2 R30510152025 Cell #3 R1 R2 R30510152025Days 1-50 Days 51-115 Days 116-165 Figure 2. Influent and effluent concentrations across three distinct operational periods: ice-forming period (days 1 -50); full ice covered period (days 51 -115) and spring thaw period (days 116 -165)). a) TAN; b) NO x; c) TN, d) alkalinity and e) TSS. a) b) c) d) Cell #3 R1 R2 R3020406080100120TSS (mg/L) Cell #3 R1 R2 R3020406080100120 Cell #3 R1 R2 R3020406080100120Days 1-50 Days 51-115 Days 116-165 e) 20 TAN removal of the nitrifying reactors in -series, R1-R2, was 88.3 ± 18.0% (134.14 ± 35.03 361 g-N·m3·d-1) (Figure 2 a, b, c). This performance is similar to results observed in a previous, 362 non-acclimatized ultra-low temperature MBBR nitrification stud y (Ahmed et al., 2019) . 363 Alkalinity consumption in the two nitrifying reactors is observed in this study and is 364 proportional to the oxidation of ammonia in the reactors (Figure 2 d). TSS concentrations 365 did not increase significantly within the nitrification train . This is consiste nt with previous 366 work , as intrinsic cell yield s for nitrifiers are low (Madigan et al., 2012; Forrest et al., 2016; 367 Young et al., 2016a) . During this period, NO 2- accounted for 11.5 ± 14.2% of NO x in R1 368 and 57.2 ± % of NO x in R2. This difference in NO x is potentially due to the fact that R1 369 had a slightly lower average operating temperature than R2 (3.9 ± 1.9°C and 3.4 ± 2.2°C 370 respectively) , causing nitrite accumulation (Delatolla et al., 2009; Wang and Li, 2015) . It 371 has also been hypothesized that sulfides may also have bled through the reactor 372 treatment train inside the pilot treatment plant, which would have caused low-level 373 inhibition of nitrification rates , but this was not verified. 374 Full ice cover period (days 51 -115) 375 Full ice -cover operation (days 51 -115) did not demonstrate a significant change in 376 average reactor temperature in the two nitrifying reactors in -series, R1 and R2 ( 2.9 ± 377 1.0°C and 3.5 ± 1.5°C respectively). The average TAN percent removal and removal rate 378 in R1 and R2 was of 45.8 ± 27.8% (54.2 ± 33.2 g-N/m3·d-1) and 41.2 ± 27.4.0% (24.7 ± 379 21.0 g-N/m3·d-1). The total treatment train performance of both nitrifying reactors in series 380 (R1 + R2 ) was slightly lower than the ice-forming period (day 1 -50), with TAN percent 381 removal of 70.5 ± 20.2% (66.6 ± 43.2 g-N/m3·d-1). Nitrification in the first nitrifying reactor 382 21 (R1) was slightly lower in the full ice cover period (days 51 -115) as compared to the ice- 383 forming period (days 1 -50). Water t emperatures did not change significantly between the 384 ice-forming and full ice cover periods , and it is therefore believed that the slight decrease 385 in observed TAN removal rates in the first nitrifying reactor (R1) may have been caused 386 by hydrogen sulfide bleed -through into R1, impacting the performance of the TAN 387 removal . No notable chang es were observed in TSS or alkalinity concentrations within 388 the nitrification train. During this period, NO 2- accounted for 60.2 ± 21.9% of NO x in R1 389 and 57.2 ± 21.4% of NO x in R2. The nitrite accumulation is believed to occur because of 390 NOB temperature inhibition (Wang and Li, 2015) . Another possi ble contributing factor 391 could be that low -level sulfide bleed -through preferentially inhibit s the (NOB) compared 392 to the ammonia -oxidizing -bacteria (AOB) (Erguder et al., 2008) in nitrifying biofilm 393 communities . 394 Spring thaw period (days 116 -165) 395 The spring thaw operation period of the study was characterized by increasing 396 water temperatures and thawing of lagoon ice cover. Average temperatures in the 397 nitrifying treatment train ( R1 + R2) rapidly increased from day 116 to day 125, from 3.1°C 398 to 12.5°C and 3.1°C to 12.7°C, re spectively. Temperatures further increased until the end 399 of the spring thaw period , reaching 20.6°C in R1 and 20.7°C in R2. The spring thaw period 400 removal performance increased in the nitrifying treatment train ( R1 + R2) as compared to 401 the full ice cover p eriod , demonstrated by higher TAN percentage removal and TAN 402 removal rates of 92.1 ± 7.7% (65.1 ± 16.0 g-N/m3·d-1) and 46.8 ± 28.7% (3.3 ± 4.1 g- 403 N/m3·d-1) during the spring thaw period . With increasing temperatures and decreasing 404 22 reactor loading due to drop in influent TAN concentrations , the majority of TAN was 405 removed by the first nitrifying reactor in -series ( R1), with the second nitrifying reactor in - 406 series ( R2) not performing signif icant removal. The spring thaw period was also 407 characterized by icemelt and snowmelt , leading to some potential dilution effect s on some 408 of the water quality parameters, including TP and alkalinity. The TAN removal 409 performance of the complete nitrifying treatment train ( R1 + R2 ) was of 9 5.7 ± 4.2% (65.6 410 ± 17.8 g-N·m3·d-1). TSS concentrations in R2 increased throughout the period to reach 411 effluent concentrations of 58.5 ± 16.3 mg/L , likely due to the start of reactor starvation 412 and sloughing of some attached biofilm. 413 Effects of temperature on nitrifying treatment train performance 414 Throughout the study, the nitrifying treatment train achieved a TAN percent 415 removal of 82.1 ± 19.6%, with an average TAN effluent concentration of 3.42 ± 3.78 mg/L. 416 The maximum nitrifying removal rate observed during this period was 157.25 g-N/m3·d. 417 During spring -melt, it is possible that the nitrifying treatment train (R1 + R2) may have 418 been TAN limited, especially during the end of the study. Near complete conversion of 419 TAN to NO x in the first nitrifying reactor ( R1) was possible at higher temperatures 420 TAN (mg-N/L) <1.50 1.51 - 2.50 2.51 - 4.00 4.01 - 15.000510152025 Temperature (°C) Figure 3. Ranges of TAN concentrations in the MBBR treatment train, by temperature. 23 observed during the spring thaw period (day 115 -165), however, at lower temperatures 421 only partial nitrification was achieved in the first nitrifying reactor ( R1), with the remainder 422 occurring in the second nitrifying reactor in -series ( R2) (Figure 3). Over the duration of 423 the study, R1’s TAN removal performance was on average 15.6% higher at operational 424 temperatures ranging between 1.5 – 2.5°C versus <1.5°C , outlining noticeable 425 operational differences occurring over the span of less than 2.5°C. At temperatures below 426 1.5°C, the average TAN removal rate of the complete nitrifying treatment train (R1 + R2 ) 427 was 52.34 ± 0.52 g-N/m3·d-1. 428 3.2.2. Denitrif ying MBBR reactor 429 The denitrifying reactor , R3, in the pilot system was installed and operated 430 downstream of the two nitrifying reactors. In an attempt to again id entify the effects of 431 temperature on process performance within the MBBR treatment train, results are 432 separated and presented based on the operational conditions of the upstream lagoon 433 system , over the t hree distinct operational periods: i) ice -forming period (days 1 -50), ii) 434 full ice cover period (days 51 -115) and iii) spring thaw period (days 116 -165). 435 Ice-forming period (days 1 -50) 436 The ice-forming period was characterized by low waste water temperatures. The 437 period average temperature was of 3.2 ± 2.6°C. The ice-forming period NO x removal 438 performance in the denitrifying reactor, R3, was 20.8 ± 23.2% (38.0 ± 49.9 g-N·m3·d-1). 439 During this period, R3 produced alkalinity at an average rate of 1.97 mg as CaCO 3/mg of 440 N denitrified. This is slightly lower than values observed by other in studies of 2.85 – 3.93 441 mg-CaCO 3/mg-N (Jeris and Owens, 1975; Hamlin et al., 2008) , but may be explained by 442 24 the partial denitrification of nitrite and nitrate , hence producing less alkalinity to the system 443 (Asano et al., 2007; Chung et al., 2014) . TSS concentrations increase d slightly in the 444 denitrifying reactor ( R3) due to the relatively high cell yield of denitrif iers compared the 445 nitrifiers (Strohm et al., 2007) , leading to ice-forming period -average effluent TSS 446 concentrations of 39.8 ± 30.5 mg/L. 447 Full ice cover period (days 51 -115) 448 The full ice cover period (days 51 -115) was characterized by no significant change 449 in average reactor temperature in the denitrifying reactor, R3, (4.4 ± 2.0°C). The period 450 NO x removal performance in R3 was 90. 8 ± 18.9% (66.9 ± 26.8 g-N·m3·d-1). During this 451 period, the denitrifying reactor ( R3) produced alkalinity at an average rate of 2.01 mg as 452 CaCO 3/mg of N denitrified. This is slightly lower than values observed by other in studies 453 of 2.85 – 3.93 mg -CaCO 3/mg-N due to the relatively large proportion of NO x being in the 454 NO 2- instead of the NO 3- form in R2 (57.2 ± 21.4% for NO 2-). Similarly to what was 455 observed during the first period, TSS concentrations increased slightly in the denitrifying 456 reactor ( R3), leading to a period -average effluent TSS concentration of 3 6.8 ± 28.3 mg/L. 457 Spring thaw period (days 116 -165) 458 The spring thaw period of the study was characterized b y increasing water 459 temperatures and thawing of lagoon ice cover. The a verage temperature in the 460 denitrifying reactor ( R3) rapidly increased from 1.6°C to 13.0°C between days 116 -125. 461 Temperatures further increased until the end of the period, reaching 20.6°C by the end of 462 the study . The period NO x removal performance in the denitrifying reactor ( R3) was 9 1.4± 463 25 14.7% (62.2 ± 18.3 g-N·m3·d-1). Lagoon icemelt led to a global dilution eff ect of most 464 water quality parameters, including nitrogen and alkalinity. During this period, the 465 denitrifying R3 produced alkalinity at an average rate of 3.68 mg as CaCO 3/mg of N 466 denitrified . Stoichiometrically, heterotrophic denitrification theoretically yields 467 approximately 3.57 mg -CaCO 3/mg-N (Jeris and Owens, 1975; Asano et al., 2007) . The 468 close ness of observed alkalinity yield to theoretical values is explained by the near total 469 proportio n of NO x as NO 3- (82.9 ± 28.4%). TSS concentrations in the denitrifying reactor 470 R3 increased throughout the period to reach effluent concentrations of 60.9 ± 78.0 mg/L 471 TSS. TSS concentrations were notably higher than in the first two periods (ice-forming 472 and full ice cover) , likely due to biofilm thinning in warmer operational temperatures , 473 causing sloughing (Bjornberg et al., 2012) . 474 475 476 477 478 479 Effects of temperature on denitrifying reactor performance 480 26 Throughout the study, there was significant removal of NO x in the denitrifying 481 reactor ( R3), with an average removal efficiency of 94.3 ± 0.1% of NO x at temperatures 482 from <1.5°C to 4°C. At higher temperatures (4.0°C to 15.0°C), removal efficiencies 483 decreased to 80.3 ± 19.0% (Figure 4). This decrease in removal efficiency was due to an 484 upset in the operati on of the denitrifying reactor over a period of approximately one and 485 a half weeks (failure of a carbon source pump tubing). At temperatures below 1.5°C, the 486 average NO x removal rate of R3 was of 69.75 ± 24.49 g-N/m3·d-1. 487 488 489 490 NOx (N-mg/L) <1.50 1.51 - 2.50 2.51 - 4.00 4.01 - 15.000510152025 Temperature (°C) Figure 4. Ranges of NO x concentrations in the MBBR treatment train, sorted by temperature. 27 3.2.3. Phosphorus cycling within reactors 491 TP concentrations within the lagoon’s cell #3, the nitrifying reactors in -series (R1 492 + R2) remained similar throughout the ice -forming (day 1 -50) and full ice cover (day 51 - 493 115) periods , with average TP concentrations of 0.33 ± 0.09 mg -P/L, 0.29 ± 0.08 mg -P/L 494 and 0.29 ± 0.02 mg -P/L, respectively ( Figure 5). The denitrifying reactor, R3, exhibited an 495 interesting trend throughout the study, as the denitrifying org anisms assimilated TP , 496 substantially decreasing R3’s TP concentrations , rarely exceeding 0.10 mg P/L. 497 498 499 This behaviour is well understood and com mon for post - 500 nitrification denitrif ying treatment (Boltz et al., 2012; Mases et al., 2012; Wilson et al., 501 2012) , outlining an additional benefit of deploying denitrifying MBBR systems as an end - 502 of-pipe phosphorus polishing system in constant -discharge or seasonally -discharged 503 lagoon treatment systems due to th e phosphorus requirements of heterotrophic 504 denitrifying bacteria to sustain metabolic activity (Kuba et al., 1996; Aspegren et al., 505 Cell #3 R1 R2 R30.00.10.20.30.40.5TP (mg/L) Cell #3 R1 R2 R30.00.10.20.30.40.5 Cell #3 R1 R2 R30.00.10.20.30.40.5Days 1-50 Days 51-115 Days 116-165 Figure 5. Progression of influent and in -reactor TP concentrations throughout the ice -forming period (days 1 -50), full ice cover period (day 51 -115) and spring thaw period (116 -165). 28 1998) . Meanwhile, minimal decre ases in TP were observed in the two nitrifying reactors 506 (R1 + R2), likely due to the intrinsically low cell yield of autotrophic nitrifying organisms 507 (Gee et al., 1990) . 508 509 510 511 512 513 514 515 516 517 518 519 520 521 522 523 524 525 526 Studies 527 have also demonstrated a similar trend in denitrifying bacterial communities, where they 528 29 are considered phosphorus limited at concentrations between 0.03 to 0.10 mg -P/L. 529 Heterotrophic denitrification in attached -growth systems ha s a relatively high cell yield 530 and requires a minimum TP concentration to occur uninhibited (Nordeidet et al., 1994; 531 deBarbadillo et al., 2014) 532 30 4. Conclusion s and recommendations 533 Following the completion of this study, the conclusions outlined a re as follow: 534 1. The MBBR treatment technology has emerged as a suitable and sustainable upgrade 535 technology for existing lagoon and waste stabilization pond facilities operating in 536 temperate, northern, and cold climate countries . 537 2. Implementation of nitrifying and denitrifying MBBRs in -series at the outlet of a lagoon 538 in a NIT -NIT-DENIT configuration provide d treatment r obustness in terms of both total 539 nitrogen and phosphorus removal at temperatures as low as 0.6°C. 540 3. Throughout the 165 -day study, the lagoon effluent treated by the pilot treatment plant 541 demonstrates that TN concentrations decreased by an average of 69.0 ± 24.5% , 542 where the average TN concentration exiting the pilot treatment plant were 7.60 ± 5.70 543 mg-N/L. 544 4. The nitrifying MBBR reactors achieved TAN removal of 82.1 ± 19.6%, reducing TAN 545 concentrations to an average of 3.42 ± 3.78 mg/L . The denitrifying MBBR reactor 546 achieved NO x removal of 69.5 ± 37.8%, reducing NO x to an average of 3.63 ± 5.44 547 mg/L. 548 5. It is hypothesized that sulfide toxicity in lagoons might be another important limitation 549 to cold -temperature wintertime nitrification, leading to the hypothesis that nitrification 550 in lagoons situated in northern or temperate climates may not only be inhibited by 551 temperature effects, but also by significant sulfide inhibition events. 552 31 6. The use of the denitrification technology treat lagoon effluent demonstrates significant 553 phosphorus removal due to phosphorus assimilation concomitant ly occurring with 554 total nitrogen removal through denitrification . 555 7. Throughout the 165 -day study, the lagoon effluent treated by the pilot treatment plant 556 saw TP concentrations decrease by 74.7 ± 20.1%. In addition, the average TP 557 concentrations exiting the pilot treat ment plant were 0.05 ± 0.02 mg -P/L. 558 8. It is hypothesized that the embedded biofilm microorganisms can possibly perform 559 phosphorus homeostasis and increasing bulk -water phosphorus concentrations in the 560 reactors, as a response to low influent phosphorus stres s induced by cessation of in - 561 lagoon sulfate -reduction and dilution effects. 562 9. The results of the study suggest that nitrifying + denitrifying MBBR systems can be 563 installed at the outlets of traditional facultative lagoons to achieve reliable TAN, TN 564 and TP treatment, thereby providing a pathway for substantially reducing the 565 environmental impact of municipal wastewater lagoon discharge on receiving waters 566 in Canada and other northern communities during wintertime operation . 567 32 Declaration of competing interests 568 The authors confirm that there are no known conflicts of interest associated with 569 this publication and there has been no significant financial support for this work that could 570 have influenced its outcome. 571 Acknowledgements 572 The authors acknowledge the support of the Municipality of Casselman, Veolia 573 Water Technologies Canada and the Natural Sciences and Engineering Research 574 Council of Canada (NSERC) for their supp ort. The authors also thank Mr. Alain 575 Castonguay and Dr. Bradley Young for their continued ai d in field operations and logistical 576 support. 577 Data availability statement 578 The data that support the findings of this study are available from the 579 corresponding aut hor upon reasonable request. 580