Constructed wetlands for landfill leachate treatment: A review

Constructed wetlands for landfill leachate treatment: A review

Ecological Engineering 146 (2020) 105725 Contents lists available at ScienceDirect Ecological Engineering journal homepage:

3MB Sizes 0 Downloads 64 Views

Ecological Engineering 146 (2020) 105725

Contents lists available at ScienceDirect

Ecological Engineering journal homepage:


Constructed wetlands for landfill leachate treatment: A review a

Reza Bakhshoodeh , Nadali Alavi Jan Vymazalg, Pooya Paydaryh




T e,f

, Carolyn Oldham , Rafael M. Santos , Ali Akbar Babaei ,


Department of Civil, Environmental and Mining Engineering, The University of Western Australia, Perth, Australia Environmental and Occupational Hazards Control Research Center, Shahid Beheshti University of Medical Sciences, Tehran, Iran Department of Environmental Health Engineering, School of Public Health, Shahid Beheshti University of Medical Sciences, Tehran, Iran d School of Engineering, University of Guelph, Guelph, ON, Canada e Environmental Technologies Research Center, Ahvaz Jundishapur University of Medical Sciences, Ahvaz, Iran f Department of Environmental Health Engineering, School of Public Health, Ahvaz Jundishapur University of Medical Sciences, Ahvaz, Iran g Czech University of Life Sciences Prague, Faculty of Environmental Sciences, Czech Republic h Department of Civil and Environmental Engineering, Northeastern University, Boston, MA, USA b c



Keywords: Constructed wetlands Landfill leachate Nutrient treatment Heavy metals mobility Organics biodegradation

Constructed wetlands (CWs) are engineered systems that are constructed to mimic natural wetlands. These systems simulate the processes that happen in natural wetlands and remove pollutants from wastewater. CWs have been previously used to treat a wide range of waste streams, including landfill leachate, and they have shown relatively good removal efficiencies. Although the literature on using CWs to treat different waste streams has been previously reviewed, there has been no literature review of constructed wetlands for leachate treatment. This critical analysis of available literature on landfill leachate treatment by CWs will help optimize future research in the field. The primary objective of this article is to present a comprehensive overview of the diverse range of practices, applications and research into the use of CW systems for removing contaminants from landfill leachate. This review of 85 papers across 20 countries focusses on the treatment performance of three different types of CWs (free water surface flow; subsurface flow (horizontal and vertical); and hybrid systems), using data from field- and pilot-scale studies, and discusses the impact of design criteria on CW treatment performance. The reported average % removal efficiencies of BOD5, COD, TP, PO4, Ammonia-N, TKN, TN, and TSS, for Horizontal/ Vertical/Hybrid/Free water surface CWs were, respectively: 60.1/79.7/72.2/80.6%; 54.5/59.2/56.2/45.4%; 63.5/46.2/−6.4/5.5%; 67.7/62.1/5.2/−1.6%; 67.2/66.7/68.9/70.0%; 45.4/64.2/64.9/10.6%; 72.1/88.2/ 67.3/81.7%; and 69.3/55.5/51.8/59.5%.

1. Introduction Growing population and increasingly consumptive lifestyles have caused a significant growth in the production of municipal solid waste (MSW). Leachate produced by MSW landfill, and its composting, is a major problem for solid waste facilities. Leachate usually contains high concentrations of pollutants and toxic compounds; as a result, it can cause severe threats to the environment and human health, and requires appropriate treatment before release. Treatment of landfill leachate is typically undertaken by up flow anaerobic sludge blankets (UASB) (Castillo et al., 2007), activated sludge, reverse osmosis, ion exchange, ultrafiltration, sequencing batch reactors (SBR) (He et al., 2007) and chemical flocculation (Marañón et al., 2008) Each of these methods requires high construction, maintenance and operation costs. In contrast, constructed wetlands (CW) ⁎

may be considered more sustainable and environmentally friendly solutions for landfill leachate treatment. CWs are engineered systems that are designed and constructed to mimic the processes in natural wetlands. CWs have been shown to use these processes to remove contaminations from different wastewater streams (Vymazal, 2005). In CWs, wastewater is treated by physical processes (e.g., sedimentation, filtration), chemical (e.g., precipitation, adsorption), as well as biological and chemical processes (e.g., microbial degradation, uptake from the water column, and root zone) (Midhun et al., 2016). CWs have been used to treat a wide range of waste streams, including mine drainage (Sheoran and Sheoran, 2006; Sheridan et al., 2018), stormwater runoff (Kadlec and Wallace, 2009; Choi et al., 2015; Li et al., 2017), domestic wastewater (Vymazal and Kröpfelová, 2008; Kadlec and Wallace, 2009; Matamoros et al., 2017), industrial effluents (Vymazal, 2014), agricultural drainage (Tanner et al., 2005;

Corresponding author at: Environmental and Occupational Hazards Control Research Center, Shahid Beheshti University of Medical Sciences, Tehran, Iran. E-mail address: [email protected] (N. Alavi). Received 26 May 2019; Received in revised form 18 December 2019; Accepted 19 January 2020 0925-8574/ © 2020 Published by Elsevier B.V.

Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

other hand, HF CWs provide suitable conditions for denitrification, i.e., removal nitrate. VF CWs are more prone to clogging and have higher capital, operation and maintenance costs (Bakhshoodeh et al., 2017a). Sub-surface flow CWs are better suited for cold climates, due to the lack of surface water and usually require less land than FSW CWs. Hybrid CWs combine different types of wetland systems to optimize treatment performance. Most common hybrid systems are those combining HF and VF CWs staged in series. The VF-HF CW was designed in the 1960s (Seidel, 1966) while HF-VF CW was developed in mid-1990s (Brix et al., 2003). Hybrid CWs usually have higher treatment efficiency than non-hybrid systems, especially for total nitrogen, but they typically require more space and are more expensive to build (Bakhshoodeh et al., 2017b).

Fig. 1. Free surface water CW.

Kynkäänniemi et al. 2013; Sheridan et al., 2014) as well as landfill leachate (Mulamoottil et al., 1999; Yin et al., 2017). The purpose of this study is to review the use of CWs for landfill leachate treatment, summarize treatment performance measured in field and pilot-scale studies, and finally discuss the impact of design criteria on CW treatment performance.

3. Leachate treatment studies using CWs The previous studies that used constructed wetlands to treat landfill leachate, which are discussed in the following sub-sections, have been summarized in Tables S-1 and S-2 in the supporting information (SI). Table S-1 summarizes studies performed on non-metallic contaminants, while Table S-2 is focused on heavy metals. All the hybrid CWs (regardless of their configuration) have been categorized as hybrid, and studies where the same CW was used but a design parameter (e.g., hydraulic retention time or type of plants) was changed are presented as two different studies.

2. Types of constructed wetlands Constructed wetlands are usually categorized based on the presence/absence of water on the surface (free water surface flow or subsurface flow) or flow direction (vertical or horizontal). Different hybrid CWs can also be set up by combining different types of constructed wetlands (Vymazal, 2005). Free water surface CWs (FWS CWs, Fig. 1) are shallow (usually 5–40 cm) basins planted with wetland macrophytes (Kadlec and Knight, 1996). There is no special requirement for soil quality, the major function of soil is to support the growth of vegetation. Treatment occurs in the water column and in the layer of decaying litter on the bottom. Sub-surface flow CWs can be categorized into the horizontal flow (HF) and vertical flow (VF) CWs (Fig. 2). HF CWs are continuously fed with wastewater and, therefore the filtration bed is mostly anoxic with aerobic zones occurring only in zones adjacent to roots (Vymazal, 2005). VF CWs are fed intermittently and allow for oxygen diffusion into filtration bed once the bed is empty after the water percolated to the bottom. Therefore, VF CWs have higher oxygen transfer capacity as compared to HF CWs, resulting in better removal of ammonia. On the

3.1. Leachate characteristics The physical and chemical composition of landfill leachate depends on many parameters, including but not limited to, the composition of the solid waste, climate, age of the landfill, and the site hydrogeology (Mulamoottil et al., 1999). However, leachate usually contains high concentrations of organic matter, heavy metals, nitrogen, phosphorous, and salts, with a pH in the neutral range. Table S-1 presents the landfill leachate characteristics for the reviewed CW studies. Organic matter concentrations (COD and BOD5) cover a wide range and are typically very high (Table S-1). COD varies from 65 to 84,200 mg/L, while BOD5 varies from 15 to 64,400 mg/L, with the most frequent values being observed between 500 and 2000 mg/L for COD and 100–200 mg/L for BOD5. BOD5/COD ratios typically range between 0.03 and 0.76, where BOD5/COD ratio is used as an indicator of biodegradability of organic matter. Landfills less than 3–5 years old may have BOD5/COD ratios as high as 0.7 (Surmacz-Gorska (2001), while more mature landfills (5–10 years old), have lower BOD5/COD ratio around 0.3–0.5. Most of the studies report ammonia-N concentrations but few report TN or TKN. Ammonia-N concentrations varied between 1 and 2865 mg/ L. Comparing ammonia-N values with TKN or TN (where reported) shows that, as expected, most of the TKN or TN was in the form of NH3. Reported landfill leachate pH values ranged from 5.8–8.5 (Table S-1). TSS values covered a wide range from 22 mg/L to 12,480 mg/L, with the most frequent values in the range 100–800 mg/L. Most of the studies were conducted in Asia (45%) and Europe (41%), while few were conducted in North America (12%) and none in Africa. Most studies focused on the treatment of ammonia-N and COD (Table 1). Comparison of landfill leachate parameter values across Asia, Europe and America (Table 1) shows that landfill leachate used in the CWs studies in Asia is more contaminated in almost all the quality parameters (except TN and EC). Higher levels of urbanization directly affects the composition of waste, when associated with growing incomes and increased affluence; higher consumerism tends to generate more packaging materials, with higher paper and plastic content (Idris et al., 2004).

Fig. 2. a) Horizontal and b) Vertical subsurface CWs. 2

Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

Table 1 Parameters monitored during treatment of landfill leachate in constructed wetlands, and their range of values, in the reviewed literature. Region


BOD5 (mg/L)

COD (mg/L)


Ammonia-N (mg/L)

NO3 (mg/L)

TN (mg/L)

TKN (mg/L)

TP (mg/L)

PO4 (mg/L)

TSS (mg/L)

5275 15.5 635 ( ± 350) 918 12.25 283 ( ± 241) 792 25 272 ( ± 268) 53

9335 13 2037 ( ± 1850) 1108 65.5 635 ( ± 245) 4770 445 1815 ( ± 1333) 64

0.74 0.03 0.34 ( ± 0.19) 0.53 0.06 0.30 ( ± 0.2) 0.5 0.05 0.20 ( ± 0.15) 43

2865 4 340 ( ± 189) 253 0.78 137 ( ± 63) 286.5 23.5 331 ( ± 588) 67

64 0.21 11.4 ( ± 8.5) 166 0 20 ( ± 46) 159 0.1 91 ( ± 48) 51

400 12.03 76 ( ± 42) 384 384 384

6176 25.5 651 ( ± 495) 132 1.6 88 ( ± 61) – – –

75 0.09 12.3 ( ± 9.2) 0.8 0.8 0.8

117 3.4 26.3 ( ± 14.8) 7.3 7.3 7.3

5.5 0.07 2.2 ( ± 0.6) 22

5 0.113 2 ( ± 1.9) 21

1164 25.4 458 ( ± 369) 50.5 9.5 30 ( ± 29) 2720 22 590 ( ± 999) 29


Max Min Mean average


Max Min Mean average

8.55 5.4 7.5 ( ± 0.85) 8.2 5.3 7.2 ( ± 1.1)


Max Min Mean average

8.43 7 7.7 ( ± 0.5)

Total number of studiesa a


684 2.25 280 ( ± 189) 26


Complete data and references are found in the Supplementary Materials.

processes and process conditions (e.g. flow rate, temperature), but also by the biodegradability of various compounds and availability of oxygen. As discussed in the literature, major organic matter removal pathways in CWs are sedimentation and filtration of colloid organics, and biological decomposition by microorganisms under aerobic, facultative and anaerobic conditions (Vymazal and Kröpfelová, 2009). In other words, flowing particles can absorb the organic contaminants into the wetland, and then the settled particles are broken down by microorganisms. Organic molecules are broken down by the microbiota by fermentation and aerobic/anaerobic respiration, are mineralized as a source of energy, or are assimilated into biomass. The organic carbon degradation rate and efficiency depends on the organic compound present in the influent. Volatilization may also be a significant removal mechanism in the microbial removal of organic contaminants. Fig. 4 shows BOD5 and COD content and BOD5/COD ratios in the influent versus the effluent of different types of CWs. Fig. 4-a. shows BOD5out versus BOD5in for different types of CWs. It can be observed that for studies with higher BOD5in values (more than 300 mg/L) mostly hybrid constructed wetlands have been used and they have been able to lower BOD5 considerably. HF for higher BOD5 values was not as successful as hybrid CWs. For lower BOD5 values, all CW types almost yielded the same results. Fig. 4-b. shows that almost all types of CWs give the same COD removal efficiency and the observed removal efficiency is not a function of the influent COD concentration. Studies with hybrid CWs have had high BOD5in concentrations and CODin concentrations (lower than 2000 mg/L). Fig. 4-c shows BOD5out/CODout versus BOD5in/CODin. It can be observed that in CWs BOD5/COD ratio is lower in the effluent comparing to the influent, as most of the points fall below the y = x line. These means biodegradation pathways are active in organics removal in CWs. However, it seems like regardless of CW type BOD5/CODin versus BOD5/CODout tends to stay on a straight line which means BOD5/COD ratio gets reduced in all types of wetlands with the same ratio. Fig. 4-c shows that the influent BOD/COD ratio varies widely, from less than 0.1 up to nearly 0.8, indicating the variable biodegradability of wastewater (Scholz and Hedmark, 2009), while the BOD/COD ratio is almost always reduced at the effluent, by as much as approximately half, regardless of the CW type, which confirms that biodegradation is the most dominant organic carbon removal mechanism. BOD removal efficiency is independent of BOD/COD inlet ratio (Fig. 4-d).

3.2. Constructed wetland types and contaminants removal Although all types of CWs provide similar treatment efficiencies, hybrid CWs are more efficient in removing contaminants (Fig. 3; see Table 2 for the number of papers in each category). Hybrid CWs on average displayed removals of 72.2 ± 28% for BOD5, 56.2 ± 23% for COD, −6.4 ± 4% for TP, 61.0 ± 18% for PO4, 64.9 ± 25% for TKN, 67.3 ± 28% for TN, 68.9 ± 27% for ammonia-N, and 51.8 ± 29% for TSS. VFs showed higher removal efficiencies for heavy metals. FWS appears to be the least efficient for landfill leachate treatment. The relation between inlet and outlet of parameters in this review paper in comparison with other studies (Table S-3). 3.2.1. Organic matter FWS, HF, hybrid, and VF CWs have removals of 80.6 ± 17%, 60.1 ± 17%, 72.2 ± 28%, and 79.7 ± 13% respectively for BOD5 and 45.4 ± 36%, 54.5 ± 25%, 56.2 ± 23%, and 59.2 ± 22% respectively for COD. These ranges may not be caused only by different

3.2.2. Nitrogen Ammonia removal in constructed wetlands occurs through different mechanisms: volatilization (in FWS CWs); nitrification (under aerobic conditions); adsorption (very limited); uptake by plants and other living organisms; and anammox (under anaerobic conditions) (Dong and Sun, 2007). It is important keep in mind that ammonification of the

Fig. 3. Removal efficiency (mean average change in effluent concentration versus influent) of different types of CWs in removing different contaminants from landfill leachate. Bars represent standard deviations. 3

Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

Table 2 Number of studies used in Fig. 3, for different CW types.

FWS HF Hybrid VF Total number of studies a



















5 (8%) 26 (41%) 22 (35%) 10 (16%) 63

6 (7%) 37 (51%) 15 (21%) 16 (22%) 73

7 (7%) 51 (48%) 33 (31%) 29 (15%) 107

3 (5%) 24 (36%) 19 (29%) 20 (30%) 66

3 (13%) 8 (35%) 6 (26%) 6 (26%) 23

3 (10%) 14 (45%) 12 (39%) 2 (6%) 31

0 (−) 12 (52%) 8 (35%) 3 (13%) 23

4 (14%) 9 (31%) 9 (31%) 7 (24%) 48

7 (19%) 20 (56%) 5 (14%) 4 (11%) 36

3 (10%) 5 (17%) 9 (30%) 13 (43%) 30

2 (7%) 5 (17%) 1 (40%) 11 (37%) 30

1 (9%) 4 (36%) 3 (27%) 3 (27%) 11

0 (0%) 6 (38%) 1 (6%) 9 (56%) 16

4 (13%) 5 (17%) 9 (30%) 12 (40%) 30

0 (0%) 6 (29%) 4 (19%) 11 (52%) 21

4 (14%) 5 (18%) 8 (29%) 11 (39%) 28

4 (15%) 7 (27%) 4 (15%) 11 (42%) 26

Total number of studies-some studies have several parameter values in different situations.

Ammonia-N is oxidized to NO3 and there is not enough organics to support denitrification or the organics are recalcitrant and cannot be easily degraded. The increase in concentration of nitrate in the outflow (Fig. 3) is affected by low NO3 values in the influent. It has been shown that an important limiting factor in the conversion of Ammonia-N to nitrate (nitrification) in CWs is the presence of oxygen. It has been shown previously that presence of high organics loading leads to extremely low nitrification, mainly due to the lack of oxygen in the filtration bed which is preferentially used by bacteria to oxidize organic matter (O'luanaigh et al., 2010; Wu et al., 2011), making nitrification process the limiting step in ammonia removal. For FWS, where ammonia-N removal is about 60%, NO3 removal/ generation is small and close to zero. This might be due to the volatilization of ammonia-N in FWS as water surface is open to the atmosphere or uptake of ammonia by algae and cyanobacteria which may occur in FWS CWs in large quantities. Almost all Hybrid CWs used for landfill leachate treatment contained a VF stage (Table S-1), which is the most efficient CW in

nitrogen-containing organics increases the ammonia concentration. Nitrate is removed primarily by denitrification under anoxic conditions and plant uptake. (Alavi et al., 2011; Saeed et al., 2012). However, it is important to realize that plant uptake as a removal mechanism is effective only in case the plants are harvested and removed from the system. Otherwise nitrogen returns back in thy system during plant decomposition (Vymazal, 1999; Bakhshoodeh et al., 2017c). The removal of ammonia decreased in the order of hybrid, VF, HF, and FWS CWs. The order of Ammonia-N removal and NO3 generation confirms oxygenation efficiency ranking reported previously for different types of the CWs (Vymazal, 1999; Saeed and Sun, 2017). Looking at the NO3 removals, it can be observed that hybrid, VF, and HF have negative removal efficiencies, showing higher NO3 concentrations in the effluent compared to the influent which indicates ammonium conversion to nitrate. This conversion is an important step as denitrification is the most dominant process in nitrogen removal. The order observed in the NO3 removal rate, inversely corresponds to the order observed in the Ammonia-N removal, showing that in these wetlands

Fig. 4. a) BOD5, b) COD, and c) BOD5/COD ratio in the influent versus effluent, and d) correlation between BOD/COD ratio of influent and the BOD removal efficiency, of different types of CWs. 4

Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

Fig. 5. a) Ammonia-N, b) NO3, c) TKN, and d) TN in the influent and effluent of different types of CWs. The line indicates the 1:1 relationship.

precipitation with Al, Fe, Ca, and Mg being the most important processes (Reddy et al., 1999; Drizo et al., 2000). Biological removal is very low as uptake of phosphorus by microbiota is only temporary and uptake by macrophytes can be considered as “removal” mechanism only in case that macrophytes are harvested. If macrophytes are not harvested, phosphorus is released back in water during decomposition of the biomass and only small amount of phosphorus is not released and becomes recalcitrant (Vymazal and Kröpfelová, 2008). Particulate phosphorus can be also removed through filtration by the CWs media or in FWS CWs can be removed via sedimentation. Sorption and filtration removal mechanisms have a limited capacity and can be exhausted as CWs media may get saturated with adsorbed or filtered phosphorus (Kadlec and Wallace, 2009; Li et al., 2013; Martín et al., 2013; Hayder

oxygenation. In hybrid and VF CWs investigated, TN removal was similar then followed by HF and FWS (Fig. 3). The concentration of the investigated nitrogenic compounds in the influents and effluents are presented in Fig. 5. 3.2.3. Total phosphorus and phosphate Average removal of phosphate amounted to −1.6 ± 0.7%, 67.7 ± 28%, 5.2 ± 2%, and 62.1 ± 26% in FWS, HF, Hybrid, and VF CWs, respectively while average removal of total phosphorous amounted to 5.5 ± 3%, 63.5 ± 23%, −6.4 ± 4%, and 46.2 ± 35% in FWS, HF, Hybrid, and VF CWs, respectively (Figs. 3 and 6). It has been reported that phosphorus is removed in CWs mainly by chemical of physical-chemical pathways with sorption and

Fig. 6. a) PO4 and b) TP content ratio in the influent and effluent of different types of CWs. 5

Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

Fig. 8. Performance of different types of CWs in removing different contaminants from landfill leachate. Bars represent standard deviations.

Fig. 7. TSS content ratio in the influent and effluent of different types of CWs.

et al., 2019). After this saturation limit is reached, the CW will discharge phosphorus concentrations close to the influent values. The release of phosphorus from the sediments may also occur under anaerobic conditions - this phenomenon is well known from natural wetlands sediments (e.g., Bostrom (1982)).

acceptors for oxidation of organic matter under anoxic conditions. However, this process does not remove iron and manganese from the system and it makes heavy metals dissolved. Another biological process involved in removal of heavy metals is plant uptake and subsequent removal of aboveground biomass. However, the amount of heavy metals sequestered in the aboveground biomass is usually low, not exceeding 10% of the inflowing load (Vymazal and Březinová, 2016). The only exception is zinc, which is often found in higher amounts in the aboveground biomass (Peverly et al., 1995; Vymazal et al., 2009). Zinc plays very important role in plant metabolism such as carbohydrate metabolism, maintenance of the integrity of cellular membranes, protein synthesis, production, regulation of essential growth hormone auxin synthesis, pollen formation and tryptophan synthesis which mainly performs its function in stems (Schierup and Larsen, 1981). Fig. 8 shows the removal of eight metals in different CW types. Almost in all cases, VF CWS showed higher removal efficiencies, these amounts were 92.4 ± 6%, 90.0 ± 12%, 93.3 ± 5%, 84.1 ± 19%, 89.3 ± 26%, 93.6 ± 5%, and 77.1 ± 13% for Cd, Cr, Ni, Pb, Zn, Fe, and Mn, respectively. While the removal efficiencies for Cr, Cd, Cu, Ni, Pb, Zn, Fe, and Mn in HF CWs were 56.4 ± 34%, 31.6 ± 21%, 13.9 ± 7%, 38.1 ± 40%, 42.7 ± 30, 56.7 ± 26%, and 49.0 ± 32%, respectively. And also in hybrid CWs, the removal efficiencies for Cr, Cd, Cu, Ni, Pb, Zn, Fe, and Mn in HF CWs were 84.0 ± 39%, 39.1 ± 26%, 62.4 ± 40%, 29.8 ± 22%, 78.3 ± 27%, 56.0 ± 36%, 61.5 ± 28%, and 37.3 ± 22%, respectively. The highest removal of heavy metals in VF CWs can be cause by co-precipitation with iron and manganese under aerobic conditions which prevail in this type of CW.

3.2.4. Total suspended solids Removal of suspended solids in all types of constructed wetlands is usually very high (Vymazal and Kröpfelová, 2008; Kadlec and Wallace, 2009). In subsurface flow CWs, the major removal mechanism is filtration as the water passes through the porous media. The removal usually increases during the time of operation as filtration material gets clogged because the filtration bed gets denser. The removal of suspended solids in HF CWs is the highest in the beginning of the filtration bed and it was shown that most suspended solids are retained within the first several meters of the bed (e.g. Bavor et al. (1987)) and Wojciechowska et al. (2010))). In FWS CWs suspended solids are filtered out by dense vegetation as well as by settlement of settleable particles under quiescent conditions (Merz, 2000). FWS, HF, hybrid, and VF CWs respectively removed 59.5 ± 36%, 69.3 ± 17%, 51.8 ± 29%, and 55.5 ± 20% of TSS (Fig. 7). 3.2.5. Heavy metals Heavy metals (HMs) enter the landfill leachate as a result of the presence of a variety of consumer products such as batteries, plastics, ceramics, and electronics. The main mechanisms in HMs removal are biological pathways, chemical precipitation and co-precipitation, binding to organic matter, sorption onto soil and plant's root surfaces, and filtration of suspended solids by root and soil systems (Kadlec and Wallace, 2009; Bakhshoodeh et al., 2016). Moreover, plants can affect the biogeochemistry of a wetland which can change metal retention (Wojciechowska and Gajewska, 2013). The contribution of each of these processes to the total metal removal heavily depends on the environmental conditions and especially pH and oxidation-reduction potential as well as metal properties itself. Under aerobic conditions, the mobility and/or retention of heavy metals is mostly controlled by iron insoluble compounds as oxides, hydroxides and oxyhydroxides, namely FeOOH, which precipitate under these conditions and co-precipitate heavy metals (Wieder, 1989; Batty et al., 2002). This phenomenon may occur in zones adjacent to plant roots where oxygen leaks to the rhizosphere and precipitated iron crates so called “plaque” (Wu et al., 2019). Under anaerobic conditions, the iron precipitates dissolute and released heavy metals are precipitated as insoluble sulfides (Khalid et al., 1978). Biological processes are involved in iron and manganese reduction when bacteria use oxidized iron and manganese compounds as electron

3.3. Vegetation The presence of macrophytes is one of the most conspicuous features of wetlands and their presence distinguishes constructed wetlands from unplanted soil filters and lagoons. The macrophyte growing in CWs have several properties in relation to the treatment process that make them an essential component of the design (Brix, 1997). However, plants mostly play an indirect role in contaminant removal in CWs – they help to create suitable conditions for removal of pollutants. The only direct effect is the uptake of nutrients and heavy metals (see section on heavy metals). In order to remove nutrients and heavy metals from the landfill leachate, the aboveground biomass must be regularly harvested. In FWS CWS, macrophytes reduce wind velocity, thus reduce risk of resuspension, filter out large suspended solids, uptake nutrients, release oxygen to the water column (mostly applies only for submerged macropyhtes) and provide surface for periphyton attachment (Brix, 1997; 6

Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

Vymazal 2013). In subsurface flow CWs, the roots and rhizomes provide surface for attachment of bacteria which are involved in removal processes, (Ciria et al., 2005; Kouki et al., 2009; Taylor et al., 2011). Roots are also primary sites for nutrient uptake (Vymazal and Kröpfelová, 2008), release oxygen which is not consumed during respiration and release various antimicrobial compounds (Brix, 1993). Although choosing the suitable vegetation for the wetlands mainly depends on the climate, topography, and hydrology of the CW's site, the literature shows that some species have been used more frequently in a different part of the world. Phragmites australis (common reed), Typha latifolia (broadleaf cattail), and Chrysopogon zizanoides (vetiver grass) are the most commonly used species in CWs for leachate treatment (Table S1), while other species genera like Glyceria (mannagrass), Eleocharis (spikerush), and Scirpus (bulrush) are used as well. Using a mix of different species is also common in the literature (Table S1). Phragmites australis is a fast-growing perennial. It has an extensive rhizome system easily penetrating up to 0.6–1.0 m of depth. This plant can be as high as 8 m. It can withstand pH values between 4.8 and 8.2 (Lissner and Schierup, 1997) P. australis is suitable for poorly aerated conditions. This plant has air spaces in the aboveground section of its roots and in the rhizomes, which allow air to be transported down to the roots. This can be extremely helpful in CWs as it promotes aerobic reactions (Vymazal, 2011). Typha latifolia is a fast-growing perennial plant and can grow up to 3 m tall. It has an extensive shallow (approximately 0.3 m) rhizome system and can tolerate a wide pH range of 3.0 to 8.5. T. latifolia has been described as “quite tolerant” for salinity as a seedling and as an established plant (Choudhuri, 1968). Other commonly used Typha species is T. angustifolia (narrow leaf cattail) (Vymazal, 2011; WP, 2013). Chrysopogon zizanoides (Vetiver grass) is a fast-growing tropical plant with a complex, massive, and extensive root system, penetrating to deeper layers of soil. Vetiver grass can survive in very diverse environmental conditions. It can tolerate and survive under extreme temperatures ranging from −15 °C to 55 °C. Vetiver grass tolerates pH values in the range of 3–10. Vetiver grass is resistant to long total submergence in water as well as extreme droughts (Truong and Baker, 1996; Maffei, 2003).

3.5. Temperature Removal processes and especially microbial processes, are strongly affected by ambient temperatures (Kadlec and Wallace, 2009). Temperature affects evapotranspiration, photosynthesis, degradation of organic material, ammonification, nitrification, denitrification microbial community activity and thus affects biological removal pathways (Bulc, 2006). Laboratory studies have shown that the optimum temperature for nutrient removal is 30 °C (Wood et al., 1999), which indeed is hardly achieved under natural outdoor conditions. It has been stated that nitrification rates in wetlands become inhibited at water temperatures of about 10 °C and rates drop rapidly at 6 °C (Werker et al., 2002). Not all contaminants removal mechanisms are affected by temperature. Phosphorous removal, for example, is less sensitive to temperature change as it is dominating removal mechanisms are adsorption and precipitation (Fig. 10). Almost all of the studies on leachate treatment using CWs were performed in the temperature range of 10–35 °C. In this range, no apparent change in the removal of different contaminants with temperature change was observed (Fig. 10). Moreover, Table S-5 shows the relation between parameters removal efficiencies and temperature. 4. Conclusions A comprehensive look at the literature available on the usage of CWs to treat landfill leachate shows that CWs' ability to remove contaminants from leachate is relatively dependent on the CW type, vegetation type, temperature, and hydraulic retention time, and the variability observed is considerable. Given that standard deviations of calculated average performance observed are large, additional factors are also in effect. For leachate with very high contaminant concentrations, some of the previously well-known contaminant removal pathways are likely insufficient. For example, nitrification process in the presence of high BOD5 concentrations becomes limited to the aerobic zones on the roots, or BOD5 removal shifts from aerobic mechanism to the anaerobic mechanism (Kadlec and Wallace, 2009; Saeed and Sun, 2012; Hu et al., 2016). When influent contaminant loads are very high, other parameters become as important as the parameters most commonly investigated. Incoming leachate quality and its ability to host fragile anaerobic microorganisms, type and structure of soil and sand used in CWs and their grading, CW slope, the age of CW, and most importantly construction quality of CW become very important as well. Constructed wetlands for landfill leachate, on average, showed a removal efficiency of 60–80% for BOD5, with FWS and VF showing the highest removal. A closer look at the BOD5in range showed that FWS had been used only for very low inlet concentrations. For studies with higher BOD5in, mostly hybrid CWs have been used. COD removal efficiency, on the other hand, covered a wider range between 20 and 60%. While hybrid and VF CWs showed similar COD removal, it should be noted that HF have been used for leachates with COD values in higher range compared to VF and hybrid CWs. BOD5/COD ratio, while reduced in the effluent compared to influent, was almost predominantly similar in different types of CWs. This can also be due to the lack of reported data in the literature, as many of the studies only report BOD5 or COD as an indicator of organic matter; therefore, the BOD5/COD ratio were not always available. CWs showed 60–75% removal efficiency in removing Ammonia-N, with hybrid CWs being the most successful. It should be noted that hybrid CWs have been used for leachates with the greater NH3 concentration as well. As expected, the greatest extent of nitrification was observed in VF and hybrid CWs, while VF CWs were most successful in removing TN. In terms of TSS and TP, while overall removal ranges of 50–65% and 55–80% for TSS and TP, respectively, are observed for all CWs, HF and hybrid CWs proved to be the most successful in removing TSS and TP. The number of studies reporting heavy metal removal using CWs

3.4. Hydraulic retention time Hydraulic conditions can significantly influence the biogeochemical processes, biotic community composition, and the thus the fate of pollutants in CWs (Kadlec and Wallace, 2009; Ranieri et al. 2013). Hydraulic Retention Time (HRT), the length of time during which the pollutants are in contact with the plant rhizosphere and the substrate, is well known to be a crucial controlling factor in determining the removal efficiency and mechanisms of contaminants (Stottmeister et al., 2003). A long HRT allows for higher contaminant removal while at low HRTs, wastewater moves rapidly to the outlet reducing the contact time among the wastewater, the rhizosphere, and the microorganisms. However, longer HRTs typically require larger land space and higher capital costs. As a result, studying the effects of different HRTs on removal performance of pollutants in CW systems is vitally important. The relationship between HRT and removal efficiency of investigated contaminants in different CWs is shown in Fig. 9 and Table S4. It can be observed that for BOD5 and COD that increasing HRT results in higher removal efficiency for HF and hybrid CWs. TP removal increased in HF CWs as HRT increased. In addition, the results of paired ttest indicated that there were significant differences (p < .005) between different HRT and removal efficiencies. No such apparent trend could be observed for other water quality parameters or other CWs types, meaning that other factors may be contributing to their removal, or the number of studies with HRTs data is low and the possible trend cannot be observed. The results presented in Table S1 indicate that HRT in studies included in this study varied between 0 and 40 days (Fig. 9). 7

Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

Fig. 9. HRT versus removal efficiency of a) BOD5, b) COD, c) Ammonia-N, d) NO3−, e) TKN, f) TN, g) PO4, h) TP, i) TSS in different CWs.


Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

Fig. 10. Temperature versus removal efficiency of investigated contaminants in different CW. Eqs. (r2).


Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

was found to be small, making it harder to draw firm conclusions. VF CWs appear to have been the most successful CW type in removing different types of heavy metals. Heavy metals removals have been reported in the range of 15–95%, with Phragmites sp. plants proving to be the most successful species in removing metals. Treating landfill leachate using CWs demands careful attention and design specifications must to be investigated case by case. To enhance the performance of biodegradable organic compound removal from landfill leachate, pre-CW strategies such as using aeration and sedimentation are needed, but their quantifiable removal performance enhancement and economic assessment should be investigated further. Constructed wetland can also be modified by biochar/zeolite/adsorbent addition (Hayder et al., 2019), and combined with other methods, such as double-chamber anaerobic reactor and microbial fuel cell-coupled constructed wetland (Fang et al., 2017; Abedi and Mojiri, 2019; Montero et al., 2019), in order to increase the removal efficiency of pollutants.

in a pilot-scale constructed wetland with horizontal subsurface flow using shale as a substrate. Water Res. 34, 2483–2490. Fang, Z., Cao, X., Li, X., Wang, H., Li, X., 2017. Electrode and azo dye decolorization performance in microbial-fuel-cell-coupled constructed wetlands with different electrode size during long-term wastewater treatment. Bioresour. Technol. 238, 450–460. Hayder, A., Vanderburgt, S., Santos, R.M., Chiang, Y.W., 2019. Phosphorous runoff risk assessment and its potential management using wollastonite according to geochemical modeling. Open Agric. 4, 787–794 accepted. He, P.-J., Qu, X., Shao, L.-M., Li, G.-J., Lee, D.-J., 2007. Leachate pretreatment for enhancing organic matter conversion in landfill bioreactor. J. Hazard. Mater. 142, 288–296. Hu, Y., He, F., Ma, L., Zhang, Y., Wu, Z., 2016. Microbial nitrogen removal pathways in integrated vertical-flow constructed wetland systems. Bioresour. Technol. 207, 339–345. Idris, A., Inanc, B., Hassan, M.N., 2004. Overview of waste disposal and landfills/dumps in Asian countries. J. mater. Cycles Waste Manag. 6, 104–110. Kadlec, R., Knight, R., 1996. Treatment Wetlands. CRC, Baca Raton, FL. Kadlec, R., Wallace, S., 2009. Treatment Wetlands, 2nd edition. CRCPress, Boca Raton. Khalid, R., Patrick Jr., W., Gambrell, R., 1978. Effect of dissolved oxygen on chemical transformations of heavy metals, phosphorus, and nitrogen in an estuarine sediment. Estuar. Coast. Mar. Sci. 6, 21–35. Kouki, S., M’hiri, F., Saidi, N., Belaïd, S., Hassen, A., 2009. Performances of a constructed wetland treating domestic wastewaters during a macrophytes life cycle. Desalination 246, 452–467. Kynkäänniemi, P., Ulén, B., Torstensson, G., Tonderski, K.S., 2013. Phosphorus retention in a newly constructed wetland receiving agricultural tile drainage water. J. Environ. Qual. 42, 596–605. Li, H., Li, Y., Gong, Z., Li, X., 2013. Performance study of vertical flow constructed wetlands for phosphorus removal with water quenched slag as a substrate. Ecol. Eng. 53, 39–45. Li, Y.C., Zhang, D.Q., Wang, M., 2017. Performance evaluation of a full-scale constructed wetland for treating stormwater runoff. CLEAN–Soil, Air, Water 45, 1600740. Lissner, J., Schierup, H.-H., 1997. Effects of salinity on the growth of Phragmites australis. Aquat. Bot. 55, 247–260. Maffei, M., 2003. Vetiveria: The Genus Vetiveria. CRC Press. Marañón, E., Castrillón, L., Fernández-Nava, Y., Fernández-Méndez, A., FernándezSánchez, A., 2008. Coagulation–flocculation as a pretreatment process at a landfill leachate nitrification–denitrification plant. J. Hazard. Mater. 156, 538–544. Martín, M., Gargallo, S., Hernández-Crespo, C., Oliver, N., 2013. Phosphorus and nitrogen removal from tertiary treated urban wastewaters by a vertical flow constructed wetland. Ecol. Eng. A 61, 34–42. Matamoros, V., Rodríguez, Y., Bayona, J.M., 2017. Mitigation of emerging contaminants by full-scale horizontal flow constructed wetlands fed with secondary treated wastewater. Ecol. Eng. 99, 222–227. Merz, S.K., 2000. Guidelines for Using Free Water Surface Constructed Wetlands to Treat Municipal Sewage. Department of Natural Resources. Midhun, G., Divya, L., George, J., Jayakumar, P., Suriyanarayanan, S., 2016. Wastewater treatment studies on free water surface constructed wetland system. In: Prashanthi, R.S.M. (Ed.), Integrated Waste Management in India. Springer International Publishing, pp. 97–109. Montero, A.A.G., Serrano, E.V.P., Montiel, J.P., 2019. Sanitary landfill leachate treatment with double chamber anaerobic reactor in series with constructed wetland. Environ. Process. 1–18. Mulamoottil, G., MacBean, E.A., Rovers, F.A., 1999. Constructed Wetlands for the Treatment of Landfill Leachates. CRC Press. O’luanaigh, N., Goodhue, R., Gill, L., 2010. Nutrient removal from on-site domestic wastewater in horizontal subsurface flow reed beds in Ireland. Ecol. Eng. 36, 1266–1276. Peverly, J.H., Surface, J.M., Wang, T., 1995. Growth and trace metal absorption by Phragmites australis in wetlands constructed for landfill leachate treatment. Ecol. Eng. 5, 21–35. Ranieri, E., Gorgoglione, A., Solimeno, A., 2013. A comparison between model and experimental hydraulic performances in a pilot-scale horizontal subsurface flow constructed wetland. Ecol. Eng. 60, 45–49. Reddy, K., Kadlec, R., Flaig, E., Gale, P., 1999. Phosphorus retention in streams and wetlands: a review. Crit. Rev. Environ. Sci. Technol. 29, 83–146. Saeed, T., Sun, G., 2012. A review on nitrogen and organics removal mechanisms in subsurface flow constructed wetlands: dependency on environmental parameters, operating conditions and supporting media. J. Environ. Manag. 112, 429–448. Saeed, T., Sun, G., 2017. A comprehensive review on nutrients and organics removal from different wastewaters employing subsurface flow constructed wetlands. Crit. Rev. Environ. Sci. Technol. 47, 203–288. Saeed, T., Afrin, R., Al Muyeed, A., Sun, G., 2012. Treatment of tannery wastewater in a pilot-scale hybrid constructed wetland system in Bangladesh. Chemosphere 88, 1065–1073. Schierup, H.-H., Larsen, V.J., 1981. Macrophyte cycling of zinc, copper, lead and cadmium in the littoral zone of a polluted and a non-polluted lake. I. Availability, uptake and translocation of heavy metals in Phragmites australis (Cav.) Trin. Aquat. Bot. 11, 197–210. Scholz, M., Hedmark, Å., 2009. Constructed wetlands treating runoff contaminated with nutrients. Water Air Soil Pollut. 205, 323. Seidel, K., 1966. Reinigung von Gewässern durch höhere Pflanzen [Wastewater treatment by means of Macrophytes]. Naturwissenschaften 53, 289–297. Sheoran, A., Sheoran, V., 2006. Heavy metal removal mechanism of acid mine drainage in wetlands: a critical review. Miner. Eng. 19, 105–116.

Declaration of Competing Interest The authors declare that they have no conflict of interest. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// References Abedi, T., Mojiri, A., 2019. Constructed wetland modified by biochar/zeolite addition for enhanced wastewater treatment. Environ. Technol. Innov. 16, 100472. Alavi, N., Azadi, R., Jaafarzadeh, N., Babaei, A., 2011. Kinetics of nitrogen removal in an Anammox up-flow anaerobic bioreactor for treating petrochemical industries wastewater (ammonia plant). Asian J. Chem. 23, 5220–5224. Bakhshoodeh, R., Alavi, N., Mohammadi, A.S., Ghanavati, H., 2016. Removing heavy metals from Isfahan composting leachate by horizontal subsurface flow constructed wetland. Environ. Sci. Pollut. Res. 23, 12384–12391. Bakhshoodeh, R., Alavi, N., Majlesi, M., Paydary, P., 2017a. Compost leachate treatment by a pilot-scale subsurface horizontal flow constructed wetland. Ecol. Eng. 105, 7–14. Bakhshoodeh, R., Alavi, N., Paydary, P., 2017b. Composting plant leachate treatment by a pilot-scale, three-stage, horizontal flow constructed wetland in Central Iran. Environ. Sci. Pollut. Res. 24, 23803–23814. Bakhshoodeh, R., Soltani-Mohammadi, A., Alavi, N., Ghanavati, H., 2017c. Treatment of high polluted leachate by subsurface flow constructed wetland with vetiver. Amirkabir J. Civ. Eng. 49, 139–148. Batty, L.C., Baker, A.J., Wheeler, B.D., 2002. Aluminium and phosphate uptake by Phragmites australis: the role of Fe, Mn and Al root plaques. Ann. Bot. 89, 443–449. Bavor, H., Roser, D., McKersie, S., 1987. Nutrient Removal Using Shallow Lagoon-Solid Matrix Macrophyte Systems. Bostrom, B., 1982. Phosphorus release from lake sediment. Arch. Hydrobiol. Beih. Ergebn. Limn. 18, 5–59. Brix, H., 1993. Wastewater treatment in constructed wetlands: system design, removal processes, and treatment performance. In: Constructed Wetlands for Water Quality Improvement, pp. 9–22. Brix, H., 1997. Do macrophytes play a role in constructed treatment wetlands? Water Sci. Technol. 35, 11–17. Brix, H., Arias, C., Johansen, N., 2003. In: Vymazal, J. (Ed.), Experiments in a two-stage constructed wetland system: nitrification capacity and effects of recycling on nitrogen removal. Backhyus Publishers, Leiden, The Netherlands, pp. 237–258 WetlandsNutrients, Metals and Mass Cycling. Bulc, T.G., 2006. Long term performance of a constructed wetland for landfill leachate treatment. Ecol. Eng. 26, 365–374. Castillo, E., Vergara, M., Moreno, Y., 2007. Landfill leachate treatment using a rotating biological contactor and an upward-flow anaerobic sludge bed reactor. Waste Manag. 27, 720–726. Choi, J., Geronimo, F.K.F., Maniquiz-Redillas, M.C., Kang, M.-J., Kim, L.-H., 2015. Evaluation of a hybrid constructed wetland system for treating urban stormwater runoff. Desalin. Water Treat. 53, 3104–3110. Choudhuri, G., 1968. Effect of soil salinity on germination and survival of some steppe plants in Washington. Ecology 49, 465–471. Ciria, M.P., Solano, M.L., Soriano, P., 2005. Role of macrophyte typha latifolia in a constructed wetland for wastewater treatment and assessment of its potential as a biomass fuel. Biosyst. Eng. 92, 535–544. Dong, Z., Sun, T., 2007. A potential new process for improving nitrogen removal in constructed wetlands—promoting coexistence of partial-nitrification and ANAMMOX. Ecol. Eng. 31, 69–78. Drizo, A., Frost, C.A., Grace, J., Smith, K.A., 2000. Phosphate and ammonium distribution


Ecological Engineering 146 (2020) 105725

R. Bakhshoodeh, et al.

of Phragmites australis in horizontal flow constructed wetlands for wastewater treatment: a review. Chem. Eng. J. 290, 232–242. Vymazal, J., Kröpfelová, L., 2008. Wastewater Treatment in Constructed Wetlands with Horizontal Sub-Surface Flow. Springer. Vymazal, J., Kröpfelová, L., 2009. Removal of organics in constructed wetlands with horizontal sub-surface flow: a review of the field experience. Sci. Total Environ. 407, 3911–3922. Vymazal, J., Kröpfelová, L., Švehla, J., Chrastný, V., Štíchová, J., 2009. Trace elements in Phragmites australis growing in constructed wetlands for treatment of municipal wastewater. Ecol. Eng. 35, 303–309. Werker, A., Dougherty, J., McHenry, J., Van Loon, W., 2002. Treatment variability for wetland wastewater treatment design in cold climates. Ecol. Eng. 19, 1–11. Wieder, R.K., 1989. A survey of constructed wetlands for acid coal mine drainage treatment in the eastern United States. Wetlands 9, 299–315. Wojciechowska, E., Gajewska, M., 2013. Partitioning of heavy metals in sub-surface flow treatment wetlands receiving high-strength wastewater. Water Sci. Technol. 68, 486–493. Wojciechowska, E., Gajewska, M., Obarska-Pempkowiak, H., 2010. Treatment of landfill leachate by constructed wetlands: three case studies. Pol. J. Environ. Stud. 19, 2010. Wood, S., Wheeler, E., Berghage, R., Graves, R., 1999. Temperature effects on wastewater nitrate removal in laboratory-scale constructed wetlands. Trans. ASAE 42, 185. WP, S.F., 2013. Enhancement of Natural Water Systems and Treatment Methods for Safe and Sustainable Water Supply in India. Wu, S., Zhang, D., Austin, D., Dong, R., Pang, C., 2011. Evaluation of a lab-scale tidal flow constructed wetland performance: oxygen transfer capacity, organic matter and ammonium removal. Ecol. Eng. 37, 1789–1795. Wu, S., Vymazal, J., Brix, H., 2019. Critical review: biogeochemical networking of iron in constructed wetlands for wastewater treatment. Environ. Sci. Technol. 53, 7930–7944. Yin, T., Chen, H., Reinhard, M., Yi, X., He, Y., Gin, K.Y.-H., 2017. Perfluoroalkyl and polyfluoroalkyl substances removal in a full-scale tropical constructed wetland system treating landfill leachate. Water Res. 125, 418–426.

Sheridan, C.M., Glasser, D., Hildebrandt, D., 2014. Estimating rate constants of contaminant removal in constructed wetlands treating winery effluent: a comparison of three different methods. Process. Saf. Environ. Prot. 92, 903–916. Sheridan, C., Akcil, A., Kappelmeyer, U., Moodley, I., 2018. A review on the use of constructed wetlands for the treatment of acid mine drainage. Constructed Wetlands for Industrial Wastewater Treatment 249–262. Stottmeister, U., Wießner, A., Kuschk, P., Kappelmeyer, U., Kästner, M., Bederski, O., Müller, R.A., Moormann, H., 2003. Effects of plants and microorganisms in constructed wetlands for wastewater treatment. Biotechnol. Adv. 22, 93–117. Surmacz-Gorska, J., 2001. Degradation of Organic Compounds in Municipal Landfill Leachate. Publishers of Environmental Engineering Committee of Polish Academy of Sciences, Lublin ISBN, pp. 83–915874. Tanner, C.C., Nguyen, M.L., Sukias, J.P.S., 2005. Nutrient removal by a constructed wetland treating subsurface drainage from grazed dairy pasture. Agric. Ecosyst. Environ. 105, 145–162. Taylor, C.R., Hook, P.B., Stein, O.R., Zabinski, C.A., 2011. Seasonal effects of 19 plant species on COD removal in subsurface treatment wetland microcosms. Ecol. Eng. 37, 703–710. Truong, P., Baker, D., 1996. Vetiver grass for the stabilization and rehabilitation of acid sulfate soils. In: Proc. Second National Conf. Acid Sulfate Soils, Coffs Harbour, Australia, pp. 196–198. Vymazal, J., 1999. Nitrogen removal in constructed wetlands with horizontal sub-surface flow-can we determine the key process. In: Nutrient Cycling and Retention in Natural and Constructed Wetlands. Backhuys Publishers, Leiden, The Netherlands 17. Vymazal, J., 2005. Horizontal sub-surface flow and hybrid constructed wetlands systems for wastewater treatment. Ecol. Eng. 25, 478–490. Vymazal, J., 2011. Plants used in constructed wetlands with horizontal subsurface flow: a review. Hydrobiologia 674, 133–156. Vymazal, J., 2013. Emergent plants used in free water surface constructed wetlands: a review. Ecol. Eng. 61, 582–592. Vymazal, J., 2014. Constructed wetlands for treatment of industrial wastewaters: a review. Ecol. Eng. 73, 724–751. Vymazal, J., Březinová, T., 2016. Accumulation of heavy metals in aboveground biomass