Simultaneous Removal of Soluble Metal Species and Nitrate from Acidic and Saline Industrial Wastewater in a Pilot-Scale Biofilm Reactor

The hydrometallurgical treatment of waste printed circuit boards for the recovery of precious metals generates acidic wastewater containing nitrate, chloride and residual base metals. The scope of this work is the study of a biological treatment process for the concurrent metal sequestering, nitrate reduction and wastewater neutralization. A pilot-scale packed-bed biofilm reactor was set up, inoculated with the strain H. denitrificans and experimentally monitored. The range of operating parameters examined included: (a) nitrate concentration 750–5750 mg/L NO3−; (b) pH 3–8; (c) Cu, Ni, Zn and Fe at 50 mg/L and 100 mg/L; and (d) chloride concentration 5%–10% as NaCl. The presence of metals did not affect denitrification at the concentrations examined. H. denitrificans completely reduced nitrate and the intermediately produced nitrite at elevated chloride levels. Denitrification shifted pH towards circumneutral to alkaline values, where iron, zinc, copper and nickel were sequestered quantitatively from solution via bioprecipitation. The proposed simple, robust and low-cost biological treatment unit is advantageous compared to the conventional wastewater treatment, where metal precipitation is based on chemical neutralization and the problem of nitrate removal remains unresolved.


Introduction
Waste Electrical and Electronic Equipment (WEEE) ranks among the fastest growing waste streams in the world. The global generation of e-waste grew to 53.6 Mt in 2019 with the perspective to reach 74.7 Mt by 2030 (Forti et al. 2020). Among e-waste, printed circuit boards (PCBs) consist of 30% of base and precious metals (copper and iron 10%-20%, gold, silver, platinum) as well as rare earths. Their metal content is much higher than the natural ores, making metal recovery from PCBs a worldwide priority for circular economy (Işıldar 2018). Hydrometallurgical processing can be established in local WEEE recycling industries permitting the decentralized "green" production of pure metals (Tuncuk et al. 2012;Tunsu and Retegan 2016). These processes involve essentially two steps: (a) leaching of metals, which is generally achieved by means of strong acids such as nitric, hydrochloric and their mixtures (aqua regia); and (b) separation/ selective extraction of the metals of interest from the leachate. Thus, the effluents from the separation and recovery stages contain residual anions such as NO 3 − , Cl − as well as residual metal ions which have not been recovered. Neutralization of these streams resolves the problem of acidity, removes part of the soluble metal ions as hydroxides and part of sulfate as calcium sulfate when Ca(OH) 2 is used as neutralizing agent. However, the concentration of nitrate and chloride does not alter significantly as these anions remain mobile.
Soluble nitrogen compounds, i.e., nitrate or nitrite, can cause serious environmental and human health problems when discharged untreated into water bodies. The established discharge limit for nitrate in drinking water is 50 mg/L NO 3 − (or 11 mg/L NO 3 − -N), and for nitrite 5 mg/L NO 2 − (Ward et al. 2018). Removal of chloride, nitrate and sulfate with conventional treatment options, such as reverse osmosis and ion exchange, is costly. Biological denitrification is a promising, cost effective alternative process which has proven efficiency and selectivity (Matějů et al. 1992). It is based on the ability of the microorganisms to use nitrate and nitrite as electron acceptors instead of molecular oxygen, and involves the reduction of nitrate to nitrite and ultimately to nitrous oxide and nitrogen gas via four enzymatic steps (Reactions 1-4): In the first step (Reaction 1), the reduction of nitrate to nitrite, is catalyzed by a molybdenum-containing nitrate reductase. Nitrite is further reduced to nitric oxide (Reaction 2) that can be catalyzed by either cytochrome cd1 or copper-containing enzymes.
In the third step (Reaction 3), nitric oxide is reduced to nitrous oxide, a greenhouse gas. This is accomplished by a nitric oxide reductase that contains heme c, heme b and nonheme iron cofactors. Finally, a copper-containing enzyme (nitrous oxide reductase) drives the fourth and last step (Reaction 4) where nitrous oxide is reduced to dinitrogen (Tavares et al. 2006;Pang and Wang 2021). The overall denitrification reaction can be expressed as a single redox (Reaction 5): Heterotrophic denitrification has been demonstrated feasible and efficient for the treatment of industrial wastewater which is characterized by elevated nitrate concentrations and no or low metal content. For example, Fernández-Nava et al. (2008) reported that the biological reduction of nitrate in the presence of calcium, from rinse wastewater generated during the manufacturing process of stainless steel, was feasible by achieving 98% nitrate removal in 7 h although they observed that increased calcium concentrations had negative effect on the denitrification rate. Similarly, successful biological reduction of nitrate has been reported for wastewater from the metal-finishing industry (Gabaldón et al. 2007) and nitrocellulose industry (Ugurlu and Ozturkcu 2018), as well as for simulated mine and mill effluents (Koren et al. 2000) and mine water (Mattila et al. 2007). Since industrial wastewater contains little or no organic carbon and most of the known denitrifiers derive their energy from the oxidation of organic substrates, including single carbon compounds (Lin et al. 2009;Pang and Wang 2021), a suitable carbon source/electron donor must be provided to achieve biological denitrification, such as methanol (Gabaldón et al. 2007;Koren et al. 2000;Fernández-Nava et al. 2008;Mattila et al. 2007), ethanol Zou et al. 2014) or acetate (Constantin and Fick 1997).
In the case of industrial metal-bearing wastewater, where metal sequestering is also essential for the safe discharge of the effluent, the biological activity should be sustained in order to simultaneously achieve denitrification and metal bioprecipitation. The later may occur as a result of the alkalinity produced via the biological activity; insoluble metal hydroxides and carbonates precipitate as pH shifts towards more alkaline values and carbonates are produced due to the oxidation of organic carbon. When sulfates are also fed in the reactor, and the prevailing conditions favor sulfate reduction, metal sulfides may also form. These processes, collectively termed as bioprecipitation, are well documented in the literature (Remoudaki et al. 2003;Tsezos et al. 2007;Kousi et al. 2011). Moreover, in conjunction with bioprecipitation, the metal precipitates formed in the bulk solution may be naturally trapped or chemically absorbed by the Extracellular Polymeric Substances (EPS) which are produced by bacteria as a resistance to toxic compounds (Teitzel and Parsek 2003). EPS encase biofilms and decelerate metal diffusion within them, affecting the overall sequestering of soluble metal species. Furthermore, metal sorption on cells and cellular components is controlled by a variety of physical-chemical mechanisms and interactions, such as ion exchange, complexation, adsorption and diffusion through cell walls and membranes (Flemming 1995;Gadd 2002;van Hullebusch et al. 2003).
However, metals have been reported to inhibit denitrifiers in soil, sediments and activated sludge (Baeseman et al. 2006;Magalhães et al. 2007;Sakadevan et al. 1999;Ochoa-Herrera et al. 2011). It has also been experimentally demonstrated, via batch assays, that the denitrifying activity is affected in the case of synthetic mineral solutions containing Fe (Ramírez et al. 2018;Papirio et al. 2014), Ni (Ramírez et al. 2018;Zou et al. 2014), Co  and Cr (Ramírez et al. 2018), due to the accumulation of undesirable intermediates such as NO 2 − and N 2 O (Ramírez et al. 2018). Special attention has been given to the effect of copper (Woolfenden et al. 2013;Felgate et al. 2012;Papirio et al. 2014;Zhao et al. 2020;Cheng et al. 2019), due to the copper-containing enzyme regulating the reduction of nitrous oxide into nitrogen.
To bypass the potential adverse effects of heavy metals on the nitrogen removal capacity of microbial biomass, recent studies have focused on selected metal-resistant bacteria Sun et al. 2016). Zhang et al. (2019) reported that the inhibitory effect of heavy metals on ammonium removal decreased in the order Ni 2+ > Cr 6+ > Zn 2+ > Cu 2+ , whereas Sun et al. (2016) found that efficient removal of ammonium occurred at 20 mg/L Zn 2+ or 10 mg/L Ni 2+ or 8 mg/L Cu 2+ or 5 mg/L Cr 6+ . Few studies have investigated the potential of biofilm reactors to efficiently remove nitrate from metal-bearing wastewater (Hirata et al. 2001;Zou et al. 2015). Hirata et al. (2001) monitored a fluidized-bed bioreactor treating actual, diluted wastewater generated during the recovery process of precious metals from printed circuit boards and demonstrated that it is feasible to remove soluble nitrogen in the presence of high salinity and metal concentrations. However, their study did not involve any data on pH or the concentration of soluble metal species. Zou et al. (2015) showed that fluidized-bed denitrifying cultures tolerated soluble Ni concentrations up to 500 mg/L; however, nickel remained soluble at concentrations 60-500 mg/L Ni in the feed solution. Thus, the microbially-mediated simultaneous removal of nitrate and metals remains challenging.
Another challenging characteristic of the wastewater to be treated is the high chloride content (i.e., salinity) due to the leaching agents (i.e., hydrochloric acid or aqua regia) commonly used for the hydrometallurgical recovery of metals from WEEE (Tuncuk et al. 2012). High chloride content is expected to have negative effects on microbial activity. Hirata et al. (2001) attempted to overcome these negative effects via long-term acclimation of the microbial culture, as previous studies on the effect of salt concentration showed that the nitrification ability was greatly decreased when salt concentration exceeded 2.3% (Vredenbregt et al. 1997) whereas over 2% of sodium chloride greatly reduced denitrification rate (Yang et al. 1995).
The purpose of the present work is to experimentally investigate and demonstrate the feasibility of biological denitrification of industrial wastewater via studying the effects of key factors such as acidity, salinity and metal content which generally stress bacteria and limit the microbial activity leading to nitrate reduction. The proposed single-stage process should meet three essential goals: (a) effluent neutralization; (b) nitrate removal; and (c) soluble metal species sequestering. To the best of our knowledge, such results have not been yet reported in the literature.
To this end, a pilot-scale packed-bed biofilm reactor was set up with a novel porous packing material which embeds trace elements for supporting the micronutrient requirements of an active biofilm. The bioreactor was inoculated with Halomonas denitrificans, a halophilic denitrifier capable of completely reducing nitrate to elemental nitrogen in the presence of high chloride, nitrate and metal content. The denitrifying capacity of the reactor was tested at extreme conditions of elevated nitrate (≈ 6 g/L NO 3 − ) and chloride (10% w/v NaCl) levels in the presence of 100 mg/L Zn, Cu, Fe and Ni.

Bioreactor Set-up
A pilot-scale packed-bed reactor was selected as the bioreactor configuration for the treatment of wastewater containing nitrate and metal ions. The reactor is shown schematically in Fig. 1(a) and physically in Supplementary Material (SM) Fig. SM1. Such a bioreactor setup was previously tested and found highly efficient for the treatment of acidic, sulphate-and metal-bearing wastewater (Kousi et al. 2007;Kousi et al. 2011). The reactor was made from transparent PVC (height: 50 cm; I.D.: 9.5 cm).
Two types of packing material for biofilm support were used: (a) spherical porous sinteredglass beads of average diameter 1.5-3.0 cm (Biohome Supergravel; Fig. 1b) at the bottom of the column creating a 5 cm high layer; and (b) cylindrical porous sintered-glass pieces of average length 2.5-3.5 cm and average diameter 1 cm (Biohome Ultimate Marine; Fig. 1c) filling the rest 40 cm of the column. At the top of the column, a headspace of 5 cm facilitated the release of the produced gases, providing for safety.
The packing material was produced and provided by Aqua Bio UK Ltd. (https://filterpro.co. uk/). It was selected due to its durability, the low pressure drop, the high porosity (approximately 50% according to the technical data sheet of the supplier), the good wetting properties and the trace element content which facilitates the development of active biofilm by providing the required trace elements for microbial growth (mainly Fe and Mn). The microscopic structure of the packing material is shown in Fig. 1d and e for Biohome Supergravel and Biohome Ultimate, respectively. The relevant EDS spectra at selected sites show characteristic peaks of silicon (Si) and oxygen (O) confirming that both Biohome Supergravel (Fig. 1f) and Biohome Ultimate (Fig. 1g) consist mainly of sintered glass. Both materials are also proprietarily enriched with trace elements, such as aluminum (Al), iron (Fe), manganese (Mn), chromium (Cr), sodium (Na), bromine (Br), calcium (Ca), titanium (Ti) and magnesium (Mg).
In all experiments, the reactor was run upflow in sequencing batch mode, meaning that the effluent returned into a well-mixed continuous stirred tank reactor (CSTR) from where it was fed again by means of a peristaltic pump (Shenchen LabM6). The working volume of the CSTR was 1 L and the liquid volume retained in the bioreactor was 1.7 L, corresponding to 53% void space. Thus, in each batch, a total volume of 2.7 L was treated.
H. denitrificans was first isolated from saline water in Anmyeondo, Korea (Kim et al. 2007). The strain was selected for this study based on two specific characteristics: (a) its halotolerant nature which allows its growth in environments of high salinity, i.e., high chloride content (Miao et al. 2015); and (b) its phylogenetic traits which denote its ability to reduce completely nitrate to nitrogen gas (González-Domenech et al. 2010;Felgate et al. 2012;Wang Fig. 1 (a) Layout of the denitrification unit: (1) feed from the completely stirred tank reactor (CSTR), (2) peristaltic pump, (3) packed-bed bioreactor, (4) reactor outflow, (5/6) sampling ports, (7) gas vent; photos of (b) Biohome Supergravel and (c) Biohome Ultimate packing materials; SEM micrographs for (d) Biohome Supergravel and (e) Biohome Ultimate; EDS spectra at selected sites on the surface of (f) Biohome Supergravel and (g) Biohome Ultimate and Shao 2021). To the best of our knowledge, there are no bibliographic reports that the strain can tolerate elevated concentrations of metal ions or has been used in industrial wastewater treatment; these are investigated in this work for the first time.

Bioreactor Start-Up
Before inoculation, the column, the packing material and the silicon tubing were sterilized by recirculating 70% ethanol for 3 h followed by thorough rinsing with 3 L sterile deionized water. Bacto Marine Broth supplemented with 10% w/v NaCl and 1000 mg/L NO 3 − was used as growth medium for culture growth and transfer. An inoculum of 3 L was used for the first and subsequent inoculations of the reactor. The inoculation process was repeated four times: each phase lasted 5 days during which the reactor was run in sequencing batch mode for establishing an active biofilm on the surface and the pores of the packing material. The reactor was self-running with fresh medium (without microorganisms) after 21 days from the initial inoculation, indicating that an active biofilm was already established in the reactor. Subsequent experiments were performed in non-sterile clean conditions, while the reactor was periodically (on average, monthly) re-inoculated to ensure the dominance of H. denitrificans in the biofilm.

Experimental Set-up
In order to investigate the denitrification capacity, metal tolerance and metal sequestering capacity at high chloride concentration, different operating conditions were tested in the bioreactor. Four groups of experiments were designed and are reported in this work. All experiments were carried out at ambient temperature and in duplicate.
(a) In order to examine the tolerance and the denitrifying capacity of the biofilm, three different levels of initial nitrate concentration were tested: 750 mg/L, 2750 mg/L and 5750 mg/L NO 3 − . The ability of the microbial culture to tolerate and respire only on nitrite was also examined via feeding the reactor with solutions containing 2000 mg/L NO 2 − . (b) The effect of pH on the denitrification process was investigated at initial pH 3, 4, 5, 6 and 8 at initial concentration 1250 mg/L NO 3 − . Due to the acidity of the wastewater to be treated, it is essential to study denitrification at low pH (i.e., 3-5). However, as denitrification progresses, pH shifts towards alkaline values. Thus, it is also important to study denitrification at circumneutral pH (i.e., 6-8). The initial pH of the medium was adjusted by the addition of 0.1 N HCl or 0.1 N NaOH as appropriate. (c) The effect of salinity on the denitrification process was studied at pH 3 (the lowest of the tested pH) by adjusting the concentration of NaCl to 10% w/v (at 1250 mg/L NO 3 − ), 7.5% w/v (at 2750 mg/L NO 3 − ) and 5% w/v (at 3250 mg/L NO 3 − ). (d) To examine the effect of metal ions on the denitrification process as well as the fate of the metals during denitrification, a set of experiments was carried out at pH 3, nitrate concentration 2750-3250 mg/L NO 3 − and 5% w/v NaCl at two different initial metal concentrations: 50 mg/L and 100 mg/L. Four divalent metal ions were added separately (acclimation period 4 days before sampling): Fe 2+ was added as FeSO 4 ·7H 2 O, Cu 2+ was added as CuCl 2 ·2H 2 O, Ni 2+ was added as NiCl 2 ·6H 2 O and Zn 2+ was added as ZnCl 2 .

Sampling and Monitoring of the Reactor
Samples were collected from the reactor outflow (stream returning to the CSTR compartment) hourly for the first 6 h and once every 24 h for each run as appropriate. pH was determined on unfiltered samples. For all subsequent analytical determinations, the samples were vacuum filtered through 0.20 μm sterile membrane filters (Whatman ME 24/21 ST). The concentration of nitrate and nitrite was determined by colorimetry (Merck method 1.09713.0001 for nitrate adjusted to Hach DR/2500 at 340 nm and Merck method 1.14776.0002 for nitrite adjusted to Hach DR/2500 at 550 nm, respectively). The measuring range for nitrate and nitrite was 0-20 mg/L NO 3 − -N and 0.02-1 mg/L NO 2 − -N, respectively. Total organic carbon (TOC) was determined by colorimetry (Hach method 10129, DR/2500, measuring range: 0-20 mg/L TOC). The concentrations of soluble copper, nickel, iron and zinc were determined by flame atomic absorption spectroscopy (AAS) after acidification of the samples by adding HNO 3 . All results were plotted as mean values with error bars representing one standard deviation. The sludge generated within the reactor was examined by Scanning Electron Microscopy (SEM) using a Jeol 6380 LV microscope (accelerating voltage: 15 kV at low vacuum 30 Pa) and a backscattered electron detector. Microanalysis was performed by an Oxford INCA Energy Dispersive Spectrometer (EDS) connected to the SEM. SEM was performed on polished sections, which had been produced by vacuum impregnation of the selected sample in a low viscosity epoxy resin. After removing a small surface by cutting in micro-saw, the sample was grinded and polished with 1 μm diamond paste, on a lapping disk.

Kinetics of Nitrate and Nitrite Reduction
The kinetics of nitrate reduction are presented in Fig. 2a. Complete nitrate removal is attained within 4 h and 6 h for 750 mg/L and 2750 mg/L NO 3 − , respectively, whereas for 5750 mg/L NO 3 − , nitrate is completely removed within 25 h. The parallel monitoring of nitrite concentration verified that nitrate was reduced via the formation of intermediate nitrite. Nitrite profiles (Fig. 2b) represent typical kinetic curves of an intermediately produced and sequentially consumed in a sequence of reactions in series; thus, the peaks of the curves (Fig. 2b) correspond to the time when the nitrite reduction rate was higher than the nitrite production rate. The same profiles also show that the maximum nitrite concentration depends on the initial nitrate concentration: profile peaks at 200 mg/L, 580 mg/L and 1000 mg/L NO 2 − are observed for 750 mg/L, 2750 mg/L and 5750 mg/L NO 3 − , at 2 h, 4 h and 5 h, respectively. Complete nitrite reduction was observed within 4 h for 750 mg/L NO 3 − and less than 25 h for 2750 mg/L and 5750 mg/L NO 3 − . The results demonstrate the high denitrifying capacity of the biofilm reactor inoculated with H. denitrificans. All soluble nitrogen species (i.e., NO 3 − and NO 2 − ) were completely removed from solution within 25 h, in compliance with the WHO established discharge limit, even for initial nitrate concentrations as high as 6 g/L, verifying similar findings for biofilm reactors reported in the literature (Mattila et al. 2007;Hirata et al. 2001).
Monitoring pH vs. time (Fig. 2c) verified that pH shifted towards more alkaline values as the reduction of nitrate to nitrogen progressed via the consumption of protons (Reaction 5). The rise of pH depends on the extent of nitrate reduction: higher initial nitrate concentrations correspond to higher final pH values. Moreover, pH profiles show a consistent pattern: initial pH (6.2-7.2) rose steeply to reach a peak value from where it kept decreasing slowly towards an equilibrium value. The peak values 7.6, 8.3 and 8.6 for an initial concentration of 750 mg/L, 2750 mg/L and 5750 mg/L NO 3 − , respectively, coincide with zero residual nitrate concentration.
The ability of H. denitrificans to grow only by reducing nitrite in the absence of nitrate was investigated by conducting an experiment feeding the reactor with 2000 mg/L NO 2 − as the sole electron acceptor. The results shown in Fig. 2a prove that there is no apparent delay of nitrite reduction which is completely reduced to elemental nitrogen within 6 h. Therefore, it is confirmed that H. denitrificans can grow in the absence of nitrate by using only nitrite as terminal electron acceptor (nitrite respiration). This is consistent with the observed consumption of the nitrite produced as intermediate during nitrate reduction. It should be noted that, when only nitrite was present, the pH of the effluent reached a peak of 8.8 while maintaining the highest observed equilibrium pH value of 8.3 (Fig. 2c). Complete reduction of nitrate to the harmless elemental nitrogen is a significant prerequisite for the successful treatment of any nitrate-containing wastewater. Residual concentrations of nitrite are undesirable after wastewater treatment. It should be noted that only selected microbial species can carry out complete denitrification; several species stop the reduction process at an intermediate product such as nitrite, nitric oxide (ΝΟ) or nitrous oxide (Ν 2 Ο) or start the reduction process from these intermediates (Kim et al. 2007;Holmes et al. 2019;You et al. 2020).
The results from TOC monitoring (data not shown) indicate that about 87% of TOC originates from bacto peptone while the rest 13% from the yeast extract contained in the Fig. 2 Effect of initial nitrate/nitrite concentration on the denitrification process: (a) nitrate/nitrite concentration profiles; (b) produced nitrite concentration profiles; and (c) pH profiles growth medium. TOC degradation reached about 50% in all runs representing the conversion of the organic fraction of peptones and yeast into CO 2 . Thus, it was ensured that the carbon source was never in growth-limiting conditions and did not limit the denitrification process.

The Effect of Initial pH on Denitrification
Although the scope of this study is the treatment of acidic wastewater, the effect of the initial pH on the denitrification process was investigated over a wider range of values. To this end, the effect of the initial pH was studied at pH 3, 4, 5, 6 and 8 (Fig. 3a-f) versus the non-adjusted pH of the feed solution (pH = 6.8). All experiments were carried out with nitrate as electron acceptor at 1250 mg/L NO 3 − . These experiments also confirmed that denitrification progressed through the formation of nitrite and the subsequent reduction of nitrite to nitrogen (Fig. 3a-d). However, the initial pH of the medium significantly affected the concentration of nitrite (Fig.  3c-d). Both nitrate and nitrite were reduced within 4 h for acidic initial pH values. For pH 3, 4 and 5, the maximum nitrite concentration was lower than 75 mg/L, while for the circumneutral and alkaline initial pH values, i.e., 6, 6.8 and 8, nitrite concentration was significantly higher (120 mg/L -200 mg/L) and complete reduction was observed at 24 h. Therefore, it can be concluded that, under acidic pH, lower nitrite maxima are observed; this can be attributed either to the increased rate of nitrite reduction and/or to the decreased rate of nitrate reduction. This result is on the benefit of the proposed process aiming at the treatment of highly acidic streams with short residence times concerning nitrate reduction and low peak nitrite concentrations.
The control experiment at pH 6.8 showed a typical profile with a peak at pH 8 and a tail reaching equilibrium at pH 7.8. When the initial pH of the feed solution was adjusted to 3, 4 and 5, pH increased considerably to an equilibrium value of around 7.5 (Fig. 3e). When the initial pH was adjusted to 6 and 8, the corresponding equilibrium pH was 7.3 and 7.8, respectively (Fig. 3f). These findings are consistent with other results reported in literature for Halomonas species; pure or mixed cultures of Halomonas species retained high denitrification rates over a wide range of pH (Zhu and Liu 2017).
The steep increase of pH, especially in the case of acidic values (i.e., 3, 4 and 5), demonstrates the efficiency of the treatment unit to tolerate and neutralize acidic wastewater originating from WEEE acid leaching processes. It has been experimentally shown that biofilm reactors tolerate feed pH as low as 2.5  or 4.0 (Ugurlu and Ozturkcu 2018) with no adverse effects on denitrification. In the present work also, the established biofilm tolerated acidic feeds (pH = 3) and achieved both aims, namely wastewater neutralization and soluble nitrogen removal, within 4 h (Figs. 3a-d). Furthermore, the inherent increase of pH to 7.5-8.0, during denitrification, affects the solubility and mobility of the present metal ions as it will be discussed in the following.

The Effect of Salinity on Denitrification
The effect of salinity on the denitrification process was studied at three different chloride levels (NaCl: 5% w/v, 7.5% w/v and 10% w/v). These experiments simulated the presence of chloride in the wastewater due to the use of hydrochloric acid/aqua regia as leaching agents as well as the high ionic strength of the solution to be treated due to the presence of various dissolved species (i.e., nitrate, chloride and metal ions).
The results reported in this study prove that the increased salinity did not affect the denitrification process which was completed within 4 h for different initial nitrate concentrations (Fig. 4a). However, Fig. 4b shows that salinity affected the temporal accumulation of nitrite in the medium, implying a direct effect on the physiology of H. denitrificans, as the decrease of NaCl stress may enhance the biodiversity of the denitrifying bacteria (Miao et al. 2015). Accumulation of nitrite was also observed in an anoxic packed-bed reactor that treated metallurgical wastewater as salinity increased (Yoshie et al. 2002). This phenomenon can be attributed to the decreasing diversity of the nitrite reductase genes as salinity increases (Yoshie  Figure 4c indicates that, starting from pH 3, pH increased steeply within 2 h close to neutral values (7.0-7.7) regardless of the salinity level.
These findings are consistent with other results reported in the literature for Halomonas species; H. campisalis was shown to completely reduce nitrate at 12.5% NaCl and pH 9 (Peyton et al. 2001) whereas a recent study testing 14 Halomonas strains in a medium containing 6% NaCl demonstrated that the ability to remove nitrogen from saline media is a common property of Halomonas bacteria .

The Effects of Zn, cu, Fe and Ni on Denitrification
The effect of zinc, copper, iron and nickel on the denitrification process was studied, for each metal individually, at two initial concentrations: 50 mg/L and 100 mg/L. The synergistic effect of these metals was also studied with feed solutions containing 20 mg/L and 50 mg/L Zn, Cu, Fe and Ni ( Fig. SM2 and SM3).
Figure 5a-f present the profiles for nitrate/nitrite concentration and pH vs. time compared to the control experiment which was conducted without any metal addition in the feed solution. Complete nitrate reduction was systematically observed within 5 h for all four metals tested at both concentrations ( Fig. 5a-b) indicating that there is no significant inhibition on the denitrification process due to the presence of the metal ions. This is also indicated by the pH profiles (Fig. 5e-f) which show that, for all metals and at both concentrations, starting from pH 3, pH shifts to 7.2-7.5 within 2 h and equilibrates at a value close to 8.0. However, regarding their effect on the denitrifying physiology of H. denitrificans, the tested metals can be clustered into two groups: (a) zinc and copper; and (b) iron and nickel. Figure 5a-b indicate that, in presence of Zn and Cu, denitrification was completed within 3 h, as in the case without metals addition, while, in presence of Fe and Ni, denitrification was retarded and was completed within 5 h. The effects of these metals are more noticeable on the nitrite concentration profiles (Fig. 5c-d): nitrite concentration was much higher when Zn or Cu was added to the feed solution compared to Fe or Ni; in the latter case, nitrite reduction proceeded as in the case without metals addition. Specifically, for Zn or Cu, peak concentrations of 75 mg/L and 100-150 mg/L NO 2 − were observed (Fig. 5c-d) for 50 mg/L and 100 mg/ L metal concentration, respectively. This indicates that Ni and Fe significantly affect the microbial activity by decelerating nitrate reduction and accelerating nitrite reduction. The opposite behavior is observed in case of Zn and Cu where nitrate reduction rates are high but nitrite reduction rates are low.
Copper and iron are expected to positively affect individual steps of the overall process as they are needed in traces for supporting the catalyzing activity of the enzymes leading to the reduction of nitrite (Reaction 2, Cu-containing reductase), nitric oxide (Reaction 3, Fecontaining reductase) and nitrous oxide (Reaction 4, Cu-containing reductase). Thus, it has been shown that copper stimulates both growth and activity of denitrifying bacteria, resulting in the accumulation of nitrogen oxides in copper-deficient cultures (Granger and Ward 2003;Woolfenden et al. 2013). However, at elevated copper concentrations, accumulation of nitrite was observed (Cheng et al. 2019) and the activity of nitrate and nitric oxide reductase was suppressed (Zhao et al. 2020). This effect was reversed and the denitrifying activity recurred rapidly as the bioavailable Cu decreased due to immobilization (Jacinthe and Tedesco 2009). Papirio et al. (2014) also found that copper inhibited the removal of nitrate whereas iron stimulated the activity of denitrifiers increasing the denitrification rate in a fluidized-bed reactor.
Moreover, ZnO nanoparticles, and thus, the released zinc ions inhibited the catalytic activity of key denitrifying enzymes, resulting in negative effects on the reduction of nitrate and N 2 O (Zheng et al. 2014). Nickel was also found to inhibit the last two steps of denitrification  or cause accumulation of intermediates (Ramírez et al. 2018). Nevertheless, nickel did not inhibit denitrification at pH 7 when its initial concentration did not exceed 50 mg/L but denitrification was repressed at soluble Ni levels between 50 mg/L and 100 mg/L; at 100 mg/L, nitrate removal only reached 30% after 6 h (Zou et al. 2013).

The Fate of Zn, cu, Fe and Ni during Denitrification
The concentration profiles of soluble Zn, Cu, Fe and Ni during denitrification at 50 mg/L and 100 mg/L initial concentrations are presented in Figs. 6a-d, respectively. Soluble metals fed in the reactor either separately (Figs. 6a-d) or concurrently (Figs. SM2 and SM3) were sequestered from solution within 6 h except nickel which exhibited an S-shaped kinetic profile (Figs. 6d and SM3d). The eventual sequestering of all metals explains the resistance of denitrifiers to initial heavy metal concentrations as high as 100 mg/L. As it has already been shown (Figs. 2c, 3e-f, 4c, 5e-f), the solution pH always shifted to 7.5-8.0 within 2 h due to the complete reduction of the negatively charged nitrate/nitrite ions to neutral elemental nitrogen. Under these conditions, the excess of hydroxyl ions can react with the soluble metal species and form the corresponding metal hydroxide precipitates (Figs. SM4a and SM4b). The formation of metal carbonates, utilizing the bicarbonate ions (HCO 3 − ) which are generated as a result of the oxidation of the carbon source, is also possible. In addition, when biological sulfate reduction is favored, the formation of metal sulfides can occur (Figs. SM2c and SM4c) and dominates the bioprecipitation mechanism due to the lower solubility of sulfides compared to metal hydroxides and carbonates (Lewis 2010). All these precipitation mechanisms have been reported for Ni in denitrifying fluidized-bed bioreactors in earlier studies (Zou et al. 2015;Zou et al. 2014).
To explain the sequestering pattern of Ni, a set of abiotic experiments were carried out (section SM.3) to elucidate the role of organic moieties in forming stable complexes with nickel ions. The results (Fig. SM5) revealed that, when the organic content of nickel solutions is high, nickel remains soluble and does not form insoluble nickel hydroxides when pH ranges from 3 to 8 due to the formation of metal-organic complexes. Similar behavior regarding the solubility of metal ions in presence of organic matter has also been observed for chromium (Remoundaki et al. 2003;Remoundaki et al. 2007). Therefore, the S-shaped kinetic profile of nickel and the slow sequestering during the first 5 h is attributed to the soluble complexes formed with the organic content of the medium. Thus, Ni 2+ bioavailability and toxicity decreased (Ramírez et al. 2018), improving the tolerance of H. denitrificans to higher initial Ni content. Moreover, as the degradation of the organic substrate proceeds, these complexes are dissociated permitting the formation of insoluble nickel species.

Conclusions
This work presents an effective biological process for the simultaneous removal of metals and nitrate from acidic, highly saline wastewater. The process is robust and capable of treating metal-bearing wastewater at pH values as low as 3, nitrate at concentrations as high as 5750 mg/L NO 3 − and tolerate salinity as high as 10% as NaCl. The experimental results demonstrated that the reactor may operate over a wide range of values of the above parameters, maintaining a high efficiency. It is anticipated that even more extreme conditions in terms of lower pH, higher NO 3 − and higher metal content could be tackled successfully which is the aim of the ongoing research activity. This work demonstrates that the proposed pilot-scale biological system achieved, in one stage, to neutralize the treated solution, completely remove nitrate without residual forms of reduced nitrogen and precipitate the metal content of the solution (i.e., Zn, Cu, Fe and Ni each at 100 mg/L) via bioprecipitation. As a result, the proposed process could be applied as a main treatment or polishing scheme for any relevant wastewater which needs to be treated for safe discharge.