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A constructed wetland is an artificial marsh or swamp, created for anthropogenic discharge such as wastewater, stormwater runoff or sewage treatment, and as habitat for wildlife, or for land reclamation after mining or other disturbance. Natural wetlands act as biofilters, removing sediments and pollutants such as heavy metals from the water, and constructed wetlands can be designed to emulate these features.
Vegetation in a wetland provides a substrate (roots, stems, and leaves) upon which microorganisms can grow as they break down organic materials. This community of microorganisms is known as the periphyton. The periphyton and natural chemical processes are responsible for approximately 90 percent of pollutant removal and waste breakdown. The plants remove about seven to ten percent of pollutants, and act as a carbon source for the microbes when they decay. Different species of aquatic plants have different rates of heavy metal uptake, a consideration for plant selection in a constructed wetland used for water treatment.
Constructed wetlands are of two basic types: subsurface-flow and surface-flow wetlands. Subsurface-flow wetlands can be further classified as horizontal flow and vertical flow constructed wetlands. Subsurface-flow wetlands move effluent (agricultural or mining runoff, tannery or meat processing wastes, wastewater from sewage or storm drains, or other water to be cleansed) through a gravel or sand medium on which plants are rooted; surface-flow wetlands move effluent above the soil in a planted marsh or swamp, and thus can be supported by a wider variety of soil types including bay mud and other silty clays. In subsurface-flow systems, the effluent may move either horizontally, parallel to the surface, or vertically, from the planted layer down through the substrate and out. Subsurface horizontal-flow wetlands are less hospitable to mosquitoes, whose populations can be a problem in constructed wetlands (carnivorous plants have been used to address this problem). Subsurface-flow systems have the advantage of requiring less land area for water treatment, but are not generally as suitable for wildlife habitat as are surface-flow constructed wetlands.
Plantings of reedbeds are popular in European constructed wetlands, and plants such as cattails or bulrushes (Typha spp.), sedges, water hyacinth and Pontederia spp. are used worldwide. Recent research in use of constructed wetlands for subarctic regions has shown that buckbeans (Menyanthes trifoliata) and pendant grass (Arctophila fulva) are also useful for metals uptake.
Additional recommended knowledge
General Contaminant Removal
Physical, chemical, and biological processes combine in wetlands to remove contaminants from wastewater. An understanding of these processes is fundamental not only to designing wetland systems but to understanding the fate of chemicals once they have entered the wetland. Theoretically, treatment of wastewater within a constructed wetland occurs as it passes through the wetland medium and the plant rhizosphere. A thin aerobic film around each root hair is aerobic due to the leakage of oxygen from the rhizomes, roots, and rootlets . Decomposition of organic matter is facilitated by aerobic and anaerobic micro-organisms present. Microbial nitrification and subsequent denitrification releases nitrogen as gas to the atmosphere. Phosphorus is coprecipitated with iron, aluminum, and calcium compounds located in the root-bed medium . Suspended solids are filtered out as they settle in the water column in surface flow wetlands or are physically filtered out by the medium within subsurface flow wetland cells. Harmful bacteria and viruses are reduced by filtration and adsorption by biological films on the rock media in subsurface flow and vertical flow systems.
Removal of Nitrogen
The dominant forms of nitrogen in wetlands that are of importance to wastewater treatment include organic nitrogen, ammonia, ammonium, nitrate, nitrite, and nitrogen gases. Inorganic forms are essential to plant growth in aquatic systems but if scarce can limit or control plant productivity. The nitrogen entering wetland systems can be measured as organic nitrogen, ammonia, nitrate and nitrite. Total Nitrogen refers to all nitrogen species. The removal of nitrogen from wastewater is important because of ammonia’s toxicity to fish if discharged into water courses. Excessive levels of nitrates in drinking water is thought to cause methemoglobinemia in infants, which decreases the oxygen transport ability of the blood. The UK has experienced a significant increase in nitrate concentration in groundwater and rivers.
Mitsch & Gosselink (1986) define nitrogen mineralisation as "the biological transformation of organically combined nitrogen to ammonium nitrogen during organic matter degradation". This can be both an aerobic and anaerobic process and is often referred to as ammonification. Mineralisation of organically combined nitrogen releases inorganic nitrogen as nitrates, nitrites, ammonia and ammonium, making it available for plants, fungi and bacteria. Mineralisation rates may be affected by oxygen levels in a wetland.
Ammonia (NH3) and Ammonium (NH4+)
The formation of ammonia (NH3) occurs via the mineralisation or ammonification of organic matter under either anaerobic or aerobic conditions (Keeney, 1973). The ammonium ion (NH4+) is the primary form of mineralized nitrogen in most flooded wetland soils. The formation of this ion occurs when ammonia combines with water as follows:
NH3 + H2O <-----> NH4+ + OH-
Upon formation, several pathways are available to the ammonium ion. It can be absorbed by the plants and algae and converted back into organic matter, or the ammonium ion can be immobilized onto negatively charged soil particles (Mitsch & Gosselink, 1986). At this point, the ammonium ion can be prevented from further oxidation because of the anaerobic nature of wetland soils. Under these conditions the ammonium ion is stable and it is in this form that nitrogen predominates in anaerobic sediments typical of wetlands (Brock & Madigan, 1991; Patrick & Reddy, 1976). Most wetland soils have a thin aerobic layer at the surface. As an ammonium ion from the anaerobic sediments diffuses upward into this layer it is converted to nitrite or nitrified (Klopatek, 1978). An increase in the thickness of this aerobic layer results in an increase in nitrification (Patrick & Reddy, 1976). This diffusion of the ammonium ion sets up a concentration gradient across the aerobic-anaerobic soil layers resulting in further nitrification reactions (Klopatek, 1978; Patrick & Reddy, 1976).
Wetzel (1983) defines nitrification as the "biological conversion of organic and inorganic nitrogenous compounds from a reduced state to a more oxidized state". Nitrification is strictly an aerobic process in which the end product is nitrate (NO3-); this process is limited when anaerobic conditions prevail (Patrick & Reddy, 1976). Nitrification will occur readily down to 0.3 ppm dissolved oxygen (Keeney, 1973). The process of nitrification (1) oxidizes ammonium (from the sediment) to nitrite (NO2-), and then (2) nitrite is oxidized to nitrate (NO3-). The overall nitrification reactions are as follows:
(1) 2 NH4+ + 3 O2 <----> 4 H+ + 2 H2O + 2 NO2-
Two different bacteria are required to complete this oxidation of ammonium to nitrate. Nitrosomonas sp. oxidizes ammonium to nitrite via reaction (1) , and Nitrobacter sp. oxidizes nitrite to nitrate via reaction (2) (Keeney, 1973).
According to Wetzel (1983) " Denitrification by bacteria is the biochemical reduction of oxidized nitrogen anions, nitrate-N and nitrite-N, with concomitant oxidation of organic matter." The general sequence as given by Wetzel (1983) is as follows:
NO3- ---> NO2- ---> N2O ---> N2
The end products, N2O and N2 are gases that re-enter the atmosphere. Denitrification occurs intensely in anaerobic environments but will also occur in aerobic conditions (Bandurski, 1965). A deficiency of oxygen causes certain bacteria to use nitrate in place of oxygen as an electron acceptor for the reduction of organic matter (Patrick & Reddy, 1976). The process of denitrification is restricted to a narrow zone in the sediment immediately below the aerobic-anaerobic soil interface (Mitsch & Gosselink, 1986; Nielson et al., 1990). Denitrification is considered by Richardson et al. (1978) to be the predominant microbial process that modifies the chemical composition of nitrogen in a wetland system and the major process whereby elemental nitrogen is returned to the atmosphere (Patrick & Reddy, 1976). To summarize, the nitrogen cycle is completed as follows: ammonia in water, at or near neutral pH is converted to ammonium ions; the aerobic bacterium Nitrosomonas sp. oxidizes ammonium to nitrite; Nitrobacter sp. then converts nitrite to nitrate. Under anaerobic conditions, nitrate is reduced to relatively harmless nitrogen gas, that is given off to the atmosphere.
Nitrogen removal in constructed wetlands used to treat domestic sewage
In a review of 19 surface flow wetlands (US EPA, 1988) it was found that nearly all reduced total nitrogen. In a review of both surface flow and subsurface flow wetlands Reed (1995) concluded that effluent nitrate concentration is dependent on maintaining anoxic conditions within the wetland so that denitrification can occur. He found that subsurface flow wetlands were superior to surface flow wetlands for nitrate removal. The 20 surface flow wetlands reviewed reported effluent nitrate levels below 5 mg/L; the 12 subsurface flow wetlands reviewed reported effluent nitrate ranging from <1 to < 10 mg/L. Results obtained from the Niagara-On-The-Lake vertical flow systems show a significant reduction in both total nitrogen and ammonia (> 97%) when primary treated effluent was applied at a rate of 60L/m²/day. Calculations made showed that over 50% of the total nitrogen going into the system was converted to relatively harmless nitrogen gas. Effective removal of nitrate from the sewage lagoon influent was dependent on medium type used within the vertical cell as well as water table level within the cell (Lemon et al.,1997).
Removal of Phosphorus
Phosphorus occurs naturally in both organic and inorganic forms. The analytical measure of biologically available orthophosphates is referred to as soluble reactive phosphorus (SR-P). Dissolved organic phosphorus and insoluble forms of organic and inorganic phosphorus are generally not biologically available until transformed into soluble inorganic forms (Mitsch and Gosselink, 1986).
In freshwater aquatic ecosystems phosphorus has been described as the major limiting nutrient. Under undisturbed natural conditions, phosphorus is in short supply. The natural scarcity of phosphorus is demonstrated by the explosive growth of algae in water receiving heavy discharges of phosphorus-rich wastes. Because phosphorus does not have an atmospheric component as does nitrogen, the phosphorus cycle can be characterized as closed. The removal and storage of phosphorus from wastewater can only occur within the constructed wetland itself. According to Mitsch and Gosselink (1986) phosphorus may be sequestered within a wetland system by the following:
1) The binding of phosphorus in organic matter as a result of incorporation into living biomass, and
2) precipitation of insoluble phosphates with ferric iron, calcium, and aluminum found in wetland soils.
Incorporation into biomass
Higher plants in wetland systems may be viewed as transient nutrient storage compartments absorbing nutrients during the growing season and releasing large amounts at senescence (Bernard and Solsky, 1976; Guntensbergen, 1989). Generally, plants from nutrient-rich habitats accumulate more nutrients than plants found in nutrient-poor habitats, a phenomenon referred to as luxury uptake of nutrients (Guntensbergen, 1989; Kadlec, 1989). Aquatic vegetation may play an important role in phosphorus removal and, if harvested, extend the life of a system by postponing phosphorus saturation of the sediments (Breen, 1990; Guntensbergen, 1989; Rogers et al., 1991). According to Sloey et al. (1978) vascular plants may account for only a small amount of phosphorus uptake with only 5 to 20% of the nutrients detained in a natural wetland being stored in harvestable plant material. Bernard and Solsky (1976) also reported relatively low phosphorus retention, estimating that a sedge (Carex sp.) wetland retained 1.9 g of phosphorus per square metre of wetland . Bulrushes (Scirpus sp.) in a constructed wetland system receiving secondarily treated domestic wastes contained 40.5% of the total phosphorus influent. The remaining 59.0% was found to be stored in the gravel substratum (Sloey et al., 1978). Phosphorus removal in a surface flow wetland treatment system planted with one of Scirpus sp., Phragmites sp. or Typha sp. was investigated by Finlayson and Chick (1983). Phosphorus removal of 60%, 28%, and 46% were found for Scirpus sp., Phragmites sp. and Typha sp. respectively. More recent work by Breen (1990) may prove this to be a low estimate. His work on an artificial wetland indicated that vascular plants are a major phosphorus storage compartment accounting for 67.3% of the influent phosphorus. Thut (1989) attributed plant adsorption with 80% phosphorus removal. Only a small proportion (<20%) of phosphate removal by constructed wetlands can be attributed to nutritional uptake by bacteria, fungi and algae (Moss, 1988). Swindell et al., (1990) found that the lack of seasonal fluctuation in phosphorus removal rates suggests that the primary mechanism is bacterial and alga fixation. However, Richardson (1985) dismisses this mechanism as temporary saying that although the initial removal of dissolved inorganic phosphorus from the water under natural loading levels is due largely to microbial uptake and adsorption, the microbial pool is small and quickly becomes saturated at which point the soil medium takes over as the major contributor to phosphate removal. There are more indirect ways in which plants contribute to wastewater purification. Plants create a unique environment at the attachment surface of the biofilm. Certain plants transport oxygen which is released at the biofilm/root interface perhaps adding oxygen to the wetland system (Pride et al., 1990). Plants also increase soil or other root-bed medium hydraulic conductivity. As roots and rhizomes grow they are thought to disturb and loosen the medium increasing its porosity which may allow more effective fluid movement in the rhizosphere. When roots decay they leave behind ports and channels known as macropores which are effective in channeling water through the soil (Conley et al., 1991). Whether or not wetland systems act as a phosphorus sink or source seems to depend on system characteristics such as sediment and hydrology. Kramer et al., (1972) indicated that there seems to be a net movement of phosphorus into the sediment in many lakes. In Lake Erie as much as 80% of the total phosphorus is removed from the waters by natural processes and is presumably stored in the sediment. According to Klopatek (1978) marsh sediments high in organic matter act as sinks. He has also shown that phosphorus release from a marsh exhibits a cyclical pattern. Much of the spring phosphorus release comes from high phosphorus concentrations locked up in the winter ice covering the marsh; in summer the marsh acts as a phosphorus sponge. Simpson (1978) found that phosphorus was exported from the system following dieback of vascular plants. It has been demonstrated by Klopatek (1978) that phosphorus concentrations in water are reduced during the growing season due to plant uptake but decomposition and subsequent mineralisation of organic matter releases phosphorus over the winter and accounts for the higher winter phosphorus concentrations in the marsh (Klopatek, 1978; Mitsch, 1986).
Phosphorus retention by soils or root-bed media
Two types of phosphate retention mechanisms may occur in soils or root-bed media: chemical adsorption onto the medium (Hsu, 1964) and physical precipitation of the phosphate ion (Faulkner and Richardson, 1989). Both result from the attraction between phosphate ion and ions of Al, Fe or Ca (Hsu, 1964; Cole et al., 1953) and terminates with formation of various iron phosphates (Fe-P), aluminum phosphates (Al-P) or calcium phosphates (Ca-P) (Fried and Dean, 1955). Redox potential (Eh) of soil or water is a measure of its ability to reduce or oxidize chemical substances and may range between -300 and +300 millivolts (mV) (Hammer, 1992). Though the oxidation state of phosphorus is unaffected by redox reactions, the redox potential is important because of Fe reduction. Severely reduced conditions in the sediments may result in phosphorus release (Mann, 1990). Typical wetland soils may have an Eh of -200 mV (Hammer, 1992). Under these reduced conditions Fe3+ (Ferric iron) may be reduced to Fe2+ (Ferrous iron) and may release the bound phosphate ion back into solution (Faulkner and Richardson, 1989; Sah and Mikkelson, 1986). The introduction of oxygen causes the Fe2+ to be oxidized to Fe3+ producing a simultaneous reduction of phosphate (Wetzel,1983). The solubility of phosphorus may be affected by the amount of oxygen present in the sediment because saturation by water and subsequent loss of oxygen generally cause wetland soils to have negative redox potentials (Hammer, 1992). A well documented occurrence in the hypolimnion of lakes is the release of soluble phosphorus when conditions become anaerobic (Burns & Ross, 1972; Williams & Mayer, 1972). This phenomenon also occurs in natural wetlands (Gosselink & Turner, 1978) and Kramer et al., (1972) report that oxygen concentrations of less than 2.0 mg/l result in the release of phosphorus from sediments.
Phosphorus removal in constructed wetlands used to treat domestic sewage
Adsorption to binding sites within the sediments was identified as the major phosphorus removal mechanism in the surface flow constructed wetland system at Port Perry, Ontario (Snell, unpublished data). Release of phosphorus from the sediments occurred when anaerobic conditions prevailed. The lowest wetland effluent phosphorus levels occurred when oxygen levels of the overlying water column were above 1.0 mg / L. Removal efficiencies for total phosphorus were 54-59% with mean effluent levels of 0.38 mg P/L. Wetland effluent phosphorus concentration was higher than influent levels during the winter months. Lantzke et al., (1999) investigated phosphorus removal in a VF wetland in Australia and found that the quantity of phosphorus removed over a short term was stored in the following wetland components in order of decreasing importance: substratum> macrophyte >biofilm but over the long term phosphorus storage was located in macrophyte> substratum>biofilm components. They also found that medium iron-oxide adsorption provides additional removal for some years. Mann (1990) investigated the phosphorus removal efficiency of two large-scale, surface flow wetland systems in Australia which had a gravel substratum. He then compared these results to laboratory phosphorus adsorption experiments. For the first two months of wetland operation the mean phosphorus removal efficiency of system 1 and 2 was 38% and 22%, respectively. Over the first year a decline in removal efficiencies occurred. During the second year of operation release of phosphorus from the system was often recorded such that more phosphorus came out than was put in. This release was attributed to the saturation of phosphorus binding sites. Close agreement between the phosphorus adsorption capacity of the gravel as determined in the laboratory and the adsorption capacity recorded in the field was found. The phosphorus adsorption capacity of a subsurface flow constructed wetland system containing a predominantly quartz gravel was investigated by Breen (1990). The adsorption characteristics of this gravel as determined by laboratory adsorption experiments and using the Langmuir adsorption isotherm was 25 mg P / g gravel. Close agreement between calculated and realized phosphorus adsorption was found. Because of the poor adsorption capacity of the quartz gravel, plant uptake and subsequent harvesting were identified as the major phosphorus removal mechanism.
by: Lloyd R. Rozema, M.Sc. (excerpt form Master of Science thesis, Brock University, St. Catharines, ON, 2000)
Many reedbed systems aerate the water after the final reedbed using cascades such as Flowforms before holding the water in a shallow pond.
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