# Pathways and efficiency of nitrogen attenuation in wastewater effluent through soil aquifer treatment

Alex Abu, Raúl Carrey, Cristina Valhondo, Cristina Domènech, Albert Soler, Lurdes Martínez-Landa, Silvia Diaz-Cruz, Jesus Carrera, Neus Otero

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Alex Abu, Raúl Carrey, Cristina Valhondo, Cristina Domènech, Albert Soler, et al.. Pathways and efficiency of nitrogen attenuation in wastewater effluent through soil aquifer treatment. Journal of Environmental Management, 2022, 321, pp.115927. 10.1016/j.jenvman.2022.115927. hal-04891124

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Research article

# Pathways and efficiency of nitrogen attenuation in wastewater effluent through soil aquifer treatment

<figure><img src="image_2.png" /></figure>

Alex Abu<sup>a,b,*</sup>, Raúl Carrey<sup>a,b</sup>, Cristina Valhondo<sup>c,d,e</sup>, Cristina Domènech<sup>a,b</sup>, Albert Soler<sup>a,b</sup>, Lurdes Martínez-Landa<sup>f,g</sup>, Silvia Diaz-Cruz<sup>e,g</sup>, Jesús Carrera<sup>e,g</sup>, Neus Otero<sup>a,b,h</sup>

## ARTICLE INFO

Keywords:

Isotopic fractionation

Enhanced biological denitrification

Permeable reactive barrier

Soil aquifer treatment

Nitrification

Denitrification

## ABSTRACT

Soil Aquifer Treatment (SAT) is used to increase groundwater resources and enhance the water quality of wastewater treatment plant (WWTP) effluents. The resulting water quality needs to be assessed. In this study, we investigate attenuation pathways of nitrogen (N) compounds (predominantly NH₄⁺) from a secondary treatment effluent in pilot SAT systems: both a conventional one (SAT-Control system) and one operating with a permeable reactive barrier (PRB) to provide extra dissolved organic carbon to the recharged water. The goal is to evaluate the effectiveness of the two systems regarding N compounds by means of chemical and isotopic tools. Water chemistry (NO₃⁻, NH₄⁺, Non-Purgeable Dissolved Organic Carbon (NPDOC), and O₂) and isotopic composition of NO₃⁻ (δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻) and NH₄⁺ (δ¹⁵N–NH₄⁺) were monitored in the inflow and at three different sections and depths along the aquifer flow path. Chemical and isotopic results suggest that coupled nitrification-denitrification were the principal mechanisms responsible for the migration and distribution of inorganic N in the systems and that nitrification rate decreased with depth. At the end of the study period, 66% of the total N in the solution was removed in the SAT-PRB system and 69% in the SAT-Control system, measured at the outlet of the systems. The residual N in solution in the SAT-PRB system had an approximately equal proportion of N–NH₄⁺ and N–NO₃⁻ while in the SAT-Control system, the residual N in solution was primarily N–NO₃⁻. Isotopic data also confirmed complete NO₃⁻ degradation in the systems from July to September with the possibility of mixing newly generated NO₃⁻ with the residual NO₃⁻ in the substrate pool.

## 1. Introduction

Nitrogen (N) pollution has become an important threat to water resources (Ren et al., 2014; Shi et al., 2019). Anthropogenic activities involving intensive fertilizer application and animal waste disposal represent the main global sources of N contamination (Arauzo, 2017; Zirkle et al., 2016). Furthermore, domestic, and industrial wastewater and septic-system effluents also serve as a significant source of N and

contaminants of emerging concern (CECs). In manure and wastewater, N is mainly found as ammonium (NH₄⁺) which, under aerobic conditions, is oxidized to nitrate (NO₃⁻) (nitrification) as

$$NH_4^+ + 2O_2 \rightarrow NO_3^- + 2H^+ + H_2O \quad (1)$$

Nitrate contamination impairs its quality (Bourke et al., 2019). This fact has gained global attention in recent years as NO₃⁻ concentrations have been broadly reported above the 50 mgNO₃⁻/L drinking water

<footnote><sup>a</sup> Grup MAIMA, SGR Mineralogia Aplicada, Geoquímica I Geomicrobiologia, Departament de Mineralogia, Petrologia I Geologia Aplicada, Facultat de Ciències de La Terra, Universitat de Barcelona (UB), 08028, Barcelona, Catalonia, Spain</footnote>

<footnote><sup>b</sup> Institut de Recerca de L'Aigua (IdRA), Universitat de Barcelona (UB), 08001, Barcelona, Catalonia, Spain</footnote>

<footnote><sup>c</sup> Université de Montpellier. UMR 5243 Géosciences Montpellier. 300 Avenue Emile Jeanbrau CC MSE. 34095, Montpellier, France</footnote>

<footnote><sup>d</sup> Université de Montpellier. UMR 5569 HydroSciences Montpellier. 15 Avenue Charles Flahault-BP 14491. 34093, Montpellier, France</footnote>

<footnote><sup>e</sup> Institute of Environmental Assessment and Water Research (IDAEA). Severo Ochoa Excellence Center. Spanish National Research Council (CSIC), Jordi Girona 18-24, 08034 Barcelona, Spain</footnote>

<footnote><sup>f</sup> Department of Civil and Environmental Engineering, Universitat Politecnica de Catalunya (UPC), Jordi Girona 1-3, 08034 Barcelona, Spain</footnote>

<footnote><sup>g</sup> Hydrogeology Group (UPC-CSIC), Associate Unit, Jordi Girona, 08034 Barcelona, Spain</footnote>

<footnote><sup>h</sup> Serra Húnter Fellowship. Generalitat de Catalunya, Catalonia, Spain</footnote>

<footnote>* Corresponding author. Faculty of Earth Sciences, Department of Mineralogy, Petrology and Applied Geology, University of Barcelona, Martí i Franquès, s/n, 08028, Barcelona, Spain.</footnote>

<footnote>E-mail address: secretaria.ciencies.terra@ub.edu (A. Abu).</footnote>
threshold value (European Union, 2006; 1991). The ingestion of NO₃⁻ contaminated water is linked to health hazards such as methemoglobi-nemia in infants and young children (WHO, 2004) and stomach cancer (Volkmer et al., 2005). Furthermore, nitrate is usually the limiting factor for algal growth, so increasing its concentration leads to eutrophication of groundwater dependent water bodies (Withers et al., 2014).

The main natural process to remove NO₃⁻ in aquifers is the reduction of NO₃⁻ to dinitrogen gas (N₂(g)) through a sequence of microbial reduction reactions (denitrification) (Korom, 1992; Matchett et al., 2019) (Fig. 1).

$$NO_3^- + 12H^+ + 10e^- \rightarrow N_2 + 6H_2O \quad (2)$$

Natural attenuation of NO₃⁻ can occur in anaerobic conditions in groundwater with oxygen levels below 1–2 mg/L (Singleton et al., 2007). Denitrifying bacteria responsible for these transformations are ubiquitous in surface and subsurface waters (Beauchamp et al., 1989) and they are predominantly facultative anaerobic denitrifying hetero-trophs generating energy for growth and cell synthesis via the oxidation of organic compounds

$$CH_2O + 2NO_3^- + 7H^+ + 6e^- \rightarrow N_2 + HCO_3^- + 4H_2O \quad (3)$$

Autotrophic denitrifiers in groundwater use reduced inorganic compounds such as pyrite, sulfide, sulfite, S⁰, HS⁻, Fe(II), Mn(II) and H₂ as electron donors. Nonetheless, numerous studies have shown that the availability of electron donors is the main limiting factor for denitrifi-cation under natural aquifer conditions (King et al., 2012; Pabich et al., 2001). Other nitrogen attenuation pathways include (1) ammonium (NH₄⁺) oxidation to nitrogen gas (N₂(g)) using nitrite (NO₂⁻) as the elec-tron acceptor in anaerobic conditions (anaerobic ammonium oxidation, or Anammox)) (2) anaerobic respiration by microorganisms using ni-trate as electron acceptor (dissimilatory nitrate reduction to ammonium (DNRA)) and (3) plant uptake (assimilation). The main reactions involving the N transformation pathways are summarized in Fig. 1.

Microbial-mediated reactions tend to alter the distribution of the isotopes between the substrate and product pools. For this reason, stable isotopes of N in compounds such as NO₃⁻ and NH₄⁺ (δ¹⁵N–NO₃⁻, δ¹⁸O–NO₃⁻ and δ¹⁵N–NH₄⁺) have been successfully used to study and characterize natural and induced denitrification processes (Adebowale et al., 2019; Carrey et al., 2013; Margalef-Martí et al., 2019) as well as nitrification processes (Castellano-Hinojosa et al., 2020).

Biologically enhanced denitrification (BED), which involves the biological stimulation of heterotrophic denitrification, has become a strategy to create optimized conditions by providing an external elec-tron donor source in the aquifer (Critchley et al., 2014; Gibert et al., 2019; Margalef-Martí et al., 2019). This technology has evolved over the past decade in an attempt to clean up NO₃⁻ in groundwater to regulatory limits at a low cost. BED has been evaluated in several contexts and

approaches like ex-situ groundwater remediation techniques (Baú and Mayer, 2008) and in situ bioremediation techniques such as in (I) arti-ficial recharge ponds (Grau-Martínez et al., 2018), (II) permeable reac-tive barriers (PRB), reactive media installed perpendicular to flow path (Gibert et al., 2019; Valhondo et al., 2014) and (III) carbon injection systems (Critchley et al., 2014; Margalef-Martí et al., 2019).

Soil Aquifer Treatment (SAT) is a technique often employed to augment groundwater resources and improve water quality in water-stressed zones by promoting the infiltration of effluents from wastewater treatment plants (WWTPs) (Drewes, 2009; Maeng et al., 2010). Conventional wastewater treatment removes a large fraction of the biochemical oxygen demand and suspended solids from wastewater (Jokela et al., 2017; Naidoo and Olaniran, 2013) but is deficient in other water quality indicators such as N and organic compounds, such as CECs (Sonune and Ghate, 2004; Valhondo et al., 2014). N species and CECs are frequent in WWTP effluents (Díaz-Cruz and Barceló, 2008; Kuster et al., 2010; Maeng et al., 2010). The implementation of a PRB based on organic substrates at the bottom of infiltration basins has proven to in-crease the removal of NO₃⁻ and CECs from the recharged water SAT-PRB system using plant-based compost as reactive barriers has proven to successfully promote BED in artificial recharge ponds (Grau-Martínez et al., 2018). The barrier operates by releasing dissolved organic carbon, which acts as the electron donor to stimulate the consumption of diverse electron acceptors, including NO₃⁻ and favouring the CECs’ biotrans-formation, improving the recharged water quality, and reducing NO₃⁻ leakage to the aquifer (Valhondo et al., 2015). The nitrogen cycle and the attenuation of organic contaminants across PRBs must be evaluated during the application SAT-PRB to assess its effectiveness. The rate of denitrification in SAT systems applying PRBs has shown to vary over time and is a function of the NO₃⁻ concentration reaching the reactive media, the residence time of NO₃⁻ within the reactive media and tem-perature (Robertson et al., 2008). Furthermore, wastewater often con-tains N as NH₄⁺ rather than NO₃⁻, affecting the efficiency of the SAT-PRB.

This work evaluates two pilot SAT systems that receive secondary treatment effluent from a WWTP located in the North of Catalonia, Spain. One system has a PRB consisting of plant compost blended with sand and a small fraction of clay (SAT-PRB system, T4) and the other has no PRB (SAT-Control system, T2) Fig. 2. The biochemical processes controlling the migration and speciation of the N compounds in the wastewater effluent are investigated to:

(1) understand the nitrogen cycle during SAT operations, so as to

(2) understand the prevailing processes affecting N in SAT, and to

(3) evaluate the potential of SAT and SAT-PRB to remove N

<figure><img src="image_3.png" /><figcaption>Fig. 1. Biogeochemical processes controlling the fate of N compounds in groundwater. Each coloured arrow indicates a specific N transformation pathway.</figcaption></figure>
<figure><img src="image_4.png" /><figcaption>**Fig. 2.** A) General view of the experimental site, including recharge and discharge areas, piezometer sections and replicates T1 to T6; B) Cross-section and plan view of each system and monitoring points: O, A, B and C are the sampling sections along the flow path, INF = inflow water and E = outflow water; 1, 2 and 3 are the three sampling piezometers at each section (Deep, Middle and Shallow, respectively). A detailed description can be found in (Valhondo et al., 2020a,b).</figcaption></figure>

## 2. Materials and methods

### 2.1. Site description

Six mesoscale SAT systems, named from T1 to T6, were built in the facility of a WWTP (Fig. 2A and B). Each system consists of a 15 × 2.38 × 1.5 m (length × width × height) sand canal, emulating the aquifer, coupled to a 1.5 × 2.38 × 1.15 m box emulating the infiltration basin. The infiltration box was filled with fine sand (φ = 0.1–0.2 mm) in the SAT-Control system (T2) and with a 1 m thick PRB in the remaining tanks. Plants were left to grow naturally at the surface of the infiltration areas and sand canals. Different plants, primarily grama grass, grew immediately after construction. The bottom 40 cm of each system was

filled with coarse sand to favour the vertical flow to the discharge pipe (Fig. 2B). The systems were fed, during recharge episodes, with the effluent of the secondary treatment of the WWTP using PRIUS dosing pumps (EMEC, Rieti, Italy). The elevation of the discharge pipe, which collected the water from the bottom of the system, controlled the head in the outlet.

Ten piezometers were installed in each system for sampling and monitoring three distances from the infiltration area and at three depths. Piezometer “O” is a completely screened crosswise piezometer located just below the infiltration basin (Fig. 2B). Three 10-cm screened piezometers (1 at 1.3–1.4 m, 2 at 0.8–0.9 m, and 3 at 0.3–0.4 m depth) were placed in three sections (A, B, and C, at 1.5, 6.5, and 12.5 m from the recharge area, respectively) (Fig. 2B). Additionally, the inlet and outlet
of each system were sampled and monitored frequently. A detailed description of the systems can be found in (Valhondo et al., 2020b).

## 2.2. System performance and sampling collection

Four recharge episodes were performed during 2018. The first one lasted 27 days (from January, 5th through February, 1st) and aimed at conditioning the systems. The second, third and fourth recharge episodes (RE-2, RE-3 and RE-4) lasting 126 days (February to June), 65 days (July to September), and 74 days (October to January 2019), respectively, aimed at studying the performance of the systems.

An average flow rate of 0.4 m/d fed each system in all four recharge episodes. During these episodes, water samples were collected from the inlet, the monitoring points and the outlets of the systems for chemical and isotopic analyses. Field parameters (O₂, pH, Eh, Electrical Conductivity, and temperature) were measured during sampling using a multiparameter probe (YSI, Yellow Spring, OH, USA).

## 2.3. Analytical methods

Samples for Non-Purgeable Dissolved Organic Carbon (NPDOC) and chemical analyses were collected in muffled glass bottles, filtered through 0.22 µm membrane filters, acidified with HCl (for NPDOC analysis) or HNO₃ (for cations determination) and stored at 4 °C. NPDOC was measured with a TOC-VCSH analyser Shimadzu (Kyoto, Japan). Cations were analysed by inductively coupled plasma optical emission spectroscopy (ICP-OES) using an iCAP 6500 instrument (Thermo Fisher Scientific, Massachusetts, USA). NH₄⁺ was analysed using an ORION Ion-Selective Electrode (ISE, Thermo Fisher Scientific, Massachusetts, USA). The main anions were measured by ion chromatography (IC) using a Dionex AQUION (Dionex, Sunnyvale, CA, USA) with an Ionpack AS9 2 × 250 mm column and Na₂CO₃ 9 mM solution as eluent. Samples for NH₄⁺, δ¹⁵N–NH₄⁺, δ¹⁵N–NO₃⁻, and δ¹⁸O–NO₃⁻ analysis were kept frozen until analysed.

## 2.4. Isotopic methods

The nitrogen and oxygen isotopic abundance in nitrate (δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻) were analysed using the cadmium (Cd) and sodium azide (NaN₃) reduction methods (McIlvinn and Altabet, 2005; Ryabenko et al., 2009). The nitrogen isotopic abundance of NH₄⁺ (δ¹⁵N–NH₄⁺) was measured using the sodium hypobromite (NaBrO) and sodium azide reduction methods (Ryabenko et al., 2009; Zhang et al., 2007).

The standards used were the Vienna Standard Mean Oceanic Water (V-SMOW) for δ¹⁸O–NO₃⁻ and δ¹⁸O–NO₂ analyses and AIR (Atmospheric N₂) for δ¹⁵N–NO₃⁻ analyses. To correct the δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ values, three international standards (USGS 32, 34 and 35) and one internal laboratory standard (CCIT-IWS-NO3) were used. To correct δ¹⁵N–NH₄⁺, two international standards (USGS-25 and IAEA-N2) and two internal laboratory standards (CCIT-IWS-NO2 and CCIT-IWS-NH4)

<caption>**Table 1**</caption>

Standards and reproducibility for isotopic analysis. International and laboratory (CCIT) standards used for the normalization of the results.

<table><thead><tr><th>Analysis</th><th>Standard</th><th>Reference value [%o]</th><th>Reproducibility (1σ)</th></tr></thead><tbody><tr><td rowspan="3">δ¹⁵N–NO₃⁻</td><td>USGS-32,</td><td>+180.0</td><td rowspan="3">±1.0‰</td></tr><tr><td>USGS-34,</td><td>-1.8</td></tr><tr><td>USGS-35</td><td>+2.7</td></tr><tr><td rowspan="3">δ¹⁸O–NO₃⁻</td><td>CCIT-IWS</td><td>+16.9</td><td rowspan="3">±1.5‰</td></tr><tr><td>USGS-32,</td><td>+25.3</td></tr><tr><td>USGS-34,</td><td>-27.9</td></tr><tr><td rowspan="3">δ¹⁵N–NH₄⁺</td><td>USGS-35</td><td>+57.3</td><td rowspan="3">±1.0‰</td></tr><tr><td>CCIT-IWS</td><td>+28.5</td></tr><tr><td>USGS-25</td><td>-30.4</td></tr><tr><td></td><td>IAEA-N2</td><td>+20.3</td><td></td></tr><tr><td></td><td>CCIT-IWS-NO2</td><td>-28.5</td><td></td></tr><tr><td></td><td>CCIT-IWS-NH4</td><td>-0.8</td><td></td></tr></tbody></table>

were employed (Table 1). The samples for chemical and isotopic analyses were prepared at the laboratory of the MAIMA-UB research group and analysed at the Centres Científics i Tecnològics of the Universitat de Barcelona (CCiT-UB).

## 3. Results

Results of NH₄⁺, NO₃⁻, NPDOC and O₂ concentration and isotopic data are available as Supplementary information (Table S1).

### 3.1. Chemical and isotopic results of inflow water

The average NH₄⁺ concentration of the inflow water during RE-2, RE-3 and RE-4 varied significantly ranging from 0.93 to 5.54 mM. The measured NH₄⁺ in the inflow water was higher during RE-3 (average of 4.2 mM, SD = 0.7, n = 20) while RE-2 and RE-4 measured similar NH₄⁺ concentrations (Fig. 3). NO₃⁻ in the wastewater was below the detection limit (<0.01 mM) during all recharge episodes. The average O₂ measured was 0.27 mM (SD = 0.14, n = 18), 0.12 mM (SD = 0.09, n = 7) and 0.20 mM (SD = 0.03, n = 2) during RE-2, RE-3 and RE-4 respectively. RE-2 measured maximum and minimum NPDOC concentrations of 2.11 mM and 1.08 mM, respectively with an average of 1.41 mM and RE-3 measured maximum and minimum concentrations of 1.48 mM and 1.10 mM, respectively with an average concentration of 1.31 mM and RE-4 measured an average NPDOC concentration of 0.94 mM.

The δ¹⁵N–NH₄⁺ of the inflow water showed considerable constant values with an average of +9.6‰ (n = 8 and SD = 0.80).

### 3.2. Chemical and isotopic results along the flow path of the systems

The average concentration of NH₄⁺, NO₃⁻, NPDOC and O₂ measured at the sampling points are shown in Fig. 3. RE-2 and RE-4 had similar NH₄⁺ concentrations whereas slightly higher values were observed in RE-3 (Fig. 3) in both systems. For the 3 recharge episodes studied in both systems, NH₄⁺ concentration progressively decreased from O to C and slightly increased from C to the outlet (E) for all recharge episodes. RE-2 showed a slight increase of NH₄⁺ from O to A (Fig. 3). The NO₃⁻ concentrations displayed similar behaviour in both systems; measured NO₃⁻ concentrations were very low at O, A and B and significantly increased from C to the outlet (E) during all the recharge episodes (Fig. 3). NPDOC concentration showed, overall, a decrease from the inflow water, which showed no significant differences between recharge episodes, to the outlet (Fig. 3). The NPDOC concentration in the SAT-PRB increases at O compared to the inflow due to the DOC release from the RB, especially during RE-2 (2.065 ± 0.12 mM) and RE-3 (1.965 ± 0.46 mM) (Fig. 3). NPDOC concentration in the SAT-PRB system decreased progressively from O to E due to the oxidation of DOC. The concentration of O₂ decreased from the inlet (INF) to O in both systems during the three recharge episodes (Fig. 3). The O₂ concentration remained practically the same from the O to C and showed a slight increase from C to E except in RE-4 in both systems (Fig. 3).

### 3.3. Chemical and isotopic results with depth in the systems

Detailed profiles of NH₄⁺, NO₃⁻, NPDOC and O₂ with depth at 1, 2 and 3 points of A, B and C sections are presented in the supporting information (Table S2) and summarized in Fig. 4, where boxplots represent samples belonging to all the different depths at each section for each recharge episode. The concentration of NH₄⁺ displayed an increasing profile with depth in monitoring points of section C of both systems and section A of the SAT-Control system, with an increase and a slight decrease in section B of both systems. NH₄⁺ in the SAT-PRB system at section A which was high at the top showed decreasing and an increasing profile with depth (Fig. 4A). NO₃⁻ concentration showed no important change with depth at all sections in both systems (Fig. 4B). The NPDOC decreased from the top to the bottom at all sections of the
<figure><img src="image_5.png" /><figcaption>**Fig. 3.** Boxplots of NH₄⁺, NPDOC, O₂ and NO₃⁻ aqueous concentration for all recharge episodes along the flow path. A scheme showing sections along the systems is also provided. Concentrations are indicated by the box plots and the outliers are represented by dots. Values in A, B, and C were calculated with the samples of the different depths at each section. INF = inflow water concentration and E is the outflow water concentration. First boxplot at each section (light grey) = SAT-Control system and the second boxplot (dark grey) = SAT-PRB system.</figcaption></figure>

SAT-PRB system while exhibiting no significant decrease at all the sections of the SAT-Control system. The NPDOC in SAT-PRB systems was significantly higher due to the release of DOC from the RB (Fig. 4C). O₂ also slightly decreased from the top to the bottom of the systems at all the sampling points (Fig. 4D). The isotopic composition of NH₄⁺ (δ¹⁵N–NH₄⁺) in the SAT-Control system ranged from +8.4‰ to +29.4‰ where the lowest values were found in sample points O (just below the PRB) and the highest values were observed in sample points B and slight decrease afterwards at C and the outlet (Table S1). δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ ranged from +4.6‰ to +27‰ and from +9.2‰ to +41.9‰, respectively, where the highest values were observed in sample point O and decreased along the flow path with the lowest values observed in the outlet (Table S1). The isotopic composition of δ¹⁵N–NH₄⁺ in the SAT-PRB system ranged from +9.6‰ to +34.7‰ where lower values were found in sample points O and increased along the flow path with the highest values observed in sample points C and slightly decreased at the outlet (Table S1). δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ ranged from +5.6‰ to +22.9‰ and from +10.5‰ to +38.2‰, respectively, where the highest values were observed in sample points A and decreased along the flow path with the lowest values observed in the outlet (Table S1).

# 4. Discussion

## 4.1. Ammonium migration and transformations: spatial variability

The inflow water employed in the systems showed NH₄⁺ concentrations ranging from 0.93 to 5.54 mM, with δ¹⁵N–NH₄⁺ ranging from +8.2‰ to +10.4‰. These values are in agreement with an organic pollution source (Smith and Miller, 2006) as expected since the inflow water is urban wastewater and a wastewater discharge from nearby hospitals. NH₄⁺ concentration in the inflow water showed an annual maximum in summer attaining a peak concentration in September (Fig. 6) which could be attributed to the population increase that doubles during summer months, linked to tourism, which implies a higher sewage input of organic N.

Microbial oxidation of NH₄⁺ to NO₃⁻ (nitrification) appears to be the main transformation mechanism of the inorganic N (NH₄⁺ + NO₃⁻) species in the systems. During nitrification, NH₄⁺ concentration decreases coupled with an O₂ decrease and an increase of NO₃⁻ concentration along the flow path. Considering nitrification as the only reaction that produces the decrease of NH₄⁺ in the systems, an increase of NO₃⁻ is expected following the stoichiometry of the reaction (Eq. (1) and (1) mol of NO₃⁻).
<figure><img src="image_6.png" /><figcaption>**Fig. 4.** The evolution of $NH_4^+$ (A figure), $NO_3^-$ (B figure), NPDOC (C figure) and $O_2$ (D figure) aqueous concentration with depth at A, B and C sections in the systems (Concentrations are indicated by the box plots and the outliers represented by the dots plotted using all temporal series. First boxplot at each depth (light grey) = SAT-Control system and the second boxplot (dark grey) = SAT-PRB system.</figcaption></figure>

<figure><img src="image_7.png" /><figcaption>**Fig. 5.** Comparison of the % distribution of inflow $NH_4^+$, residual $NO_3^-$, and nitrogen removed along the flow path in the systems for all depth profiles during each recharge episode for SAT-Control (top) and SAT-PRB (bottom) systems.</figcaption></figure>
However, the measured NO₃⁻ concentration in the field does not match this stoichiometry. Total N measured in water is below the expected value, suggesting that other processes may be responsible for N removal. The deficit in N (N<sub>deficit</sub>) along the flow path in the systems (Fig. 5) can be calculated as the difference between the total N in the inflow water and the residual N dissolved in solution (Eq. 4).

$$N_{\text{deficit}} = N - \text{NH}_4^+ \text{INF} - (N - \text{NH}_4^+ \text{res} + N - \text{NO}_3^- \text{res}) \quad (4)$$

where N–NH₄⁺NF is the NH₄⁺ concentration in the inflow water (nitrate was negligible in the inflow, Table S1), N–NH₄⁺res and N–NO₃⁻res are the residual concentration of NH₄⁺ and NO₃⁻ in each sample point in the system. The observed N deficit can be explained either by NH₄⁺ consumption or by another process such as sorption of NH₄⁺ (Alshameri et al., 2018; Fidel et al., 2018; Wang et al., 2015) or plant assimilation (Desimone and Howes, 1998; Lusby et al., 1998) or due to contemporaneous production and consumption of nitrates (coupled nitrification-denitrification reaction) (Kim et al., 1997; Lusby et al., 1998; Smith et al., 2006) or Anammox reactions (Burgin and Hamilton, 2007; Castro-barros et al., 2017). In fact, genetic analyses of the microbial communities present in the barriers demonstrate a potential for all these microbially mediated processes to occur (Hellman et al., 2022). Clay is a component of the PRB and it could provide sorption sites for cation exchange. Since no isotopic fractionation is expected during sorption, comparing the e¹⁵N–NH₄⁺ for the different recharge periods cannot confirm or discard its occurrence. Still, sorption will not affect dramatically long-term N removal processes because its main effect is to buffer inflow fluctuations, which is confirmed by the fact that similar e¹⁵N–NH₄⁺ was observed in the SAT-PRB system in all recharge episodes

(Table S1) and explained further in the section dealing with isotopic fractionation.

Plant assimilation could be relevant in the recharge area, especially during plant growth at the beginning of the study. Nevertheless, plant assimilation is reduced once plants are fully grown, and plant decay can also contribute as a source of organic N in the systems. Denitrification can also be important, especially after infiltration, when O₂ concentration is low. The latter hypothesis is confirmed using the δ¹⁵N and δ¹⁸O of NO₃⁻ since the highest values for these parameters were observed at O and A in both the SAT-Control and SAT-PRB systems. This suggests that part of the NO₃⁻ produced by nitrification is quickly reduced due to the low O₂ and high NPDOC. A plot of ¹⁵N vs. ¹⁸O confirms that denitrification is taking place in the systems (Fig. 7 C).

During the RE-2 and RE-3, NH₄⁺ concentration removed along the SAT-PRB system was significantly higher than in the SAT-Control system (Fig. 3A). More efficient nitrification was expected in the SAT-Control system, than in the SAT-PRB system where O₂ is used to degrade its higher NPDOC, thus inhibiting nitrification as reported in previous groundwater and nitrifying biofilm studies (Jie et al., 2009; Ling and Chen, 2005). The unexpected higher removal rate of NH₄⁺ may be linked to sorption but the higher δ¹⁵N–NH₄⁺ values measured in the SAT-PRB system support that nitrification was higher in this system. We attribute this paradox to the NO₃⁻ produced from nitrification of NH₄⁺ is then denitrified upon reaching the PRB due to reducing conditions created at the reactive barrier base (O) as the NPDOC consumes O₂ in the inflow water.

<figure><img src="image_8.png" /><figcaption>Fig. 6. (A) % distribution of residual NH₄⁺ in solution at E; (B) % distribution of residual NO₃⁻ in solution at E; (C) the % of N (NH₄⁺ + NO₃⁻) removed from the solution and the average temperature in both systems. Grey circles represent the SAT-Control system, black circles represent the SAT-PRB system, and orange circles represent the average temperature in both systems. The dotted line represents the point below which we have zero N removal from the systems. (For interpretation of the references to colour in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>
<figure><img src="image_9.png" /><figcaption>**Fig. 7.** Evolution of δ¹⁵N–NH₄⁺ versus [NH₄⁺] in SAT-Control and SAT-PRB systems (A and B). Relationship between δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ during NO₃⁻ degradation in systems, (C).</figcaption></figure>

<figure><img src="image_10.png" /></figure>

## 4.2. Nitrate production and reduction: spatial variability

No significant change in NO₃⁻ concentration along the flow path from O to C was observed but a significant increase at the outlet was detected in both systems. The increase in NO₃⁻ at the outlets (E) could be explained by an increased nitrification rate favored by flux of O₂, which we attribute to the grama grass prevalent in the canals. The isotopic evolution of NO₃⁻ along the flow path showed typical denitrification trends suggesting that this reaction could be taking place. Both nitrification and organic carbon oxidation favour anoxic conditions in the systems; the observed NO₃⁻ reduction could, therefore, be linked to NPDOC consumption providing energy for denitrifying bacteria. Nonetheless, dissimilatory NO₃⁻ reduction to NH₄⁺ (DNRA) might have contributed to NO₃⁻ removal as well. However, the N_deficit (Eq. (4)) observed cannot be explained by DNRA suggesting that denitrification was the principal NO₃⁻ removal process.

At each particular section, maximum NH₄⁺ accumulation was observed at the bottom of the systems (Fig. 4) suggesting that the rate of nitrification decreased with depth. More aerobic conditions prevailing in the shallower part of the systems favored nitrification resulting in higher nitrification occurring at the top of the systems which is consistent with studies on nitrification-denitrification rates reported in the literature (Kim et al., 1997; Smith et al., 2006). The upper points of our systems are in contact with the atmosphere so oxygen diffusion can account for these higher values at the top. Maximum NO₃⁻ was not observed at the top of our systems due to the contemporaneous production and consumption of NO₃⁻. The higher rate of nitrification at the top of the systems produced a higher consumption of O₂ creating anoxic conditions for denitrification to occur. This assumption is consistent with the high concentration of NH₄⁺ and low O₂ observed at the bottom of the systems (Fig. 4).

In the SAT-Control system, NPDOC decreases with depth but the variability of SAT-PRB is due to the release of DOC from the barrier and

the different travel times of the water at the bottom of the systems. The water flow in the bottom of the systems is slower than in the middle and the top of the systems where flow is faster (based on the tracer tests performed data not shown). Therefore, the water flowing through the bottom of the systems has more time for the DOC to degrade.

## 4.3. Temporal variability of NO₃⁻ and NH₄⁺ production and reduction rates at the outlet

It was observed that the processes responsible for N attenuation in the systems were more effective during RE-3 (July to September) as NH₄⁺ concentration decreased from 4.19 mM in the inflow to 0.69 mM at the C section in the SAT-Control system and 0.58 mM in the outlet of the SAT-PRB system (Fig. 3). Low residual NO₃⁻ concentration was measured suggesting a high denitrification rate during RE-3. For example, 31.5% (RE-2), 7.2% (RE-3) and 28.9% (RE-4) of the total N were measured in the SAT-Control system and 18.7% (RE-2), 8.5% (RE-3) and 10.8% (RE-4)) of the total N was measured in the SAT-PRB system at the outlet. Temperatures measured at the infiltration water increased from 12.0 °C to 24.2 °C (March to May) during RE-2, from 26.1 °C to 29.3 °C (July to September) during RE-3 and decreased to 15.4 °C from October to December during RE-4. High N attenuation during RE-3 could be attributed to the high temperature and the rapid acclimatization of the latent denitrifying bacteria at the beginning of RE-3 (Margalef-Martí et al., 2019). The dry periods between the recharge episodes did not significantly influence N attenuation in the systems. At the start of RE-2, RE-3 and RE-4, the latent biomass community might have quickly adapted when recharge resumed.

During 3 days in March 2018, residual N at the outlet (E) of the systems was higher than the total inorganic N (NH₄⁺ + NO₃⁻) in the inflow water (INF) (Fig. 6C). This reflects the sudden drop in inflow concentration due to a storm (a unitary sewage system feeds the WWTP so that wastewater is diluted during rain events). It is therefore assumed
zero N removal in the systems during this period (February 16, 2018 to March 27, 2021) as depicted in Fig. 6C.

Residual inorganic N progressively decreased in the outflow of the systems beyond the March event (dotted line in Fig. 6C). Biomass growth and adaptation in the systems coupled with the increasing temperatures during the warm period may have increased the N removal rates, thereby, decreasing the residual inorganic N along time. The average N removal rate increased from 31% of the total discharged N (RE-2) to 72% (RE-3) and from 38% (RE-2) to 65% (RE-3) in the SAT-Control and the SAT-PRB systems respectively (Fig. 6C). In both cases, the % increase of the N removal rate from RE-2 to RE-3 suggests that temperature played a significant role in the N removal from the systems (Fig. 6C). Still, while acknowledging the impact of temperature, the fact that N removal continued to increase during RE-4 when temperatures dropped, suggests that microbial community growth and adaptation also played a significant role. The NH₄⁺ progressively decreased from 72% to 4% of total N inflow in the SAT-Control system and 67%–16% in the SAT-PRB system while NO₃⁻ decreased from 54% to 27% (0.80 mM) of total N inflow in the SAT-Control system and from 75% to 18% in the SAT-PRB system measured at the outlets at the end of the study period (Fig. 6B). The significant fluctuations observed may be attributed to the composition of the recharge water during the recharge episodes, fluctuations in the inflow concentrations (population seasonality and rainfall), and variations in biomass during non-recharge periods, which affect the prevailing reactions. The residual N measured at the outlet of both systems are very similar. However, differences were observed in the dominant N species measured at the outlet of the two systems and the recharge periods. For the SAT-Control system, during RE-2, NH₄⁺ was the main N form (1% NO₃⁻ and 29% NH₄⁺), during RE-3 an approximately equal proportion of NO₃⁻ and NH₄⁺ was observed (14% NO₃⁻ and 13% NH₄⁺) and finally, during RE-4 NO₃⁻ was the dominant species (27% NO₃⁻ and 4% NH₄⁺) (Fig. 6A and B). For the SAT-PRB system, NH₄⁺ was also the dominant species (2% NO₃⁻ and 32% NH₄⁺) in RE-2, and during both RE-3 and RE-4 an approximately equal proportion of NO₃⁻ and NH₄⁺ was observed (16–18% NO₃⁻ and 13–16% NH₄⁺) (Fig. 6A and B).

## 4.4. Isotopic fractionation

NH₄⁺ and NO₃⁻ transformations are often characterized by isotopic fractionation (Casciotti et al., 2003; Mariotti et al., 1981), which reflects that reaction rates depend on the isotopes (Casciotti et al., 2011). The degree of fractionation is affected by the extent of substrate consumption, and processes that compete for the substrate (Wankel et al., 2007). The magnitude of fractionation is characterized using the Rayleigh distillation equation, which expresses a power law relationship between the evolution of the isotopic ratio (R = δ + 1) and that of the species concentration, or more conveniently.

$$ \mathrm{Ln}\left(\frac{\mathrm{R}_{\mathrm{residual}}}{\mathrm{R}_{\mathrm{initial}}}\right) = \varepsilon \mathrm{Ln}\left(\frac{\mathrm{C}_{\mathrm{residual}}}{\mathrm{C}_{\mathrm{initial}}}\right) \quad (5) $$

where ε is the isotope enrichment factor. This equation can be linearized as

$$ \delta_{\text{residual}} = \varepsilon (1 + \delta_{\text{initial}}) \text{Ln}(\text{C}_{\text{residual}} / \text{C}_{\text{initial}}). $$ Indeed, a linear correlation was observed in both systems between δ¹⁵N–NH₄⁺ and the natural logarithm (Ln) of the residual NH₄⁺ (Fig. 7A). The resulting enrichment factor (ε¹⁵N–NH₄⁺) in the SAT-Control and the SAT-PRB systems was −5‰ (R² = 0.8). That is, the transformation of NH₄⁺ to NO₃⁻ resulted in an increased δ¹⁵N–NH₄⁺ in the residual NH₄⁺ pool. The lighter isotopes (δ¹⁴N–NH₄⁺) are preferentially incorporated into the generated NO₃⁻ resulting in δ¹⁵N–NH₄⁺ enriched residual NH₄⁺ and low δ¹⁵N–NO₃⁻ in the generated NO₃⁻. This is consistent with similar studies of NH₄⁺ migration and reaction in contaminated groundwater where a decrease in NH₄⁺ concentration yielded an increase in δ¹⁵N–NH₄⁺ (Smith and Miller, 2006; Venkiteswaran et al., 2019). ε¹⁵N–NH₄⁺ values reported in the literature for nitrification range from −10‰ to −28‰ in laboratory culture and

field-scale studies (Sebilo et al., 2006). Lower ε¹⁵N–NH₄⁺ (in absolute values) could be explained by some potential contribution of plant assimilation in the systems. Hellman et al. (2022) suggested that the mixed vegetation at the infiltration area contributed to the NH₄⁺ attenuation in the SAT-PRB system at the beginning of the experiment. The assimilation of N–NH₄⁺ by different plant species reported by Emmerton et al. (2001), estimated an increase in δ¹⁵N–NH₄⁺ of 2‰. The study on the uptake of ionic NH₄⁺ by two plant species by Yoneyama et al. (2001) showed that the magnitude of the uptake of NH₄⁺ depended on the concentrations of NH₄⁺ and the plant species with ε¹⁵N–NH₄⁺ ranging from −2 to −8.1‰ and −3.9 and −24.1‰ for low concentrations of NH₄⁺ (0.04–0.2 mM) and −9.4 to −18‰ and −13.4 to −28.9‰ at high concentrations (0.5–4 mM). Consequently, it is somewhat difficult to distinguish between nitrification and plant assimilation isotopically. In the systems, evaluated plant assimilation must be mainly located in the infiltration area and limited to the shallowest portions of the canals, where vegetation is limited and surficial. The vertical transport of atmospheric O₂ into the systems resulted in a high rate of nitrification between sections A, B and C. The diffused O₂ was consumed during nitrification keeping its content similar to the O₂ in the infiltrated water. Note, however, that the conditions (relatively high concentrations of both electron donors and acceptors) favour a highly dynamic biological activity, with relatively fast growth of biofilms and parallel changes in redox conditions (Carrera et al., 2022). Results of δ¹⁵N–NH₄⁺ in our systems showed constant isotopic fractionation from O to the outlets (Fig. 6B). This can be explained by two hypotheses. The first is that assimilation and nitrification have similar fractionation. The second hypothesis is that fractionation related to assimilation is limited and the main process affecting the isotopic composition of δ¹⁵N–NH₄⁺ would be nitrification.

The low δ¹⁵N–NO₃⁻ in the generated nitrate compared to δ¹⁵N–NH₄⁺ can be explained by the preferential incorporation of the lighter isotopes to the nitrate substrate produced by nitrification. The increase in δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ in the systems was associated with a corresponding decrease in the residual nitrate concentration linked to denitrification. This was evidenced as the NO₃⁻ concentration decreased from 1.24 mM to 0.8 mM in the SAT-Control system and from 1.72 mM to 0.52 mM in the SAT-PRB system measured at the outlets (end of the monitoring period). A linear correlation between δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ evolution in the systems is observed with slopes of 0.6 and 0.5 in the SAT-Control and the SAT-PRB systems, respectively (Fig. 7C). The deviation of δ¹⁵N–NO₃⁻:δ¹⁸O–NO₃⁻ from the 1:1 linear relationship can be explained by (I) the variance of the δ¹⁸O of the O atom donors (O₂: H₂O = 1:2) during the oxidation of NH₄⁺ to hydroxylamine NH₂OH to NO₂⁻ and NO₂⁻ to NO₃⁻ (Kumar et al., 1983) (II) O isotope exchange and fractionation (Martin and Casciotti, 2016).

Several authors have reported diverse ε¹⁵N–NO₃⁻ and ε¹⁸O–NO₃⁻ for groundwater studies ranging from −8.6‰ to −22.9‰ and −5.5‰ to −18.3‰ for ε¹⁵N–NO₃⁻ and ε¹⁸O–NO₃⁻ respectively (Carrey et al., 2013; Fukada et al., 2003). However, in the present experiment, the linear relationship typical for microbial nitrate reduction between δ¹⁵N–NO₃⁻, δ¹⁸O–NO₃⁻ and the natural logarithm of the residual NO₃⁻ was not observed (Fig. S3). As there is NO₃⁻ production and reduction at the same time ε¹⁵N–NO₃⁻ and ε¹⁸O–NO₃⁻ could not be modelled as a closed system using the Rayleigh model. This limits the use of isotopes to quantify NO₃⁻ degradation in coupled nitrification-denitrification systems. Fig. 8 shows δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ and the substrate remaining fraction from the present study using literature ε¹⁵N–NO₃⁻ and ε¹⁸O–NO₃⁻ produced by compost (ε¹⁵N–NO₃⁻ = −10.4‰, ε¹⁸O–NO₃⁻ = −9‰) (Grau-Martínez et al., 2017) and low ε¹⁵N–NO₃⁻ and ε¹⁸O–NO₃⁻ of −3‰. Some samples follow the trend expected for denitrification as NO₃⁻ fraction decreases. However, some samples of a very low fraction of residual NO₃⁻ showed lower δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ close to initial isotopic composition. This can be explained considering that in some areas of the systems, denitrification was able to completely remove NO₃⁻ and, after that, nitrification continues producing new NO₃⁻ that would have lower
<figure><img src="image_11.png" /><figcaption>**Fig. 8.** Isotopic fractionation for ε¹⁵N–NO₃⁻ = -10.8‰ (solid line) and -3.0‰ (dashed line) (A) and ε¹⁸O–NO₃⁻ = -9‰ (solid line) and 3.0‰ (dashed line) (B) using the Rayleigh equation. The shaded region shows the evolution of δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ between ε -10.8‰ and -3.0‰ (A) and -9‰ and -3.0‰ (B) normally observed in denitrification studies in similar conditions (Grau-Martínez et al., 2017). Dots and square boxes are δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ experimental data in MAR-PRB and SAT-Control systems. The asterisks show δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ in the inflow (INF) and the box highlights the δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻ values that do not fall between the fractionation (extreme deviation).</figcaption></figure>

δ¹⁵N–NO₃⁻ and δ¹⁸O–NO₃⁻. The linearity of δ¹⁵N and δ¹⁸O and the natural logarithm of the residual NO₃⁻ during reduction can, therefore, be affected by the mixing of newly generated NO₃⁻ with the residual NO₃⁻ in the substrate pool for nitrification-denitrification processes in the present study.

**5. Conclusion**

We studied the removal of dissolved N from N-contaminated inflow water in two SAT mesocosm systems. Conventional SAT and SAT-PRB systems were compared. Results showed that both systems are efficient in removing N since the net N removed reached up to 69% in the conventional SAT and 66% in the SAT-PRB, and this removal seems to grow with time and be temperature-dependent. Contemporaneous production and consumption of NO₃⁻ (nitrification-denitrification) were the principal processes active in the systems. The NH₄⁺ and NO₃⁻ isotopic characterization confirmed that nitrification and denitrification were accompanied by the enrichment of the residual substrate concentration in heavy isotopes. The isotopic fractionation associated with NO₃⁻ degradation could not be obtained in this study due to constant production and reduction of NO₃⁻.

The plant compost in the SAT-PRB system installed to favour CECs removal did not introduce a significant change in the global N removal. The isotopic fractionation calculated for δ¹⁵N–NH₄⁺ was –5‰ in both systems. The residual inorganic N remained in the SAT-Control system predominantly as NO₃⁻ and in the SAT-PRB as equal proportions of NH₄⁺ and NO₃⁻, suggesting that the plant compost in the PRB increased the electron donor source and contributed to the higher rate of denitrification in the SAT-PRB system.

The concentration of NO₃⁻ in inflow water was below the EU nitrate threshold limit (50 mg/L - NO₃⁻) throughout the whole experiment in the SAT-PRB system except at the beginning of the first recharge episode (average = 33.2 mg/L). In the SAT-Control system, the concentration of NO₃⁻ in the inflow water was above the EU nitrate threshold limit during the entire experiment except during the warm periods when the temperature was high (average = 51.4).

The concentration of NH₄⁺ in the inflow water in both systems was above the maximum EU threshold value for NH₄⁺ (5 mg/L - NH₄⁺) throughout the entire experiment (average = 55.0 mg/L in the SAT-PRB system, and 54.5 mg/L in the SAT-Control system). The presence of PRB favored the removal of NO₃⁻ with time. The SAT-PRB had NO₃⁻ below safe water drinking limit, but NH₄⁺ above safe water drinking limit and the SAT-Control system had both NO₃⁻ and NH₄⁺ above safe water drinking limit, rendering water from both systems inappropriate for direct human consumption.

**Funding**

This work has been financed by the following projects: EU project ACWAPUR (Water JPI -WaterWorks 2014 code PCIN-2015-239), PACE-ISOTEC (CGL2017-87216-C4-1-R) financed by the Spanish Government and AEI/FEDER from the UE, the Water JPI (MARadentro-PCI2019-103603), the project RESTORA (CA210/18/00040) from the Catalan Water Agency, and the Consolidated Research Group MAG (2017-SGR-1733), financed by the Catalan Government. IDAEA-CSIC is a Center of Excellence Severo Ochoa (Spanish Ministry of Science and Innovation, Project CEX2018-000794-S).

The EU Water JPI project MARADENTRO (PCI2019-103603- WW2017), the Catalan Research Project RESTORA (ACA210/18/ 00040) and AGAUR (AQU - 2017 SGR 1485).

**Authorship statement**

All authors certify that they have participated sufficiently in the work to take public responsibility for the content, including participation in the concept, design, analysis, writing, or revision of the manuscript. Furthermore, each author certifies that this material or similar material has not been and will not be submitted to or published in any publication before its appearance in the *Journal of Environmental Management*.

Conception and design of study: C. Valhondo, L. Martínez-Landa, R. Carrey, J. Carrer, N.Otero; acquisition of data: R. Carrey, C. Valhondo, L. Martínez-Landa; analysis and/or interpretation of data: A. Abu, R. Carrey, C. Valhondo, L. Martínez-Landa; Drafting the manuscript: A. Abu, R. Carrey; revising the manuscript critically for important intellectual content: C. Valhondo, C. Domènech, A. Soler, S. Diaz-Cruz, L. Martínez-Landa, R. Carrey, J. Carrera, N. Otero; Approval of the version of the manuscript to be published (the names of all authors must be listed): C. Domènech, A. Soler, S. Diaz-Cruz, A. Abu, L. Martínez-Landa, C. Valhondo, R. Carrey, J. Carrera, N. Otero.

**Declaration of competing interest**

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

**Data availability**

Data will be made available on request.
## Acknowledgements

The authors are grateful to Consorcii de la Costa Brava Girona (CCBGi), the staff of the Palamós WWTP for their unconditional help, and the AGAUR-SGR2017-1485 research group. Abu Alex would like to thank The Agency for Management of University and Research Grants of the Generalitat de Catalunya for the Ph.D. grant (2019 FI_B 01059) and the CCiT of University of Barcelona for the analytical support.

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.jenvman.2022.115927.

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