Ammonium-Nitrogen (NH₄⁺-N) Removal from Groundwater by a Dropping Nitrification Reactor: Characterization of NH₄⁺-N Transformation and Bacterial Community in the Reactor

Abstract

A dropping nitrification reactor was proposed as a low-cost and energy-saving option for the removal of NH₄⁺-N from contaminated groundwater. The objectives of this study were to investigate NH₄⁺-N removal performance and the nitrogen removal pathway and to characterize the microbial communities in the reactor. Polyolefin sponge cubes (10 mm × 10 mm × 10 mm) were connected diagonally in a nylon thread to produce 1 m long dropping nitrification units. Synthetic groundwater containing 50 mg L⁻¹ NH₄⁺-N was added from the top of the hanging units at a flow rate of 4.32 L day⁻¹ for 56 days. Nitrogen-oxidizing microorganisms in the reactor removed 50.8–68.7% of the NH₄⁺-N in the groundwater, which was aerated with atmospheric oxygen as it flowed downwards through the sponge units. Nitrogen transformation and the functional bacteria contributing to it were stratified in the sponge units. Nitrosomonadales-like AOB predominated and transformed NH₄⁺-N to NO₂⁻-N in the upper part of the reactor. Nitrospirales-like NOB predominated and transformed NO₂⁻-N to NO₃⁻-N in the lower part of the reactor. The dropping nitrification reactor could be a promising technology for oxidizing NH₄⁺-N in groundwater and other similar contaminated wastewaters.

1. Introduction

Groundwater, an important source of domestic drinking water, poses serious environmental and public health concerns when contaminated by pollutants such as ammonium-nitrogen (NH₄⁺-N). Other contaminants include arsenic as well as microorganisms such as Giardia, Campylobacter, Cryptosporidium, Salmonella, and Escherichia coli. NH₄⁺-N groundwater contamination is usually caused by anthropogenic activities [1,2,3,4,5]. Groundwater NH₄⁺-N levels of ≤ 120 mg L⁻¹, 69.8 mg L⁻¹, 10 mg L⁻¹, and 120 mg L⁻¹ have been reported for Australia [6], Vietnam [7], China [8], and the Kathmandu valley of Nepal [9], respectively. These concentrations can vary temporally in areas of intensive groundwater/surface water interactions [1]. The threshold proposed in the World Health Organization (WHO) guideline for drinking water of 1.5 mg L⁻¹ [10] is substantially less than the levels for these different locations. High NH₄⁺-N levels in water supplies lead to unpleasant odors and taste. These high levels are a major cause for reducing the disinfection efficacy of chlorine and other halogens, as well as increasing the risk of pathogen contamination during water treatment and distribution; hence, the levels of NH₄⁺-N in groundwater must be decreased before consumption.

The in situ permeable reactive barrier (PRB) is a promising groundwater NH₄⁺-N removal technology. Several laboratory- and full-scale PRBs have effectively removed NH₄⁺-N from groundwater [6,8,11,12,13]. Although a PRB is cost-effective in the long term, it, nonetheless, requires large-scale construction and incurs a high initial cost [14]. Therefore, low-cost, energy-sparing, simple, and compact alternative technologies are needed to remove groundwater NH₄⁺-N, particularly in developing countries and small rural communities.

In the present study, a dropping nitrification reactor was proposed for the removal of NH₄⁺-N from contaminated groundwater (Figure S1). This configuration simulated a down-flow hanging sponge (DHS) system. The DHS system is a low-cost, energy-saving option that is simple in construction and easy in operation and maintenance compared to the PRB. The DHS reactor was originally developed for the treatment of domestic sewage. DHS systems efficiently reduce organic compounds, chemical oxygen demand (COD), biochemical oxygen demand (BOD), and nitrogen in sewage [15,16,17,18,19,20]; they have also been used for the treatment of several kinds of industrial wastewater and for various other purposes [21]. The DHS reactor is a trickling filter using polyurethane sponge as the microbial carrier. Wastewater trickles down from the top of the sponge under the influence of gravity. In the down-flow process, wastewater is exposed to oxygen in the air; thus, this reactor does not require any external aeration [17,22,23]. The organic compounds and NH₄⁺-N in wastewater can be oxidized by the microorganisms on the surface and in the interior of the sponge. During DHS wastewater treatment and other biological nitrogen removal processes, NH₄⁺-N is subjected to ammonia-oxidizing bacteria (AOB) that generate nitrite-nitrogen (NO₂⁻-N). The latter, in turn, is oxidized to nitrate-nitrogen (NO₃⁻-N) by nitrite-oxidizing bacteria (NOB) [22]. To the best of our knowledge, however, no prior study has investigated the application of DHS for the treatment of NH₄⁺-N-contaminated groundwater containing little or no organic carbon. Here, we hypothesized that a dropping nitrification reactor can effectively remove NH₄⁺-N from contaminated groundwater. This reactor will be a low-cost, low-energy consuming, easily operable, and environmentally friendly option for NH₄⁺-N removal from groundwater.

The aims of this study were to test the potential of the dropping nitrification reactor for the removal of NH₄⁺-N from contaminated groundwater to elucidate the single axis, top-to-bottom nitrogen removal pathway and to identify the characteristics of the microbial community, including AOB and NOB in the reactor.

2. Materials and Methods

2.1. Synthetic NH₄⁺-N-Contaminated Groundwater

Synthetic NH₄⁺-N-contaminated (50 mg L⁻¹ NH₄⁺-N) groundwater (104.5 mg Na₂HPO₄·12H₂O, 17 mg KH₂PO₄, 37.5 mg NaCl, 17.5 mg KCl, 23 mg CaCl₂·2H₂O, 25.6 mg MgSO₄·7H₂O, 236 mg (NH₄)₂SO₄, and 353 mg of NaHCO₃ per L tap water; pH 8.0 ± 0.1) was prepared for this study. The constituents and their concentrations in the synthetic groundwater were determined based on the chemical composition of NH₄⁺-N-contaminated groundwater sampled from the Kathmandu Valley [24].

2.2. Setup of the Laboratory-Scale Dropping Nitrification Reactor

Polyolefin sponges (10 mm × 10 mm × 10 mm; Sekisui Aqua Systems Company, Osaka, Japan) served as the medium for the reactor. Polyolefin sponges are superior to polyurethane sponges in DHS (more durable, firmer shape, and higher load bearing capacity). Approximately 82 sponge cubes were connected diagonally in a nylon thread to form a hanging sponge unit with an effective length of 1 m (Figure S1) and a dry weight of approximately 2.94 g. The sponge units were incubated with 10 L of a synthetic NH₄⁺-N-contaminated groundwater mixed with the activated sludge (10:1, v/v). The activated sludge from a municipal wastewater treatment plant in Kofu, Yamanashi, Japan, served as the bacterial inoculum for the sponge units. The suspension was aerated with a pump at a flow rate of 1 L min⁻¹ for 10 days before setting up the reactor to establish microbial communities in the sponge media. A laboratory-scale reactor was prepared using a stainless-steel frame (1.2 m width, 0.45 m depth, 1.5 m height) and set up in a greenhouse at the University of Yamanashi, Kofu, Yamanashi, Japan. The reactor consisted of four separate hanging sponge units (Figure S1). The reactor was covered with a sun-shielding sheet to prevent microalgal growth.

2.3. Operating Conditions of the Laboratory-Scale Dropping Nitrification Reactor

The groundwater NH₄⁺-N removal reactor was operated for 56 days. The contaminated groundwater was fed on the top of each hanging sponge unit at a flow rate of 4.32 L day⁻¹ (3 mL min⁻¹). Water samples were collected from the influent (0 m from the top of the unit), effluent (1 m from the top), and at intermediate sampling points (0.25, 0.50, and 0.75 m from the top) to monitor NH₄⁺-N, NO₂⁻-N, and NO₃⁻-N concentrations. Influent and effluent samples were collected to measure water temperature, pH, redox potential (ORP), and dissolved oxygen (DO) every day or every alternate day. The sponge material was sampled at 0.00, 0.25, 0.50, 0.75, and 1.00 m from the top of the unit every 2 weeks for microbial community analyses.

2.4. Physicochemical Parameters and Nitrogen Concentration in Water Samples

Water temperature, pH, ORP, and DO of the influent and effluent samples were measured on-site with a pH/Temp meter (AS600, AS ONE Corporation, Osaka, Japan), a waterproof ORP meter (ORP-6041, Custom Corp., Chiyoda-ku, Japan), and a DO meter (PDO-520, Fuso Inc., Chuo-ku, Japan), respectively.

Before determining the nitrogen concentration, water samples were filtered through a membrane filter (polypropylene, 0.45 μm pore size; Membrane Solutions Co. Ltd., Minato-ku, Japan). NH₄⁺-N, NO₂⁻-N, and NO₃⁻-N concentrations were measured in accordance with the Standard Methods for the Examination of Water and Wastewater [25]. The indophenol method was used to determine NH₄⁺-N concentration, and the N-(1-naphthyl) ethylenediamine and UV adsorption (220 and 275 nm) methods were used to determine NO₂⁻-N and NO₃⁻-N concentrations, respectively.

2.5. Kinetics of NH₄⁺-N Removal Along the Single Axis of the Dropping Nitrification Reactor

The NH₄⁺-N profiles along the single axis of the reactor from the top (influent, 0 m) to the bottom (effluent, 1 m) were evaluated according to zero- and first-order kinetic reactions as shown in Equations (1) and (2), respectively:

$$\frac{dC}{dt}=-k_0\times C^0=-k_0$$

(1)

$$\frac{dC}{dt}=-k_1\times C^1=-k_1\times C$$

(2)

where C is the NH₄⁺-N concentration, $\frac{dC}{dt}$ is the rate of change in NH₄⁺-N concentration per unit time, and $k_0$ and $k_1$ are the zero- and the first-order kinetic rate constants, respectively.

2.6. DNA Extraction from Sponge Samples and Quantification of Functional Microbial Genes

The sponge samples (approximately 100 mg wet weight = approximately 7.6 mg dry weight) were cut from the corners of the sponges and used for the microbial community analyses. Microbial DNA was extracted from the sponge samples with NucleoSpin Tissue (MACHEREY-NAGEL GmbH, Duren, Germany) according to the manufacturer’s protocol.

Bacterial 16S rRNA and the ammonia monooxygenase (amoA), nitrite oxidoreductase (nxrA), and nitrite reductase (nirK and nirS) genes were quantified by real-time quantitative polymerase chain reaction (RT-qPCR) based on functional gene-targeted primer sets (Table S1) in a Thermal Cycler Dice Real-Time System II (TaKaRa Bio Inc., Shiga, Japan). Each qPCR assay was conducted in a 25 μL reaction mixture including 12.5 μL SYBR Premix Ex Taq (TaKaRa Bio, Shiga, Japan), 0.5 μM of each forward and reverse primer (Table S1), 2 μL template DNA, and 9.5 μL deionized H₂O. The reaction conditions were as follows: initial denaturation by preheating at 95 °C for 30 s, 40 cycles of 98 °C for 5 s, annealing at the specified temperatures (which varied with primer type; Table S1) for 50 s, and an extension at 72 °C for 1 min, followed by a dissociation stage (95 °C for 15 s, 60 °C for 30 s, and 95 °C for 15 s). A standard curve was plotted for each gene using a synthetic plasmid carrying the target sequence. All qPCRs were conducted in duplicate. The gene abundances in the nitrification unit (copies per unit) were calculated using the gene abundances in the sponge material (copies per mg) and the total weight of sponge materials in the unit (mg per unit).

2.7. Phylogenetic Analysis of the Bacterial Community

The extracted bacterial DNA samples were subjected to Illumina MiSeq 16S rRNA gene sequencing (Illumina, San Diego, CA, USA). The V4 region of the 16S rRNA gene was amplified by PCR with the universal primers 515F (5′-Seq A-TGT-GCC-AGC-MGC-CGC-GGT-AA-3′) and 806R (5′-Seq B-GGA-CTA-CHV-GGG-TWT-CTA-AT-3′). PCR amplicons were sequenced in an Illumina MiSeq Sequencer (Illumina, San Diego, CA, USA). Sequence reads were analyzed with Sickle v. 1.33 (https://github.com/najoshi/sickle), Fastx Toolkit v. 0.0.13.2 (http://hannonlab.cshl.edu/fastx_toolkit/download.html), FLASH v. 1.2.10 (https://sourceforge.net/projects/flashpage/files/), and USEARCH v. 8.0.1623_i86linux64 (https://www.drive5.com/usearch/). In these analyses, contigs were formed, and error sequences and chimeras were removed. All operational taxonomic units (OTUs) were clustered at a cutoff of 0.03 (97% similarity). OTUs with < 1% relative abundance among all sequences in all samples were summed as “others”. Sequencing and sequence-read analyses were conducted at FASMAC (Kanagawa, Japan).

2.8. Statistical Analysis

The mean and standard deviation of the physicochemical parameters and nitrogen concentrations were calculated from the data of four replicates of 1 m long dropping nitrification units. Gene abundances (± SD) were calculated using duplicate experiments for each sample. A t-test was used to compare the pairs of groups for significant differences (P < 0.05). The data were processed in SPSS v. 20 (IBM Corp., Armonk, NY, USA).

3. Results and Discussion

3.1. Changes in Physicochemical Parameters

Changes in pH, ORP, and DO in the influent and effluent samples over 56 days of experimentation are shown in Figure 1. Water temperature in the influent and effluent samples ranged from 22 to 32 °C. The pH of the influent was 8.0 ± 0.1 and that of the effluent ranged from 5.7 to 7.4. The pH of the effluent was significantly lower (P < 0.05) than that of the influent. The pH decrement from the top to the bottom of the reactor indicated that H⁺ ion was released (acidification) as a result of microbial nitrification [26,27].

The ORP measurements in the influent and effluent samples were in the ranges of 109–155 and 127–174, respectively. The DO measurements in the influent and effluent samples were in the ranges of 6.0–8.3 and 6.5–8.5, respectively. The ORP and DO after 1 week of operation were significantly higher (P < 0.05) in the effluent than in the influent. Groundwater was fed from the top to the bottom of the reactor. During downward flow, oxygen diffusion into the groundwater from the air caused an increase in the DO concentration in the groundwater. Similar results have been obtained in the DHS reactors treating anaerobic wastewater [19,28,29]. Elevated DO levels are favorable for microbial nitrification [28].

3.2. Removal Efficiency of NH₄⁺-N and Nitrogen Transformations in the Dropping Nitrification Reactor

Changes in NH₄⁺-N, NO₂⁻-N, and NO₃⁻-N in the influent and effluent samples over 56 days of experimentation are shown in Figure 2. The NH₄⁺-N concentration in the effluent gradually decreased after initiating operation, reached steady state after 1 week of operation, and then stabilized in the range of 15.0–25.2 mg L⁻¹ (50.8–68.7% NH₄⁺-N removal). The reactors achieved a NH₄⁺-N removal steady state within a short period of operation. The total inorganic nitrogen (NH₄⁺-N, NO₂⁻-N, and NO₃⁻-N) in the effluent samples was approximately the same as the influent NH₄⁺-N concentration after 3 days of operation. The decrease in NH₄⁺-N concentration from the influent to the effluent was equivalent to the increase in the NO₂⁻-N and NO₃⁻-N concentrations in the effluent where the latter two compounds accumulated; thus, NH₄⁺-N removal was caused exclusively by biological nitrification in which NH₄⁺-N was oxidized to NO₂⁻-N and NO₃⁻-N. Ammonia volatilization was not significant in the reactor, because pH value was < 9.3 in the synthetic groundwater [30]. For up to 35 days, the NO₂⁻-N concentration was higher than the NO₃⁻-N concentration in the effluent. Notably, after 35 days, the opposite trend was observed; therefore, the oxidation of NH₄⁺-N to NO₂⁻-N was effective and fast after starting operation of the reactor when compared to the oxidation of NO₂⁻-N to NO₃⁻-N.

3.3. Variations in the Abundances of the Functional Microbial Genes Involved in the Nitrogen Transformations

The abundances of bacterial 16S rRNA, amoA, nxrA, nirK, and nirS in the reactor over 56 days of operation are shown in Figure 3. The bacterial 16S rRNA gene abundance ranged from 8.2 × 10¹² to 4.5 × 10¹³ copies per unit. The abundance of amoA significantly increased (P < 0.05) from 1.5 × 10¹⁰ to 3.3 × 10¹¹ copies per unit within the initial 14 days and then reached and maintained a peak at 5.5 × 10¹¹ copies per unit. The abundance of nxrA gradually and steadily increased from 5.2 × 10⁷ to 1.3 × 10¹² copies per unit over 56 days; thus, the AOB population and activity increased faster in the reactor than those of NOB. These changes in amoA and nxrA abundance corresponded to the changes in NH₄⁺-N, NO₂⁻-N, and NO₃⁻-N concentration in the effluent (Figure 2). The AOB population should be larger than that of the NOB in balanced nitrifying systems [31,32]. The NOB population was smaller than the AOB population in wastewater treatment systems, such as in sequencing batch reactors [33] and in combined activated sludge-rotating biological contactor and anaerobic-anoxic-oxide (A2O) processes [34]. NOB growth is relatively slow in the absence of nitrite [35]. Similar to these previous reports, the AOB population increased more quickly in this reactor after the onset of the operation compared to the NOB population.

In contrast, the abundances of nirK and nirS were 4.3 × 10⁹–2.6 × 10¹⁰ and 1.2 × 10¹¹–3.1 × 10¹¹ copies per unit, respectively, and did not change significantly over 56 days. Functional denitrification genes (nirK and nirS) were detected in the sponge media. Nevertheless, no denitrification was observed in this experiment as the reactor was highly aerobic (Figure 1c) and the organic carbons (i.e., electron donors) for heterotrophic denitrification were not included in the synthetic groundwater.

3.4. NH₄⁺-N, NO₂⁻-N, and NO₃⁻-N Concentration and Functional N-Transformation Gene Profiles along the Single Axis of the Dropping Nitrification Reactor

The biweekly NH₄⁺-N, NO₂⁻-N, and NO₃⁻-N concentrations and the abundances of the functional bacterial genes (bacterial 16S rRNA, amoA, nxrA, nirK, and nirS) along the single axis of the nitrification units (from the top (influent) to the bottom (effluent; 1 m)) are shown in Figure 4. The NH₄⁺-N concentration rapidly decreased from the top (0 m) to the middle (0.25–0.75 m) of the nitrification reactor. The NO₂⁻-N concentration sharply increased from the top to the middle and then decreased gradually. The NO₃⁻-N concentration gradually increased from the top to the bottom of the sponge units. The abundance of amoA was higher at the top than at the bottom. In contrast, the abundance of nxrA in the sponge units after 14 days of operation was higher in the middle to the bottom parts than in the top (Figure 4). Ammonium oxidation almost occurred in the upper parts, while nitrite oxidation almost occurred from the middle to the bottom of the units.

The decrease in NH₄⁺-N concentration along the single axis of the nitrification units more nearly approximated a first-order than a zero-order kinetic reaction (Figure S2). The first-order kinetic constant (k₁) significantly increased (P < 0.05) from 0.35 to 1.55 m⁻¹ within the first 14 days and then ranged from 1.01 to 1.28 m⁻¹ (Figure 5). The first-order kinetic constant for the increase in NH₄⁺-N removal within the initial 14 days corresponded to the increase in amoA gene abundance (Figure 3).

3.5. Bacterial Community Structure Profiles along the Single Axis of the Dropping Nitrification Reactor

Bacterial community compositions at the class and order levels along the single axis of the reactor on the 56th day are shown in Figure 6. At the class level, Betaproteobacteria (37.9% of all classes), Alphaproteobacteria (24.2%), and Gammaproteobacteria (17.2%) predominated at the top of the nitrification reactor, whereas Alphaproteobacteria (16.2%), Betaproteobacteria (15.1%), and Nitrospira (NOB class; 25.6%) predominated at the bottom.

At the order level, Nitrosomonadales (AOB order; 34.5%) and Xanthomonadales (16.6%) predominated at the top of the nitrification unit; however, their abundances decreased to 11.4% and 7.3%, respectively, at the bottom part. On the other hand, the abundance of Nitrospirales (NOB order) increased along with the downward flow starting from the middle and reached 25.7% at the bottom of the unit. Nitrosomonas spp. (Nitrosomonadales) and Nitrospira spp. (Nitrospirales) were the dominant AOB and NOB species, respectively, in aerobic biological wastewater treatment plants [36] and DHS systems [37]. Nitrosomonadales and Nitrospirales members were also detected in the nitrification units of the present study. Moreover, the abundance of Nitrosomonadales-like AOB was relatively higher in the upper part, while the abundance of Nitrospirales-like NOB was higher in the lower part of the nitrification units. These spatial differences in AOB and NOB along the single axis of the nitrification unit were consistent with the profiles for the NH₄⁺-N, NO₂⁻-N, and NO₃⁻-N concentrations and the abundances of amoA and nxrA in the same system (Figure 4). Thus, NH₄⁺-N is efficiently oxidized and transformed to NO₂⁻-N by Nitrosomonadales-like AOB in the upper part, and NO₂⁻-N is oxidized and transformed to NO₃⁻-N by Nitrospirales-like NOB in the middle to lower part. In addition, Nitrobacter-like NOB (Order; Rhizobiales) might be responsible for NO₂⁻ oxidation in the reactor, because Rhizobiales (2.8%-6.3% of all orders) was detected in the reactor. The coexistence of AOB and NOB has been reported for various environments [38,39], wastewater treatment plants [40,41,42], and drinking water treatment plants [43]. AOB and NOB function independently, but their synergistic relationships mutually benefit their growth and activity [44]. AOB and NOB are in close proximity within biofilms, but the active NH₄⁺-oxidizing zone is separate from the NO₂⁻-oxidizing zone at the micrometer colony diameter scale [45]. In the present study, we demonstrated that NH₄⁺ and NO₂⁻ oxidation in the nitrification unit treating contaminated groundwater was distinctly and widely separated from the top to the bottom of the flow.

This study demonstrated the highly efficient NH₄⁺-N removal from groundwater by using the dropping nitrification reactor and revealed the characteristics of the reactor for the first time. In future studies, we will examine the NH₄⁺-N oxidation efficiency of the reactors in series and the operational lifetime of the sponge media. However, NO₂⁻-N and NO₃⁻-N were accumulated in the effluent of the reactor. Excess NO₃⁻-N (>11 mg L⁻¹ for drinking water [10]) has a negative impact on human health; thus, the effluent is still not recommended for drinking purposes. Therefore, we plan to conduct another study of NH₄⁺-N removal from groundwater by combining nitrification with a denitrification or anammox system for complete nitrogen removal.

4. Conclusions

The dropping nitrification reactor consisting of polyolefin sponge hanging units treated synthetic groundwater containing 50 mg L⁻¹ NH₄⁺-N over 56 days at a flow rate of 4.32 L day⁻¹. The groundwater was aerated with atmospheric oxygen as it flowed downwards from the top to the bottom of the units. The reactor removed 50.8–68.7% of the NH₄⁺-N. Nitrosomonadales-like AOB predominated in the upper part of the reactor and transformed NH₄⁺-N to NO₂⁻-N. Nitrospirales-like NOB predominated in the lower part of the reactor and transformed NO₂⁻-N to NO₃⁻-N. The dropping nitrification reactor is a promising technology for the removal of NH₄⁺-N from contaminated groundwater.