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Article

Use of Zeolite (Chabazite) Supplemented with Effective Microorganisms for Wastewater Mitigation of a Marine Fish Farm

by
Mauro Lenzi
1,*,
Marco Leporatti Persiano
1,
Maurizio Ciarapica
1 and
Antonella D’Agostino
2
1
Lagoon Ecology and Aquaculture Laboratory (LEALab), Via G. Leopardi 9, 58015 Orbetello, Italy
2
Department of Economics and Statistics, University of Siena, 53100 Siena, Italy
*
Author to whom correspondence should be addressed.
Sustainability 2024, 16(4), 1353; https://doi.org/10.3390/su16041353
Submission received: 28 December 2023 / Revised: 26 January 2024 / Accepted: 31 January 2024 / Published: 6 February 2024
(This article belongs to the Special Issue Nature-Based Technologies for Wastewater Treatment: Removal of Emerging Organic Contaminants and Potentially Toxic Metals)
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Abstract

:
A study was conducted to assess the efficacy of chabazite zeolite in mitigating ammonia levels in wastewater from a land-based marine fish farm in southern Tuscany (Italy). The fish farm discharges effluent into a lagoon, constituting an important eutrophication source. The experimental setup involved a pond/canal that received wastewater from three sea bream tanks (40 L/s). A 50 m canal section was divided into two parallel halves (T and B), each about 3 m wide. In T, a chabazite bed (granules about 3 cm ϕ) was placed that was about 6 cm thick; B was untreated and used as a control. Five sampling trials were conducted in both T and B to determine N-NH4, N-NO3, and P-PO4 levels, in surface and near-bottom waters at both input and output. Prior to the zeolite addition, T and B sediments were sampled for TN and TP determination. Results indicated the not-managed canal system released nutrients and the output values were higher than the input, overshadowing the zeolite effects. Significant zeolite effects were observed by comparing B and T for differences between input and output: in T, nitrate increased (p = 0.05), demonstrating a resumption of nitrification, and ammonium (p = 0.07) and SRP (p = 0.06) decreased, in contrast to B.

1. Introduction

Over the past 35 years, marine aquaculture production has grown from one million tons per year to about 55 million tons [1]. Onshore and offshore fish farming certainly has a very critical aspect in the impact of farm wastes on coastal marine areas.
According to Pearson and Black [2], intensive aquaculture can have environmental impacts both on land and offshore. Phosphorus release is 19.6–22.4 kg/tons of product, of which 34–41% is released in dissolved form [3], while nitrogen dispersion is 52–95% of the nitrogen in the feed [4].
In seabream and seabass farm wastewater, ammonium nitrogen is more than 80% of the dissolved inorganic nitrogen, and depends mainly on the feed, pointing out the importance of adequate and proper nutrition [5]. Intensive land-based fish farms produce high particulate organic matter (food residues, feces, algal debris, etc.) that is discharged from wastewater and leaks into the final receiving environment. This process leads to turbidity, enrichment in nutrients and organic matter, alteration of sediment biogeochemical processes, disruption of benthic communities, and anoxic condition dominance. Wastewater from intensive aquaculture released into low water turnover environments, such as estuaries and lagoons, especially non-tidal lagoons, can exacerbate these issues.
The impact of aquaculture remains an unsolved problem to this day, which is addressed and continues to attract many researchers to this realm [6,7,8,9,10,11]. However, methods of reducing the nutrient load of wastewater from land-based fish farms, which move a large mass of water, are mainly based on the use of opportunistic nitrophilous macroalgal species. The management of these species presents actual difficulties and requires expensive efforts [7]. If the commercialization of algal masses and their processed products were viable, there would be an economic return to contribute to the costs of aquaculture wastewater mitigation, but this is rarely the case, especially in countries with high industrial development. Hence, there is a pressing need to explore low-cost alternative methods. Furthermore, the impact of eutrophication often requires costly downstream interventions in order to mitigate its consequences on the final receiving environment [12,13].
Zeolite has been used as an ion exchanger to absorb ammonium [14]. It was extensively tested in agriculture [15,16,17,18], freshwater basins [19], sewage effluent treatment for the reduction of ammonia and heavy metals [20,21,22,23], and more recently in the breeding of seabass juveniles [24]. In a seawater experiment, Lopez-Ruiz and Goméz-Garrudo [25] showed that more zeolite must be used to achieve the same amount of ammonium subtraction as obtained in freshwater experiences.
The types of zeolites vary in relation to the ionic ratios of the crystal structure between silicon, aluminum, and iron, e.g., clinoptilolite and mordenite have high Si/(Al + Fe) ratios, heulandite, chabazite, phillipsite, and erionite have intermediate ratio values, while analcime and laumontite have low values. The chelate ions within the voids of the crystal structure normally follow the following affinity order: K+ > NH4+ > Na+ > Ca++ > Mg++. Therefore, it is the value of the ionic concentration that determines the metal cation substitution with NH4+: it is the most abundant cation that occupies the chelation sites, in relation to affinity [26]. The NH4+ removal consequently leads to a reduction of undissociated ammonia (NH3) with which it is in a chemical equilibrium that depends on pH, T, and salinity [27]. Once the ion replacement pores are saturated by the NH4+ ion, the action of the mineral against this ion ends; in essence, there is a saturation of the mineral [26].
Promoting bacterial processes of nitrification and denitrification is the most effective means of regeneration in the natural environment. Nitrifying bacteria oxidize NH4+ to nitrate (NO3−) [28] while denitrifying bacteria act on the nitrate by eliminating it from the system as N2O and N2 gaseous [29]. As soon as the flow of NH4+ enters a system, the ion enters the bacterially acting biogeochemical cycle. If the load is high relative to the capacity of the bacteria to convert it, the presence of zeolite may be able to absorb the excess. When the NH4+ concentration decreases, the chelated NH4+ is released into the water, replaced by another cation, and can be converted by bacteria. New NH4+ can then be trapped in the crystal structure, at the next flow. This process can occur spontaneously in the natural environment due to the bacterial pools in the sediment. Bacterial strains are found in the natural environment and could naturally form films around the zeolite crystals. Nevertheless, initial enrichment can expedite and enhance the process. Bacteria can be introduced to the treatment by mixing them with the zeolite crystals.
For semi-intensive aquaculture, small ponds, and wastewater from land-based rearing tanks, the nutrient load affecting ponds and tanks, as well as that conveyed with wastewater, could be mitigated through the use of mineral zeolite due to its high affinity for ammonium ion.
The aim of this study was precisely to verify the effectiveness of zeolite in reducing the nitrogen load in effluent from a marine land-based fish farm released into a non-tidal Mediterranean coastal lagoon. Looking ahead, the ultimate goal is to make these wastewaters more compatible with the lagoon environment.

2. Materials and Methods

2.1. Study Area

Between July and September 2023, a field experiment was conducted to assess zeolite effectiveness in reducing the ammonia load from wastewater of a land-based fish farm (FF1) producing the marine species Sparus aurata and Dicentrarchus labrax. This facility is located along the southern coast of Tuscany (Ansedonia, Orbetello, Italy) [30].
FF1 encompasses 42 breeding tanks with a total volume of 22,400 m3. The fish farm is equipped with a system of settlement ponds and canals, covering an area of about 20,000 m2 (about 12,000 m3). Within the pond/canal system, detritivorous fish, such as Mugil cephalus and Chelo ramada, are housed, and microalgae and macroalgae develop. The pond/canal system is a complex network collecting wastewater from the tanks and then directing it into the final collector, which convoys all the wastewater into the eastern basin of the Orbetello lagoon. The effluent from FF1 is combined with that of another fish farm (FF2), which also has its own 2-hectare settlement pond/canals that discharge into the same lagoon basin (Figure 1). FF1 releases 560 L/s with 291 ± 16 μM dissolved inorganic nitrogen (DIN) and 344 ± 13 μM total dissolved nitrogen (TDN); FF2 releases 420 L/s with 162 ± 3 μM DIN and 654 ± 3 μM TDN. Dissolved organic nitrogen (DON) and dissolved organic phosphorus (DOP) are the most abundant components of the dissolved nitrogen and phosphorus in the water column, varying between 87% and 91% of TDN and between 46% and 79% of TDP [30]. Overall, fish production of the two fish farms amounts to 700–800 tons annually.
In spite of the canal/pond system as a mitigation measure for the two fish farms, the waters discharged into the lagoon constitute a major source of eutrophication in the eastern basin, which has significant microphyte and nitrophilous macroalgae developments and frequent anoxic and dystrophic crises [30].

2.2. Zeolite and Effective Microorganisms

Zeolites stand out as crucial porous materials owing to their diverse physicochemical properties, unlocking a broad spectrum of potential applications. As the utilization of zeolites continues to rise, concerns regarding increased exposure have prompted studies investigating their toxicity. Findings indicate low toxicity, primarily associated with the inhalation of dust and potentially linked to the generation of reactive oxygen species (ROS) through reactions on the surface of particles [31].
NH4+ absorption efficiency by zeolites depends on the pore size within the crystalline body. The pore size can vary among different zeolites, varying from 20% to 50% of the volume of the crystal body. For zeolite chabazite (CHA), these are 30% [32]. CHA, which has K+ as the most related cation [25], was the zeolite adopted in this field experiment. Zeolite-CHA subtracts NH4+, which is chelated within the crystal structure by substituting the other cation, due to the effect of higher concentration in the environment. The decrease in NH4+ concentration affects the chemical equilibrium with the undissociated species (NH3), which in turn decreases. Ammonia is the most toxic component in intensive fish farms, which is released mainly through the gill respiration of farmed fish [33].
Zeolite-CHA was sourced from the processing residues at a quarry in Sorano (South Tuscany, Italy). Sorano zeolite showed 97% of chabazite content (68.5 wt%). Chemical analysis revealed main components percentages of 52.61 SiO2, 17.12 Al2O3, 6.14 K2O, and 5.32 CaO; furthermore, this mineral displayed a high cation exchange capacity (CEC: total 2.17, Ca 1.46, K 0.60) [34].
To facilitate zeolite regeneration and enable the repetition of wastewater mitigation treatment processes, we employed effective microorganisms (EM). The use of EM as a technology began in the 1980s [35]. The EM utilized in our study comprised a mixture of 88 microbial strains (Lactobacillus spp., yeast, photosynthetic bacteria, nitrifying bacteria, and denitrifying bacteria) commercially available from EMRO Japan (https://emrojapan.com/aquaculture/, accessed on 11 November 2023). This mixture had the purpose of enhancing and expediting the organic matter decomposition and mineralization and increasing the processes of nitrification and denitrification [36].

2.3. Experimental Design

To mitigate the nutrient load carried by fish farm effluent, a bed of CHA was placed at the bottom of a canal receiving wastewater from only three breeding tanks. One of the critical issues with this operational choice was the constrained ionic exchange surface between the water mass and the zeolite. Indeed, ionic exchange could only occur at the bottom surface and the interstitial surface of the granules constituting the zeolite bed thickness. Consequently, part of the water mass would have been unaffected by the treatment, although we assumed the damage was limited because of the relatively low water column along the canals, ranging between 40 and 60 cm.
The area where zeolite effectiveness was tested is shown in Figure 1. It affected a section about 50 m long and 6 to 10 m wide of a canal that carried wastewater from three breeding tanks at a total flow of 40–50 L/s. This area was partitioned into two parallel hemi-canals by a septum comprising plastic sheeting supported by poles; one hemi-canal was designated for treatment (T) and the other for comparison (B). The two parallel areas received a total flux of about 45 L/s, roughly divided between the two areas. The treatment area was 150 m2 (50 m × 3 m). A grain size of 35 mm was selected for the zeolite to ensure good water permeability within the artificial bottom thickness, and to prevent clogging of the interstices by organic detritus fallout, consequently maintaining the highest possible ionic exchange surface area. A total of 9 m3 of CHA was used (weighing 5.81 tons; specific gravity g = p/V 0.6457 Newton/m3, for a residual humidity of 15%), resulting in an average bottom cover thickness of about 0.06 m. Given the quantity used, it was acknowledged that some material might remain submerged in the bottom, which initially consisted of a thick layer of soft mud with a high organic load.
During the grain size selection stage, the material was enriched with a commercial EM blend (2.5 mL per kg of CHA). This choice was supported by the results obtained in a seawater microcosm experiment, where the addition of EM prolonged the zeolite efficacy in counteracting ammonium development under anaerobic conditions throughout the experimental period, compared to the control without EM [37].
The distribution of the material was conducted manually with the objective of achieving as homogeneous a distribution as possible. The parallel area B did not undergo treatment with zeolite.

2.4. Sampling and Analytical Determinations

The experimental design centered on assessing the nitrogen load mitigation along the water pathway on the zeolite bed. Consequently, sampling was planned at two key points: at the water entry into the system and at its exit, for both treatment canal T and control canal B (Figure 1). Additionally, initial sediment sampling was conducted to check the load of sediment total nitrogen (TN) and total phosphorus (TP).
Top water (−20 cm from the surface) and bottom water (5 cm above the sediment) were sampled, the latter collected gently using a suction tube. Samplings were carried out in duplicate and occurred just before the insertion of CHA on 15 July (t0), 2 days after insertion (17 July, t2), 1 week later (22 July, t7), 12 days later (27 July, t12), 34 days later (18 August, t34), and 74 days after the start of treatment (28 September, t74). This sampling timeline was devised to discern any temporal effects of CHA, evaluate whether the EM added during mineral preparation and any local bacterial pools were capable of prolonging the effectiveness of the treatment, i.e., whether a flora proficient in nitrification and denitrification had been established, placing the zeolite in a position to adsorb additional ammonium beyond the first saturation of the pores present in the crystalline body. The collected samples, refrigerated and transported to the laboratory, were filtered at 0.45 mm, and then nitrate nitrogen (N-NO3), ammonia nitrogen (N-NH4), and soluble reactive phosphorus (SRP) were determined. The analyses were conducted according to APAT IRSA-CNR [38]. Nitrate analysis was based on its reduction to nitrite and the subsequent formation of a colored diazo compound. Ammonium determination was based on a series of reactions, catalyzed photochemically, that led to the formation of indophenol blue. SRP determination was based on the formation of a blue-colored phosphomolybdic complex. The concentrations of the three colored compounds were then assessed spectrophotometrically.
Sediment sampling occurred before the addition of zeolite to the T canal. Subsequent trials did not repeat sediment sampling due to the thick layer of zeolite in T that would not have allowed the sampling of mud. Eight sediment samples were taken by a horizontal sampler [30], four from each area, and distributed in two separate locations (Figure 1). The collected sediment (about 4 cm thick) was transferred to polyethylene containers, refrigerated, and subsequently frozen pending analysis. The samples were dried to constant weight at 75 °C, and then total nitrogen (TN) and total phosphorus (TP) were determined. TN was determined using an elementary analyzer (CHN Thermoquest, model 1110), TP according to Aspila et al. [39].

2.5. Statistical Analysis

Graphical representation, employing boxplots, was employed to process data. In Table 1, the sample size is reported. Statistical significance was assessed using the Mann–Whitney U test to compare single variables between the entry (1) and exit (2) points in the two systems (T, B). The comparison was conducted by examining the differences (Δ) between output and input for both systems (B and T) in both the top (a) and bottom (b) water (ΔBa vs. ΔTa and ΔBb vs. ΔTb). The choice of a non-parametric test was motivated by the absence of assumptions regarding the normality of the data, as advocated by Hollander et al. [40]. Statistical analysis was performed using STATA [41].

3. Results and Discussion

Before the zeolite was placed on the T bottom (t0), the waters showed an increase in the variable values between input and output, in both T and B systems. Inorganic dissolved nitrogen (DIN as N-NO3− + N-NH4+) increased from 281 ± 25 μM to 345 ± 60 μM (23%) in B, and from 305 ± 7 μM to 546 ± 195 μM (79%) in T. SRP increased from 2.75 ± 0.46 μM to 2.83 ± 0.17 μM (2.93%) in B, and from 2.70 ± 0.04 μM to 3.44 ± 0.43 μM (27.46%) in T. This suggests that the fish farm pond/canal system had evolved into a source of nutrients. In the surface sediment analysis conducted before treatment, TN was 0.551 ± 0.083% and TP 0.509 ± 0.091%, with an atomic ratio N:P of 2.4, highlighting a relative N-limitation [42]. This was likely a consequence of the different rates of accumulation of the two nutrients, with N being more easily released as ammonia and/or eliminated as gaseous N2O and N2 [43]. Settling basins can improve effluent quality, and significantly reduce N concentration in the water column [44,45,46,47]. However, it is crucial to manage to minimize nutrient release, which has accumulated mainly as debris organic matter [47].
Table 1 shows the means (±SD) in μM of the variables examined for the trials t2–t74, for both T and B systems. In Table 2, the increases or decreases between output and input from the system, for both T and B, are given by applying the formula: [(output − input) × 100/input].
The sample size in the t0 trial is low and not extended over time; nevertheless, notable trends emerge. T showed a decrease in nitrate nitrogen (−5%) and an increase in ammonia nitrogen (85%), while in subsequent trials, after the insertion of zeolite + EM, there was an increase in nitrate nitrogen and a decrease in ammonia nitrogen in the bottom water (Table 2). These observations suggest an enhancement in the nitrification process. At t0, B turned out to increase nitrate (6%) and keep ammonia nitrogen essentially stable (+1.1%). Later, between t2 and t74, B displayed stability in nitrate levels, but increased ammonium (Table 2). This suggests that the nitrification process likely remained inefficient overall in all trials. In t0, DIN increased in both B and T, while, in the subsequent trials, it increased even more in output from B, and decreased in output from T (Table 2). SRP values at the output of B exhibited minimal variation in trials t2–t74 (Table 2), compared with those at the beginning of t0, when there was an increase of about 3%. In contrast, in T, SRP went from an increase at the output of 27% in t0 to a decrease in subsequent trials of about 10% and 9% for top and bottom water, respectively (Table 2).
However, statistical analysis showed no significant differences between input and output from the two systems for any variable. Statistical significance was obtained by comparing the differences (Δ) between output and input, between the two systems B and T at different significance levels (5% or 10%). The results obtained using the Mann–Whitney U test are shown in Table 3, whereas in Figure 2, Figure 3, Figure 4 and Figure 5 the boxplots are reported. Specifically, significant differences were observed in system T versus B, in bottom water (ΔBb vs. ΔTb) where N-NO3 increased (p = 0.05), and in top water (ΔBa vs. ΔTa) where both N-NH4 (p = 0.07) and SRP (p = 0.06) decreased.
The increase in nitrate nitrogen at the bottom in T during the t2–t74 trials suggests that the bacterial nitrification process may have been more active compared to B, although this was matched by a significant decrease in ammonium nitrogen only in T top water. Considering that the water column averaged only 50 cm, the decrease in ammonia nitrogen could express partly the subtraction of the ion by zeolite and partly its oxidation to nitrate. Conversely, in B, where nitrate decreased and ammonium increased, though not significantly, it is very likely that nitrification and denitrification were never efficient. Confirmation of an enhanced oxidative state of the zeolite-containing sediment layer is provided by the decrease in SRP output from T, likely due to precipitation and blockage by iron and manganese oxides–hydroxides [48].
It can be assumed that in the treated area, the application of zeolite + EM resulted in a modest, barely detectable improvement in effluent quality. This improvement persisted throughout the experiment, signifying that zeolite saturation did not occur permanently and bacterial activity promoted regeneration of the mineral’s crystalline pores, allowing activation of an ionic pump. Furthermore, it is likely that the effect of zeolite was partially masked by the bottom nutrient releases, which were evident in t0 for both B and T and in subsequent trials for B.
Addressing the impact of wastewater from aquaculture practices is imperative, considering it is a bottleneck that hinders the adequate development of future fish production. Although in small plant systems, effluent mitigation solutions are feasible [49], solutions to date are energetically intensive and economically unviable for fish farms with a large volume of water and high fish production, as in the case of the farm in this study that uses over 45,000 m3 per day. For these farms, it is necessary to study systems with low economic and environmental impact. Contrary to the intended purpose, the pond/canal system, designed to settle suspended solids and enhance bacterial processes to break down organic detritus, droppings, and leftover feed, and to promote the processes of oxidation and N removal from the system, had proven ineffective, ultimately contributing to increase nutrient load in the effluent. The use of zeolite partly succeeds in counteracting this process. Effective sediment management is crucial to reduce anaerobic mineralization. Several options can address this issue, such as periodic removal of organic sludge accumulated on the bottom, or aerobic mineralization through frequent sediment resuspension in the water column [28,50]. This process, conducted along the settling basin system, makes orthophosphates insoluble, facilitates nitrification [50], and, consequently, promotes denitrification [29,51].

4. Conclusions

The results obtained in this field experiment in a small stretch of canal demonstrate, albeit subtly, mitigation of eutrophication components conveyed by wastewaters from a land-based fish farm that uses more than 45,000 m3 of marine water per day and produces about 500 tons of sea bass and sea bream annually. The placement of the zeolite bed in the studied canal section notably enhanced the oxidative state of sediments and facilitated orthophosphate sedimentation, nitrification, and ammonia nitrogen reduction, in comparison to the untreated parallel section of the canal. These results, albeit of modest magnitude, were maintained throughout the entire 74-day study period, suggesting that the added pool of microorganisms to the zeolite likely contributed to sustaining the active ammonium ion chelation capacity of the mineral.
For a more comprehensive exploration of zeolite + EM performance, wastewater should again be tested on a pristine surface or within a system where a periodic jet stream of air or water is applied to the zeolite bed. This approach can help to maintain a heightened oxidative condition and prevent the deposition of excess debris fallout on the mineral.
We believe that the results are encouraging and give hope for the possible use of zeolite in the aquaculture of marine species as well, allowing for greater sustainability of the impact of wastewater through low energy consumption management. But there is still much to be studied, e.g., how to arrange the zeolite by studying low labor use systems that allow more exchange surface between the wastewater and the mineral, and how to reduce the problem of detritus fallout on the zeolite layer with the least possible energy expenditure, considering the high detritus load in effluents of intensive aquaculture farms.

Author Contributions

Conceptualization and methodological approach, M.L., M.L.P. and M.C.; sampling and analysis, M.L. and M.L.P.; data curation, M.L.; statistical analysis, A.D.; writing—original paper preparation, M.L.; writing—review and editing, the whole group. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the Cave Piandirena Company.

Institutional Review Board Statement

The study did not require ethical approval.

Informed Consent Statement

Informed consent was obtained from all subjects involved in the study.

Data Availability Statement

Data are contained within the article.

Acknowledgments

We thank Cave Piandirena Company for supplying the mineral used and funding for the study, and we also thank the fish farm Il Vigneto for providing the plant facilities and the boat for sampling.

Conflicts of Interest

The authors declare that this study received funding from Cave Piandirena Company. The funder was not involved in the study design, collection, analysis, interpretation of data, the writing of this article or the decision to submit it for publication.

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Figure 1. Study area and experimental design: (a) location in central Italy of fish farm; (b) proximity of fish farm to the Orbetello lagoon east basin, where wastewater is released; (c) extension of the fish farm and location of the study canal; bleu line on the right, lagoon coast; green line, collector canal of all fish farm effluents; (d) section of the canal where the experiment was conducted; red line, septum dividing the canal into two halves; T, canal part where the zeolite chabazite bed was placed; B, untreated canal part used as control; yellow circles, water sampling stations; red triangles, sediment sampling stations.
Figure 1. Study area and experimental design: (a) location in central Italy of fish farm; (b) proximity of fish farm to the Orbetello lagoon east basin, where wastewater is released; (c) extension of the fish farm and location of the study canal; bleu line on the right, lagoon coast; green line, collector canal of all fish farm effluents; (d) section of the canal where the experiment was conducted; red line, septum dividing the canal into two halves; T, canal part where the zeolite chabazite bed was placed; B, untreated canal part used as control; yellow circles, water sampling stations; red triangles, sediment sampling stations.
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Figure 2. Boxplots for nitric nitrogen (N—NO3). Comparison of the differences between input and output, between T and B: ΔBa vs. ΔTa, left (gray); ΔBb vs. ΔTb, right (green).
Figure 2. Boxplots for nitric nitrogen (N—NO3). Comparison of the differences between input and output, between T and B: ΔBa vs. ΔTa, left (gray); ΔBb vs. ΔTb, right (green).
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Figure 3. Boxplots for ammonium nitrogen (N—NH4). Comparison of the differences between input and output, between T and B: DBa vs. DTa, left (gray); DBb vs. DTb, right (green).
Figure 3. Boxplots for ammonium nitrogen (N—NH4). Comparison of the differences between input and output, between T and B: DBa vs. DTa, left (gray); DBb vs. DTb, right (green).
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Figure 4. Boxplots for dissolved inorganic nitrogen (DIN). Comparison of the differences between input and output, between T and B: DBa vs. DTa, left (gray); DBb vs. DTb, right (green).
Figure 4. Boxplots for dissolved inorganic nitrogen (DIN). Comparison of the differences between input and output, between T and B: DBa vs. DTa, left (gray); DBb vs. DTb, right (green).
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Figure 5. Boxplots for soluble reactive phosphorus (SRP). Comparison of the differences between input and output, between T and B: ΔBa vs. ΔTa, left (gray); ΔBb vs. ΔTb, right (green).
Figure 5. Boxplots for soluble reactive phosphorus (SRP). Comparison of the differences between input and output, between T and B: ΔBa vs. ΔTa, left (gray); ΔBb vs. ΔTb, right (green).
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Table 1. Means (±SD), in μM, among all values of the five water sampling trials (t2, t7, t12, t34, t74), of nitrate nitrogen (N-NO3), ammonia nitrogen (N-NH4), total dissolved nitrogen (DIN), and soluble reactive phosphorus (SRP), in the treatment system (T) and control system (B), in input (T1, B1) and output (T2, B2), for top water (T1a, T2a, B1a, B2a) and bottom water (T1b, T2b, B1b, B2b). N, sample size; min, minimum value; max, maximum value; median.
Table 1. Means (±SD), in μM, among all values of the five water sampling trials (t2, t7, t12, t34, t74), of nitrate nitrogen (N-NO3), ammonia nitrogen (N-NH4), total dissolved nitrogen (DIN), and soluble reactive phosphorus (SRP), in the treatment system (T) and control system (B), in input (T1, B1) and output (T2, B2), for top water (T1a, T2a, B1a, B2a) and bottom water (T1b, T2b, B1b, B2b). N, sample size; min, minimum value; max, maximum value; median.
Mean ± SDNMinMaxMedian
μM
T1aN-NO385.9 ± 39.21029.43125.29108
N-NH4182.2 ± 53.110122.86265.86183.25
DIN268.1 ± 87.310152.29371299.96
SRP2.65 ± 0.461023.232.71
T1bN-NO376.6 ± 34.11031.07109.6493.71
N-NH4250.4 ± 150.910118.21603.64211.21
DIN327.0 ± 177.310158.14711.71307.21
SRP2.25 ± 0.37101.522.772.31
T2aN-NO393.1 ± 35.91048.57128.86112.57
N-NH4184.3 ± 42.610125.43248.36193.61
DIN277.4 ± 75.610174.86368.93313.75
SRP2.39 ± 0.24101.972.682.45
T2bN-NO380.5 ± 35.11031.64116.5799.18
N-NH4216.9 ± 99.010103.43408.57199.39
DIN297.5 ± 126.910135.07512.79303.79
SRP2.05 ± 0.21101.652.322.1
B1aN-NO388.9 ± 38.01038.29134.57111
N-NH4182.9 ± 59.610110.07285.71186.43
DIN271.8 ± 91.310159.5395303.71
SRP2.43 ± 0.24101.972.712.5
B1bN-NO382.8 ± 30.11047.21115.4399.39
N-NH4218.2 ± 31.810127.5441.86180.25
DIN300.9 ± 136.310174.71557.29287.43
SRP2.34 ± 0.40101.842.812.29
B2aN-NO392.4 ± 37.71039.36139.14112.5
N-NH4396.3 ± 282.710124.93780.21270.57
DIN488.7 ± 313.110170.5895.29399.39
SRP2.38 ± 0.19102.062.652.4
B2bN-NO379.2 ± 33.01038.64120.6490.14
N-NH4238.7 ± 150.110101.14585.21195.61
DIN317.9 ± 176.810144.79705.86296.21
SRP2.36 ± 0.29102.13.062.27
Table 2. Percentage increase or decrease between output (2) and input (1) from the systems T and B, for top water (a) and bottom water (b), for the variables nitrate nitrogen (N-NO3), ammonia nitrogen (N-NH4), dissolved inorganic nitrogen (DIN), and soluble reactive phosphorus (SRP).
Table 2. Percentage increase or decrease between output (2) and input (1) from the systems T and B, for top water (a) and bottom water (b), for the variables nitrate nitrogen (N-NO3), ammonia nitrogen (N-NH4), dissolved inorganic nitrogen (DIN), and soluble reactive phosphorus (SRP).
N-NO3N-NH4DINSRP
T2a-T1a8.395.293.4−9.964
T2b-T1b5.03−13.35−9.04−8.621
B2a-B1a4.02116.6279.82−1.86
B2b-B1b−4.269.445.60.97
Table 3. U–Mann-Whitney test p-values. Comparison of differences (Δ) between inputs and outputs of B versus those of T, for the four variables examined.
Table 3. U–Mann-Whitney test p-values. Comparison of differences (Δ) between inputs and outputs of B versus those of T, for the four variables examined.
N-NO3N-NH4DINSRP
ΔBa vs. ΔTa0.380.070.380.06
ΔBb vs. ΔTb0.050.550.290.36
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Lenzi, M.; Leporatti Persiano, M.; Ciarapica, M.; D’Agostino, A. Use of Zeolite (Chabazite) Supplemented with Effective Microorganisms for Wastewater Mitigation of a Marine Fish Farm. Sustainability 2024, 16, 1353. https://doi.org/10.3390/su16041353

AMA Style

Lenzi M, Leporatti Persiano M, Ciarapica M, D’Agostino A. Use of Zeolite (Chabazite) Supplemented with Effective Microorganisms for Wastewater Mitigation of a Marine Fish Farm. Sustainability. 2024; 16(4):1353. https://doi.org/10.3390/su16041353

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Lenzi, Mauro, Marco Leporatti Persiano, Maurizio Ciarapica, and Antonella D’Agostino. 2024. "Use of Zeolite (Chabazite) Supplemented with Effective Microorganisms for Wastewater Mitigation of a Marine Fish Farm" Sustainability 16, no. 4: 1353. https://doi.org/10.3390/su16041353

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