Pergamon PH: S0273-1223(97)00046-2
Wal. Sci. T«h. Vol. 35, No.5, pp. 1-10, 1997. C 1997IAWQ. Published by Elsevier Science Ltd Printed in Great Britain. 0273-1223197 $17'00 + 0'00
BIOGEOCHEMICAL INDICATORS TO
EVALUATE POLLUTANT REMOVAL
EFFICIENCY IN CONSTRUCTED WETLANDS K. R. Reddy and E. M. D'Angelo
University of Florida, Soil and Water Science Department, Wetland Biogeochemistry
Laboratory, 106 Newell HaI~ Gainesville, FL 32611, USA
ABSTRACf Wetlands support several aerobic and anaerobic biogeochemical processes that regulate removal/retention of pollutants, which has encouraged the intentional use of wetlands for pollutant abatement. The purpose of this paper is to present a brief review of key processes regulating pollutant removal and identify potential indicators that can be measured 10 evaluate treatment efficiency. Carbon and toxic organic compound removal efficiency can be determined by measuring soil or water oxygen demand, microbial biomass, soil Ell and pH. Similarly, nitrate removal can be predicted by dissolved organic C and microbial biomass. Phosphorus retention can be described by the availability ofreactive Fe and AI in acid soils and Ca and Mg in alkaline soils. Relationships between soil processes and indicators are useful tools to transfer mechanistic infonnation between diverse types of wetland treatment systems. @ 1997 IAWQ. Published by Elsevier Science Ltd
KEYWORDS Aquatic macrophytes; decomposition; microbial biomass; nutrient cycling; water quality.
INTRODUcnON Many wetlands are open systems that receive allochthonous inputs of organic matter, nutrients, metals, and toxic organic compounds from adjacent agricultural watersheds, industrial and urban areas. While many wetlands are sensitive to these contaminants, others are capable of attenuating pollutants. Differences between wetland systems are largely due to differences in physical, chemical, and biological conditions that affect transformations and transport processes and treatment efficiency in the soil-water-plant system. While individual fate processes are known, little is known about how processes are influenced by diverse conditions in wetlands, and conversely, how pollutants influence the functioning of wetlands. Evaluation for pollutant removal efficiency has considerable merit when wetlands are intentionally used for pollutant abatement. However, pollutant impacts on wetlands are a serious concern because of associated eutrophication problems and resulting adverse effects on the overall value and functioning of wetlands. Wetlands are considered low-cost alternatives for treating municipal, industrial and agricultural effluents. At present, there are thousands of wetland-based wastewater treatment systems around the world. During the past two decades a vast amount of literature has been published on various topics inclUding: (i) potential use 1
K. R. REDDY and E. M. D'ANGELO
of these systems to remove biochemical oxygen demand (BOD), nutrients, and to a limited extent metals and toxic organic compounds, (ii) engineering assessment and design criteria, and (ill) optimization and cost• effectiveness. Numerous studies on various aspects of constructed wetlands are available (Reddy and Smith, 1987; Hammer, 1989; Cooper and Findlater, 1990; Mitsch and Cronk, 1992; Moshiri, 1993). Design procedures are described in detail by the U. S. Environmental Protection Agency (EPA, 1988); WPCF (1990); Kadlec and Knight (1995); and Reed et al. (1995). To evaluate pollutant removal efficiency of constructed wetlands, we need to consider the following issues: (i) What biogeochemical processes can be used to assess the pollutant removal efficiency of wetlands? (ii) What soil and environmental factors regulate the retention and movement of pollutants in wetlands? (iii) Can the extent and rates of pollutant fate processes be predicted by easily measured soil and environmental indicators? (iv) Can conditions in constructed wetlands be optimized for pollutant removal? REDOX GRADIENTS
One of the main distinguishing features of wetlands is the presence of aerobic and anaerobic interfaces, which create steep redox gradients in the range of +700 to -300 mV. These gradients are affected by at least three different conditions, namely hydrologic fluctuations, presence of electron acceptors (such as 02, N03and SOi-), and transport of 02 by plants into the root zone (Fig. 1). Under drained conditions, anaerobic microsites within the soil aggregate can result in steep redox gradients. Under flooded conditions, diffusion of O 2 through floodwater maintains aerobic conditions at the soil-floodwater interface ranging in thickness of a millimetres to about 2 cm. In wetlands with high algal activity, photosynthetic production of 02 may result in a thicker aerobic soil zone during daytime, shrinking as 02 is consumed during the night. Redox gradients may also be setup as a result of buffering by alternate electron acceptors used during anaerobic respiration. For example, denitrification, FeOOH and S042- reduction and methanogenesis are distinguished by distinctly different redox potentials. Thus, redox gradients may be used as an indicator of potential nitrification-denitrification reactions, iron-oxide regulated precipitation of P, and oxidation of CH. and sulfides, and breakdown of toxic organic compounds. [A] Dissolved Oxygen (mg/L) 20
0 -10 -20
"g2 > '0
-80 -100 -120 -140
.••. SltamSecIimaU -
-160 0 100.200 300 400 SOO 600 700
Redox Potential (mY)
Figure 1. Aerobic-anaerobic interfaces in wetland loils. Dissolved oxygen gradients (A] at the soil-floodwater interface and (B] within a saturated soil aggregate. (D'Angelo and Reddy. unpublished results). (e] Redox profiles of stream sediments in Okeechobee Basin, Florida (Reddy, unpublished results).
Wetland plants have unique characteristics of adaptation to anaerobic soil conditions, such as developing internal air spaces (aerenchyma) for transporting 02 into the root zone. Depending on plant species, these air spaces can occupy up to 60% of the total tissue (Brix, 1994), permitting O 2 transport through molecular diffusion as a result of partial pressure gradients and mass flow as a result of temperature and humidity -
Pollutant removal efficiency in constructed wetlands
induced pressurization (Armstrong and Annstrong, 1990). Oxygen transport into the root zone plays an important role in BODs and NH4+ removal by promoting oxidation reduction reactions in the rhizosphere (Reddy et al., 1990). BIOGEOCHEMICAL INDICATORS Carbon. Wetlands are major sinks for e, that are typically characterized by accretion of organic matter (Fig. 2). Net accumulation of organic C is a result of the balance between primary production and heterotrophic respiration. Organic matter produced is deposited seasonally on the soil surface and may be eventually converted to a new soil layer providing long-term storage of C and nutrients. In addition to internal production of e, effluents containing particulate and dissolved C are added to constructed wetlands, usually measured as BOD (biochemical oxygen demand). The BOD added to wetlands may be removed by (i) settling of particulate BOD, and (ii) breakdown of soluble BOD during microbial respiration. Much of the added effluent e is removed in the water column and close to inflow.
Dissolved Inorganic QDlC) Di••olved organic C (DOC) Particulate organic C (POC)
fron pi nl
Organic matter accretion
Figure 2. Carbon transformations in the soil and water column of wetlands. [I) fragmentation and leaching, 2) mineralization, 3) plant/microbial uptake, 4) precipitation and solubilization, S) respiration, 6) methanogenesis, 7) methane oxidation, 8) burial, and 9) volatilization].
Decomposition of plant litter and particulate e is the main pathway for e removal and involves stepwise conversion of complex organic molecules to simple organic and inorganic constituents through: (i) abiotic leaching and fragmentation, (ii) extracellular enzyme hydrolysis, and (iii) aerobic and anaerobic catabolic activity of heterotrophic microorganisms. Step (i) is a physical process while, steps (ii) and (iii) are microbially mediated reactions which are affected by substrate quality (e.g. cellulose, lignin nutrient content), electron acceptors, environmental factors such as pH, temperature, and nutrient availability. It is generally believed tliat steps (i) and (ii) are rate limiting steps in the overall decomposition process. The residual undecomposed organic matter is detached from the plant and deposited on the soil surface, resulting in accretion of peat. One of the most important regulators of organic matter mineralization is supply of 02 and alternate electron acceptors. In most constructed wetlands receiving high loading of organic substrates, the electron donor to acceptor ratio is high, so decomposition with 02 as electron acceptor is restricted only to the water column. Depending on soil type and effluent composition, reduction of alternate electron acceptors N0l", 5°42- and He0l" and to a lesser extent FeOOH and Mn02' are coupled to organic matter decomposition (Burgoon et al., 1995).
K. R. REDDY and E. M. D'ANGELO
Table 1. A list of processes and respective biogeochemical indicators of the fate of C, N, P and toxic organic compounds in constructed wetlands Carbon
Indicators DIC + CH. in porewater Soil oxygen demand (SOD) Carbonaceous BOD Microbial biomass C Extracellular enzymes (Cellulases) Plant litter composition (Ugnocellulose index; C/N, CIP ratio) Soil Redox potential and pH Electron acceptor availability
NH.-N + N03-N (dissolved + exchangeable) Microbial biomass N Extracellular enzymes (proteases, peptidase) Plant litter composition (C, N and P content) Soil Redox potential and pH
NH.-N Dissolved oxygen
NH.-N (dissolved + exchangeable) pH Alkalinity Temperature
N03-N Microbial biomass C Plant litter composition Dissolved organic C Redox potential
Bicarbonate extractable P Microbial biomass P Extracellular enzyme (Phosphatase) Plant litter composition (C, N and P content) Redox potential
Soil organic matter Oxalate extractable Fe and AI HCI-extractable Ca and Mg Particle size distribution (clay. silt and sand) Redox potential and pH Algal species composition
Parameters shown for carbon
Organic matter content (particulate and dissolved) Redox potential and pH
Toxic organic compounds
Pollutant removal efficiency in constructed wetlands
A number of approaches can be used to evaluate C removal in constructed wetlands. The common approach is to monitor inflow-outflow effluent for BOD or total organic C, and calculate removal efficiency by the wetland. A second approach is to measure microbial decomposition of organic substrates, such as particulate (litter) and dissolved organic C in the soil and water column. This approach involves determining respiration potential by monitoring CO 2 and CH4 and calculating turnover of organic matter. This provides a more mechanistic approach, however it is limited because of the complexity associated with the experimental techniques. An alternate approach (a compromise between 'black box' and 'mechanistic approach') is to measure selected independent parameters (indicators) involved in organic matter mineralization, and develop empirical relationships between indicators and respective processes. Resulting equations can be used to predict C turnover rate within the wetland. A list of potential indicators is shown in Table 1. Microbial respiration rates may be adequately determined from 02 demand of water and soil (BOD and SOD). Indices based on microbial biomass, activity, and soil organic C have been proposed to provide an operationally-defined response of soil microbial populations to substrate quality and environmental conditions (Sparling, 1992). For example, the ratio of microbial respiration rate per unit of microbial biomass C can be used as microbial index to evaluate efficiency of constructed wetlands to process organic matter. Plant litter quality or composition, such as C/N or C/P ratios, cellulose and lignin content, have been significantly correlated to microbial breakdown of litter. The ligno-cellulose index (LCI = ligninlIignin + cellulose) has been used to characterize substrates for their decomposability. For example, during decomposition, the LCI for cattails was shown to increase from 0.2 to 0.8, with low values observed in easily decomposable live tissue and high values in the recalcitrant litter incorporated into soil organic matter (DeBusk, 1996). An empirical relationship between microbial respiration rates in soils/plant litter and microbial biomass C in peat wetlands was found to be: Cm = 0.057 [MBC] R2 = 0.881(n=34), where C m = microbial respiration rate, mg C g-I day-I; MBC = microbial biomass C, mg g-I (DeBusk, 1996; McLatchy and Reddy, 1996). Similarly, soil microbial respiration rates were also correlated to soil oxygen demand (SOD): Cm = 0.364 [SOD] - 0.002 R2 = 0.939 (n=20) where: Cm = microbial respiration rate, mmoles C g-I day -I; SOD=soil oxygen demand, mmoles 02 g-I day -I. This equation indicates that only 36% of SOD was accounted for by microbial respiration, with other pathways also involved in 02 consumption (e.g. oxidation ofNH4 +, CH 4, Fe 2+, and S2-). Nitrogen. Nitrogen entering constructed wetlands is present in particulate and dissolved organic and inorganic (NH4 + and N03-) forms. The relative proportion of each depends on type of wastewater and pretreatment. Particulate forms are removed through settling and burial, while the removal of dissolved forms is regulated by various biogeochemical reactions functioning in soil and water column (Fig. 3). Nitrogen reactions in wetlands effectively process inorganic N through nitrification and denitrification, ammonia volatilization, and plant and microbial uptake. Ammonification of dissolved organic N derived from detrital plant tissue or soil organic matter may be a source of inorganic N to the water column (D'Angelo and Reddy, 1994). Detailed reviews of N processes functioning in wetlands are presented by Reddy and D'Angelo (1994); and Howard-Williams and Downes (1993). A wide range of ammonification rates are reported in the literature, with values ranging between 0.0040.357 g N m-2 day -I (Martin and Reddy, 1996). Organic N mineralization can be described as a function of C/N ratio, extracellular enzyme (such as protease), microbial biomass and soil redox conditions (McLatchy and Reddy, 1996). Floodwater NH4-N maybe lost through ammonia volatilization, regulated by temperature, vegetation density, air movement above water surface, mixing in the water column, NH 4+ concentration,
K. R. REDDY and E. M. D'ANGELO
algal activity and associated pH fluctuations. In constructed wetlands, this process is most significant when the influent water contains high levels of NH4-N and pH exceeds 8.0. Nitrification occurs in aerobic zones of the water column. soil water interface and root zone. The relative importance of these zones in overall nitrification depends on 02 availability and NH4-N concentration. Nitrification rates were reported to be in the range of 0.01-0.161 g N m- 2 day-I (Martin and Reddy. 1996). These values are lower than the values reported for ammonification. suggesting that 02 and NH4-N availability limit nitrification. Measurement of 02 or Eh values in soil-water column can provide a reliable indication of potential nitrificatioh.
Inflow Soluble organic (SON), NII.-N and NOJ-N Particulate N
ON,NH.-N and NOJ-N Particulate N
._---• --.. _........_"'...... 0,
....... ~..... ... ~.~ .
Figure 3. Nitrogen transformations in soil and water column of wetlands. [I) volatilization. 2) plant and microbial uptake, 3) denitrification, 4) nitrification, 3) mineralization, 6) nitrogen fixation. 7) fragmentation and leaching. 8) sorption and desorption. 9) burial. and 10) nitrate reduction to ammonium).
floodwater NOf diffuses into anaerobic soil, where it is reduced to gaseous end products (N20 and N or NH4-N. Reported NOf removal rates were in the range of 0.003-1.02 g N m-2 day-I (Martin and Reddy, 1996). Denitrification rates are usually limited by NOf concentration and diffusion of NOf from aerobic zones to anaerobic sites (Martin and Reddy, 1996). Denitrification rates are usually higher in soils receiving steady loading of N03- than soils receiving low or negligible levels (Cooper, 1990; Gale et al., 1992). In· a system with active denitrification, N0 3-levels are usually low, thus measurement ofN0 3- in soil and water column does not provide a good indication of this process. Since denitrification is mediated by heterotrophic microorganisms, its rate may be indicated by available organic C (Reddy et al., 1982). Similarly, microbial biomass C and denitrification enzyme assay (DEA) can also serve as potential indicators of denitrification rates, Vegetation can playa significant role in N removal by: (i) assimilating N into plant tissue, and (ii) providing environment in the root zone for nitrification-denitrification to occur. Plants derive most of their N from soil porewater with only a small amount from floodwater. Measurement of plant biomass and tissue N can provide an indication of N removal efficiency by wetlands.
Phosphorus. Both biotic and abiotic processes regulate P removal by wetlands (Fig, 4). Biotic processes include: (i) uptake by vegetation, periphyton, and microbes, and (ii) mineralization of plant litter, and soil organic P (Gachter and Meyer, 1993). Abiotic processes include: (i) sedimentation and burial, (ii) adsorption
Pollutant removal efficiency in constructed wetlands
and precipitation and (iii) exchange processes between soil and overlying water column (Caraco et al., 1991; Reddy et al., 1996). Innow Soluble P Particulate P
Litter detached from the plan
B - -
Particulate P-'---'_ _ _...J
SolubleP Particulate P
Phosphorus .... accretion ......
~I ~ A7l
(Fe, AI or
Microbial f~ Biomass P . - Soluble ~
Particulate In ......nieP
Figure 4. Phosphorus transformations in soil and water column of wetlands. [1) adsorption and desorption; precipitation and dissolution. 2) plant and microbial uplake. 3) fragmentation and leaching, 4) mineralization, 5) sedimentation and burial].
Water column P can be readily removed by periphyton, followed by deposition of dead biomass on soil surface. Uptake of P from soil porewater by emergent macrophytes may drive downward diffusion from floodwater, enhancing biotic and abiotic retention mechanisms. Upon plant senescence, a substantial portion of P present in above ground tissue is translocated and stored back into below ground biomass (Davis and van der Valk, 1983). The remaining tissue P deposited on the soil surface is subject to enzyme hydrolysis and released back to the bioavailable pool (Hoppe et al., 1988; Chrost, 1991) or becomes an integral part of the soil for longer term storage. Organic P associated with humic substances accounts for a substantial portion (>40%) of the total P (Stewart and Tiessen, 1987). In aerobic zones, this fraction was shown to be rapidly available to plant. However, under anaerobic conditions, organic P forms are relatively resistant to enzyme hydrolysis, and hence are considered an important P sink. Inorganic P discharged into wetlands, or resulting from mineralization of organic P, may be retained by oxyhydroxides of Fe and AI in acid soils and by calcium minerals in alkaline soils. In soils dominated by Fe oxides, P can be readily immobilized through sorption and precipitation by ferric oxyhydroxide, and formation of ferric phosphate in the oxidized zones at the soil-water interface. In calcareous systems, P can be precipitated as Ca mineral bound-P, especially when pH of the floodwater is altered diurnally by photosynthetic activities of algae. The rate of adsorption is controlled by soil pH and Eh, adsorptive surface area (active iron and aluminium or calcium carbonate), and temperature. These regulators can be used as indicators to evaluate P retention capacity of wetland soils. The following empirical relationship was derived for a number of wetland soils and stream sediments in Florida (Reddy et al., 1996); Smax = 0.24 [oxalate Fe + AI]; R2 = 0.872 (n=6O); where Smax = maximum P retention capacity of soils; [oxalate Fe + AI] = ammonium oxalate extractable Fe and AI. The above equation is applicable for [Fe +AI] concentration in the range of 0-100 mmoles kg-t. To retain one mole ofP, about 4 moles of (Fe + AI) are needed.
K. R. REDDY and E. M. D'ANGELO
Toxic organics, Toxic organics added to wetlands undergo similar pollutant removal processes as natural
organic matter, including aerobic and anaerobic microbial breakdown, vegetative uptake, volatilization, photolysis, chemical hydrolysis, sorption and burial in the soil. The extent of these processes depends on the type of compound as well as biological and chemical conditions in the soil water column including pH, temperature, light intensity, nutrient and electron acceptor availability, organic matter content.
It is generally believed that highly chlorinated chemicals such as pentachlorophenol (pCP), PCBs, pentachloronitrobenzene (PCNB), hexachlorobenzene (HCB), tetrachloroethylene (PCE) typically undergo anaerobic dechlorination more readily than aerobic degradation, however, this results in accumulation of resistant, less chlorinated species. In contrast, less chlorinated organics, such as oils, phenols, and polyaromatic hydrocarbons (PAHs), tend to undergo aerobic degradation more readily. Hale et al (1991) found that 83% of variation in the half life of dichlorophenol in pond sediments was predicted from environmental conditions, sediment pH, Eh, and concentration of electron acceptors NOf and S042-. In some cases limitations to microbial degradation may be overcome through careful wetland management strategies such as alternately flooding and draining (Fogel et ai., 1982) and addition/removal of appropriate electron acceptors, donors, and nutrients (Bae and Rittman, 1995). Indicators to evaluate toxic organic removal include monitoring soluble and sorbed phases of the parent chemical and known degradation intermediates in the soil/water column. Passive sampling devices for pesticides (Zabik et al., 1992) may be useful for measuring in situ pesticide gradients and diffusive fluxes in treatment wetlands. The amount of toxic organic compound sorbed to soil substrates is a critical indicator not only as a sink, but may also limit bioavailability and toxicity to microbial degradation (Mihelcic et al., 1993). Environmental parameters which regulate sorption include soil pH, particulate and dissolved organic matter, as well as information about chemical octanol-water partition coefficients (Shimizu et al., 1992; Pardue et al., 1993; Koelmans and Lijklema, 1992; Lyman, 1982). Microbial biomass and activity measurements (e.g. respiration determined from CO 2 production or 02 consumption rates) may be valuable indicators to assess toxic organic removal, since transformations are largely microbially mediated. For example, Simon et al. (1992) developed an empirical equation to predict mineralization of fenamiphos insecticide as a function of soil pH and microbial Cltota! carbon ratio. In cases where toxic organics are overloaded to wetlands, microbial activity and degradation may be greatly curtailed. Microbial activitylbiomass measurements may also provide information about whether toxicity of certain organics adversely affects other desirable microbial transformations including nitrification (Sayler et al., 1982), and cellulose and lignin degradation (Katayama and Kuwatsuka, 1991; McKinley et al., 1982). For toxic organics that are deactivated by enzymatic hydrolysis or oxidation, such as organophosphorus pesticides, acyl anilides, phenols, and anilines (Bollag, 1992) enzyme assays may indicate the ability of wetlands to ameliorate these pollutants. Together with measurements of Eh, electron acceptor, and nutrient availability, a combination of measurements may more accurately depict the dominant types and extent of microbially mediated transformations of toxic organics in wetlands (Suflita and Sewell, 1991; Dolfmg and Harrison, 1993). CONCLUSIONS Wetlands offer sharp gradients in Eh and dissolved 02 creating aerobic-anaerobic interfaces at the soil• floodwater interface and the root zone. Gradients in alternate electron acceptors and pollutants are also observed in vertical and horizontal scales along aerobic-anaerobic interfaces. These interfaces support many aerobic-anaerobic transformations that are critical to pollutant removal. Many of these processes are often difficult to measure and can not be included in routine monitoring schemes to evaluate treatment efficiency. Several indicators which predict the processes regulating pollutant removal were identified. For example, microbial activity and biomass measurements provide direct indication of organic C, N and P removal. Similarly, N0 3- removal is related to soluble organic C in soils. Phosphorus precipitation can be predicted by extractable Fe, AI, Ca and Mg. Plant biomass composition may provide an indication of the role of
Pollutant removal efficiency in constructed wetlands
vegetation in pollutant removal. The relationships between indicators and respective processes provide useful tools for inclusion in predictive models to evaluate treatment efficiency of constructed wetlands.
REFERENCES Armstrong. J. and Armstrong. W. (1990). Light enhanced convective through flow increases oxygenation in rhizomes and rhizosphere of Phragmites australis (Cav.) Trin. ex. stend. N. Phytologist. 114. 123. Bac. W. and Rittman. B. E. (1995). Accelerating the rate of cometabolic degradations requiring an intracellular electron source _ Model and Biofilm Application. War. Sci. Tech .• 31(1) 29·39. Bollag. J. M. (1992). Decontaminating soil with enzymes. Environ. Sci. Technol.26. 1876-1881. l~rix. H. (1994). Functions of macrophytes in constructed wetlands. Wat Sci. Tech. 29(4}. 71-78,) . l!!~goon. P. S.• Reddy. K. R. and DeBusk. T. A.. (1995). Performance of sub-surface flow in wetlands. War. Environ. Res. 67. 8SS-862.
Caraco. N.• Cole. J. 1. and Likens. G. E. (1991). A cross-system study of phosphorus release from lake sediments. In: Comparative Analysis of Ecosystems. Springer Verlag. pp. 241-258. Chrost. R. J. (1991). Environmental control of the synthesis and activity of aquatic microbial ectoenzymes. Chap. 3. pp. 29-S9.1n R. J. Chrost (ed.) Microbial Enzymes in Aquatic Environments. Springer-Verlag. New York. NY. Cooper. A. B. (1990). Nitrate depletion in the riparian zone and stream channel of a small headwater catchment Hydrobiologia. 202, 13-26. Cooper. P.F. and Findlater. B. C. (1990). Constructed Wetlands in Waler Pollution Control (Adv. Wat. Pol1ut. Control no. 11). Pergamon Press. Oxford. England. D·Angelo. E. M. and Reddy. K. R. (1994). Diagenesis of organic matter in a wetland receiving hypereutrophic lake water. I. Distribution of dissolved nutrients in the soil and water column. J. Environ. Qual. 23. 937-943. Davis. C. B. and van der Valk. A. G. (1983). Uptake and release of nutrients by lining and decomposing Typha glauca Godr.• tissues at Eagle Lake. Iowa. Aquat. Bot. 16.75-89. DeBusk. W. F. (1996). Organic matter turnover along a nutrient gradient in the Everglades. Ph.D. dissertation. University of Florida, Gainesville. Fl. Dolfing. 1. and Harrison. B. K. (1993). Redox and reduction potentials as parameters to predict the degradation pathway of chlorinated benzenes in anaerobic environments. FEMS Microbiol Ecol. 13. 23·30. EPA (1988). Design manual· Constructed wetlands and aquatic plant systems for municipal wastewater treatment EPAl62S/I881022. Fogel. S.• Lancione. R. L. and Sewall. A. E. (1982). Enhanced biodegradation of methoxychlor in soil under sequential environmental conditions. Appl. Environ. Microbiol. 44(1}. 113-120. Gachter. R. and Meyer. J. S. (1993). The role of microorganisms in mobilization and fixation of phosphorus in sediments. Hydrobiologia 253. 103-121. Gale. P. M.• Reddy. K. R. and Graetz, D. A. (1992). Denitrification potential of soils from constructed and natural wetlands. EcoL Eng. 2. 229-130. Hale. D. D.• Rogers. J. E. and Wiegel. J. (1991). Environmental factors correlated to dichlorophenol dechlorination in anoxic freshwater sediments. Environ. Toxicol. and Chem. 10. 12SS-126S. Hammer. D. (1989). Constructed Wetlands for Wastewater Treatment: Municipal, Industrial and Agricultural. Lewis Pub!,. Chelsea, MI. Hoppe. H. G .• Kim. S. J. and Gocke. K. (1988). Microbial decomposition in aquatic environments: Combined processes of extracellular activity and substrate uptake. Appl. Env. Microbial. 54. 784-790. Howard-Williams. C. and Downes. M. T. (1993). Nitrogen cycling in wetlands. In Nitrat~: Processes. Parurns and Management. (T. P. Hurt. A. L. Heathwaite and S. T. Trudgill. eds.) John Wiley & Sons. pp. 100-140. Kadlec. R. H. and Knight, R. L. (199S). Trearment Wetlands. Lewis Publ. Chelsea, MI. Katayama, A. and Kuwatsuka, S. (1991). Effect of pesticides on cellulose degradation in soil under upland and flooded conditions. Soil Sci. Plant Nutri. 37(1). 1-6. Koelmans. A. A. and Lijklema, L. (1992). Sorption of 126.96.36.199 Tetrachlorahenzene to sediments: The application of a simple three phase mode!. Chemosphere 25. 313-325. Lyman. W. J. (1982). Handbook o/Chemical Property Estimation Methods. McGraw-Hill. Inc. Martin. J. F. and Reddy. K. R. (1996). Interaction and spatial distribution of wetland nitrogen processes. Ecol. Modeling (in review). McKinley. V. L.• Federle. T. W. and Vestal. J. R. (1982). Effects of petroleum hydrocarbons on plant litter microbiota in an arctic lake. AppL Environ. Microbiol. 43(1). 129-135. McLatchy. G. P. and Reddy. K. R. (1996). Rcgulation of organic mattcr decomposition and nutrient release in a wetland soil. J. Environ. Qual. (in review). Mihelcic. J. R.. Leuking. D. R.. Mitzell. R. J. and Stapleton. J. M. (1993). Bioavailability of sorbed- and separate phase chemicals. Biodegradation 4. 141-IS3. Mitsch. W. J. and Cronk, J. K. (1992). Creation and restoration of wetlands: some design considerations for ecological engineering. Adv. Soil Sci. 17. 217-2S5. Moshiri. G. A. (ed) (1993). Construct~d Wetlands/or Water Quality Improvement. Lewis Pub!,. Ann Arbor. MI
K. R. REDDY and E. M. D'ANGELO
Pardue, 1, H" Masscheleyn, P.H., Delaune, R. D. and Patrick, Jr., W. H. (1993). Assimilation of hydrophobic chlorinated organics in freshwater wetlands: Sorption and sediment-water exchange. Env. Sci. Technol. 27(5), 875-882. Reddy, K. R., Rao, P. S. C. and Jessup, R. E. (1982). The effect of carbon mineralization and denitrification kinetics in mineral and organic soils. Soil Sci. Soc. Am.I. 46, 62-68. Reddy, K. R. and Smith, W. H. (1987). Aquatic Pklnts for Water Treatment and Resource Recovery. Magnolia Publ. Inc., Orlando, Fl. pp. 1032. Reddy, K. R., D'Angelo, E. M. and DeBusk, T. A. (1990). Oxygen transport through aquatic macrophytes: The role in wastewater treatment. Journal of Env. Qual. 19,261-267. Reddy, K. R. and D'Angelo, E. M. (1994). Soil processes regulating water quality in wetlands. In. Global Wetlands - Old World and New (W. Mitsch ed). pp. 309-324. Elsevier Publ. New York. Reddy, K. R., Kadlec, R. H., Flaig, E. and Gale, P. M. (1996). Phosphorus assimilation in streams and wetlands: Critical reviews. Env. Science and Tech. (in review). Reed, S. C., Crites, R. W. and Middlebrooks, E. 1. (1995). Natural Systems for Waste Management and Treatment. 2nd Ed. New York: McGraw Hill. Sayler, G. S., Shiaris, M. P.. Beck, W. and Held, S. (1982). Effects of polychlorinated biphenyls and environmental biotransformation products on aquatic nitrification. AppL Environ. Microbiol. 43(4), 949-952. Shimizu, Y., Yamazaki, S. and Terashima, Y. (1992). Sorption of anionic pentachlorophenol (PCP) in aquatic environments: The effect of pH. Wat. Sci. Tech. 25(11),41-48. Simon, L., Spiteller, M., Haisch, A. and Wallnofer, P. R. (1992). Influence of soil properties on the degradation of the nematocide fenamiphos. Soil BioL Biochem. 24(8), 769-173. Sparling, G. P. (1992). Ratio of microbial biomass carbon to soil organic carbon as a sensitive indicator of changes in soil organic matter. Aust. J. Soil Res. 30, 195-207. Stewart. 1. W. B. and Tiessen, H. (1987). Dynamics of soil organic phosphorus. Biogeochemistry 4, 41-60. Suflita, 1. M. and Sewell, G. W. (1991). Anaerobic biotransformations of contaminants in the subsurface. EPN6OOIM-90/024. 9 pp. WPCF (1990). Natural Systems for Wastewater Treatment. Manual of Practice FD-16. Water Pollution Control Federation. pp. 270. Zabik, J. M., Aston, L. S. and Seiber, J. N. (1992). Rapid characterization of pesticide residues in contaminated soils by passive sampling devices. Env. ToxicoL and Chemistry. 11(6),765-170.