CONSTRUCTED WETLAND NITROGEN REMOVAL FROM CATTLE FEEDLOT WASTEWATER

M.E. Ancell, C.B. Fedler, and N.C. Parker

ABSTRACT

A 23-tank, 43 m2, pilot-scale constructed wetland system was loaded daily with 136.2 liters of cattle feedlot wastewater to measure the nitrogen removal effectiveness. The 23-tanks were separated into six different treatment series, and the effects of four different total nitrogen (TN) loading rates were investigated with three different series surface areas and detention times. The four TN loading rates were 11.4, 8.0, 2.3, and 0.5 g TN/day. All four loading rates were tested in treatment series consisting of four tanks. Additionally, the 2.3 g TN/day loading rate was tested in a series with two tanks and a series with five tanks.

The removal of nitrogen constituents from wastewater is dominated by maximizing the permanent removal processes inherent to the nitrogen cycle. Although the nitrogen cycle is a complex interaction of biological and chemical phenomena, maximizing its inherent removal processes is attainable in the wetland environment. The primary facilitator of this nitrogen removal is the root-zone aeration of the predominantly anaerobic environment surrounding the wetland soil. Given proper amounts of dissolved oxygen, the microbiota of nitrification can oxidize ammonia to nitrate, and denitrification can take place in the anaerobic environment, ultimately removing nitrogen from the wastewater in the form of nitrogen gas. An additional permanent nitrogen removal pathway in wetlands is defined by the plant uptake of ammonia and/or nitrate. However, maximizing this removal pathway requires plant harvesting, which can be costly in the full scale wetland treatment setting and does not always yield an appreciable amount of nitrogen removal.

Of the series with four tanks, the series loaded with 11.4 g TN/day removed an average of 54.9 percent of the applied TN; the 8.0 g TN/day series removed 60.8 percent of the applied TN; the series loaded with 2.3 g TN/day removed an average of 78.0 percent of the a TN applied; and the series loaded with 0.5 g TN/day removed an average of 33.4 percent of the applied TN. Additionally, the 2.3 g TN/day series with two tanks removed 44.0 percent of the applied TN, and the 2.3 g TN/day series with five tanks removed 82.8 percent of the applied TN.

The wetland plants removed 14.9 percent of the applied TN in the 0.5 g TN/day series. This nitrogen removal in the biomass was the largest of all series tested. The smallest plant nitrogen removal percentage was observed in the 11.4 g TN/day series (4.87 percent). Additionally, a model of dry-weight biomass production with increasing TN loading is presented. The biomass yield as a function of TN loading rate was highly significant ( = 0.010) with an R2 = 0.978.

INTRODUCTION

The passage of the 1987 Water Quality Act Amendments to the CWA established a new direction in the dispersment of funds for the treatment of wastewater. These amendments returned the focus of water pollution control back to the states. This translates into less federal grants for wastewater treatment improvements, and thus, a greater need for state participation and funding for water quality control. This problem is magnified by the current need for water quality improvement in the United States. Smith (1988) outlines many of the water quality problems facing America:

  • Treatment works construction needs for small communities (under 10,000 population) - with the least ability to pay - amount to between $10 and $15 billion nationwide.
  • Agricultural activities generate large amounts of nonpoint source water pollution and are causing serious water quality problems in at least 24 states.
  • Approximately one-third of existing lakes and reservoirs have water quality impairment from nonpoint source pollution, principally from agricultural activities.
  • Large treatment needs exist in the industrial sector. For example, acid mine drainage affects over 11,800 miles of streams in Appalachia alone.
  • Leachate from landfill disposal of both industrial and municipal solid waste is affecting water quality.
  • Industrial toxic chemicals and hazardous waste treatment and disposal have become major water quality concerns, with high potential costs and uncertain future treatment requirements and alternatives.

As federal funds for the treatment of wastewater decline, and more pressure is placed on municipalities and industries to discharge lesser quantities of pollutants, the need for low-cost and effective wastewater treatment increases (Hammer 1989). One possible low-cost and effective means of treating wastewater is with the use of constructed wetlands.

Natural And Constructed Wetlands

Wetlands provide many services and perform many functions for our natural environment. Wetlands provide fish and wildlife habitat, drinking water supply, ground water recharge, flood control, protection from erosion, sites for outdoor recreation, improvement of water quality, nutrient re-cycling, metals removal, point source and nonpoint source water pollution control, and opportunities for education and research (Dennison and Berry, 1993; Hammer, 1996). Although all of these functions of natural wetlands are important to our environment, Hammer (1991) states that the most important of these wetland functions is water quality improvement. Unfortunately, this function and the related functions of nutrient recycling, metals removal, and water pollution control are the least understood. Consequently, in order to maximize the effectiveness of these functions, an analysis of the current knowledge base on this subject is required.

Scientists first realized the advantages of having wetlands in the water environment when in the 1960's and 1970's the EPA began monitoring the water quality characteristics of natural wetlands that received wastewater discharges. It seemed that natural wetlands had been used as wastewater discharge sites for as long as sewage has been collected (for more than 100 years in some locations). It was not until the 1960's that these wetlands were monitored for their water quality characteristics and the water purification potential of these wetlands was finally realized (Kadlec and Knight, 1995).

Hammer (1994) states that constructed wetlands are formerly terrestrial environments that have been modified to create a system of undrained soils and wetland emergent and submergent vegetation whose primary purpose is the removal of contaminants or pollutants from wastewater. The processes in and the functions of natural wetlands define the water purification of constructed wetlands. Additionally, if constructed wetlands are designed properly, they serve to maximize the water purification potential of natural wetlands.

There are two primary types of constructed wetlands, free water surface (FWS) constructed wetlands, and subsurface flow (SF) constructed wetlands. Influent water in free water surface constructed wetlands flows over and largely above the surface of the soil and through the emergent stems and leaves of wetland vegetation. Water depth in FWS wetlands can range from 2 cm to 0.8 m or more, depending on the use and location of the wetland. Typically, FWS wetland depths are approximately 0.3 m (Reed et al., 1995). Water of subsurface flow (SF) wetlands passes entirely through the soil substrate, leaving no visible surface water flow. Typically, the soil substrate in SF wetlands is 0.3 to 0.6 m deep and made up of various sizes of gravel, crushed rock, and soil (Reed et al., 1995).

Although there are only a few SF wetlands in the U.S., European municipal systems frequently employ soil-based SF wetlands. Many operational SF systems have experienced serious substrate clogging, and therefore SF constructed wetlands are not recommended for any wastewater treatment greater than tertiary polishing of effluents with low nitrogen loading. Additionally, SF systems are not reliable treatment techniques for effective nitrogen removal (Hammer, 1994).

Constructed wetlands accomplish water quality improvement or purification through a variety of physical, chemical, and biological processes. Some of these processes are independent, and others depend upon the products of a previous process to facilitate pollutant removal. Wetland vegetation serves four purposes: (1) obstruction of wastewater flow, which reduces flow velocity and allows for particulate settling, (2) creation and maintenance of a thin-film bioreactor on the surface area of live and

decomposing plant material, (3) transportation of oxygen from the atmosphere to the largely anaerobic environment beneath the surface of the water, and (4) assimilation of wastewater nutrients and carbonaceous material. Because many organic and nitrogen compounds enter wetlands through wastewater absorbed to particulate matter, as flow velocity decreases these particles settle to the bottom of the wetland, providing additional surface area and nutrients to microbial biota. Additionally, as nutrients accumulate in the wetland environment, wetland plants grow more rapidly, produce more surface area for microbial growth, and provide more oxygen to their root structure and the attached microbiota. The result is a constructed wetland environment with excess nutrients and carbonaceous material, abundant microbial life, and aerobic and anaerobic sites for corresponding microbial-mediated pollutant removal.

Constructed Wetlands For Wastewater Treatment

Worldwide, there are approximately 1000 managed FWS wetland systems in operation for various purposes, and at least half of these systems are in the U.S. (Reed et al., 1995). Constructed wetlands have been used to adequately treat many different types of wastewater, including: (1) municipal wastewater, (2) acid mine drainage wastewater, (3) industrial wastewater, and (4) agricultural wastewater (Hammer, 1988; Kadlec and Knight, 1995; Reed et al., 1995; WPCF, 1990). Constructed wetlands for agricultural wastewater treatment for both nonpoint and single discharge pollution have been investigated extensively (Baldwin and Davenport, 1994; Brenton, 1994; Bankson, 1994; Campbell, 1995; Cathcart et al., 1994; Dennison and Berry, 1993; DuBowy and Reaves, 1994; Godfrey et al., 1985; Hammer, 1997; Hammer, 1994; Hammer, 1988; Healy and McCloud, 1994; Holmes et al., 1994; Hubbard et al., 1994; Hunt et al., 1994; Jann, 1994; Kadlec and Knight, 1995; Kent, 1994; MacMaster et al., 1994; McCaskey et al., 1994; Moshiri, 1993; Reaves et al., 1994; Reed et al., 1995; Sikora, 1994; Skarda et al., 1994; Toor and Eddleman, 1994; Tanner, 1995; WPCF, 1990).

Three main attributes of wetlands define their water purification potential: wetland hydrology, wetland soils, and wetland plants (Hammer, 1988; Kadlec and Knight, 1995; Reed et al., 1995). These three attributes in turn define the environment inside a wetland that facilitates microbial growth and nitrogen pollutant removal.

In the wetland environment, water is found in excess. This excess water presents a plant stress because it inhibits the diffusion of gases to and from plant roots (oxygen diffusion is about 10,000 times slower in water than in air) and because the presence of oxygen-demanding constituents in the water tend to lower the amount of dissolved oxygen (DO) available to supply root metabolism (Whitlow and Harris, 1979). Wetland plants, namely the vascular plants called hydrophytes (plants that thrive in the presence of excess water), have adapted to the flooded conditions found in wetlands through the development of various plant tissues and transport mechanisms necessary for growth in this hostile environment. Lenticels are small openings on the above water portions of wetland plants that provide an entry point for atmospheric oxygen that is then transported by an aeranchymous tissue network to and from the roots through the vascular tissues of the plant that are above water and in contact with the atmosphere (Armstrong, 1978; Jackson and Drew, 1984; Zimmerman, 1988). Lenticel surface area may be increased through plant growth, plant height increases, or the formation of swollen buttresses in trees and woody herbs and in cypress knees. Another adaptation to flooding shared by many hydrophytes is the growth of adventitious roots from flooded stem tissue. These roots can potentially extract DO and plant nutrients from water, where gases and nutrients may be more available than in the anaerobic soil zone (Kozlowski, 1984).

Oxygen transport into the root zone of wetland plants has been measured and found to be at various levels: 2.08 g of O2/m2 of wetland/day (Brix and Schierup, 1990) and between 5 and 12 g O2/m2/day (Armstrong et al., 1990) for Phragmites australis grown in gravel beds; and between 5 and 45 g O2/m2/day (Boon, 1985; Lawson, 1985), depending on plant density and oxygen stress levels in the root zone (Kadlec and Knight, 1995; Reed et al., 1995). Sometimes this oxygen transport seems to be great enough to only offset root metabolism, and therefore not result in additional aeration of the surrounding soil and sediment (Brix, 1990). These aerobic microsites around the roots and rhizomes of wetland plants can serve to initialize removal of wastewater nitrogen constituents in wetlands. Therefore, wetland plants can support aerobic bacteria that under appropriate conditions achieve pollutant removal (Reed et al., 1995).

Nitrogen Cycle In Wetlands

The nitrogen cycle is a very complex environmental interaction that defines the extent and fate of the nitrogen element in our environment. This cycle in a wetland includes six basic chemical transformations of nitrogen: 1) biological nitrogen fixation, 2) ammonification, 3) ammonia volatilization, 4) nitrification, 5) denitrification, and 6) biological assimilation. In a wetland, the magnitude of these transformations is directly related to wetland hydrology, wetland soil, wetland plants, temperature, pH, and dissolved oxygen content in the wetland water column.

Elemental nitrogen has an atomic weight of 14.01 g/mol with five electrons in its outer shell of atomic structure. Three electrons of this outer shell are available to form nitrogen-based compounds. These compounds occur in nature with varying stability and have oxidation states ranging from +5 to -3. Many organic and inorganic nitrogen compounds are essential for biological life. The most important inorganic forms of nitrogen are ammonia (NH3), nitrite (NO2-), nitrate (NO3-), nitrous oxide (N2O), and dissolved elemental nitrogen or dinitrogen gas (N2). Typically, inorganic forms of nitrogen are expressed in terms of elemental nitrogen with the terminology of nitrate-nitrogen (NO3- -N) or ammonia-nitrogen (NH3-N), for example. Organic nitrogen compounds include urea, amino acids, amines, purines, and pyrimidines. An understanding of each of these nitrogen compounds is essential to analysis of the nitrogen cycle.

Nitrate-nitrogen is the most highly oxidized form of nitrogen found in wetlands. Nitrate-nitrogen’s high oxidation state (+5) and overall negative charge dictate its existence and persistence in water. The resulting chemical nature of nitrate-nitrogen is chemically stable, highly mobile, and highly soluble in water. Nitrate-nitrogen’s high mobility and water solubility are a result of the negative charge (-1) of the compound. Because soil particles are generally negatively charged as well, nitrate-nitrogen has no affinity for these particles and thus is not bound by the soil. Therefore, nitrate-nitrogen would persist in the wetland environment were it not for two processes for nitrate- nitrogen removal, denitrification and plant uptake. Detrimental effects arise from excess levels of nitrate-nitrogen. The first of these effects is termed eutrophication, which becomes a problem in surface waters with high nitrate-nitrogen concentrations. The over-abundance of nitrate allows plants an unlimited supply of nitrogen, resulting in uncontrolled growth of algae and other plant species. Nitrate-nitrogen and subsequently nitrite-nitrogen are of further importance in water quality control because they are toxic to infants (they result in a potentially fatal disease known as methylglobanemia) when present in drinking water supplies of polluted surface or ground water (Kadlec and Knight, 1995). As a consequence of nitrate-nitrogen’s chemical nature and detrimental environmental effects, the removal of this compound from wetlands is important, and design procedures for its removal are necessary.

A number of important processes that serve to transport and translocate nitrogenous compounds without transforming them exist in wetlands. These processes include: 1) particulate settling and re-suspension, 2) diffusion of dissolved nitrogen compounds, 3) plant uptake and translocation, 4) litterfall, 5) sorption of soluble nitrogen substrates, 6) seed release, and 7) organism migrations (Kadlec and Knight, 1995). Although these processes do not significantly reduce nitrogen compound concentrations, the nitrogen cycle would exhaust itself without their aid.

Nitrification

Nitrification is the principal transformation pathway for the reduction of ammonia-nitrogen concentrations in wetlands. Nitrification involves a two-step microbial process that ultimately converts ammonia-nitrogen to nitrate-nitrogen via oxidation. The rate of nitrification is directly dependent upon DO concentration. If DO concentrations remain above 0.3 mg/L (Reddy and Patrick, 1984), two bacteria genera, Nitrosomonas and Nitrobacter, are able to oxidize ammonia (NH4+) to nitrate (NO3-). The overall nitrification process can be summarized by a single equation:

NH4+ + 2.0 O2 NO3- + 2 H+ + H2O (1)

As defined here, approximately 4.6 mg of O2 is required to oxidize a mg of ammonia-nitrogen to nitrate-nitrogen (Kadlec and Knight, 1995).

Denitrification

If not for denitrification, the biogeochemical processes that cause nitrogen fixation would ultimately deplete the atmosphere of nitrogen gas. Denitrification is an energy-requiring process that reduces nitrate or nitrite-nitrogen (NOx-N) to nitrogen gas, nitrous oxide, or nitric oxide. A stepwise reaction summary of denitrification is as follows (Tchobanoglous, 1991):

NO3- NO2- NO N2O N2 (2)

Denitrification is an essential and complementary process to heterotrophic metabolism in soil and aquatic environments when dissolved or free oxygen is absent. Aerobic respiration is an energy deriving mechanism that utilizes oxygen as the final electron acceptor in the most energetically positive step of respiration, the electron transport chain. In denitrification, an enzyme called nitrate reductase allows certain genera of bacteria to use the tightly bound oxygen atoms in nitrate and nitrite molecules as the final electron acceptors, in the absence of free oxygen (Kadlec and Knight, 1995).

Nitrate-nitrogen can also be reduced in wetlands through plant uptake. However, nitrate uptake by wetland plants is presumed to be less favored than ammonium uptake. Subsequently, nitrate-nitrogen loss in wetlands is often attributed to denitrification. However, other known and studied candidate mechanisms for nitrate loss in wetlands include assimilation by plants, assimilation by microbiota, and dissimilatory reduction to ammonium nitrogen (Kadlec and Knight, 1995).

Total nitrogen (TN) is the sum of all water soluble nitrogen constituents: nitrate-nitrogen, ammonia-nitrogen, and organic nitrogen. The removal or persistence of each of these nitrogen species is dependent upon each of the nitrogen cycle components. Therefore design models for the removal of TN in wetlands must recognize these nitrogen flux processes and integrate them into an overall wetland nitrogen removal scenario.