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Article

Pollutant Removal Efficiency of Pilot-Scale Horizontal Subsurface Flow Constructed Wetlands Treating Landfill Leachate

by
Ioannis Ntountounakis
1,
Ioanna-Eirini Margaritou
1,
Ioannis Pervelis
1,
Pavlos Kyrou
1,
Paraskevas Parlakidis
2 and
Georgios D. Gikas
1,*
1
Laboratory of Ecological Engineering and Technology, Department of Environmental Engineering, School of Engineering, Democritus University of Thrace, 67100 Xanthi, Greece
2
Laboratory of Agricultural Pharmacology and Ecotoxicology, Faculty of Agricultural Development, Democritus University of Thrace, 193 Pantazidou, 68200 Orestias, Greece
*
Author to whom correspondence should be addressed.
Appl. Sci. 2025, 15(5), 2595; https://doi.org/10.3390/app15052595
Submission received: 1 February 2025 / Revised: 22 February 2025 / Accepted: 26 February 2025 / Published: 27 February 2025
(This article belongs to the Special Issue Promising Sustainable Technologies in Wastewater Treatment)
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Abstract

:
Landfill leachate contains various organic and inorganic substances resulting from the decomposition of solid waste. The treatment of this complex mixture is an imperative need for environmental protection. This study used five pilot-scale horizontal subsurface flow (HSF) constructed wetland (CW) units to treat landfill leachate. The main objective was the evaluation of the performance of CW units in the removal of pollutants. The effect of porous media (gravel and zeolite), plants (common reed and cattail), and hydraulic residence time (HRT, 8 and 10 days) were investigated. Two pilot-scale CW units differed in HRT, two in porous media, and three in planting. The results showed that the planted CW units had higher organic matter (OM) and nitrogen (TKN, NH4-N) removal compared with the unplanted unit. The 10-day HRT CW unit had higher average removal rates for all pollutants compared with the CW unit with an 8-day HRT. Finally, the CW unit with zeolite (25%, by volume) in the fill material showed higher average removal rates for OM and nitrogen compared to the unit with gravel.

1. Introduction

Landfill leachate is a highly contaminated liquid that forms when precipitation (e.g., rainwater) filters through waste materials in landfills. As a result of solid waste breaking down, this liquid comprises a variety of organic and inorganic compounds, as well as additional contaminants [1]. The landfill leachate’s composition usually varies depending on the landfill age, the type of waste deposited, meteorological conditions, and the landfill operation methods [2]. Typical contaminants found in landfill leachate include ammonia, nitrates, phosphates, metals, and organic compounds [3]. Furthermore, compounds such as pesticides, pharmaceuticals, personal care products, endocrine-disrupting chemicals (EDCs), and industrial chemicals have been identified in raw landfill leachates, raising interest and attention due to their potential bioaccumulation and environmental risk [4,5,6]. The management and treatment of this complex mixture is a critical environmental challenge worldwide, as untreated landfill leachate can cause severe pollution to soil and water bodies, posing a serious threat to the environment and, by extension, to human health [7].
In Greece, studies on medium-sized landfills indicate that existing secondary treatment methods fail to meet regulatory discharge limits, highlighting the need for improved treatment technologies [8]. Globally, landfill leachate contains high concentrations of both biodegradable and persistent toxic compounds, making effective treatment essential for environmental protection [9]. Traditional methods of treating leachate include various technologies such as physicochemical and biological ones, membrane technology, and advanced oxidation [10,11]. The physicochemical processes include coagulation/flocculation, adsorption, chemical precipitation, and chemical oxidation. These processes are usually used in combination with other processes as a pre- or post-treatment to improve the quality of landfill leachate that is discharged to water bodies [12].
Coagulation and flocculation are processes applied to solutions where suspended solids are present, aiming to reduce the forces that hold suspended solids in the aqueous solution. In the coagulation process, chemical compounds such as ferric chloride or aluminum sulfate are added to neutralize or reduce the electrical charge of the suspended particles. In the flocculation process, to increase the formation of flocs, chemical compounds such as polymers are added. From the above, it is clear that the amount of chemicals used in these processes affects the cost of treating landfill leachate. In addition, the coagulation–flocculation process produces large amounts of sludge [2,12,13].
Biological treatment technologies refer to the use of microorganisms to degrade organic pollutants and are categorized into aerobic and anaerobic processes. They can be used alone or in combination with other traditional methods to increase the effectiveness of the treatment [2]. These methods include aerated lagoons, activated sludge processes, trickling filters, sequence batch reactors (SBRs), etc., [14,15]. The effectiveness of biological treatment technologies depends on their operating conditions being suitable for the growth of microorganisms. While these systems perform well in pollutant removal, they have several disadvantages, including the production of large quantities of sewage sludge (activated sludge), high electricity requirements (activated sludge), the need for an extensive land area (aerated lagoons), significant clogging and odor issues (trickling filters), and the requirement for specialized personnel (SBRs) [16,17,18].
Membrane technology is a fairly effective method used to treat landfill leachate by removing contaminants such as heavy metals, organic compounds, suspended solids, and microorganisms. Membrane technology includes reverse osmosis, nanofiltration, ultrafiltration, and membrane bioreactors [12,19]. However, their main disadvantages are high energy requirements for their operation, high maintenance costs, and the production of a concentrated waste stream [2].
In recent years, increasing interest has been given to the treatment of landfill leachate using advanced oxidation processes (AOPs). The characteristic of these methods is that they produce hydroxyl radicals, which can break down a wide range of compounds. AOPs are particularly effective as they can oxidize many types of organic compounds into simpler ones that are biodegradable. Thus, they can be used as pretreatment methods for landfill leachate before further treatment in traditional biological treatment plants or natural systems (e.g., constructed wetlands) [20]. The main methods are the Fenton process, ozonation, photocatalysis, and electrochemical oxidation [21]. These methods remove ammonia and COD to a satisfactory degree. However, they present some disadvantages, such as large amounts of by-products (the Fenton process, electrochemical oxidation), a high energy consumption (ozonation, electrochemical oxidation), limited full-scale application, and potential electrode degradation over time (electrochemical oxidation) [2,12]. Furthermore, although the aforementioned methods can be effective in treating landfill leachates, they often require sophisticated operational controls and maintenance, which can be challenging for some landfill operators, particularly in developing regions [22]. In response to these challenges, constructed wetlands (CWs) have gained attention as a sustainable and cost-effective alternative for effective pollutant removal. These engineered systems belong to biological treatment technologies, and they are similar in their mode of operation to natural wetlands, utilizing synergistic interactions between water, vegetation, microorganisms, and the substrate to remove pollutants from leachate [23,24]. Constructed wetlands can remove a wide variety of pollutants through various processes occurring in their environment, primarily including sedimentation, adsorption, plant uptake, microbial degradation, and chemical precipitation.
CWs used for wastewater treatment can be categorized into two main types: free-water surface (FWS) and subsurface flow (SF) wetlands. In FWS CWs, the water flows on the surface, supporting a variety of aquatic plants. SF CWs, on the other hand, have water flowing beneath the surface through a gravel or soil medium, which supports plant roots and provides a habitat for microorganisms. SF CWs are further divided into horizontal subsurface flow (HSF) and vertical flow (VF) systems, each with distinct design characteristics and treatment mechanisms [25,26].
CWs offer numerous advantages over traditional treatment methods [27]. Their advantages are as follows: (a) They generally have lower construction and operating costs than conventional treatment systems [28,29]. (b) They utilize natural processes, reducing the need for expensive and energy-intensive mechanical equipment. (c) These systems rely on natural energy sources, such as sunlight, wind, and plant growth, minimizing the need for external energy inputs. (d) CWs are characterized by simplicity and robustness as they are simple to design and operate, making them suitable for use in remote or underserved areas where advanced treatment technologies may be impractical. Additionally, CWs provide ancillary benefits such as habitat creation, biodiversity enhancement, and landscape aesthetics. They also have the potential to sequester carbon, thereby contributing to climate change mitigation efforts [30].
Despite their advantages, constructed wetlands also face several challenges and limitations. Their performance and efficiency can be influenced by climatic conditions, particularly in colder regions. Their proper design and operation and ensuring their optimum performance require the careful selection of hydraulic loading rates, hydraulic residence times, and vegetation. Subsurface flow systems, in particular, are prone to clogging, necessitating regular monitoring and maintenance. Another challenge is the potential accumulation of heavy metals and other toxic substances in the wetland substrate and vegetation, which may pose risks for long-term sustainability and disposal [31].
Ongoing research and development in the field of CWs for landfill leachate treatment focuses on improving the design, efficiency, and robustness of these systems. Innovations include hybrid systems combining different types of wetlands, integration with other treatment technologies, and the use of novel plant species and substrates to enhance pollutant removal. Therefore, optimizing CWs’ performance requires further investigation into key design parameters, which is the objective of the present study. Specifically, this study aims at the investigation of the effect of substrate material, plant presence, vegetation species, and hydraulic residence time (HRT) on pollutant removal. To achieve this objective, the operation of five pilot-scale horizontal subsurface flow (HSF) CW units in landfill leachate treatment is investigated, and their performance in reducing pollutant concentrations is evaluated. The ultimate goal is to develop recommendations for optimizing CW units to achieve maximum pollutant removal and to guide future research in the field of landfill leachate treatment. Furthermore, this study seeks to highlight CWs as a sustainable solution for mitigating the environmental impacts of landfill leachate.

2. Materials and Methods

2.1. Leachate Qualitative Characteristics

The leachates used in this study were taken from a sanitary landfill located in the Xanthi area in Northern Greece, and the solid waste was received from the Xanthi prefecture. The landfill has a total area of approximately 21 hectares, with an active area of 6 hectares, and it receives approximately 2500 tons of solid waste per month. The characteristics of the landfill leachate based on the age of the landfill and the average concentrations of the main landfill leachate parameters are presented in Table 1. The landfill has been operating since 1993, and according to the data in Table 1, it can be characterized as an “old” landfill. The results (e.g., standard deviation values) show that the quality characteristics of the leachate varied during the experimental period. This variation was expected, as the quality characteristics of landfill leachate are influenced by factors such as the age and design of the landfill, rainfall, temperature, the type of waste it contains, and management practices [2]. Based on Table 1, the BOD and pH values, as well as the BOD/COD ratio, of the landfill leachate in the present study correspond to those of an old landfill (age > 20 years), while the values of the remaining parameters align with those of an average-aged landfill. This is likely due to the fact that the landfill continues to receive new waste, and thus, the waste decomposition process is ongoing.

2.2. System Configuration and Operation

Five pilot-scale HSF CW units were used for the experiment (Figure 1). The units were placed outside the Ecological Engineering and Technology Laboratory (location: 41°08′47″ N, 24°55′09″ E). Each CW was rectangular, with dimensions of 1.10 m long, 0.45 m wide, and 0.60 m deep, and was filled with porous material to a thickness of 45 cm. The porous media used were (a) fine gravel (code name: FG) with D50 = 8 mm, originating from igneous rocks and taken from the bed of a local river, and (b) fine-grained zeolite (code name: FZ) with D50 = 8 mm, the main component being clinoptilolite. The proven effectiveness of zeolite and fine gravel to improve pollution removal served as the basis for their usage as fill materials. In addition to supporting plant root growth, zeolite and gravel also promote microbial colonization, which is critical for biodegradation processes. Because of its high cation exchange capacity, especially for ammonium adsorption, which improves nitrogen removal, zeolite was chosen and added to one unit at a rate of 25% by volume [34]. The plants used for planting the CWs were common reed (Phragmites australis) and cattail (Typha latifolia), with code names R and C, respectively, collected from watercourses in the area. The plant species were selected due to their proven effectiveness in wetland treatment systems. According to Akinbile et al. [35], both species support microbial communities engaged in pollutants degradation, improve oxygen transmission in the rhizosphere, and demonstrate excellent flexibility to changing water conditions. Both plants were tolerant to a high EC of 14–16 mS/cm and exhibited good nutrient (N and P) uptake in CWs [36]. To ensure viability for large-scale applications, their selection was also based on their local availability. The design and operational characteristics of the CW units are presented in Table 2. In the CW units with code names W1FG-R, W2FG-U, W3FG-R, and W4FG-C the porous media was FG, and in the unit codenamed W5FGZ-R, it was FG 75% and FZ 25% (by volume). Furthermore, the CW units W1FG-R, W3FG-R, and W5FGZ-R were planted with Phragmites australis, and the unit W4FG-C was planted with Typha latifolia. The W2FG-U unit was used as a control unit and therefore remained unplanted.
The operation of the wetlands began in June 2017 and continued until June 2018. The landfill leachate was transferred from the landfill to the laboratory using a tank of 1.0 m3. Because the electrical conductivity of the leachate was high (Table 1), and to prevent any adverse effects on the plants, the charging of the CW units was performed with diluted leachate [37]. Dilution was performed with tap water in a tank of 250 L, so that the electrical conductivity was approximately 7 mS/cm. In the tap water, the mean EC value was about 0.55 mS/cm, and the average concentration (mg/L) of the main ions was Ca: 90; Na: 28; Cl: 42; SO4: 38; and Mg: 14. The residual chlorine was 0.2 mg/L. In the landfill leachate, the mean value of EC was 28.5 mS/cm (Table 1), and the typical average concentration (mg/L) of the main ions was Ca: 240-2330; Na: 85-3800; Cl: 47-2400; SO4: 20-730; and Mg: 4-780 [33]. It is clear that the EC value and ion concentrations in tap water were much lower than those in the leachate. Therefore, adding water to dilute the leachates is considered unlikely to affect the results. During the first operational period of the CW units (approximately one month), the units were charged with diluted wastewater in which the EC was less than 7 mS/cm to avoid any possible shock to the plants and allow for acclimatization. Normal charging began in early July and was performed daily. The units were loaded twice daily (approximately every 12 h) with equal amounts of leachate. The total daily volume of loading leachate was 9.5 L/d for the W1FG-R unit and 7.6 L/d for the rest of the units, thus, achieving a hydraulic residence time (HRT) of 8 and 10 days and a hydraulic loading rate (HLR) of 19.4 mm/d and 15.3 mm/d, respectively (Table 2). The selection of HRT was made considering previous studies related to the removal of organic matter and nutrients from domestic wastewater and landfill leachate [38,39,40]. Longer HRTs facilitate greater nitrification–denitrification and organic degradation by extending the period that leachate, microbes, and substrate are in contact. The selected values achieve a balance between performance and practicality, because overly lengthy HRTs may result in operational inefficiencies, larger land needs, and a higher capital cost [41]. With this CW unit configuration, it is possible to evaluate the effect of the presence and type of plants, the existence of zeolite in the substrate, and the applied HRT on the removal of pollutants, by comparing units W2FG-U, W3FG-R, and W4FG-C with each other, W3FG-R with W5FGZ-R, and W1FG-R with W3FG-R, respectively.
A meteorological station (model: ELE MM900/950) was installed outside the laboratory for monitoring and measuring meteorological parameters. Air temperature, humidity, precipitation depth, and wind speed and direction were recorded on-site, at 5 min intervals.

2.3. Sampling and Monitoring

Sampling and measurements began in mid-July. Temperature (T), dissolved oxygen (DO), pH, and electrical conductivity (EC) were measured in situ every three or four days. The measurements were taken at the inflow and outflow of the CWs units using the portable instrument, the HQ30D Field Case (HACH, USA). Leachate samples were collected every seven or eight days from the same locations of the CW units where the physicochemical parameter measurements were performed. The samples were transported to the chemical laboratory and analyzed immediately for the determination of BOD, COD, TKN (total Kjeldahl nitrogen), and NH4-N (ammonia nitrogen), according to standard methods [42].

2.4. Statistical Analysis

The t-test was applied to evaluate the effect of HRT and fill material on the removal efficiency of the CW unit. A one-way between-groups ANOVA at a 95% confidence interval (p < 0.05) with Tukey’s Honestly Significant Difference (HSD) test was used to investigate if there were significant differences in the mean removal of pollutants between the W2FG-U, W3FG-R, and W4FG-C units. The t-test and one-way ANOVA were chosen because the data met the assumption of normality, tested with the Shapiro–Wilk statistical criterion, and the assumption of homogeneity of variances tested with the Levene test. If the normality test and homogeneity of variances (Levene’s test) are taken into account, the t-test and one-way ANOVA accurately evaluate the experimental data. Statistical analyses were performed using SPSS Statistics 25.0 for Windows.

3. Results and Discussion

3.1. Physicochemical Parameters of Leachate Landfill in Pilot-Scale CW Units

The temporal variation and statistics of the physicochemical parameters of the inflow and outflow of the pilot-scale CW units are presented in Figure 2 and Table 3, respectively. The temperature of the landfill leachate at the inlet of the pilot-scale CW units ranged between 7.0 °C and 36.2 °C and did not show a remarkable difference from that at the outlet of the units, which ranged between 7.0 °C and 36.1 °C (Figure 2a, Table 3). In general, the variation in the leachate temperature in all units showed the same pattern as that of the air temperature during the experimental period. The mean DO concentration in the effluent of the five pilot-scale CW units ranged from 4.27 mg/L to 5.14 mg/L and was higher than that in the influent (3.24 mg/L), probably due to the dissolution of atmospheric oxygen in the landfill leachate (Table 3). In the planted units, the average concentration of DO was higher than that in the unplanted unit (Figure 2b), a fact attributed to the action of plants, which transfer oxygen from the atmosphere to the wetland in the root zone [36].
The mean pH value at the inlet of the CW units was 8.11, while at the outlet of all the units it was slightly lower, ranging from 7.53 (W5FGZ-R unit) to 7.91 (W2FG-U unit) (Figure 2c, Table 3). Lower pH values were measured in the planted units, likely due to the interaction of the porous media and the biofilm that develops on the plant root system. According to Kadlec and Wallace [36], the plant life cycle generates organic compounds that are a source of natural acidity. Additionally, the nitrification process that occurs in CWs produces hydrogen ions that contribute to the reduction in pH. Similar results for the reduction in pH in CWs have been reported in other studies [35,40,43].
In the planted units, the mean effluent EC value was higher than that in the influent, in contrast to the unplanted unit, where the average effluent EC value was lower than that in the influent (Figure 2d, Table 3). In addition, higher EC values were measured in the planted units during the warmer months when evapotranspiration was higher. Therefore, evapotranspiration is an important factor affecting the EC value due to the condensation of leachate [44]. The increased EC value in the planted units compared to the unplanted ones is likely due in part to a higher concentration of ions in the leachate, resulting from the interaction of the plant root system with the substrate [40]. In the planted units, the maximum EC values ranged from 24.50 to 27.70 mS/cm and were observed on the hottest days of the experimental period. However, no problems were observed in the vegetation, indicating that the plants selected for planting the units are resistant to relatively high EC values.

3.2. Overall Performance of Pilot-Scale CW Units

Table 4 presents the statistics of pollutants in the influent and effluent of the five pilot-scale CW units. The temporal variation in pollutant removal rates in the pilot-scale CW units during the experimental period is presented in Figure 3. The concentrations of the pollutants studied at the outlets of the five units were lower than those at the inlets, indicating that removal occurred (Table 4). Regarding organic matter (BOD and COD), the mean effluent concentrations for the CW units W1FG-R, W2FG-U, W3FG-R, W4FG-C, and W5FGZ-R for BOD were 105.4, 133.5, 102.0, 85.7, and 77.0 mg/L, respectively, and for COD they were 1875, 2446, 1516, 1403, and 1354 mg/L, respectively (Table 4). The average removal of BOD was 45.9%, 32.2%, 47.7%, 55.4%, and 60.4% and of COD was 40.6%, 24.1%, 48.3%, 53.3%, and 56.7% in the CW units W1FG-R, W2FG-U, W3FG-R, W4FG-C, and W5FGZ-R, respectively (Figure 4). These results showed that the planted units achieved higher organic matter removal than the unplanted unit. Organic matter (OM) removal in CWs occurs by various processes, such as sedimentation and biodegradation by aerobic and anaerobic microorganisms [41]. The plants provide a surface for the growth of the biofilm and, as mentioned above, provide oxygen for additional oxygenation of the CWs’ substrate (Table 3). The moderate removal of OM is probably due to the very low (<0.1; Table 1) BOD/COD ratio of the landfill leachate, indicating that the leachate contains a high concentration of biologically recalcitrant organic substances.
The mean effluent concentrations of TKN and NH4-N at the five pilot-scale CW units were lower than the mean influent concentration, indicating the removal of these pollutants in the CW units (Table 4). The main mechanisms of nitrogen removal in CWs are nitrification, denitrification, removal as ammonia gas (volatilization), adsorption on porous media, and uptake by plants [36]. Nitrification and denitrification occur under aerobic and anoxic conditions, respectively, converting ammonia nitrogen first to nitrate (nitrification) and finally to nitrogen gas (denitrification). Ammonia volatilization occurs when the pH value is between 8 and 9. Adsorption requires an adsorbent material, such as zeolite, which has cation exchange sites.
The mean removal efficiencies for TKN were 40.9%, 36.6%, 45.5%, 45.9%, and 50.9%, and for NH4-N they were 49%, 40.8%, 49.4%, 50.4%, and 63.8%, in CW units W1FG-R, W2FG-U, W3FG-R, W4FG-C, and W5FGZ-R, respectively (Figure 4). For both TKN and NH4-N, the highest removal was achieved by the W5FGZ-R unit, and the lowest removal was achieved by the W2FG-U unit (unplanted unit). The removal of TKN and NH4-N in the pilot-scale CW units is mainly attributed to the following: (a) nitrification–denitrification as aerobic and anaerobic conditions in the CW units prevail, with the DO concentrations ranging from 0.33 to 8.58 mg/L (Table 3), (b) uptake by the plants in the planted units, and (c) adsorption of ammonia nitrogen by the zeolite of the W5FGZ-R unit. Ammonia volatilization was considered limited, as the mean pH values in the CW units ranged between 7.53 and 7.91. Research on CWs has shown that nitrogen removal through plant uptake accounts for 10% of total removal, with nitrification and denitrification being the primary removal processes [45].
EU Directive 1991/271/EEC imposes the following general effluent limits from wastewater treatment plants and for the discharge of treated effluent into natural water bodies: BOD 25 mg/L; COD 125 mg/L; TN 15 mg/L. Alternatively, this directive imposes minimum reduction rates, i.e., for BOD 70–90%, for COD 75%, and for TN 70–80% of the inflowing concentration. The average effluent values of all parameters from the CW units were higher than the limit set by the EU. However, large-scale CW systems usually involve three or more treatment stages. The first stage usually involves an anaerobic tank, followed by the remaining stages involving horizontal subsurface flow and vertical flow CWs. The various pollutants are effectively removed as the landfill leachate passes sequentially through the above stages, so that the final effluent meets the requirements of the legislation.
The performance of CWs is affected by temperature, precipitation, and seasonal variations, which can affect plant performance, microbial activity, and overall pollutant removal. Microbial activity is almost constant throughout the year in temperate and Mediterranean climates, including the one at the study site, and plant species like Typha latifolia and Phragmites australis thrive [36]. However, the efficacy of nitrogen removal by nitrification and denitrification may be diminished in colder climates, as microbial decomposition and plant growth slow down considerably (Weerakoon et al., 2013). Longer HRTs or hybrid wetland systems are needed in colder climates to compensate for decreased biological activity, according to studies [41]. Higher temperatures in tropical regions may promote microbial decomposition, but they also raise evapotranspiration, which causes leachate condensation in the wetland system.
The age of the landfill, the type of waste, and the regional environmental conditions affect the composition of the leachate. A major problem in mature landfills is the relatively low biodegradability (BOD/COD < 0.1), which is reflected in the leachates used in this investigation. According to Yulcuk and Ugurlu (2009), leachates from newer landfills usually contain more biodegradable organic matter (BOD/COD > 0.5), which can promote microbial decomposition but can also result in excessive biofilm formation and the clogging of CWs. Furthermore, increased levels of heavy metals and persistent organic pollutants in landfills receiving industrial waste may require pretreatment or the use of specialized adsorbents (such as activated carbon or modified zeolite) in order to avoid toxicity effects on microbial communities [46].
Several issues that affect treatment efficiency and system sustainability, such as clogging, plant death, and media aging, may affect the long-term operation of CWs. Reduced hydraulic conductivity and flow short-circuiting are the results of clogging, a major problem caused by excessive organic matter accumulation, biofilm overgrowth, and fine particle deposition [36,47]. Clogging can be mitigated by maintaining appropriate hydraulic loading rates, routinely checking flow patterns, and pretreating the leachate in an anaerobic tank or in an Imhoff settling tank [41,48]. The treatment’s effectiveness may also be lowered by plant death brought on by harsh weather, hazardous pollutants, or inadequate oxygen transmission [49]. A stable system performance can be guaranteed by choosing hardy plant species like Typha latifolia and Phragmites australis and by replanting the area on a regular basis.

3.3. Effect of Vegetation

The effect of plants on pollutant removal was tested by comparing the removal efficiency of three units with the same porous media (fine gravel): one unplanted (W2FG-U), one planted with common reed (W3FG-R), and one planted with cattail (W4FG-C) (Table 2). The unit with cattail achieved the best performance in removing all pollutants, while the unplanted unit had the worst. The one-way between-groups ANOVA and Tukey’s HSD test (Table 5) showed a statistically significant difference in the average OM removal between units W2FG-U and W3FG-R and between W2FG-U and W4FG-C. In addition, there is a statistically significant difference in the average removal of nitrogen forms (TN and NH4-N) between units W2FG-U and W4FG-C (Table 5). Therefore, we conclude that the presence of plants creates favorable conditions in the CW, enhancing its effectiveness in removing these pollutants. As mentioned previously, plants transport oxygen to the rhizosphere, which is necessary for the growth of microorganisms, the decomposition of organic compounds, and nitrification. They also serve as a support medium by providing a surface for biofilm growth on the stems and roots [35,50]. A crucial stage in the removal of nitrogen is the conversion of ammonium to nitrate, which is made easier by the oxygen expelled from plant roots. Furthermore, organic substances expelled by plant roots provide heterotrophic bacteria with carbon sources, facilitating denitrification and the breakdown of organic compounds. By stabilizing substrate particles, vegetation also improves adsorption and sedimentation processes, increasing the overall effectiveness of pollutant removal [36,51].
The broad root and rhizome systems of both species offer a large surface area for pollutant absorption, sediment stabilization, and microbial colonization. The thick root systems of Typha latifolia and Phragmites australis improve sedimentation and filtration by physically trapping suspended particles [35]. Typha latifolia creates favorable circumstances for denitrification in deeper, anoxic zones, and while it provides enough aeration to support nitrifying bacteria, it is somewhat less effective at transferring oxygen [52]. The dynamics of microbial communities are greatly influenced by the presence of Typha latifolia and Phragmites australis in CW. Numerous microbial consortia, such as nitrifiers (i.e., Nitrosomonas and Nitrobacter) and denitrifiers (i.e., Pseudomonas and Paracoccus), which are essential for the nitrogen cycle, are found in root-associated biofilms [41]. Furthermore, aerobic heterotrophic bacteria, which decompose organic matter and improve COD and BOD removal, are supported by plant-mediated oxygenation. Studies have shown that, compared to unplanted systems, planted CWs support greater microbial diversity and enzymatic activity, which improves pollutant degradation [50].

3.4. Effect of Hydraulic Residence Time and Porous Media

Previous studies have shown that hydraulic residence time affects the efficiency of CW units in removing various pollutants [36]. In this study, the effect of HRT on pollutant removal was examined by comparing the removal efficiency of the W1FG-R and W3FG-R units, which have the same construction characteristics but differ in HRT (8 and 10 days, respectively) (Table 2). In addition, the two units were in the same area and operated simultaneously under the same climatic conditions. The average removal rates for each unit and the t-test results are presented in Table 6. The W3FG-R unit with an HRT of 10d exhibited higher average removal rates for all pollutants, indicating that longer HRTs increase both OM and nitrogen removal. Statistical analysis showed that the COD removal at an HRT of 10 days is statistically significantly higher (p < 0.05) than that of 8 days. HRT values longer than ten days may further improve unit performance. According to previous studies, a longer HRT contributes to the creation of microbial communities and at the same time increases the contact time of pollutants with the biofilm, enhancing the performance of the CW [41,49,52,53]. Chiemchaisri et al. [54] reported removal efficiencies of 55%, 42%, and 41% for BOD, COD, and TKN, respectively, at an HRT of 10 days in HSF CWs, which are comparable to those of the present study. By increasing the HRT to 28 days, the BOD, COD, and TKN removal efficiencies increased to 71%, 58%, and 46%, respectively, while decreasing the HRT to 5 days resulted in reductions to 44%, 33%, and 20%, respectively. Yalcuk and Ugulu [40] also reported comparable BOD and NH4-N removal efficiencies of 35.7% and 38.3%, respectively, in HSF CWs at an HRT of 12.5 days. Furthermore, similar results to the BOD, COD, and NH4-N removal efficiencies of the present study are reported in the literature [55,56]. These results show that HRT plays an important role in pollutant removal. With longer HRT values (e.g., 28 days vs. 8 days), more thorough biological, chemical, and physical treatment processes are possible. Since ammonium is oxidized to nitrate in aerobic zones, and nitrate is reduced to nitrogen gas in anaerobic conditions, an extended HRT improves nitrification–denitrification cycles [49]. Furthermore, because microbial populations have more time to decompose complex organic substances, a prolonged HRT promotes the decomposition of organic compounds [53,57].
The effect of porous media on pollutant removal was examined by comparing the removal efficiency of the W3FG-R and W5FGZ-R units. Both units were planted with common reed and had the same operational characteristics, but the porous media differed. In the W3FG-R unit, the porous medium was FG, while in the W5FGZ-R unit, the media were composed of 75% FG and 25% FZ (Table 2). The W5FGZ-R unit exhibited higher average removal rates than the W3FG-R unit for all pollutants (i.e., BOD, COD, TKN, and NH4-N). The t-test showed statistically significant differences (p < 0.05; Table 6) for the average removal values of BOD, COD, and NH4-N. These results suggest that the presence of zeolite in the substrate enhances the removal of both OM and nitrogen. Zeolite is a low-cost material and a very good adsorbent for cations such as ammonium ions. It improves cation exchange and ammonium adsorption because of its high surface area and ion exchange capabilities. Furthermore, it serves as a site for microbial biofilm growth, as the porosity of the material may favor and enhance the growth of microorganisms [34]. According to studies [46], zeolite enhances nitrogen removal by temporarily binding ammonium, delaying rapid leaching, and making it available for microbial decomposition. As nitrifying bacteria within the biofilm use ammonium adsorbed to the zeolite particles, the ion exchange capacity of the zeolite is bio-regenerated [58]. These properties make zeolite a popular material for use as a filler in constructed wetland units for the removal of ammonium and metal ions. It has been used in CW systems by many researchers for the removal of ammonia and metals in high-strength wastewaters, such as landfill leachate [46,59].

4. Conclusions

The performance of the units was affected by the design and operating conditions. Key factors such as vegetation, HRT, and the composition of the porous media were found to significantly impact the effectiveness of the CWs. These factors should be carefully considered during the design of CW systems. An HRT greater than 10 days is likely to enhance the performance of the units. Additionally, the presence of zeolite in the porous media delivered promising results, improving the performance of the unit in the removal of all the pollutants studied. Therefore, CW systems can be considered a sustainable alternative to conventional treatment methods, providing environmental, economic, and social benefits.

Author Contributions

Conceptualization, G.D.G.; Methodology, G.D.G.; Formal Analysis, G.D.G., I.N., I.-E.M., I.P., P.K. and P.P.; Investigation, G.D.G., I.N., I.-E.M., I.P., P.K. and P.P.; Resources, G.D.G.; Writing—Original Draft Preparation, I.N., G.D.G. and P.P.; Writing—Review and Editing, G.D.G., I.N. and P.P.; Visualization, G.D.G.; Supervision, G.D.G. All authors have read and agreed to the published version of the manuscript.

Funding

This research received no external funding.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The data supporting the research findings of this study are available from the corresponding author upon request.

Conflicts of Interest

The authors declare no conflicts of interest.

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Applsci 15 02595 g001
Figure 1. (a) View of the five pilot-scale HSF CWs and (b) schematic description of CW units.
Figure 1. (a) View of the five pilot-scale HSF CWs and (b) schematic description of CW units.
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Figure 2. Temporal variation in physicochemical parameters of CW units during the experimental period for (a) temperature (T), (b) dissolved oxygen (DO), (c) pH, and (d) electrical conductivity (EC).
Figure 2. Temporal variation in physicochemical parameters of CW units during the experimental period for (a) temperature (T), (b) dissolved oxygen (DO), (c) pH, and (d) electrical conductivity (EC).
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Figure 3. Temporal variation in removal rate % of CW units during the experimental period for (a) BOD, (b) COD, (c) TKN, and (d) NH4-N.
Figure 3. Temporal variation in removal rate % of CW units during the experimental period for (a) BOD, (b) COD, (c) TKN, and (d) NH4-N.
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Figure 4. Pollutant (BOD, COD, TKN and NH4-N) removal chart for pilot-scale CW units.
Figure 4. Pollutant (BOD, COD, TKN and NH4-N) removal chart for pilot-scale CW units.
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Table 1. Characteristics of landfill leachate depending on landfill age and comparison with the leachate of the present study [10,12,32,33].
Table 1. Characteristics of landfill leachate depending on landfill age and comparison with the leachate of the present study [10,12,32,33].
ParametersLandfill AgePresent Study
YoungMiddle-AgedOld
Age (years)<10 years10–20 years>20 years
pH3–66–7>7.58.24 3 (±0.67)
1 EC (mS/cm)15–41.56–14 28.5 (±5.2)
BOD (mg/L)10,000–25,0001000–400050–1000685 (±102)
COD (mg/L)15,000–40,00010,000–20,0001000–50009344 (±1956)
BOD/COD0.6–0.70.1–0.2<0.10.07 (±0.005)
2 TKN (mg/L)1500–4500400–80075–300762 (±151)
NH4-N (mg/L)1500–4250250–70050–200579 (±97)
1 EC = electrical conductivity; 2 TKN = total Kjeldahl nitrogen; 3 (±n) = standard deviation.
Table 2. Design characteristics and operational conditions of pilot-scale CWs.
Table 2. Design characteristics and operational conditions of pilot-scale CWs.
CW UnitPorous MediaPlant SpeciesHRT (Days)Qin (L/Day)HLR (mm/Day)
W1FG-RFGPhragmites australis89.519.4
W2FG-UFGNone107.615.3
W3FG-RFGPhragmites australis107.615.3
W4FG-CFGTypha latifolia107.615.3
W5FGZ-R75% FG and 25% FZPhragmites australis107.615.3
FG = fine gravel, FZ = fine zeolite, HRT= hydraulic residence time, Qin = influent rate, HLR = hydraulic loading rate.
Table 3. Statistics of the physicochemical parameters in each pilot-scale CW unit.
Table 3. Statistics of the physicochemical parameters in each pilot-scale CW unit.
ParametersInfluent Effluent
W1FG-RW2FG-UW3FG-RW4FG-CW5FGZ-R
Τ (°C)mean21.521.020.820.721.121.1
SD (n) 16.4 (106)6.4 (106)6.6 (106)6.6 (106)6.8 (106)6.7 (106)
min7.07.07.17.07.27.7
max36.235.836.133.534.834.3
DO
(mg/L)		mean3.244.494.274.504.415.14
SD (n)1.18 (106)1.40 (106)1.10 (106)1.39 (106)1.49 (106)1.19 (106)
min0.331.661.780.571.372.13
max5.018.586.917.657.748.14
pHmean8.117.737.917.707.657.53
SD (n)0.87 (106)0.71 (106)0.81 (106)0.71 (106)0.610.66 (106)
min6.116.116.116.116.235.98
max9.478.688.778.578.478.45
EC
(mS/cm)		mean7.379.806.0410.459.6110.38
SD (n)0.88 (106)4.71 (106)1.05 (106)5.55 (106)4.50 (106)6.31 (106)
min4.054.043.693.543.563.12
max9.5027.709.4825.8024.5027.00
1 SD = standard deviation, n = number of measurements.
Table 4. Statistics of pollutant concentrations in each pilot-scale CW unit.
Table 4. Statistics of pollutant concentrations in each pilot-scale CW unit.
ParametersInfluentEffluent
W1FG-RW2FG-UW3FG-RW4FG-CW5FGZ-R
BOD
(mg/L)		mean195.9105.4133.5102.085.777.0
SD (n) 129.8 (46)31.2 (46)32.8 (44)36.5 (42)19.5 (44)20.0 (44)
min146.050.756.328.051.547.5
max300.0180.0200.0197.0130.5131.0
COD
(mg/L)		mean311518752446151614031354
SD (n)652 (46)418 (45)706 (41)422 (45)318 (43)438 (44)
min206510651065710898900
max520727954312284220853751
TKN
(mg/L)		mean345.3204.1220.1202.3201.9172.8
SD (n)60.7 (46)75.2 (45)74.1 (45)72.3 (45)87.6 (43)65.8 (43)
min210.0101.0119.084.091.084.0
max483.0350.0385.0317.3396.7308.0
NH4-N (mg/L)mean231.6117.3134.9115.4114.081.4
SD (n)38.9 (46)42.8 (44)39.8 (45)40.4 (43)45.2 (43)25.8 (42)
min140.035.049.021.041.042.0
max308.0219.3266.0210.0196.0137.0
1 SD = standard deviation, n = number of measurements.
Table 5. One-way ANOVA and Tukey HSD test results for comparison of pollutant removal efficiencies of W2FG-U, W3FG-R, and W4FG-C units.
Table 5. One-way ANOVA and Tukey HSD test results for comparison of pollutant removal efficiencies of W2FG-U, W3FG-R, and W4FG-C units.
One-Way ANOVAPost Hoc (Tukey HSD) Test
ParameterFpCW (I)–CW (J)Mean Difference (I–J)p
BOD25.1060.001W2FG-U–W3FG-R−15.510.001 *
W2FG-U–W4FG-C−23.260.001 *
W3FG-R–W4FG-C−7.740.061
COD54.4120.001W2FG-U–W3FG-R−24.230.001 *
W2FG-U–W4FG-C−29.200.001 *
W3FG-R–W4FG-C−4.960.209
TKN3.5480.032W2FG-U–W3FG-R−8.930.063
W2FG-U–W4FG-C−9.260.045 *
W3FG-R–W4FG-C−0.320.996
NH4-N3.7220.027W2FG-U–W3FG-R−8.590.071
W2FG-U–W4FG-C−9.520.040 *
W3FG-R–W4FG-C−0.930.969
* The mean difference is significant at the 0.05 level.
Table 6. t-test results between mean values of pollutant removal: of W1FG-R and W3FG-R units; of W3FG-R and W5FGZ-R units.
Table 6. t-test results between mean values of pollutant removal: of W1FG-R and W3FG-R units; of W3FG-R and W5FGZ-R units.
ParameterMean Values of Pollutant Removal (%)t-Test for Equality of Means
W1FG-RW3FG-Rtdfp (2-Tailed)
BOD45.947.7−0.486860.628
COD40.648.3−2.802880.006 *
TKN40.945.5−1.214880.228
NH4-Ν49.049.4−0.119850.905
W3FG-RW5FGZ-R
BOD47.760.4−3.740630.001 *
COD48.356.7−3.091870.003 *
TKN45.550.9−1.356860.179
NH4-Ν49.463.8−4.334780.001 *
* Statistically significantly different mean values.
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Ntountounakis, I.; Margaritou, I.-E.; Pervelis, I.; Kyrou, P.; Parlakidis, P.; Gikas, G.D. Pollutant Removal Efficiency of Pilot-Scale Horizontal Subsurface Flow Constructed Wetlands Treating Landfill Leachate. Appl. Sci. 2025, 15, 2595. https://doi.org/10.3390/app15052595

AMA Style

Ntountounakis I, Margaritou I-E, Pervelis I, Kyrou P, Parlakidis P, Gikas GD. Pollutant Removal Efficiency of Pilot-Scale Horizontal Subsurface Flow Constructed Wetlands Treating Landfill Leachate. Applied Sciences. 2025; 15(5):2595. https://doi.org/10.3390/app15052595

Chicago/Turabian Style

Ntountounakis, Ioannis, Ioanna-Eirini Margaritou, Ioannis Pervelis, Pavlos Kyrou, Paraskevas Parlakidis, and Georgios D. Gikas. 2025. "Pollutant Removal Efficiency of Pilot-Scale Horizontal Subsurface Flow Constructed Wetlands Treating Landfill Leachate" Applied Sciences 15, no. 5: 2595. https://doi.org/10.3390/app15052595

APA Style

Ntountounakis, I., Margaritou, I.-E., Pervelis, I., Kyrou, P., Parlakidis, P., & Gikas, G. D. (2025). Pollutant Removal Efficiency of Pilot-Scale Horizontal Subsurface Flow Constructed Wetlands Treating Landfill Leachate. Applied Sciences, 15(5), 2595. https://doi.org/10.3390/app15052595

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