Natural wastewater treatment systems (2nd edition)(1)

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Natural Wastewater Treatment Systems: An Overview

Natural Treatment Processes

Wastewater management relies on natural processes, such as gravity for sedimentation and biological organisms for treatment While conventional systems often incorporate energy-intensive mechanical equipment, a natural system primarily utilizes these inherent components to function effectively Typically, it may include pumps and piping for wastewater conveyance, but it does not solely rely on external energy sources for its primary treatment processes.

Following the passage of the Clean Water Act of 1972, there was a renewed interest in natural waste treatment methods in the United States Initially, the focus was on achieving the law's "zero discharge" mandate through advanced mechanical wastewater treatment units However, the high energy demands and costs associated with this approach led to the exploration of alternative solutions Consequently, land application of wastewater emerged as the first rediscovered natural technology for effective waste treatment.

In the 1840s, England prioritized avoiding water pollution and returning nutrients from wastewater to the land, which was the only accepted waste treatment method until modern devices emerged Research demonstrated that land treatment could meet the goals of PL 92-500 while effectively reusing nutrients, minerals, and organic matter from waste Following the passage of PL 92-500, land treatment gained recognition among engineers as a viable wastewater treatment option and is now routinely considered in project planning and design Additionally, natural treatment methods like lagoon systems and land application of sludges remain in use; lagoons replicate the interactions found in natural ponds, while sludge application follows traditional farming practices involving animal manure.

Aquatic and wetland concepts represent innovative approaches in the United States for utilizing wastewaters and sludges, offering cost-effective wastewater treatment alternatives This article discusses various sludge management techniques, including conditioning, dewatering, disposal, and reuse, which leverage natural components and processes The biosolids management procedures outlined align with the current U.S Environmental Protection Agency (EPA) regulations and guidelines for sewage sludge handling, ensuring compliance with 40 CFR Parts 257, 403, and 503.

1.1.2 W asteWater t reatment c oncepts and p erformance e xpectations

Effective wastewater treatment can be achieved through three primary natural systems: aquatic, wetland, and terrestrial Each system relies on natural physical and chemical reactions, along with distinct biological elements that contribute to the treatment process.

Table 1.1 outlines the design features and performance expectations for natural aquatic treatment units, highlighting that the primary treatment responses are attributed to biological components The process design chapters further categorize aquatic systems into pond and wetland systems for clarity.

Chapter 4 focuses on systems reliant on microbial life and lower forms of plants and animals, contrasting with the higher plant and animal systems discussed in Chapters 6 and 7 The performance and water quality in many pond systems, as shown in Table 1.1, are heavily influenced by the presence of algae, which play a crucial role in providing oxygen and facilitating nitrogen removal through algal–carbonate reactions However, removing algae can be challenging, particularly when strict suspended solids limits are imposed, necessitating alternative solutions to facultative ponds Controlled discharge systems have been developed to retain treated wastewater until the pond's water quality aligns with that of the receiving water Additionally, hyacinth ponds, mentioned in Table 1.1, inhibit algal growth by shading the water surface, thereby reducing sunlight penetration Other vegetation and animal life utilized in aquatic treatment units are further explored in Chapters 6 and 7.

Wetlands are areas where the water table is at or above the ground surface for extended periods, leading to saturated soil and supporting specific vegetation Numerous studies across different geographical locations have confirmed wetlands' effectiveness in wastewater treatment These wetlands can include natural ecosystems such as marshes, swamps, and bogs, as well as specially constructed systems designed for wastewater management.

Table 1.2 summarizes the design features and expected performance of the three primary wetland categories A significant limitation on the utilization of many natural marshes is their classification as receiving waters by regulatory authorities Consequently, wastewater must comply with discharge standards before being applied to the wetland, which hinders the full utilization of the wetland's renovative potential.

Design Features and Expected Performance for Aquatic Treatment Units

Organic Loading (lb/ac-d; kg/ ha-d)

TSS, 40–100 Partial-mix aerated pond

TSS, 30–60 Storage and controlled- discharge ponds

Rooted aquatic plants play a significant role in the treatment of wastewater and sludge when utilized in sand and gravel basins This method, highlighted in a workshop at Clemson University, emphasizes the effectiveness of low-cost wastewater treatment solutions The research conducted by Banks and Davis, along with contributions from Middlebrooks et al., showcases the potential of these natural systems to enhance water quality while providing an environmentally friendly alternative to traditional treatment methods.

Pollut Control Fed., 53(7), 1172–1198, 1981; and USEPA, Process Design Manual for Land Treatment of Municipal Wastewater: Supplement on Rapid Infiltration and Overland Flow, EPA

625/1-81-013a, U.S Environmental Protection Agency, Washington, DC, 1984.

The initial cell in the advanced water treatment system functions as a facultative or aerated treatment unit, effectively managing biochemical oxygen demand (BOD), total nitrogen (TN), total phosphorus (TP), and total suspended solids (TSS) for optimal water quality improvement.

Constructed wetlands offer advantages over natural marshes by eliminating specific influent quality requirements and providing more reliable hydraulic control The three primary types of constructed wetlands include free water surface (FWS) wetlands, which mimic natural marshes with exposed water surfaces; subsurface flow (SSF) wetlands, which utilize a permeable medium with water levels maintained below the surface; and vertical flow (VF) wetlands, where water flows vertically through sand and gravel media For comprehensive details on these systems and their variations, refer to the relevant chapters.

6 and 7 Another variation of the concept used for sludge drying is described in Chapter 9.

Design Features and Expected Performance for Four Types of Wetlands

Organic Loading (lb/ac-d; kg/ha-d)

TN, 5–20 Subsurface flow Secondary to

TN, 5–20 Vertical flow Primary to

Rooted aquatic plants play a significant role in the treatment of wastewater and sludge within sand and gravel basins According to Banks and Davis (1983), these plants effectively enhance the purification process, offering a low-cost solution for wastewater management The integration of such natural systems can improve water quality while promoting sustainable practices in environmental engineering.

Pollut Control Fed., 53(7), 1172–1198, 1981; and Reed, S.C et al., Wetlands for wastewater treatment in cold climates, in Proceedings of the Water Reuse III Symposium, American Water Works Association, August 1984, Denver, CO, 1984.

Note: AWT, advanced water treatment; BOD, biochemical oxygen demand; TN, total nitrogen; TSS, total suspended solids.

The three basic terrestrial concepts for wastewater treatment, detailed in Table 1.3, rely on physical, chemical, and biological reactions within the soil matrix Both slow rate (SR) and overland flow (OF) methods necessitate vegetation, with SR capable of utilizing diverse plant types, including trees and row-crop vegetables, while OF relies on perennial grasses for vegetative cover Soil aquifer treatment (SAT) systems, or rapid infiltration (RI) systems, often have hydraulic loading rates too high for supporting vegetation but can still produce high-quality effluent The SR process can be designed to achieve drinking water quality in the percolate, and while all three concepts allow for treated water reuse, recovery is most straightforward with OF due to its surface discharge to ditches Most SR and SAT systems require underdrains or wells for recovery, and SAT has proven effective in removing organic chemicals known as constituents of emerging concern (CEC).

On-site systems designed for single-family homes, schools, public facilities, and commercial operations typically involve a preliminary treatment phase followed by in-ground disposal For a detailed discussion of these on-site concepts, refer to Chapter 10 Additionally, Chapters 6 and 7 provide insights into small-scale constructed wetlands utilized for the preliminary treatment process.

Project Development

When developing a waste treatment project, whether municipal or industrial, it is crucial to address both institutional and social factors alongside technical considerations, as these elements can significantly influence decision-making during the planning and design phases Understanding current regulatory requirements at federal, state, and local levels is essential, and engineers should identify these requirements early in the project development process to ensure institutional feasibility Valuable insights on the institutional and social dimensions of project development can be found in the works of Deese (1981), Forster and Southgate (1983), and USEPA (2006), with additional summary guidance provided in Table 1.5.

Characterize waste Define the volume and composition of the waste to be treated

Not covered in this text; see Metcalf and Eddy

(1981, 2003) Concept feasibility Determine which, if any, of the natural systems are compatible for the particular waste and the site conditions and requirements

Design limits Determine the waste constituent that controls the design

Collection network in the community, pump stations, transmission piping, etc.

Not covered in this text; see Metcalf and Eddy,

The technical requirements for project development are outlined in the work of Metcalf and Eddy (1981, 2003), which specifies the necessary criteria across various chapters This text does not cover detailed information on waste characterization or civil and mechanical engineering design, as these aspects are not exclusive to natural systems.

2003) are recommended for that purpose.

Banks, L and Davis, S (1983) Wastewater and sludge treatment by rooted aquatic plants in sand and gravel basins, in Proceedings of a Workshop on Low Cost Wastewater

Treatment, Clemson University, Clemson, SC, April 1983, pp 205–218.

Bastian, R.K and Reed, S.C., Eds (1979) Aquaculture Systems for Wastewater Treatment, EPA 430/9-80-006, U.S Environmental Protection Agency, Washington, DC.

Crites, R.W (2009) Soil Aquifer Treatment for Microconstituents Removal Proceedings of WateReuse Symposium, Seattle, WA, September 2009.

Deese, P.L (1981) Institutional constraints and public acceptance barriers to utilization of munic- ipal wastewater and sludge for land reclamation and biomass production, in Utilization of

Municipal Wastewater and Sludge for Land Reclamation and Biomass Production, EPA

430/9-81-013, U.S Environmental Protection Agency, Washington, DC.

Drewes, J.E., Sedlak, D., Snyder, S., and Dickenson, E (2008) WateReuse Foundation and WERF, Development of Indicators and Surrogates for Chemical Contaminant Removal during Wastewater Treatment and Reclamation.

Forster, D.L and Southgate, D.D (1983) Institutions constraining the utilization of municipal wastewaters and sludges on land, in Proceedings of Workshop on Utilization of Municipal

Wastewater and Sludge on Land, University of California, Riverside, February 1983, pp 29–45.

In their 1979 report, Jewell and Seabrook explore the historical context of land application as a viable treatment alternative for wastewater, published by the U.S Environmental Protection Agency Additionally, Metcalf and Eddy's 1981 publication, "Wastewater Engineering: Collection and Pumping of Wastewater," provides comprehensive insights into the engineering practices related to wastewater management Together, these works underscore the evolution and significance of effective wastewater treatment methodologies.

Metcalf and Eddy (2003) Wastewater Engineering: Treatment and Reuse, 4th ed., McGraw- Hill, New York.

Middlebrooks, E.J., Middlebrooks, C.H., and Reed S.C (1981) Energy requirements for small wastewater treatment systems, J Water Pollut Control Fed., 53(7), 1172–1198. Middlebrooks, E.J., Middlebrooks, C.H., Reynolds, J.H., Watters, G.Z., Reed, S.C., and George, D.B (1982) Wastewater Stabilization Lagoon Design, Performance and

Reed et al (1979) provided insights into the cost of land treatment systems in a report by the U.S Environmental Protection Agency, highlighting the economic aspects of these systems Additionally, Reed and colleagues (1984) discussed the effectiveness of wetlands for wastewater treatment in cold climates during the Water Reuse III Symposium, emphasizing their potential in enhancing water quality in challenging environments.

USEPA (2006) Process Design Manual for Land Treatment of Municipal Wastewater, EPA/625/R-06/016, U.S Environmental Protection Agency, Cincinnati, OH.

USEPA (1984) Process Design Manual for Land Treatment of Municipal Wastewater:

Supplement on Rapid Infiltration and Overland Flow, EPA 625/1-81-013a, U.S

Environmental Protection Agency, Washington, DC.

Planning, Feasibility Assessment, and Site Selection

Concept Evaluation

Natural systems can be categorized into discharging and nondischarging systems Discharging systems include treatment ponds, constructed wetlands, and overland flow land treatment, often permitted under the National Pollutant Discharge Elimination System (NPDES) In contrast, nondischarging systems encompass reuse wetlands, slow rate land treatment, onsite methods, and biosolids treatment and reuse Key factors for constructing discharging systems include site topography, soils, geology, and groundwater conditions, while these elements are critical to the treatment process in nondischarging systems Design features and performance expectations for both system types are detailed in Tables 2.1 through 2.3, with special site requirements outlined for planning in Tables 2.1 and 2.2 It is assumed that percolate from nondischarging systems interacts with existing groundwater, and compliance is typically measured by the quality of this percolate as it reaches the project boundary.

Special Site Requirements for Discharge Systems

When designing treatment ponds, it is essential to ensure proximity to surface water for discharge, utilize impermeable soils or liners to minimize percolation, avoid steep slopes, and position them outside of flood plains while ensuring no bedrock or groundwater is present within the excavation depth Similarly, constructed wetlands require proximity to surface water for discharge, impermeable soils or liners, gentle slopes of 0%–6%, and must also be located outside of flood plains with no bedrock or groundwater within the excavation depth For overland flow systems, the use of relatively impermeable soils, such as clay and clay loams, is crucial, along with appropriate slope management.

0%–12%, depth to groundwater and bedrock not critical but 0.5–1 m desirable, must have access to surface water for discharge or point of water reuse

Underdrained slow rate (SR) and soil aquifer treatment (SAT)

For SR, the system is similar to the tables in Chapter 1 and Table 2.2, but it includes an impermeable layer or high groundwater, necessitating the use of underdrains to eliminate percolating water In the case of SAT, wells or underdrains can be utilized to manage and discharge percolating water effectively.

Special Site Requirements for Nondischarging System

Slow rate (SR) Sandy loams to clay loams: >0.15 to 1.5 m, slopes 1500 Injection or surface application at all of the above

The site requirements outlined in Tables 2.1 and 2.2 will determine the feasibility of the intended concept If conditions are favorable, the site can be deemed suitable, allowing for a shift towards final design and potential detailed field tests In cases where multiple sites or concepts remain feasible, a preliminary cost analysis is essential to identify the most cost-effective option The design criteria from Chapters 4 to 10 should guide the preliminary design, while planning-level equations should be avoided for final design Cost-effectiveness must be assessed through a comprehensive estimate of all treatment system components, including pumping, pre-application treatment, storage, land, and final water disposition Additionally, non-financial factors, such as social and institutional acceptance, will significantly influence the final selection.

Black, A., Ed (1965) Methods of Soil Analysis Part 2 Chemical and Microbiological

Properties, Agronomy 9, American Society of Agronomy, Madison, WI.

Bouwer, H (1978) Groundwater Hydrology, McGraw-Hill, New York.

Childs, E.C (1969) An Introduction to the Physical Basis of Soil Water Phenomena, John Wiley & Sons, London.

Crites, R.W and Tchobanoglous, G (1998) Small and Decentralized Wastewater Management

Systems, McGraw-Hill, New York.

Crites, R.W., Reed, S.C., and Bastian R.K (2000) Land Treatment Systems for Municipal and

Industrial Wastes, McGraw-Hill, New York.

Duke, H.R (1972) Capillary properties of soils: influence upon specific yields, Trans Am

Jackson, M.L (1958) Soil Chemical Properties, Prentice-Hall, Englewood Cliffs, NJ.

McKim, H.L., Cole, R., Sopper, W., and Nutter, W (1982) Wastewater Applications in Forest

Ecosystems, CRREL Report 82-19, U.S Cold Regions Research and Engineering

NOAA, Climatic Summary of the United States (a 10-year summary), National Oceanic and Atmospheric Administration, Rockville, MD.

NOAA, Local Climatological Data, annual summaries for selected locations, National Oceanic and Atmospheric Administration, Rockville, MD.

NOAA, Monthly Summary of Climatic Data, National Oceanic and Atmospheric Administration, Rockville, MD.

Richards, L.A., Ed (1954) Diagnosis and Improvement of Saline and Alkali Soils, Agricultural Handbook No 60, U.S Department of Agriculture, Washington, DC.

In 1980, G.L Taylor presented a preliminary method for evaluating sites designated for the treatment of municipal wastewater through spray irrigation on forest land This research was shared during the Conference of Applied Research and Practice on Municipal and Industrial Waste held in Madison, Wisconsin The findings emphasize the importance of site assessment in optimizing wastewater treatment processes in forested areas.

USDOI (1978) Bureau of Reclamation: Drainage Manual, U.S Department of Interior, U.S Government Printing Office, Washington, DC.

USEPA (1981) Process Design Manual: Land Treatment of Municipal Wastewater, EPA 625/1-81-013, Center for Environmental Research Information, U.S Environmental Protection Agency, Cincinnati, OH.

USEPA (1983) Process Design Manual: Land Application of Municipal Sludge, EPA 625/1- 83-016, Center for Environmental Research Information, U.S Environmental Protection Agency, Cincinnati, OH.

USEPA (1995) Process Design Manual: Land Application of Sewage Sludge and Domestic

Septage, EPA 625/R-95/001, Center for Environmental Research Information, U.S

Environmental Protection Agency, Cincinnati, OH.

USEPA (2002) Onsite Wastewater Treatment Systems Manual, EPA 625/R-00/008, Center for Environmental Research Information, U.S Environmental Protection Agency, Cincinnati, OH.

USEPA (2006) Process Design Manual: Land Treatment of Municipal Wastewater Effluents, EPA 625/R-06/016 Center for Environmental Research Information, U.S Environmental Protection Agency, Cincinnati, OH.

Basic Process Responses and Interactions

Water Management

Effective water management must address key issues such as the movement of contaminants through groundwater, the risk of leaks from ponds and aquatic systems, the possibility of groundwater mounding under land treatment systems, the necessity for proper drainage, and the maintenance of optimal flow conditions in ponds, wetlands, and other water bodies.

Chapter 2 introduced some of the hydraulic parameters (e.g., permeability) that are important to natural systems and discussed methods for their determination in the field or laboratory It is necessary to provide further details and definitions before undertaking any flow analysis.

Field and laboratory test results can vary in depth and area, even when the same soil type is present across a site For systems relying on water infiltration and percolation, the soil layer with the lowest permeability is used as the design basis In instances of significant data variability, it is essential to establish a "mean" permeability for effective design.

In uniform soil, vertical permeability (K v) remains consistent across depth and area, with any test result discrepancies attributed to variations in testing procedures Consequently, K v can be accurately represented as the arithmetic mean.

K K K K K am = 1+ 2+n 3+ n (3.1) where K am is the arithmetic mean vertical permeability, and K 1 through K n are indi- vidual test results.

Where the soil profile consists of a layered series of uniform soils, each with a distinct K v generally decreasing with depth, the average value can be represented as the harmonic mean:

D = soil profile depth d n = depth of nth layer

In the absence of a discernible pattern or preference from statistical analysis, it is essential to assume a random distribution of K values for a given layer Under these circumstances, the geometric mean serves as the most conservative estimate of the true K value.

The geometric mean permeability, denoted as K gm, is calculated using the formula K gm = [(K 1 )(K 2 )(K 3 )(K n )] 1/n This equation can also be applied to ascertain the lateral permeability, K h, by utilizing relevant data Additionally, Table 2.17 provides typical values for the ratio of K h to K v, offering valuable insights for understanding permeability relationships.

To determine the actual flow velocity in a groundwater system, it is essential to integrate Darcy’s law, the fundamental velocity equation from hydraulics, along with the soil's porosity, as groundwater movement is restricted to the pore spaces within the soil.

K h = horizontal saturated permeability, mid (ft/d; m/d) ΔH/ΔL = hydraulic gradient (ft/ft; m/m) n = porosity (as a decimal fraction; see Figure 2.4 for typical values for in situ soils)

Equation 3.4 can also be used to determine vertical flow velocity In this case, the hydraulic gradient is equal to 1 and K v should be used in the equation.

The transmissivity of an aquifer is determined by multiplying its permeability by the saturated thickness, indicating how effectively a unit width of the aquifer can transmit water The flow rate of water through this unit width can be calculated using the formula q = K b w H h L.

∆ (3.5) where q = volume of water moving through aquifer (ft 3 /d; m 3 /d) b = depth of saturated thickness of aquifer (ft; m) w = width of aquifer, for unit width w = 1 ft (1 m) ΔH/ΔL = hydraulic gradient (ft/ft; m/m)

Well pumping tests are essential for determining aquifer properties, particularly the transmissivity, which can be estimated through the analysis of pumping rate and draw-down data (Bouwer 1978; USDOI 1978).

Groundwater contamination dispersion results from molecular diffusion and hydrodynamic mixing, leading to reduced contaminant concentration but an increased contact zone at downgradient locations This dispersion occurs both longitudinally (D x) and transversely (D y) to the flow path Studies using dye in homogeneous and isotropic granular media have shown that dispersion forms a cone shape approximately 6° from the point of application.

In 1953, it was observed that stratification and areal differences significantly increase both lateral and longitudinal dispersion in geological formations For instance, in fractured rock, the divergence of the cone can exceed 20° Additionally, the dispersion coefficient is closely linked to the seepage velocity, as outlined in previous studies.

The dispersion coefficient (D) is measured in both longitudinal and transverse dimensions (ft²/d; m²/d), while dispersivity (a) is also defined in these dimensions (ft; m) Seepage velocity in a groundwater system is calculated as v = V/n, where V represents Darcy's velocity from Equation 3.5 and n denotes porosity For typical values of in situ soils, refer to Figure 2.4.

Measuring dispersivity in the field and laboratory poses significant challenges, often requiring the addition of a tracer at the source and monitoring its concentration in nearby wells Field experiments, such as those conducted at Fort Devens, Massachusetts, have yielded an average dispersivity value of 10 m²/d However, increasing the assumed dispersivity by 100% or more resulted in minimal changes to predicted contaminant transport levels Additionally, many reported values in the literature are site-specific and "fitted," limiting their reliability for broader applications.

Hydrodynamic dispersion impacts contaminant concentrations uniformly; however, adsorption, precipitation, and chemical interactions with groundwater constituents slow the movement of affected contaminants This phenomenon is quantified by the retardation factor (R_d), which can vary from 1 to 50 for organic substances commonly found at field sites Conservative substances like chlorides exhibit the lowest R_d values, moving at the same velocity as surrounding water, with concentration changes attributed solely to dispersion The R_d is influenced by soil and groundwater characteristics, leading to variability across locations; for instance, metals may have an R_d close to 1 in clean sandy soils but approach 50 in clayey soils The R_d for organic compounds is determined by their sorption to soil organic matter, along with volatilization and biodegradation processes, which are less dependent on soil type, resulting in more uniform removal across different sites Insoluble compounds such as DDT and PCBs are effectively removed by most soils, while highly soluble compounds like chloroform and benzene are removed less efficiently Retardation factors for various organic compounds are summarized in Table 3.1, based on literature estimates.

Retardation Factors for Selected Organic Compounds

The migration of pollutants in groundwater is influenced by various factors, which can impact ponds and aquatic systems, particularly in the context of SR and rapid infiltration land treatment methods Understanding the subsurface zone of influence is crucial, as illustrated in Figure 3.1, which depicts a rapid infiltration basin system or treatment pond with significant seepage It is often essential to assess pollutant concentrations in the groundwater plume at specific distances downgradient from the source Additionally, determining the distance at which a certain concentration will occur over time or the timeline for a concentration to reach a particular point is important for effective management.

FIGURE 3.1 Subsurface zone of influence for SAT basin.

Biodegradable Organics

Biodegradable organic contaminants in wastewater, whether dissolved or suspended, are defined by their Biochemical Oxygen Demand (BOD) This article outlines typical BOD removal expectations for various natural treatment systems, as detailed in Tables 1.1 to 1.3.

BOD loading is a critical design factor for pond, aquatic, and wetland systems, primarily influencing the maintenance of aerobic conditions and odor control The primary sources of dissolved oxygen (DO) in these environments are surface reaeration, which is enhanced by wind or mechanical turbulence, and photosynthetic oxygenation Research indicates that DO levels in unaerated wastewater ponds fluctuate with photosynthetic activity, peaking in the early afternoon and dropping at night Algal photosynthetic responses are driven by light availability, water temperature, and the presence of essential nutrients and growth factors.

Algae can be challenging to eliminate and may lead to excessive suspended solids in effluent, prompting some pond and aquaculture systems to employ mechanical aeration for oxygen supply In partially mixed aerated ponds, the increased depth and partial mixing of the turbid water inhibit algae growth more effectively than in facultative ponds Additionally, FWS wetland systems help control algae proliferation, as the vegetation restricts light penetration into the water column.

Emergent plant species in wetlands treatment play a crucial role in enhancing oxygen transfer from leaves to roots, facilitating organic decomposition While these plants do not directly remove biological oxygen demand (BOD), they support a variety of attached growth organisms responsible for this process The structures of these emergent plants, including stems, stalks, roots, and rhizomes, provide essential surfaces for microbial activity To optimize interactions between wastewater and microbial growth, a shallow reactor and low flow velocity are necessary.

Wu et al (2001) found that Typha latifolia released minimal oxygen from its roots in constructed wetlands, with atmospheric diffusion being the primary oxygen source Their findings suggested species-specific differences, as Spartina pectinata could potentially release oxygen up to 15 times more than T latifolia They concluded that the oxygen supplied to wetlands via macrophyte roots and atmospheric diffusion was insufficient for the oxidation of ammonia.

In municipal land treatment processes, the biochemical oxygen demand (BOD) of wastewater or sludge is rarely the critical design factor, as highlighted in Idaho DEQ (2005) Conversely, for high-strength industrial wastewaters, BOD loading rates often serve as the primary design constraint (USEPA 2006; Coody et al 1986) Additional design considerations include nitrogen levels, metals, toxic substances, and the hydraulic capacity of soils, which typically prevent municipal systems from reaching their maximum organic biodegradation limits For reference, Table 3.2 outlines the typical organic loadings found in natural treatment systems.

The management of suspended solids in wastewater treatment is crucial to prevent process failures, despite not being a primary design constraint Proper distribution of solids within treatment reactors is essential for both aquatic and terrestrial systems Techniques such as inlet diffusers in ponds, step feeding in wetland channels, and high-pressure sprinklers in industrial overland-flow systems enhance solid distribution and mitigate anaerobic conditions In pond systems, suspended solids removal primarily relies on gravity sedimentation, with algae potentially complicating the process Additionally, sedimentation and microbial entrapment play significant roles in wetland and overland-flow processes, while soil matrix filtration is key in SR and SAT systems Overall, removal efficiencies generally surpass secondary treatment levels, although some pond systems may exhibit algal solids in their effluents.

Typical Organic Loading Rates for Natural Treatment Systems

Process Organic Loading (kg/ha/d)

Land application of municipal sludge 27–930 a a These values were determined by dividing the annual rate by 365 days.

Organic Priority Pollutants and CECs

Many organic priority pollutants are highly resistant to biological decomposition, leading to prolonged environmental persistence Some of these pollutants are toxic or hazardous, necessitating specialized management approaches Additionally, constituents of emerging concern (CECs), including pharmaceutical compounds, behave similarly to other trace organics in natural systems Further details on these topics can be found in Chapter 8, which discusses soil aquifer treatment for groundwater recharge, and Chapter 9, which covers biosolids.

Volatilization, adsorption, and biodegradation are key processes for removing trace organics in natural treatment systems Volatilization takes place at the surfaces of ponds, wetlands, and SAT basins, as well as from water droplets in land treatment sprinklers, liquid films in overland-flow systems, and exposed sludge surfaces Adsorption mainly occurs on organic matter within the treatment system that interacts with waste Subsequently, microbial activity often degrades the adsorbed materials, enhancing the overall removal of contaminants.

The evaporation of volatile organic compounds from water surfaces follows first-order kinetics, assuming an atmospheric concentration of zero The fundamental kinetic equation, represented as Equation 3.17, outlines this process, while Equation 3.18 facilitates the calculation of the contaminant's half-life For a deeper understanding of the half-life concept and its relevance to sludge organics, refer to Chapter 9.

C t = concentration at time t (mg/L or g/L)

C 0 = initial concentration at t = 0 (mg/L or g/L) k vol = volatilization mass transfer coefficient (cm/hr) = (k)(y) k = overall rate coefficient (hr −1 ) y = depth of liquid (cm) t y k vol

/ =( )( ) (3.18) where t 1/2 is the time when concentration C t = (1/2)(C 0 ) (hr), and the other terms are as defined previously.

The volatilization mass transfer coefficient is influenced by the contaminant's molecular weight and the air/water partition coefficient, which is defined by Henry’s law constant.

( )( ) / (3.19) where k vol = volatilization coefficient (hr −1 )

H = henry’s law constant (10 5 atmãm 3 ãmol −1 )

M = molecular weight of contaminant of concern (g/mol)

The coefficients B 1 and B 2 are specific to the physical system of concern Dilling

(1977) determined values for a variety of volatile chlorinated hydrocarbons at a well- mixed water surface:

Jenkins et al (1985) experimentally determined values for a number of volatile organics on an overland flow slope:

In the case of overland flow, the coefficients are significantly lower due to the nonturbulent nature of the liquid flow down the slope, which can be nearly classified as laminar flow with a Reynolds number ranging from 100 to 400 Research by Jenkins et al (1985) indicates that the average depth of the flowing liquid on this slope is approximately 1.2 cm.

Parker and Jenkins (1986) utilized a variation of Equation 3.19 to assess volatilization losses from large-droplet wastewater sprinklers operating at low pressure, with the y term representing the average droplet radius Their coefficients are specific to the sprinkler system studied but the methodology is applicable to other sprinklers and pressures Additionally, Equation 3.20 was formulated by Parker and Jenkins for the organic compounds detailed in Table 3.3.

Aeration in pond systems is effective for removing volatile organics from water Clark et al (1984a) formulated Equation 3.21 to calculate the necessary air volume for stripping specific quantities of volatile organics through this process.

S = saturated condition of the compound of concern equal to 0, for unsatu- rated organics, 1, for saturated compounds

M = molecular weight (g/mol) s = solubility of organic compound (mg/L)

The values in Table 3.4 can be used in Equation 3.21 to calculate the air-to-water ratio required for some typical volatile organics.

The primary mechanism for the removal of trace organics in treatment systems is the sorption of these compounds to the organic matter present (USEPA 1982a) The relationship between the concentration of sorbed trace organics and their concentration in solution is quantified by the partition coefficient Kp, which is influenced by the solubility of the chemical This coefficient can be estimated using the octanol-water partition coefficient (Kow) and the percentage of organic carbon in the system, as expressed in the equation: log Koc = (1.00)(log Kow) − 0.21 (3.22).

K oc = sorption coefficient expressed on an organic carbon basis equal to K sorb /O c

K sorb = sorption mass transfer coefficient (cm/hr)

Volatile Organic Removal by Wastewater Sprinkling

Substance Calculated K΄ vol for Equation 3.20 (cm/min)

Source: Parker, L.V and Jenkins, T.F., Water Res., 20(11), 1417–1426, 1986

O c = percentage of organic carbon present in the system

K ow = octanol-water partition coefficient

Hutchins et al (1985) presented other correlations and a detailed discussion of sorption in soil systems.

Jenkins et al (1985) determined that sorption of trace organics on an overland- flow slope could be described with first-order kinetics with the rate constant defined by Equation 3.23:

( )( ) / (3.23) where k sorb = sorption coefficient (hr −1 )

B 3 = coefficient specific to the treatment system, equal to 0.7309 for the overland- flow system studied y = depth of water on the overland-flow slope (1.2 cm)

K ow = octanol–water partition coefficient

B 4 = coefficient specific to the treatment system = 170.8 for the overland-flow system studied

M = molecular weight of the organic chemical (g/mol)

The removal of trace organics often results from a combination of sorption and volatilization processes The overall process rate constant (k sv ) is calculated as the sum of the coefficients defined in Equations 3.19 and 3.23, with the combined removal process described by Equation 3.24.

Properties of Selected Volatile Organics for Equation 3.21

Source: Love, O.T et al., Treatment of Volatile Organic Chemicals in Drinking

Water, EPA 600/8-83-019, U.S Environmental Protection Agency,

Municipal Engineering Research Laboratory, Cincinnati, OH, 1983.

C t = concentration at time t (mg/L or μg/L)

C 0 = initial concentration at t equal to 0 (mg/L or μg/L) k sv = overall rate constant for combined volatilization and sorption equal to k vol + k sorb

Table 3.5 presents the physical characteristics of a number of volatile organics for use in the equations presented above for volatilization and sorption.

In an overland-flow system designed for toluene removal, a 30-meter-long terrace is utilized, with a hydraulic loading rate of 0.4 m³/hr/m The mean residence time on the slope is 90 minutes, and wastewater is applied using a low-pressure, large-droplet sprinkler system The physical characteristics of toluene, including a K_w value of 490, are critical for understanding its behavior in this treatment process.

Physical Characteristics for Selected Organic Chemicals

2,4-Dinitrophenol 34.7 0.001 — 184 a Octanol-water partition coefficient. b Henry’s law constant, 10 5 atm-m 3 /mol at 20°C and 1 atm. c At 25°C. d Molecular weight (g/mol).

H = 515, M = 92; depth of flowing water on the terrace = 1.5 cm; concentration of toluene in applied wastewater = 70 μ g/L.

1 Use Equation 3.20 to estimate volatilization losses during sprinkling:

2 Use Equation 3.19 to determine the volatilization coefficient during flow on the overland-flow terrace: k B y

3 Use Equation 3.23 to determine the sorption coefficient during flow on the overland-flow terrace: k B y

4 The overall rate constant is the sum of k vol and k sorb : k t = 0.0178 + 0.0377 = 0.055

5 Use Equation 3.24 to determine the toluene concentration in the overland- flow runoff:

Land treatment systems are the only extensively studied natural treatment systems for removing priority-pollutant organic chemicals, primarily due to concerns about groundwater contamination Research indicates that more soluble compounds, like chloroform, move through the soil more quickly than less soluble materials, such as certain PCBs, yet the amount escaping these systems in percolate or effluent is minimal Performance data for major land treatment concepts show that in the SR system, removal was observed after 1.5 meters of vertical soil travel using a low-pressure, large-droplet sprinkler In the OF system, removals were measured after flow across a 30-meter terrace with gated pipe application at a hydraulic loading of 0.12 m³/ha/hr, while SAT data were collected from wells approximately 200 meters down-gradient of the application basins.

Removal of Organic Chemicals in Land Treatment Systems

Diethylphthalate — — — 90.75 a Parker and Jenkins (1986). b Jenkins et al (1985). c Love et al (1983). d Not reported.

Table 3.6 illustrates the removal efficiencies of various systems, showing that SR systems handle wastewater concentrations between 2 and 111 μg/L, with percolate levels from 0 to 0.4 μg/L In contrast, the overland-flow system exhibits applied concentrations from 25 to 315 μg/L and effluent concentrations ranging from 0.3 to 16 μg/L Additionally, the SAT system displays applied substance concentrations varying from 3 to 89 μg/L, while percolate concentrations are between 0.1 and 0.9 μg/L.

The SR system demonstrated greater consistency and higher removal rates for trace organics compared to two other treatment concepts, likely due to the use of sprinklers and enhanced sorption opportunities in finer-textured soils Chloroform was the only compound consistently detected in the percolate, albeit at low concentrations While the other two systems were slightly less effective than SR, they still achieved significant removal rates; the inclusion of sprinklers in the OF system could potentially enhance these results further Overall, all three treatment concepts outperformed activated sludge and conventional mechanical systems in trace organic removal Although quantitative relationships for trace organic removal from natural aquatic systems remain undeveloped, volatilization removal in pond and free water surface wetland systems can be estimated, with significant removal rates expected despite longer detention times Organic removal in subsurface flow wetlands may also align with the SAT values presented, depending on the media utilized.

6 and 7 for data on removal of priority pollutants in constructed wetlands.

Wang et al (1999) demonstrated that hybrid poplar trees (H11-11) effectively remove carbon tetrachloride (15 mg/L) from soil by degrading and dechlorinating it, releasing chloride ions into the soil and carbon dioxide into the atmosphere Additionally, research by Lombi et al (2001) highlights the use of Indian mustard and maize for extracting metals from contaminated soils, while Russelle et al (2001) showed that alfalfa can remediate fertilizer spills.

Bankston et al (2002) found that trichloroethylene (TCE) can be effectively reduced in natural wetlands, suggesting that constructed wetlands may yield similar outcomes Additionally, the presence of broad-leaved cattails enhances the mineralization rate of TCE beyond what is achieved by native soil microorganisms.

Pathogens

Pathogenic organisms can be found in wastewaters and sludges, making their control essential for effective waste management Regulatory agencies often impose bacterial limits on discharges to protect surface waters Additional risks include groundwater contamination from treatment systems, potential crop contamination, and infection of grazing animals at land treatment sites, as well as the dispersal of aerosolized organisms from pond aerators or sprinklers Research indicates that natural aquatic, wetland, and land treatment methods are highly effective in controlling pathogens.

Pathogen removal in pond-type systems occurs through natural die-off, predation, sedimentation, and adsorption Parasitic cysts and eggs, such as those from helminths and Ascaris, settle in the quiescent zones of ponds Facultative ponds with three cells and approximately 20 days of detention time, along with aerated ponds featuring a dedicated settling cell before discharge, effectively eliminate helminths and protozoa.

Pond effluents pose minimal risk of parasitic infection, even when utilized in agricultural practices However, some risk may occur during the disposal of sludges To mitigate this, sludges can be treated, and temporary restrictions on public access and agricultural use can be implemented at the disposal site.

The removal of bacteria and viruses in multi-cell pond systems is highly effective for both aerated and unaerated types, as evidenced by the data in Tables 3.7 and 3.8 Notably, the effluent in all instances detailed in Table 3.8 remained undisinfected, focusing on naturally occurring enteric viruses rather than seeded types Seasonal averages indicate that viral concentrations in the effluent were consistently low; however, there was a slight decrease in removal efficiency during the winter months across all three locations For comprehensive details, refer to Bausum (1983).

Numerous studies have shown that the removal of fecal coliforms in ponds depends on detention time and temperature Equation 3.25 can be used to estimate

Enteric Virus Removal in Facultative Ponds

Source: Bausum, H.T., Enteric Virus Removal in Wastewater Treatment Lagoon Systems,

PB83-234914, National Technical Information Service, Springfield, VA, 1983. a PFU/L, plaque-forming units per liter.

Fecal Coliform Removal in Pond Systems

Fecal Coliforms (No./100 mL) Location Number of Cells Detention Time (d) Influent Effluent

Source: USEPA, Design Manual Municipal Wastewater Stabilization Ponds, EPA 625/1-83-015, U.S

Environmental Protection Agency, Center for Environmental Research Information, Cincinnati,

In 1983, research focused on the removal of fecal coliforms in pond systems revealed that the actual detention time, measured through dye studies, can be significantly lower than the theoretical design detention time, often as low as 45% due to flow short-circuiting When dye studies are impractical, it is advisable to conservatively estimate the actual detention time as 50% of the design residence time for calculations.

C f = effluent fecal coli concentration (number/100 mL)

C i = influent fecal coli concentration (number/100 mL) t = actual detention time in the cell (d) k T = temperature-dependent rate constant (d −1 ), equal to ( )( )2 6 1 19 ( T w − 20 )

T w = mean water temperature in pond (°C) n = number of cells in series

To determine the pond's temperature, refer to Chapter 4 Generally, the water temperature can be assumed to approximate the mean monthly air temperature, with a minimum threshold of 2°C.

Equation 3.26 in the form presented assumes that all cells in the system are the same size See Chapter 4 for the general form of the equation when the cells are different sizes The equation can be rearranged and solved to determine the opti- mum number of cells needed for a particular level of pathogen removal In gen- eral, a three-or four-cell (in series) system with an actual detention time of about

In just 20 days, fecal coliform levels can be effectively reduced to desired standards Research on polio and Coxsackie viruses suggests that their removal follows a first-order reaction model, as outlined in Equation 3.26 Additionally, aquatic systems such as hyacinth ponds are expected to operate in alignment with this same first-order reaction principle.

Pathogen removal in wetland systems shares similarities with pond systems, with Equation 3.26 applicable for estimating bacteria or virus removal in surface-flow wetlands Although constructed wetlands typically have shorter detention times than ponds, they offer enhanced opportunities for adsorption and filtration Subsurface-flow wetland systems operate similarly to land treatment systems in pathogen removal A comprehensive study of over 40 constructed wetlands in Colorado, funded by the Colorado Governor’s Office of Energy Management and Conservation in 2000, evaluated the performance and deficiencies of these systems, with a detailed report available through the referenced office.

Land treatment systems in the United States usually involve preliminary treatment or storage ponds, minimizing concerns about parasites Research indicates that infections in grazing animals arise primarily from direct ingestion of raw wastewater or irrigation with it The effective removal of bacteria and viruses in these systems is achieved through various processes, including filtration, desiccation, adsorption, radiation, and predation.

The major concerns relate to the potential for the contamination of surface vegeta- tion or off-site runoff, as the persistence of bacteria or viruses on plant surfaces

Pathogen Removal in Constructed Wetland Systems

Iselin, Pennsylvania (cattails and grasses) c

Source: Reed, S.C et al., in Proceedings AWWA Water Reuse III, American

In the United States, it is advised that agricultural land treatment sites avoid growing raw-consumable vegetables to mitigate infection risks to humans and animals The primary concern lies with grazing animals on pastures irrigated with untreated wastewater, necessitating a grazing delay of 1 to 3 weeks after the application of undisinfected effluent These systems are organized into smaller paddocks, allowing for rotational grazing Proper runoff control is essential in both Subsurface Flow (SR) and Soil Aquifer Treatment (SAT) systems to prevent pathogenic hazards Overland flow systems are designed to manage treated effluent runoff, achieving approximately 90% removal of fecal coliforms, with regulatory agencies determining the necessity for final disinfection of this runoff Additionally, overland flow slopes can capture varying precipitation, and even during intense rainfall, the resulting runoff often maintains or improves water quality compared to standard runoff conditions.

The potential for pathogenic contamination of groundwater aquifers from SR and SAT land treatment necessitates careful consideration Research indicates that finer-textured agricultural soils in SR systems can effectively remove bacteria and viruses A notable 5-year study conducted in Hanover, New Hampshire, showed nearly complete elimination of fecal coliforms within the top 5 feet of the soil profile.

Research from 1979 and a study in Canada (Bell and Bole, 1978) revealed that fecal coliform bacteria were predominantly retained in the upper 8 cm (3 inches) of soil Approximately 90% of these bacteria died within the first 48 hours, with the remaining bacteria eliminated over the following two weeks Additionally, virus removal in these soils is highly effective, primarily relying on the process of adsorption.

The coarse-textured soils and high hydraulic loading rates used in SAT sys- tems increase the risk of bacteria and virus transmission to groundwater aquifers

Extensive research on viral movement in Soil Aquifer Treatment (SAT) systems indicates a minimal risk under typical conditions High viral concentrations can lead to movement only when wastewater is applied at elevated loading rates on very coarse-textured soils, a scenario unlikely to occur frequently Additionally, chlorine disinfection before wastewater application is discouraged, as the resulting chlorinated organic compounds pose a greater risk to groundwater than the potential transmission of bacteria or viruses.

Pathogen levels in raw and digested sludge can be significantly high, posing risks when used in agriculture or when public exposure is a concern The EPA's sludge utilization guidelines, detailed in Chapter 9, address these issues Vermin-stabilization with earthworms may reduce pathogenic bacteria, while the freeze-dewatering process lowers pathogen concentration through enhanced drainage The reed-bed drying method effectively reduces pathogens due to desiccation and extended detention time Additionally, pathogens decrease after land application of sludge, following similar mechanisms as wastewater land application When adhering to the criteria outlined in Chapter 9 for system design, the risk of pathogen transmission to groundwater or surface waters is minimal.

Metals

Trace-level concentrations of metals are present in all wastewaters and sludges, primarily originating from industrial and commercial activities, although private residences can also contribute significantly The most concerning metals include copper, nickel, lead, zinc, and cadmium due to their potential entry into the food chain or water supply During wastewater treatment, a substantial portion of these metals accumulates in the resulting sludges, making metal concentration a critical factor for land application of sludge While metals are typically not a primary design consideration for wastewater treatment or reuse, exceptions exist for specific industries A comparison of metal concentrations in untreated municipal wastewaters and the standards for irrigation and drinking water is provided in Table 3.13.

Metal Concentrations in Wastewater and Requirements for Irrigation and Drinking Water Supplies

Irrigation (mg/L) Continuous b Short-Term c

Source: USEPA, Process Design Manual Land Treatment of Municipal Wastewater, EPA

The U.S Environmental Protection Agency's 1981 report (625/1-81-013) provides median values for typical municipal wastewater It emphasizes that these values apply to waters used indefinitely across various soil types Additionally, for fine-textured soils intended for sensitive crop cultivation, the guidelines are relevant for water usage up to 20 years.

Trace metals generally do not pose significant concerns in the design or performance of pond systems treating typical municipal wastewater, as their removal primarily relies on adsorption to organic matter and precipitation However, the effectiveness of metal removal in pond systems is often lower compared to activated sludge systems, where over 50% of metals can be transferred to the sludge quickly Despite this, pond sludges can accumulate high metal concentrations due to prolonged retention times and infrequent sludge removal Data indicates that metal concentrations in older lagoon sludges align with those found in unstabilized primary sludges, which do not hinder further digestion or land application Notably, sludge metal concentrations may be elevated in lagoons located in warmer climates with substantial industrial wastewater contributions.

Metal Concentrations in Sludges from Treatment Lagoons

Metal Facultative Lagoons a Partial-Mix Aerated Lagoons b

Source: Schneiter, R.W and Middlebrooks, E.J., Cold Region Wastewater Lagoon Sludge: Accumulation, Characterization, and Digestion, Contract Report

The study conducted by the U.S Cold Regions Research and Engineering Laboratory in Hanover, NH, in 1981, analyzed data from two facultative lagoons in Utah and two partial-mix aerated lagoons in Alaska The findings indicate that while the organic content and sludge mass are reduced, the metal concentrations are expected to rise over time.

In sub-tropical climates, utilizing water hyacinths in shallow ponds can effectively facilitate metal removal, as demonstrated by full-scale tests conducted in Louisiana and Florida (Kamber 1982) These studies reveal that the plants significantly uptake metals, with tissue concentrations reaching hundreds to thousands of times greater than those found in surrounding water or sediment, highlighting the bioaccumulation of trace elements Notably, a pilot hyacinth system in central Florida has shown impressive metal removal results, while hyacinths have proven particularly efficient in extracting metals from photoprocessing wastewater in Louisiana (Kamber 1982).

Constructed wetlands have shown remarkable effectiveness in metal removal, as evidenced by tests conducted in southern California These pilot wetlands, with a hydraulic residence time of approximately 5.5 days, achieved removal rates of 99% for copper, 97% for zinc, and 99% for cadmium (Gersberg et al 1985) Notably, plant uptake contributed to less than 1% of the metal removal, with the primary mechanisms being precipitation and adsorption interactions with the organic benthic layer.

Removal of metals in land treatment systems can involve both uptake by any veg- etation and adsorption, ion exchange, precipitation, and complexation in or on the

Metal Removal in Hyacinth Ponds

Metal Influent Concentration Percent Removal a

Source: Kamber, D.M., Benefits and Implementation Potential of

Wastewater Aquaculture, EPA Contract Report 68-01-6232,

U.S Environmental Protection Agency, Office of Water Reg- ulations and Standards, Washington, DC, 1982. a Average of three parallel channels, with a detention time about 5 days. soil As explained in Chapter 9, zinc, copper, and nickel are toxic to vegetation long before they reach a concentration in the plant tissue that would represent a risk to human or animal food chains Cadmium, however, can accumulate in many plants without toxic effects and may represent some health risk As a result, cadmium can be the limiting factor for application of sludge on agricultural land.

The near-surface soil layer in land treatment systems is highly effective for metal removal, with most retained metals located within this zone A study of a rapid-infiltration system in Cape Cod, Massachusetts, which operated for 33 years, revealed that nearly all applied metals were found in the top 50 cm (20 in.) of sandy soil, with over 95% concentrated in the top 15 cm (6 in.) (Reed 1979).

While metal concentrations in typical wastewaters are low, concerns exist about their long-term accumulation in soil, potentially impacting future agricultural viability Research by Hinesly et al (as cited by Reed, 1979) suggests that most metals retained in soil over time are not readily available to plants Current vegetation responds primarily to metals applied during the growing season, with minimal effects from past soil accumulations For instance, after 76 years of raw sewage application in Melbourne, Australia, cadmium levels in grass were only slightly elevated compared to a control site Similarly, newer systems in California show comparable cadmium levels, indicating that vegetation responds mainly to recent metal applications However, significantly higher lead levels in California sites compared to Melbourne are attributed to motor vehicle exhaust from nearby highways.

Metals do not pose a threat to groundwater aquifers, even at the very high hydrau- lic loadings used in rapid-infiltration systems Experience at Hollister, California,

Metal Content of Grasses at Land Treatment Sites

Zinc 50.0 63.0 93.0 161.0 103.0 demonstrates that the concentration of cadmium in the shallow groundwater beneath the site is not significantly different than normal offsite groundwater quality (Pound and Crites 1979) After 33 years of operation at this site, the accumulation of metals in the soil was still below or near the low end of the range normally expected for agricultural soils Had the site been operated in the slow rate mode, it would have taken over 150 years to apply the same volume of wastewater and contained metals.

Nutrients

Effective nutrient management is crucial to prevent negative health and environmental impacts while ensuring the optimal functioning of natural biological treatment systems Key nutrients, particularly nitrogen, phosphorus, and potassium, play a vital role in this balance Nitrogen, in particular, is a critical factor in designing land treatment and sludge application systems, as detailed in Chapters 6, 7, and 9 This section explores the potential for nutrient removal through various treatment concepts and outlines the nutrient needs of different system components.

Nitrogen levels in drinking water are regulated to safeguard infant health, while surface waters may also face restrictions to protect fish populations and prevent eutrophication As outlined in Chapter 8, land treatment systems are designed to comply with the 10-mg/L nitrate-nitrogen standard for any groundwater or percolate exiting the project site Additionally, nitrogen removal may be required before discharging into surface waters, particularly the oxidation or removal of ammonia, which is toxic to fish and contributes significantly to oxygen depletion in streams.

Nitrogen in wastewater exists in various forms due to different oxidation states, with total nitrogen concentrations in municipal wastewaters typically ranging from 15 to over 50 mg/L Approximately 60% of this nitrogen is found in ammonia, which can exist as either molecular ammonia (NH3) or ammonium ions (NH4+) The balance between these forms is significantly influenced by pH and temperature; at pH 7, ammonium ions dominate, while at pH 12, only dissolved ammonia gas is present This pH-dependent relationship is crucial for air-stripping processes in advanced wastewater treatment plants and plays a key role in nitrogen removal in wastewater treatment ponds.

Nitrogen removal in pond systems occurs through various mechanisms, including plant and algal uptake, nitrification, denitrification, adsorption, sludge deposition, and ammonia gas volatilization In facultative wastewater treatment ponds, volatilization is the primary removal method, with up to 80% of total nitrogen lost under optimal conditions The efficiency of nitrogen removal is influenced by factors such as pH, temperature, and detention time At near-neutral pH levels, gaseous ammonia is minimal; however, when some ammonia is volatilized, additional ammonium ions convert to ammonia to maintain equilibrium Despite low conversion rates, the extended detention time in these ponds leads to significant nitrogen removal over time.

Effective nitrogen removal in ponds is crucial for wastewater treatment design, as it often serves as the primary design parameter By reducing nitrogen levels in pond effluent, significant decreases in the land area required for wastewater application can be achieved, leading to substantial cost savings for projects.

Nitrogen removal in hyacinth ponds is primarily achieved through nitrification/denitrification and plant uptake, although the latter does not ensure permanent removal without regular harvesting Complete harvesting is challenging since hyacinth plants also provide shade, limiting algal growth by restricting light penetration Typically, only 20 to 30% of the plants are removed during harvesting, preventing the full nitrogen-removal potential from being realized Nonetheless, nitrification and denitrification can still occur in shallow hyacinth ponds, even with mechanical aeration, due to the presence of aerobic and anaerobic microsites in the plant's dense root zone and the availability of necessary carbon sources Nitrogen removal rates in these ponds can vary significantly, ranging from less than 10 to over 50 kg/ha/d (9 to 45 lb/ac/d), influenced by seasonal changes and harvesting frequency, particularly in well-managed pilot-scale or research facilities.

Nitrogen removal in wetland systems can occur through volatilization of ammonia, denitrification, and plant uptake, particularly when vegetation is harvested (Gersberg et al 1983) Research conducted in Canada (Wile et al 1985) revealed that even with regular harvesting of cattails, only about 10% of the nitrogen was removed by the system These results have been corroborated by additional studies, highlighting that the primary pathway for nitrogen removal is nitrification followed by denitrification.

Nitrogen is usually the limiting design parameter for slow-rate land treatment of wastewater, and the criteria and procedures for nitrogen are presented in Chapter

8 Nitrogen can also limit the annual application rate for many sludge systems, as described in Chapter 9 The removal pathways for both types of systems are similar, and include plant uptake, ammonia volatilization, and nitrification/denitrification Ammonium ions can be adsorbed onto soil particles, thus providing a temporary control; soil microorganisms then nitrify this ammonium, restoring the original adsorptive capacity Nitrate, on the other hand, will not be chemically retained by the soil system Nitrate removal by plant uptake or denitrification can occur only dur- ing the hydraulic residence time of the carrier water in the soil profile The overall capability for nitrogen removal will be improved if the applied nitrogen is ammonia or other less well-oxidized forms Nitrification and denitrification are the major fac- tors for nitrogen removal in rapid-infiltration systems, and crop uptake is a major method for both slow rate and overland flow systems Volatilization and denitrifica- tion also occur with the latter two types of system and may account for from 10 to over 50% of the applied nitrogen, depending on waste characteristics and application methods, as described in Chapter 8 Design procedures based on nitrogen uptake of agricultural and forest vegetation can be found in Chapter 8.

Phosphorus, while not directly linked to health concerns, is a significant contributor to the eutrophication of freshwater systems due to its presence in wastewater as polyphosphates, orthophosphates, and organic phosphorus from various sources, including industrial discharges Natural treatment systems can remove phosphorus through processes such as vegetation uptake, biological processes, adsorption, and precipitation In slow-rate and overland flow land treatment processes, harvested vegetation can remove 20 to 30% of applied phosphorus However, in wetland systems, vegetation is not a major factor in phosphorus removal, as unharvested plants release phosphorus back into the water upon decomposition Additionally, the phosphorus removal capacity of aquatic plants like water hyacinths is limited to their growth needs, typically achieving only 50 to 70% removal, even with optimal management and regular harvesting.

Adsorption and precipitation reactions are key processes for phosphorus removal in wastewater when it interacts with substantial soil volumes, particularly in slow-rate and rapid infiltration systems, as well as certain wetland systems In contrast, overland flow processes face limitations due to the use of relatively impermeable soils Soil reactions are influenced by factors such as clay content, iron and aluminum oxides, calcium compounds, and soil pH Finer-textured soils exhibit the highest phosphorus sorption potential due to their greater clay content and increased hydraulic residence time, while coarse-textured, acidic, or organic soils demonstrate the lowest capacity Although peat soils are acidic and organic, some possess significant sorption potential owing to the presence of iron and aluminum.

A laboratory-scale adsorption test provides an estimate of the phosphorus removal capacity of soil during short application periods, typically showing values that are two to five times lower than actual field retention While the sorption potential of a soil layer can be depleted over time, phosphorus removal remains highly effective until that point Additionally, the adsorption capacity can be restored through chemical precipitation of phosphorus.

Since operations began, the phosphorus adsorption capacity of soils has increased by two to four times Typically, phosphorus concentrations in percolate from slow-rate systems reach background levels of native groundwater within 2 meters of travel in the soil While phosphorus is generally not a significant concern for groundwater quality, its emergence in nearby streams or ponds can lead to eutrophication issues To estimate phosphorus concentration along the infiltration/percolation groundwater flow path, Equation 3.29 can be utilized, which was initially derived from rapid infiltration system responses, providing a conservative estimate applicable to all soil systems (USEPA 1981).

P x = total P at a distance x on the flow path (mg/L)

P 0 = total P in applied wastewater (mg/L) k p = 0.048 at pH 7 (d −1 ) (pH 7 gives the lowest value) t = detention time (d) = (x)(W)/(K x )(G), where: x = distance along flow path (ft; m)

W = saturated soil water content; assume 0.4

K x = hydraulic conductivity of soil in direction on x (ft/d; m/d); thus, K v = vertical and K h = horizontal

G = hydraulic gradient for flow system:

The equation is solved in two steps, first addressing the vertical flow component from the soil surface to any existing subsurface flow barrier, followed by the calculation of lateral flow to adjacent surface water These calculations assume saturated conditions, leading to the minimum possible detention time However, since actual vertical flow is typically unsaturated, the real detention time is likely to be significantly longer than the calculated values If the equation indicates acceptable removal rates, it suggests the site should function reliably, minimizing the need for extensive preliminary tests Nevertheless, detailed testing is essential for the final design of large-scale projects.

Potassium is generally not harmful to health or the environment when found in wastewater; however, it plays a crucial role as a nutrient for plant growth Typically, potassium is not present in wastewater in the ideal ratios with nitrogen and phosphorus In systems that rely on vegetation for nitrogen removal, it may be necessary to add extra potassium to enhance plant nitrogen uptake To determine the required supplemental potassium for both aquatic and land systems with low natural potassium levels, Equation 3.30 can be utilized.

K = annual supplemental potassium needed (kg/ha)

U = estimated annual nitrogen uptake of vegetation (kg/ha)

K ww = amount of potassium in the applied wastewater (kg/ha)

Plants need essential nutrients like magnesium, calcium, and sulfur, and deficiencies can occur based on soil characteristics Micronutrients such as iron, manganese, zinc, boron, copper, molybdenum, and sodium are also crucial for vegetative growth Wastewater typically provides adequate levels of these elements; however, excessive amounts can cause phytotoxicity issues Additionally, high-rate hyacinth systems may necessitate supplemental iron to ensure robust plant growth.

Design of Wastewater Pond Systems

Introduction

Over the past two decades, advancements in pond system design have primarily focused on the incorporation of floating plastic partitions to enhance hydraulic performance, as well as modifications to existing designs that utilize rational multivariate growth kinetics The evolution of more efficient aeration equipment has also contributed to these improvements While various complete-mix and partial-mix combinations have been developed and assessed, the fundamental design principles remain consistent Notable examples of these innovations include the BIOLAC® processes, Rich aerated lagoon design procedures, and the LemTec™ biological treatment process, all of which are explored in this chapter.

A recent conference in Brazil featured a compelling series of lectures on the history, performance, hydraulic characteristics, and biology of pond systems (IWA 2009) While the content is primarily presented from an international perspective, the performance data is valuable for anyone engaged in pond systems Relevant articles from the conference are highlighted in the respective sections of this book.

The IWA Water Science and Technology (WS&T) series includes several books focused on lagoon studies For the latest volumes, visit IWA Publishing, a branch of the International Water Association (IWA), which has been active since 1986.

2007) It is likely that the Brazil conference referenced above will be available as a WS&T volume in the near future.

The significance of hydraulic characteristics in lagoon design has been highlighted in both the 1983 and 2011 EPA lagoon manuals, yet many pond systems built in the last two decades exhibit poor hydraulic traits, suggesting a lack of adherence to these guidelines by some designers Recent advancements in hydraulic design have shown notable improvements, and there is optimism that this positive trend will persist Additionally, valuable insights into hydraulic studies of lagoon systems were shared at the recent conference in Brazil.

For comprehensive insights into the operation and design of lagoon systems, refer to the annual Water Environment Research literature review available in electronic format on CDs from the Water Environment Federation (WEF).

The ongoing trend of minimizing redundancy in lagoon system design has been noted, with an emphasis on the importance of considering future maintenance needs Often, the operating costs of aerated lagoon systems are overlooked when communities evaluate their options While the initial investment in systems lacking redundancy is lower, the long-term environmental impact and maintenance costs for owners can be significantly higher.

Since the publication of the 1983 design manual, numerous design procedures have been proposed and implemented, with several examples highlighted in the 2011 EPA manual While some modifications have achieved significant success, others have seen only moderate results, particularly when applied in varying climates The effectiveness of these systems has often been inconsistent, primarily due to a lack of reliable design information and extensive experimentation by designers There is a pressing need for a comprehensive description of these systems, alongside a comparison to traditional design methods The following sections will outline and describe the major procedures, processes, and design methods, both old and new.

• ASM1, ASM2, and ASM3 models

• High-performance aerated pond systems (Rich design)

• Advanced Integrated Wastewater Pond Systems ® (AIWPS ® ) (Oswald Design)

• BIOLAC process (activated sludge in earthen ponds)

• Nitrogen removal in pond systems (including proprietary systems)

• Modified high-performance aerated pond systems for nitrification and denitrification

• Nitrogen removal in ponds coupled with wetlands and gravel bed nitrifica- tion filters

• Control of algae and design of settling basins

Facultative Ponds

Facultative pond design primarily focuses on biochemical oxygen demand (BOD) removal, with most suspended solids being eliminated in the primary cell of the system The feedback from sludge fermentation significantly impacts the pond's performance by reintroducing organic compounds into the water Seasonal thermal overturn during spring and fall can resuspend substantial amounts of benthic solids Additionally, sludge accumulation rates are influenced by liquid temperature, necessitating extra volume for sludge in colder climates While total suspended solids (TSS) play a crucial role in the effectiveness of pond systems, design equations often simplify their impact by employing an overall reaction rate constant Typically, effluent TSS comprises suspended organism biomass, excluding suspended waste organic matter.

Various empirical and rational models have been developed for the design of ponds, including ideal plug-flow and complete-mix models Notable contributions to this field have been made by researchers such as Fritz et al (1979), Gloyna (1971), Larson (1974), Marais (1970), McGarry and Pescod (1970), Oswald et al (1970), Thirumurthi (1974), Mara (1996), Walter and Bugbee (1974), and Gehring et al (2009), along with Middlebrooks (1987) and Ramani.

In 1976, a summary of various models for evaluating and designing facultative pond systems was presented, highlighting many referenced in previous discussions (Table 4.1) While this list is not exhaustive, most models are adaptations of those previously mentioned Although several models yield satisfactory results, their application may be restricted due to challenges in coefficient evaluation or model complexity The following paragraphs will elaborate on the commonly used methods and equations.

A review of state design standards for facultative ponds since Canter and Englande (1970) reveals that design criteria for organic loading and hydraulic detention times have seen minimal changes over the decades While these criteria are intended to guarantee satisfactory performance, frequent violations of effluent standards by pond systems adhering to state design guidelines highlight their inadequacy For the most current information, it is advisable to contact individual states directly.

The EPA evaluated four facultative pond systems, revealing their organic loading and hydraulic detention times, as summarized in Table 4.2 Notably, the Corinne, Utah system consistently met federal effluent standards for biochemical oxygen demand (BOD5), likely due to its seven-cell design compared to three cells in the other systems This design difference may reduce hydraulic short-circuiting, leading to longer detention times in Corinne Additionally, the placement of inlet and outlet structures can influence detention times Many design issues identified in these systems have been addressed since 1983.

Design Equations Developed for Facultative Ponds

Source: Middlebrooks, E.J., Water Sci Technol., 19, 12, 1987 With permission.

For the design of facultative pond systems, it is recommended to use BOD5 loading rates of 40 to 80 lb/acãd when average winter air temperatures exceed 59°F (15°C), based on extensive experience with various climatic conditions.

(45 to 90 kg/haãd) is recommended When the average winter air temperature ranges between 32°F and 59°F (0°C and 15°C), the organic loading rate should range between

For effective wastewater treatment, organic loading rates should be adjusted based on climate conditions In regions with average winter temperatures below 32°F (0°C), recommended organic loading rates range from 10 to 20 lb/acad (11 to 22 kg/ha) The biochemical oxygen demand (BOD) loading rate in the initial treatment cell is typically capped at 35 lb/acad (40 kg/ha), with a total hydraulic detention time of 120 to 180 days Conversely, in milder climates where temperatures exceed 59°F (15°C), the primary cell can accommodate loadings up to 89 lb/acad (100 kg/ha).

Gloyna (1976) proposed the following empirical equation for the design of faculta- tive wastewater stabilization ponds:

La = ultimate influent BOD or chemical oxygen demand (COD) (mg/L) θ = temperature correction coefficient = 1.085

T = pond temperature (°C) f = algal toxicity factor f′ = sulfide oxygen demand

The BOD5 removal efficiency is projected to be 80 to 90% based on unfiltered influent samples and filtered effluent samples A pond depth of 5 ft (1.5 m) is

Design Equations Developed for Facultative Ponds

Hydraulic residence time (t) is measured in days, while the influent BOD₅ concentration (C₀) is expressed in mg/L, and the effluent BOD₅ concentration (Cₑ) also in mg/L The reaction rate constant (k) varies in units, alongside the maximum reaction rate (μ) for Monod-type kinetics, as established by Monod in 1950 The substrate concentration at half the maximum reaction rate is denoted as Kₛ (0.5 μm) The equation a = 1/(4 + ktD) incorporates the dimensionless dispersion number (D) and the base of natural logarithms (e = 2.7183) Additionally, the decay rate (kₑ) is measured in days⁻¹, while the yield coefficient (Y) represents the mass of total suspended solids (TSS) or volatile suspended solids (VSS) formed per mass of BOD₅ removed Finally, areal BOD₅ removal (AR) is quantified in kg/ha·d.

The article discusses various parameters influencing biochemical oxygen demand (BOD) in aquatic systems, including areal BOD loading measured in kg/ha·d, ultimate influent BOD or chemical oxygen demand (COD) expressed in mg/L, and the temperature coefficient, which is dimensionless It highlights the significance of pond water temperature in °C, the algal toxicity factor, and the sulfide oxygen demand, both of which are also dimensionless Additionally, light intensity is measured in langleys, while pH values are considered crucial for assessing water quality The article further elaborates on reaction orders represented by variables a, b, c, d, and e, along with constants Z and various reaction rate constants (k1, k2) and temperatures (T1, T2).

Table 4.2 presents design and performance data from EPA pond studies across various locations, highlighting organic loading rates and effluent BOD exceedances In Peterborough, New Hampshire, the organic loading was 39.3 kg BOD ha–1 d–1, with actual effluent BOD exceeding 30 mg/L during several months Kilmichael, Mississippi, showed a higher loading of 56.2 kg BOD ha–1 d–1 and significant exceedances in November and July Eudora, Kansas, had an organic loading of 38.1 kg BOD ha–1 d–1, with effluent BOD issues in March, April, and August Corinne, Utah, exhibited varied data, with a primary cell loading of 45.0 kg BOD ha–1 d–1 and seasonal performance fluctuations The design recommendations suggest a pond depth of 3 ft (1 m) and assume an algal toxicity factor of 1.0 for domestic and many industrial wastes Additionally, the design temperature is determined by the coldest month's average pond temperature, while sunlight's impact can be factored into the design equation.

The Gloyna method was assessed using the data presented in Table 4.2, with the optimal equation for data fitting identified as Equation 4.2 Although the data exhibits considerable scatter, the relationship remains statistically significant.

BOD = BOD 5 in the system influent (mg/L)

The Marais and Shaw (1961) equation is based on a complete-mix model and first- order kinetics The basic relationship is shown in Equation 4.3:

C 0 = influent BOD 5 concentration (mg/L) k c = complete-mix first order reaction rate (d −1 ) t n = hydraulic residence time each cell (d) n = number of equal-sized pond cells in series

To prevent anaerobic conditions and unpleasant odors, the recommended maximum concentration of BOD 5 (C e ) in primary cells is set at 55 mg/L Additionally, the allowable depth of the pond (d) in meters is directly related to this maximum concentration (C e ) value.

1 9 8 (4.4) where (C e ) max is the maximum effluent BOD (55 mg/L), and d is the design depth of the pond (in meters).

The influence of water temperature on the reaction rate is estimated using Equation 4.5: k cT = k c35 (1.085) T−35 (4.5) where k cT = reaction rate at water temperature T(d −1 ) k c35 = reaction rate at 35°C = 1.2 (d −1 )

The basic equation for the plug-flow model is:

C 0 = influent BOD5 concentration (mg/L) k p = plug-flow first order reaction rate (d −1 ) t = hydraulic residence time (d)

The reaction rate (k p) was observed to change with varying BOD loading rates, as indicated in Table 4.3 (Neel et al 1961) Although theoretical expectations suggest that the reaction rate should remain constant regardless of loading rate, the reported findings contradict this assumption.

TABLE 4.3 Variation of the Plug-Flow Reaction Rate Constant with Organic Loading Rate

Source: Neel, J.K et al., J Water Pollut Control Fed.,

33, 6, 603–641, 1961 With permission. a Reaction rate constant at 20°C.

The influence of water temperature on the reaction rate constant can be deter- mined with Equation 4.7: k pT = k p20 (1.09) T−20 (4.7) where k pT = reaction rate at temperature T(d −1 ) k p20 = reaction rate at 20°C (d −1 )

Thirumurthi (1974) identified that the flow pattern in facultative ponds lies between ideal plug-flow and complete-mix conditions He suggested utilizing the chemical reactor equation proposed by Wehner and Wilhelm (1956) for effective chemical reactor design.

The influent BOD concentration (C₀) is measured in mg/L, and is influenced by the equation a = (1 + 4ktD)⁰.⁵, where k represents the first-order reaction rate constant in days (d⁻¹), t indicates the hydraulic residence time in days, and D is a dimensionless dispersion number calculated as H/vL = Ht/L² In this context, H refers to the axial dispersion coefficient (area per unit time), v denotes the fluid velocity (length per unit time), and L signifies the travel path length of a typical particle The base of natural logarithms, e, is approximately 2.7183.

Partial-Mix Aerated Ponds

Since the release of the USEPA's 1983 design manual, advancements in the design of partial-mix aerated ponds have primarily included the use of floating plastic partitions to enhance hydraulic performance and the introduction of more efficient aeration equipment While the significance of hydraulic characteristics has been reiterated in both the 1983 and 2011 design manuals, the continued construction of pond systems with inadequate hydraulic features suggests that many designers may not be utilizing the available literature effectively.

The alarming trend of omitting redundancy in aerated lagoon system designs raises concerns about future maintenance needs Often, communities overlook the operating costs associated with these systems when evaluating their options While systems lacking redundancy may have lower initial costs, the long-term environmental impact and maintenance expenses can be significantly higher for owners.

The partial-mix aerated pond system primarily focuses on supplying adequate oxygen without fully suspending all solids, unlike complete-mix and activated sludge systems While some mixing occurs to maintain partial suspension of solids, anaerobic degradation of settled organic matter also takes place This system is often referred to as a facultative aerated pond system.

While partially mixed, ponds are often assessed using a complete-mix model and first-order reaction kinetics for BOD removal Research by Middlebrooks et al (1982) indicates that a plug-flow model with first-order kinetics provides a more accurate performance prediction for ponds utilizing surface or diffused aeration However, the study primarily involved lightly loaded ponds, leading to conservative reaction rates, as these rates tend to decrease with lower organic loading (Neel et al 1961) Despite the need for improved design reaction rates, partial-mix ponds continue to be designed using complete-mix kinetics.

The design model for first-order kinetics, utilizing multiple equal-sized cells arranged in series, is represented by Equation 4.11 as outlined by Middlebrooks et al (1982), the Great Lakes–Upper Mississippi River Board of State Sanitary Engineers (1990), and WEF/ASCE (1991).

C n = effluent BOD concentration in cel n (mg/L)

The influent BOD concentration (C0) is measured in mg/L, while the first-order reaction rate constant (k) is 0.276 d−1 at 20°C, assumed constant across all cells The total hydraulic residence time in the pond system is denoted as t (in days), and n represents the number of cells in the series.

To utilize varying reaction rates instead of a series of equal volume ponds, it is essential to apply the general equation outlined for this purpose.

(4.12) where k 1 , k 2 , …, k n are the reaction rates in cells 1 through n (all usually assumed equal for lack of better information) and t 1 , t 2 , …, t n are the hydraulic residence times in the respective cells.

Mara (1975) demonstrated that using a series of equal volume reactors is more efficient than employing reactors of unequal volumes However, site topography and other factors may necessitate the construction of reactors with varying volumes in certain situations.

4.3.1.1 Selection of Reaction Rate Constants

Selecting the appropriate k value is crucial for designing an effective pond system The Ten-States Standards recommend a k value of 0.276 d–1 at 20°C and 0.138 d–1 at 1°C, leading to a temperature coefficient of 1.036 Boulier and Atchinson (1975) suggest k values ranging from 0.2 to 0.3 at 20°C and 0.1 to 0.15 at 0.5°C, which also yield a temperature coefficient of 1.036 Additionally, Reid (1970) proposed a k value of 0.28 at 20°C and 0.14 at 0.5°C based on studies of partial-mix ponds in central Alaska, aligning closely with the Ten-States Standards' recommendations.

4.3.1.2 Influence of Number of Cells

The partial-mix design model reveals that the number of cells arranged in series significantly influences the size of the pond system needed to meet treatment requirements This relationship can be illustrated by rearranging Equation 4.11 to solve for t.

(4.13) All terms in this equation have been defined previously.

Compare detention times for the same BOD removal levels in partial-mix aerated ponds having one to five cells Assume C 0 = 200 mg/L, k = 0.28 d –1 , and T w = 20°C.

1 Solve Equation 4.13 for a single-cell system: t n k

2 Similarly, when: n = 2, then t = 11 d. n = 3, then t = 9.4 d. n = 4, then t = 8.7 d. n = 5, then t = 8.2 d.

3 Continuing to increase n will result in the detention time being equal to the detention time in a plug-flow reactor It can be seen from the tabulation above that the advantages diminish after the third or fourth cell.

The influence of temperature on the reaction rate is defined by Equation 4.14: k T =k 20θ T w − 20 (4.14) where k T = reaction rate at temperature T(d −1 ) k 20 = reaction rate at 20°C (d −1 ) θ = temperature coefficient = 1.036

The pond water temperature (T w ) can be estimated using the following equation developed by Mancini and Barnhart (1976):

A = surface area of pond (m 2 ) f = proportionality factor = 0.5

The surface area is estimated using a temperature-corrected Equation 4.13, followed by temperature calculation with Equation 4.15 Through multiple iterations, the water temperature used to adjust the reaction rate coefficient is aligned with the value derived from Equation 4.15, finalizing the selection of the system's detention time.

For optimal pond design based on complete-mix hydraulics, circular or square shapes are ideal In contrast, partial-mix ponds, while modeled on complete-mix principles, should have a length-to-width ratio of 3:1 or 4:1 to better emulate the plug-flow condition The dimensions of these cells can be determined using Equation 4.16.

V=[LW+(L−2sd W)( −2sd)+4(L sd W sd− )( − )]d

V = volume of pond or cell (m 3 )

L = length of pond or cell at water surface (m)

W = width of pond or cell at water surface (m) s = slope factor (e.g., for 3:1 slope, s = 3) d = depth of pond (m)

Oxygen requirements play a crucial role in determining the power input needed for partial-mix pond systems Various rational equations can be utilized to estimate these oxygen needs for pond systems, as highlighted by studies from Benefield and Randall (1980), Gloyna (1971, 1976), and Metcalf and Eddy.

In designing partial-mix systems, the biological oxygen demand (BOD) is crucial for estimating oxygen requirements Once the necessary oxygen transfer rate is established, manufacturers' catalogs can help identify effective aeration methods, such as surface, helical, or air gun aerators, and the appropriate spacing for perforated tubing Various aerator types used in pond systems are illustrated in Figure 4.2, while Figure 4.3 provides photographs of both the aeration equipment depicted and additional installed systems To estimate oxygen transfer rates, Equation 4.17 is utilized.

N = equivalent oxygen transfer to tapwater at standard conditions (kg/hr)

N a = oxygen required to treat the wastewater (kg/hr) (usually taken as 1.5 × the organic loading entering the cell) α = (Oxygen transfer in wastewater)/(oxygen transfer on tapwater) = 0.9

Complete-Mix Aerated Pond Systems

Complete-mix pond systems come in various configurations, but they generally share a similar design This article discusses several types that employ the complete-mix concept, including high-performance aerated pond systems, nitrogen removal techniques, modified aerated lagoon systems for nitrification and denitrification, the BIOLAC process, and Lemna systems Most designs of complete-mix systems are based on specific equations with slight adjustments Notably, there has been a concerning trend towards reducing redundancy in the design of aerated lagoon systems.

Many communities overlook the long-term maintenance costs associated with aerated lagoon systems when evaluating their options While systems without redundancy may present lower initial costs, the eventual expenses and environmental impact due to maintenance can be significantly higher for both the owner and the ecosystem Prioritizing flexibility in operation is essential to mitigate future costs and environmental risks.

An analysis of Example 4.5 highlights the similarities between the high-performance aerated pond system design and the complete-mix design, particularly when the last three cells of the complete-mix design are provided just enough dissolved oxygen to meet the biochemical oxygen demand (BOD) While this comparison does not suggest that the two designs are identical, it does illustrate the parallels in their design approaches.

The design model that employs first-order kinetics with multiple equal-sized cells arranged in series is represented by Equation 4.18, as referenced by Middlebrooks et al (1982), the Great Lakes–Upper Mississippi River Board of State Sanitary Engineers (1990), and WEF/ASCE (1991).

C n = effluent BOD concentration in cell n (mg/L)

The influent BOD concentration (C₀) is measured in mg/L, while the first-order reaction rate constant (k) is set at 2.5 d⁻¹ at 20°C, assumed to remain constant across all cells The total hydraulic residence time in the pond system is denoted as (t) in days, and the number of cells in the series is represented by (n).

To utilize ponds of unequal volume or to achieve varying reaction rates, it is essential to apply the general equation outlined below.

 (4.19) where k 1 , k 2 , …, k n are the reaction rates in cells 1 through n (all usually assumed equal for lack of better information) and t 1 , t 2 , …, t n are the hydraulic residence times in the respective cells.

Mara (1975) demonstrated that using a series of equal-volume reactors is more efficient than employing reactors of unequal volumes However, certain circumstances, such as site topography or other influencing factors, may necessitate the construction of reactors with unequal volumes.

4.4.1.1 Selection of Reaction Rate Constants

Selecting the appropriate k value is crucial in pond system design, with a recommended value of 2.5 d–1 at 20°C (68°F) established by Reynolds and Middlebrooks (1990) based on research from a complete-mix aerated lagoon in Colorado While some experts suggest higher k values, designs utilizing the 2.5 d–1 rate have proven effective under full design loads, unlike those with higher rates, which often underperform Designers frequently face limitations imposed by state standards, which may require reaction rates that surpass the recommended 2.5 d–1.

4.4.1.2 Influence of Number of Cells

The complete-mix design model shows that the number of cells in a series significantly influences the size of the pond system needed for effective treatment By rearranging Equation 4.18 and solving for time (t), we can illustrate this impact clearly.

All terms in this equation have been defined previously.

Compare detention times for the same BOD removal levels in complete-mix aer- ated ponds having one to five cells Assume C 0 = 200 mg/L, k = 2.5 d −1 , T w = 20°C.

1 Solve Equation 4.20 for a single-cell system: t n k

2 Similarly, when: n = 2, then t = 1.03 d. n = 3, then t = 0.75 d. n = 4, then t = 0.64 d. n = 5, then t = 0.58 d.

3 Continuing to increase n will result in the detention time being equal to the detention time in a plug-flow reactor It can be seen from the tabulation above that the advantages diminish after the third or fourth cell.

The influence of temperature on the reaction rate is defined by Equation 4.21: k T =k 20 T w − θ 20 (4.21) where k T = reaction rate at temperature T(d −1 ) k 20 = reaction rate at 20°C (d −1 ) θ = temperature coefficient = 1.036

The pond water temperature (T w ) can be estimated using the following equation developed by Mancini and Barnhart (1976):

A = surface area of pond (m 2 ) f = proportionality factor = 0.5

The surface area is estimated using Equation 4.20 and adjusted for temperature, followed by temperature calculation with Equation 4.22 After multiple iterations, the process concludes when the water temperature used to correct the reaction rate coefficient aligns with the value derived from Equation 4.22, finalizing the selection of detention time within the system.

The optimal design for a pond based on complete-mix hydraulics is either circular or square However, it is advisable to configure the cells with a length-to-width ratio of 3:1 or 4:1 This configuration is recommended because the hydraulic flow pattern in complete-mix systems more closely aligns with plug-flow conditions The dimensions of the cells can be determined using Equation 4.23.

V=[LW+(L−2sd W)( −2sd)+4(L sd W sd− )( − )]d

V = volume of pond or cell (m 3 )

L = length of pond or cell at water surface (m)

W = width of pond or cell at water surface (m) s = slope factor (e.g., for a 3:1 slope, s = 3) d = depth of pond (m)

Mixing requirements play a crucial role in determining the power input needed for complete-mix pond systems To estimate the oxygen needs for these systems, various rational equations have been developed (Benefield and Randall 1980; Gloyna 1971, 1976; Metcalf and Eddy 1991, 2003) The design of complete-mix systems begins with an estimation of the incoming BOD to assess biological oxygen demands, ensuring that sufficient power is available for thorough mixing After calculating the necessary oxygen transfer rate, manufacturers' catalogs should be consulted to identify the appropriate aeration methods, such as surface, helical, aeration chain, or air gun aerators, and the optimal spacing for perforated tubing Additionally, schematic sketches and photographs of various aerators used in pond systems are provided in the accompanying figures.

Equation 4.24 is used to estimate the oxygen transfer rates:

N = equivalent oxygen transfer to tapwater at standard conditions (kg/hr)

N a = oxygen required to treat the wastewater (kg/hr) (usually taken as 1.5 × the organic loading entering the cell) α = (Oxygen transfer in wastewater)/(oxygen transfer on tapwater) = 0.9.

The equation C sw = β(C ss) P represents the oxygen saturation value of waste in mg/L, where β is calculated by dividing the waste-water saturation value by the tap water oxygen saturation value The variable C ss denotes the tap water oxygen saturation value at a specific temperature (T w), while P indicates the ratio of barometric pressure at the pond site to sea level pressure, assuming a value of 1.0 for elevations up to 100 meters.

C L = minimum DO concentration to be maintained in the wastewater assume

C s = oxygen saturation value of tapwater at 20°C and one atmosphere pressure 9.17 (mg/L)

Equation 4.22 can be used to estimate the water temperature in the pond during the summer months that will be the critical period for design for biological activ- ity; however, with power to provide complete mixing, adequate dissolved oxygen is normally readily available The complete-mix design procedure is illustrated in Example 4.5 The four-cell system can be obtained by using floating plastic parti- tions such as those shown in Figure 4.4, and the aeration equipment can be selected from the types shown in Figures 4.2 and 4.3.

ASM1, ASM2, and ASM3 Models

The ASM1, ASM2, and ASM3 models are essential tools for designing activated sludge systems, utilizing a matrix format to connect process reactions and stoichiometric factors Despite their complexity, these activated sludge models have been effectively applied in small pilot facultative and maturation ponds, although their use in aerated lagoon systems has been limited due to insufficient operational data for evaluating kinetics.

A good description of the models along with a design example for an activated sludge system is presented in the EPA nitrogen control manual starting on page 117 (USEPA

The ASM2 model, described in detail by Henze et al (1995) and Barker and Dold (1997), is a more complex approach compared to its predecessors While efforts are underway to implement these models in aerated lagoon systems, the results have been limited thus far.

Anaerobic Ponds

Anaerobic lagoons are effective for treating municipal, agricultural, and industrial wastewaters by stabilizing high concentrations of organic solids rather than focusing on producing high-quality effluent Typically, these lagoons operate in conjunction with aerated or facultative lagoons, and a well-designed three-cell lagoon system can consistently yield stable, high-quality effluent throughout its lifespan To achieve optimal performance, it is crucial to consider the biological reactions involved in stabilizing organic waste during the design and operation of anaerobic lagoons.

In anaerobic conditions, extracellular enzymes hydrolyze insoluble organics into soluble carbohydrates like glucose, cellobiose, and xylose These soluble carbohydrates are then biologically converted into volatile organic acids, primarily acetic, propionic, and butyric acids The facultative organisms responsible for this transformation are referred to as acid formers or acid producers Subsequently, strict anaerobic bacteria, known as methane formers or methane producers, convert these organic acids into methane and carbon dioxide.

Anaerobic decomposition of carbohydrate to bacterial cells with the formation of organic acids can be illustrated as:

5 (CH2O)x→ (CH2O)x + 2CH3COOH + Energy Bicarbonate buffer present in solution neutralizes the acid formed in the above reaction: 2CH3COOH + 2NH4HCO3→ 2CH3COONH4 + 2H2O + 2CO2

During the growth of methane bacteria, ammonium acetate (CH3COONH4) is decom- posed to methane and regeneration of the bicarbonate buffer, NH4HCO3:

2CH 3 COONH 4 + 2H 2 O → 2CH 4 + 2NH 4 HCO 3

If sufficient buffer is not available, the pH will decrease, which will inhibit the third reaction.

Facultative acid formers exhibit resilience to environmental factors like pH, heavy metals, and sulfides, making them abundant in anaerobic digestion systems In contrast, the methane fermentation process is the rate-limiting step, as methane-producing bacteria are highly sensitive to variations in pH, heavy metals, detergents, alkalinity, ammonia nitrogen levels, temperature, and sulfides Additionally, these bacteria have a slow growth rate, further impacting the efficiency of methane production.

Environmental factors that affect methane fermentation are shown in Table 4.7

In addition, work by Kotze et al (1968), Chan and Pearson (1970), Hobson et al

Research by Ghosh et al (1974) indicates that the hydrolysis step can be a rate-limiting factor in the digestion of particulate and cellulosic feeds Effective design and operation of anaerobic lagoons should be based on essential biochemical and kinetic principles However, it is noteworthy that many anaerobic lagoons have been designed empirically rather than through a scientific approach.

Anaerobic lagoons often face significant odor issues, primarily due to the release of volatile organic compounds To mitigate these odors, creating an aerobic zone at the surface can effectively oxidize these compounds Additionally, recirculating effluent from an aerobic pond to the primary anaerobic pond introduces dissolved oxygen, which helps to oxidize sulfides and further reduce odor emissions.

Environmental Factors Influencing Methane Fermentation

Volatile acids (mg/L as acetic) 50–500 2000

To effectively manage wastewater and minimize odors, influent wastewater should be directed into a central anaerobic pond chamber where sludge can accumulate This design, as outlined by Oswald (1968), prevents oxygen contact with anaerobic processes Additionally, mixing the influent with active anaerobic sludge significantly enhances biochemical oxygen demand (BOD) removal efficiency, further reducing unpleasant odors (Parker et al 1959).

Anaerobic lagoons primarily serve to decompose and stabilize organic matter rather than purify water These lagoons act as sedimentation basins, effectively reducing organic loads before they reach subsequent treatment units This article provides a comprehensive overview of the design considerations for municipal anaerobic lagoons.

The design of anaerobic stabilization ponds lacks a universally accepted approach, often relying on surface loading rates, volumetric loading rates, and hydraulic detention times However, designing based on surface loading rates can lead to inaccuracies; therefore, a proper design should focus on volumetric loading rates, liquid temperature, and hydraulic detention time Global areal loading rates, as detailed in Table 4.8, reveal significant inconsistencies in anaerobic pond design To estimate volumetric loading rates, one can divide the areal loading rates by the average pond depth and convert to appropriate units In warmer climates with temperatures exceeding 22°C, adhering to specific design criteria can achieve a BOD5 removal efficiency of 50% or higher, as recommended by WHO in 1987.

The method for constructing a digestion chamber at the base of an anaerobic lagoon is illustrated in Figure 4.5 This technique is detailed in the work of Oswald, W.J., found in "Advances in Water Quality Improvement," edited by Gloyna, E.F and Eckenfelder, W.W., Jr., published by the University of Texas Press in Austin, TX, 1968.

Anaerobic lagoons are effective for treating municipal wastewater, with varying design and operational parameters The BOD5 mass loading rates range from 0.66 to 6.89 lb/ac·d, while the estimated volumetric loading rates vary between 0.51 and 6.89 lb/1000 ft³·d BOD5 removal efficiency can reach up to 89% in summer conditions, dropping to 60% in winter Depths typically range from 3 to 12 feet, with hydration detention times of 3 to 10 days, depending on the specific study referenced Notable research includes works by Parker (1970), Oswald (1968), and Cooper (1968), highlighting the importance of seasonal variations in treatment efficiency and operational design.

• Volumetric loading up to 300 g BOD 5 per m 3 per d

• Hydraulic detention time of approximately 5 d

In cold climates, achieving a 50% reduction in BOD5 may necessitate detention times of up to 50 days and volumetric loading rates as low as 40 g BOD5 per m³ per day The interplay between temperature, detention time, and BOD reduction is illustrated in Tables 4.9 and 4.10.

Oswald (1996) proposed an innovative design for anaerobic lagoons, featuring a deep anaerobic pond integrated within a facultative pond This advanced design optimizes organic loading rates, which are influenced by the water temperature in the pond.

Five-Day BOD 5 Reduction as a Function of

Temperature Detention Time BOD Reduction

Source: WHO, Wastewater Stabilization Ponds: Principles of

Planning and Practice, WHO Tech Publ 10, Regional Office for the Eastern Mediterranean, World Health Organization, Alexandria, 1987.

TABLE 4.9 Five-Day BOD 5 Reduction as a Function of Detention Time for Temperatures Greater Than 20°C

Source: WHO, Wastewater Stabilization Ponds: Principles of Planning and Practice, WHO Tech Publ 10,

In 1987, the World Health Organization's Regional Office for the Eastern Mediterranean in Alexandria outlined a design method for separate anaerobic digesters by assessing the per capita volume of anaerobic ponds This approach is exemplified in Example 4.6.

Maximum bottom water temperature in local bodies of water = 20°C.

Temperature of pond water at bottom of pond = 10°C.

BOD loading = Influent BOD × flow rate/1000 = 378.8 kg/d.

2 Design the anaerobic pond (fermentation pits) Except for systems with flows less than 200 m 3 /day, always use two ponds, so one will be available for desludging when the pond is filled The surface area of the anaerobic pond should be limited to 1000 m 2 , and it should be made as deep as pos- sible to avoid turnover with oxygen intrusion Minimum pit depth should be

Number of anaerobic ponds in parallel = minimum of two ponds = 2. BOD loading on single pond = 189.4 kg/d.

First, size pond on basis of load per unit volume:

Load per unit volume (varies with temperature of water) = 0.189 kg/m 3 /d. Volume in one pond = 1002.7 m 3

Hydraulic residence time in ponds = 2.12 d.

Pond surface area (assuming vertical walls) = 250.7 m 2

Maximum pond surface area = 1000 m 2 ; number of ponds = 0.25 (round to next largest number of ponds = 1.00).

The overflow rate in ponds is calculated by dividing the total surface area by the total flow rate, resulting in a rate of 1.89 m/d To effectively retain parasite eggs and particles as small as 20 μm, the overflow rate should be reduced to below 1.5 m/d Therefore, increasing the size of the pond is necessary to achieve this optimal overflow rate.

Check pond volume per capita:

Controlled Discharge Pond System

Complete Retention Pond System

Hydrograph Controlled Release

High-Performance Aerated Pond Systems (Rich Design)

The high-performance aerated pond system (HPAPS), also known as the dual-power multicellular (DPMC) system, consists of two aerated basins arranged in series, as described by Rich (1999) It features screens for the removal of large solids, followed by a reactor basin for bioconversion and flocculation, and a settling basin for sedimentation, solids stabilization, and sludge storage Algae growth is managed through limited hydraulic retention time and the division of the settling basin into sequential cells Finally, disinfection facilities are incorporated after the settling basin.

The dual-power, multicellular (DPMC) aerated lagoon system features a flow diagram that illustrates two configurations: (a) two basins arranged in series with floating baffles in the settling cells, and (b) a single basin employing floating baffles to segregate different unit processes This design enhances the efficiency of wastewater treatment by optimizing aeration and settling processes.

Aerated Lagoon Systems, American Academy of Environmental Engineers, Annapolis, MD,

Aeration plays a crucial role in wastewater treatment, occurring in both the reactor and the settling basin In the reactor, aeration is maintained at approximately 6 W/m³ to ensure solids remain suspended, with a minimum hydraulic detention time of 1.5 days For smaller systems, the reactor and settling basin can be housed within the same earthen basin, while larger systems benefit from having a separate reactor basin This separation facilitates system modifications for future upgrades, including nitrification and denitrification processes.

Reactor basins are typically designed based on Monod kinetics, ensuring a minimum hydraulic retention time of 1.5 days Rich (1999) advises against incorporating a safety factor in reactor design, as the settling basin offers sufficient retention time to offset any potential miscalculations regarding the necessary duration in the reactor basin.

Aeration in the settling basin should be limited to 1.8 W/m³ and evenly distributed among cells with floating plastic dividers This process is crucial for maintaining an aerobic water column and a top aerobic layer over the sludge, which helps prevent the release of reduced compounds, eliminates odors, and minimizes the resuspension of bottom solids Additionally, aeration promotes mixing that reduces dead zones where algae can thrive, while also allowing respiratory carbon dioxide to escape into the atmosphere during the night, preventing its use by algae when light is present It is essential that aeration is sufficient to enable the settling of solids.

Aerated lagoon systems may face challenges when treating wastewater with carbonaceous biological oxygen demand (CBOD5) concentrations below 100 mg/L, as this results in minimal settleable solids, especially in presettled wastewater It is advisable to avoid using High-Performance Aerated Pond Systems (HPAPS) at schools and seasonal recreational areas due to typically insufficient lagoon volumes and inadequate depth The 3:1 side slopes often prevent the use of commercially available aerators in settling basins, coupled with intermittent flow that leads to prolonged hydraulic retention times and excessive algae growth Design procedures for the HPAPS system are outlined by Rich (1999).

Performance data for HPAPS systems are primarily available from mild climates like South Carolina and Georgia, suggesting a need for additional data from regions with harsher climates to inform design considerations Figure 4.7 illustrates performance data for the DPMC system in Berkeley County, South Carolina, covering six years of operation Additionally, Rich (2000) provided three more years of similar data online, indicating the system has operated effectively for over nine years While performance has been outstanding during this time, information on sludge removal remains unavailable.

Continuous operation of the aeration system is essential to obtain maximum effi- ciency, as illustrated by Figures 4.7 and 4.8 The performance data for Berkeley

Jul-92 Oct -92 Ja n-93 Ap r-93 Jul-93

Concentration (mg/L) Oct -93 Ja n-94 Ap r-94

Jul-95 Oct -95 Ja n-96 Ap r-96 Jul-96

The dual-power, multicellular (DPMC) aerated lagoon system in Berkeley County, South Carolina, demonstrates continuous operation of aerators, showcasing its effective performance This system, as documented by Rich L.G in "High-Performance Aerated Lagoon Systems," highlights advancements in aeration technology for wastewater treatment.

FIGURE 4.8 Effluent TSS and BOD 5 from a dual-power, multicellular (DPMC) aerated lagoon system with aerators operating intermittently (From Rich, L.G., High-Performance

Aerated Lagoon Systems, American Academy of Environmental Engineers, Annapolis, MD,

In South Carolina, performance data revealed that systems utilizing continuous aeration significantly outperformed those with intermittent aeration, showing an improvement of approximately 50%.

The DPMC system at the Ocean Drive plant in North Myrtle Beach, South Carolina, has successfully operated for over 12 years, maintaining final effluent TSS concentrations below 15 mg/L, as illustrated in Figure 4.9 A flow diagram of the system is depicted in Figure 4.10, and data from a 1997 USEPA evaluation of the aerated lagoons is summarized in Table 4.11 Additionally, a comparable facility, the Crescent Beach plant in Myrtle Beach, has also demonstrated strong performance, as shown in Figure 4.11 Proper design and operation of DPMC systems lead to excellent results.

FIGURE 4.10 Sketch of a dual-power, multicellular (DPMC) aerated lagoon–intermittent sand filter system at North Myrtle Beach, South Carolina (From unpublished paper by Rich, Bowden, and Henry 1998.)

Date Effluent TSS and B OD 5 (mg/L)

FIGURE 4.9 Monthly average BOD 5 and TSS from Ocean Drive plant (From Rich, L.G., High-Performance Aerated Lagoon Systems, American Academy of Environmental Engineers, Annapolis, MD, 1999 With permission.)

Proprietary Systems

4.11.1 a dvanced i ntegrated W asteWater p ond s ystems

AIWPS, developed over 60 years by Dr William J Oswald at the University of California, Berkeley, has primarily focused on moderate climate regions Its key benefit is significantly reducing or eliminating the need for sludge disposal, as demonstrated by the facility in St Helena, California.

Performance of Dual-Power, Multicellular (DPMC) Aerated Lagoon at North Myrtle Beach, South Carolina

Source: Unpublished paper by Rich, Bowden and Henry 1998.

Date Effluent TSS and B OD 5 (mg/L)

FIGURE 4.11 Monthly average effluent BOD 5 and TSS at Crescent Beach Plant, North Myrtle Beach, South Carolina (From Rich, L.G., High-Performance Aerated Lagoon

The American Academy of Environmental Engineers highlights that certain wastewater treatment facilities have not needed to dispose of primary sludge for over 30 years, particularly in moderate climates Early evidence suggests that deep sludge pits in colder climates can significantly reduce solids These systems also benefit from the sludge blanket's ability to adsorb toxic materials, with toxicity tests indicating that AIWPS can produce compliant effluent from municipal and industrial wastewaters Furthermore, the construction and operational costs of AIWPS are considerably lower than those of conventional wastewater treatment methods For instance, the Hollister, California system, which has been operational for over 25 years, was constructed for about one-third the cost and operates for one-fifth the cost of a nearby mechanical plant Detailed information and operational data can be found in various studies and reports by Oswald and others.

The wastewater treatment facility in Hotchkiss, Colorado, operates similarly to an AIWPS system, albeit without the high-rate raceway component Situated approximately 60 miles east of Grand Junction at an elevation of 5,300 feet, this non-proprietary facility serves a local population effectively.

The facility, which accommodates 800 people, became operational in October 1997 It experiences an annual mean temperature of approximately 50°F (10°C), with winter lows reaching –20°F (–29°C) and summer highs ranging from 90 to 100°F (32 to 38°C) Additionally, the area receives an average annual precipitation of 13 inches.

The treatment facility utilizes approximately 33 cm of water, primarily sourced from spring snows, late summer rains, and fall storms Key characteristics of the facility are detailed in Table 4.12, while the effluent limits mandated by the State of Colorado are summarized in Table 4.13 Additionally, a flow diagram illustrating the treatment facility's layout is provided in Figure 4.12 Notably, this system differs from conventional AIWPS by featuring a separate anaerobic pond that precedes the aerated ponds, rather than incorporating it within the first facultative pond.

From October 1997 to 2000, the performance of the Hotchkiss wastewater treatment system demonstrated consistent functionality, with flow rates averaging 35% of the design flow of 0.494 mgd Despite occasional peaks exceeding 0.3 mgd, the system achieved impressive removal rates: BOD5 at 92.9%, TSS at 88.7%, and NH3–N at 79.8% Notably, NH3–N removal improved significantly with plant maturation, resulting in an average effluent concentration of 1.78 mg/L in 1999, with values ranging from 5.8 mg/L in February to 0.43 mg/L in June Furthermore, NH3–N removal efficiency showed a strong correlation with effluent water temperature.

Dove Creek, situated about 30 miles north of Cortez, Colorado, at an elevation of 6,750 feet, experiences air temperatures ranging from 0°F to over 90°F The community's wastewater treatment plant is designed to serve around 700 residents, with an average flow rate of 60,000 gallons per day.

(227 m 3 /d) The system is permitted for a design flow of 0.115 mgd (435 m 3 /d) and

The facility processes 288 lb (131 kg/d) of BOD 5 daily, featuring an anaerobic pond before the aerated cells, similar to the Hotchkiss, Colorado facility Following the aerated cells, a free water surface wetland is included in the system The fermentation pit, illustrated in Figures 4.18 and 4.19, has a total volume of 31,947 ft³ (239,123 gallons or 905 m³).

4.11.2 BioLac p rocess (a ctivated s Ludge in e artHen p onds )

In 1990, the U.S Environmental Protection Agency (USEPA) released a comprehensive summary of the BIOLAC processes in the United States, which has been updated with pertinent information from the Parkson Corporation Since then, over 600 BIOLAC systems have been installed globally Given that much of the data in the original USEPA report was based on a limited operating period for many plants, it is essential to re-evaluate the database of previously sampled installations and incorporate data from facilities built after the 1990 report.

Description of Hotchkiss, Colorado, Wastewater Treatment Facility

Unit Process Unit Process Features/Description

Anaerobic portion Volume = 1.75 MG, depth = 18.5–21.5 ft, t = 3.5 d 315 lb BOD 5 d –1 Aerobic portion Volume = 2.9 MG, depth = 13 ft, t = 5.9 d

Aeration 2- to 5-hp and 1- to 10-hp surface aerators, FTR =

1.40 lb O 2 hp –1 hr –1 Lagoon #2 Volume = 5.0 MG, depth = 13 ft, t = 10.0 d

Aeration 1- to 5-hp and 1- to 10-hp surface aerators, 1- to 5-hp aspirating aerator, FTR = 1.46 lb O 2 hp –1 hr –1 Polishing pond Volume = 1.74 MG, depth = 12 ft, t = 3.5 d 0.494 mgd Recirculation 0.5-hp pump rated at 100 gpm

Chlorination Two 150-lb gas cylinders, 0 to 4 lb/d and 0 to 10 lb/d; regulators, 2.5 mg/L maximum dosage Chlorine contact chamber

Serpentine basin: length = 190 ft, width = 3.5 ft, 54:1 length-to-width ratio; volume = 34,800 gal; t = 30 min

Irrigation pumping A pump (of undetermined size) will pump a portion of the effluent to an irrigation ditch supplying 70 acres of farmland

Dechlorination SO 2 gas, same equipment as gas chlorination equipment 0.494 mgd

The BIOLAC processes encompass several variations, primarily focusing on extended aeration activated sludge methods, both with and without solids recirculation The three fundamental systems include BIOLAC-R, an extended aeration process featuring solids recycling; BIOLAC-L, an aerated lagoon system without solids recycling; and the BIOLAC Wave-Oxidation © modification, designed for nitrifying and denitrifying wastewater Additionally, floating aeration chains have been implemented in existing lagoon systems to enhance their performance.

BOD 5 (mg/L) 30/45 b State effluent regulations

TSS (mg/L) 75/110 b State effluent regulations

Fecal coliform (number/100 mL) 6000/12,000 c State fecal coliform policy pH (minimum–maximum) 6.0–9.0 State effluent regulations

Oil and grease (mg/L) 10 d State effluent regulations

Salinity Report a Discharge permit regulations

Total residual chlorine (mg/L) 0.5 a State effluent regulations

Total ammonia (mg/L as N): Water quality standards

Flow < 0.36 mgd 15 a a 30-day average. b 30-day average/7-day average. c 30-day geometric mean/7-day geometric mean. d Daily maximum.

The BIOLAC-R system, illustrated in Figure 4.20, utilizes an extended aeration process within earthen embankments or various structures, adhering to recommended design criteria outlined in Table 4.15 The system employs conservative design parameters, with typical loadings ranging from 7 to 12 lb BOD5 per day per 1000 ft³ (0.11 to 0.16 kg/m³) of aeration pond Additionally, the food-to-microorganism ratios are maintained between 0.03 to 0.1 lb BOD5 per lb of mixed liquor volatile suspended solids (MLVSS) (0.014 to 0.045 kg/kg) According to a 1990 USEPA report, the average loading rate for 25 BIOLAC-R plants was documented, reflecting the system's efficiency in wastewater treatment.

The average biochemical oxygen demand (BOD5) in BIOLAC-R plants is 975 lb per dãMG, equating to 7.3 lb per 1000 ft³ or 0.1168 kg/m³ For optimal performance, these plants utilize an average of 385 diffusers per million gallons (MG) of aeration basin volume, with an airflow rate of 1350 standard cubic feet per minute (scfm).

The MG (0.01 scm/min/m³) or 3.5 scfm per diffuser indicates that the average operating horsepower at the 25 BIOLAC-R plants is 45 hp/MG (34 kW/MG) for fully nitrified effluent, which is comparable to other complete mix systems Hydraulic retention times vary between 24 to 48 hours, while solids retention times range from 30 to 70 days Although preliminary and primary treatment is generally not implemented, influent screening is recommended Aeration pond depths typically range from 8 to 20 feet (2.5 to 6.1 m), with shallower depths seen in retrofits or where deep construction is not feasible Integral clarifiers are predominantly used for solids management.

Aspirating aerator Surface splasher aerator Chlorine contact

FIGURE 4.12 Plan view of Hotchkiss, Colorado, wastewater plant (From Fagan, J., Consolidated Consulting Services, Delta, CO, personal communication, 2000.)

From 1997 to 2000, Hotchkiss, Colorado, exhibited varying performance data across several parameters The average influent flow rates ranged from 0.113 to 0.274, with a maximum of 0.339 The average effluent flow rates were slightly lower, peaking at 0.396 Influent temperatures fluctuated between 10.4°C and 19.6°C, while effluent temperatures reached up to 25°C Biochemical Oxygen Demand (BOD) values for influent peaked at 282, with effluent BOD consistently low, averaging around 10 The removal efficiency of BOD was notable, and filtered BOD levels remained low Total Suspended Solids (TSS) in influent reached a maximum of 306, with effluent TSS significantly reduced Fecal coliform counts varied widely, with a maximum of 2300 in influent Total Dissolved Solids (TDS) showed a raw influent average of 1282, decreasing in effluent Ammonia levels in influent peaked at 34.9, with effluent levels dropping to as low as 0.54 This data reflects the treatment facility's performance in managing water quality over the observed years.

The data presents a comprehensive analysis of water quality parameters across various months The average influent flow rate (Q) shows a seasonal variation, with a peak of 0.242 in October and a minimum of 0.097 in May The average effluent flow rate also fluctuates, reaching a maximum of 0.294 in December Temperature readings indicate that influent temperatures range from 10°C in February to 19.8°C in July, while effluent temperatures vary from 3°C in March to 25°C in August Biochemical Oxygen Demand (BOD) reveals a significant reduction from an influent average of 186 mg/L in March to an effluent average of just 4 mg/L, demonstrating a remarkable removal efficiency of up to 98.3% Total Suspended Solids (TSS) show a similar trend, with influent levels peaking at 358 mg/L in March and decreasing to 3 mg/L in December The data also highlights the presence of fecal coliforms, consistently below detection limits, and pH levels ranging from a minimum of 6.17 in December to a maximum of 8.99 in September Total Dissolved Solids (TDS) figures indicate influent concentrations peaking at 1660 mg/L, with effluent levels averaging around 1192 mg/L Ammonia concentrations demonstrate a notable decline from an influent average of 28.6 mg/L in March to 0.65 mg/L in December, underscoring the treatment efficiency of the system.

Nitrogen Removal in Lagoons

Lagoon systems are well-documented for their ability to remove BOD and suspended solids, leading to reliable design options However, the nitrogen removal capability of wastewater lagoons has historically received limited attention in system designs until recently Effective nitrogen removal is essential, as even low concentrations of ammonia nitrogen can negatively impact certain environments.

The Praxair In-Situ Oxygenation (I-SO™) system plays a crucial role in managing nitrogen levels in receiving waters, as excessive nitrogen can lead to eutrophication, negatively impacting young fish populations Effective nitrogen removal in preliminary lagoon units is essential, as it can lead to significant cost savings for land treatment systems This article will explore various conventional and commercial products designed specifically for nitrogen removal.

Nitrogen loss from aquatic systems, including streams, lakes, and wastewater lagoons, has been documented for decades, but comprehensive analysis was limited until the early 1980s Researchers have proposed various mechanisms for nitrogen removal, such as algae uptake, sludge deposition, adsorption by bottom soils, nitrification, denitrification, and ammonia volatilization Studies by Pano and Middlebrooks (1982), USEPA (1983), Reed (1984), and Reed et al (1995) indicate that multiple factors contribute to nitrogen loss, with volatilization to the atmosphere being the predominant mechanism under optimal conditions.

In the late 1970s, the U.S Environmental Protection Agency conducted extensive studies on facultative wastewater lagoon systems, confirming significant nitrogen removal capabilities in these systems Key findings indicated that nitrogen removal is influenced by factors such as pH, detention time, and temperature, with optimal conditions allowing for up to 95% nitrogen removal Although recent studies, including one of 178 facultative lagoons in France, reported an average nitrogen removal rate of 60 to 70%, data availability was limited Additionally, research by Wrigley and Toerien highlighted an 82% reduction in ammonia nitrogen over 21 months in small-scale facultative lagoons, but lacked the comprehensive sampling approach used in earlier EPA studies.

In a study by Shilton (1995), the removal of ammonia nitrogen from a facultative lagoon treating piggery wastewater was quantified, revealing volatilization rates ranging from 0.07 to 0.314 lb/1000 ft² per day (355 to 1534 mg/m² per day) Notably, the volatilization rate increased with higher concentrations of ammonia nitrogen and total Kjeldahl nitrogen (TKN).

In a study conducted by Soares et al (1995), ammonia nitrogen removal was monitored in a wastewater stabilization lagoon complex in Brazil, characterized by varying geometries and depths The research demonstrated that ammonia nitrogen concentrations were successfully reduced to 5 mg/L in the maturation lagoons, ensuring the effluent met discharge standards for surface waters Additionally, the ammonia removal process in both facultative and maturation lagoons was effectively modeled using equations derived from the volatilization mechanism proposed by Pano and Middlebrooks (1982).

Camargo-Valero and Mara (2009) concluded that the Pano and Middlebrooks

(1982) model fairly predicts ammonia removal in WSP but did not confirm ammonia volatilization as the main mechanism for permanent nitrogen removal.

Commercial products can significantly reduce ammonia nitrogen and total nitrogen levels Below are some options that illustrate these improvements.

Ammonia nitrogen removal in facultative wastewater stabilization lagoons can occur through the following three processes:

• Gaseous ammonia stripping to the atmosphere

• Ammonia assimilation in algal biomass

FIGURE 4.28 Nitrogen pathways in wastewater lagoons under favorable conditions.

Data Summary from EPA Facultative Wastewater Pond Studies

Corinne, Utah (first three cells)

The low levels of nitrates and nitrites in lagoon effluents suggest that nitrification plays a minimal role in ammonia nitrogen removal The assimilation of ammonia nitrogen by algal biomass is influenced by biological activity, temperature, organic load, detention time, and the characteristics of the wastewater Additionally, the loss of gaseous ammonia to the atmosphere is primarily determined by pH, temperature, and mixing conditions within the lagoon An alkaline pH promotes the conversion of ammonia to its gaseous form, while mixing conditions significantly impact the mass-transfer coefficient Furthermore, temperature affects both the equilibrium constant and the mass-transfer coefficient, highlighting its critical role in the process.

At low temperatures, biological activity diminishes, and the lagoon contents are typically well mixed due to wind effects, making ammonia stripping the primary method for removing ammonia nitrogen in facultative wastewater stabilization lagoons This process can be modeled using a first-order reaction approach, as outlined by Stratton in 1968 and 1969, leading to the establishment of a mass balance equation for the system.

VdC/dt = Q(C 0 − C e ) − kA(NH3) (4.25) where

C = average lagoon contents concentration of (NH4 ++NH3 ) (mg/L as N) t = time (d)

C 0 = influent concentration of (NH 4 + +NH 3 ) (mg/L as N)

C e = effluent concentration of (NH4 ++NH3 ) (mg/L as N) k = mass-transfer coefficient (m/d)

A = surface area of the pond (m 3 )

The equilibrium equation for ammonia dissociation may be expressed as:

[ ] (4.26) where K b is an ammonia dissociation constant.

By modifying Equation 4.26, gaseous ammonia concentration may be expressed as a function of the pH value and total ammonia concentration (NH 4 + +NH 3 ) as follows:

C pKW pK b (4.29) where pK W = –log K W , and pK b = –log K b

Assuming steady-state conditions and a completely mixed lagoon where C e = C, Equations 4.28 and 4.29 will yield the following relationship:

The relationship between pH, temperature, and hydraulic loading rate significantly impacts ammonia nitrogen removal Research by Stratton (1968, 1969) demonstrated that the ammonia loss-rate constant is influenced by both pH levels and temperature (°C).

Ammonia loss rate constant × e 1.57(pH−8.5) (4.31) Ammonia loss rate constant × e 0.13(T−20) (4.32)

In a study by King (1978), it was found that harvesting floating Cladophora fracta from the initial lagoon resulted in only 4% nitrogen removal from secondary effluents The primary method of nitrogen removal in the lagoons was identified as ammonia gas stripping The process of total nitrogen removal followed first-order kinetics, which can be represented by the plug-flow model equation: N t = N 0 e –0.03t, where N t denotes the total nitrogen concentration (mg/L), N 0 represents the initial total nitrogen concentration (mg/L), and t indicates time in days.

Large-scale facultative wastewater stabilization lagoon systems typically only reach steady-state conditions during windy seasons, which are necessary for achieving complete mixing Additionally, the removal of ammonia through biological processes becomes crucial, especially when ammonia is released from anaerobic activity at the lagoon's bottom Therefore, any calculations regarding ammonia removal must consider these factors alongside the theoretical aspects of ammonia stripping, as demonstrated in Equation 4.30.

This article presents mathematical relationships for total nitrogen removal derived from the performance of three full-scale facultative wastewater stabilization lagoons It emphasizes a theoretical approach that incorporates variables such as temperature, pH value, and hydraulic loading rate Instead of relying on the theoretical expression for ammonia nitrogen stripping, a new equation is proposed for TKN removal in facultative lagoons.

+ ⋅ (pH) (4.33) where K is a removal rate coefficient (L/t), and f(pH) is a function of pH.

K values are influenced by temperature and mixing conditions, but for similar lagoon configurations in the same climatic region, they can be simplified to depend primarily on temperature pH, which varies with temperature, impacts pK and pK b values as well as biological activity in the lagoon When analyzing the effect of pH on ammonia nitrogen stripping, it was determined that the relationship is exponential, supported by statistical analyses This aligns with the common use of exponential functions in reaction rate and temperature relationships, such as the van’t Hoff–Arrhenius equation, making it reasonable to apply this theoretical framework to practical scenarios.

Data were systematically collected from all lagoon systems over a complete annual cycle, enabling comprehensive quantitative analysis of key variables Two highly accurate models for predicting nitrogen removal in facultative lagoon systems were developed and validated with external data sources A summary of these models is provided in Tables 4.19 and 4.20, with detailed theoretical development discussed earlier Additional validation information can be found in works by Reed et al (1995), Reed (1984, 1985), and the USEPA.

Modified High-Performance Aerated Pond Systems for

SYSTEMS FOR NITRIFICATION AND DENITRIFICATION

CFID basins have been effectively utilized in Australia and the United States for many years, employing in-basin sedimentation instead of secondary clarifiers and incorporating sludge recycling Unlike sequencing batch reactors, these systems feature continuous flow into the basin while discharge occurs intermittently Although generally successful, CFID basins have faced challenges such as short-circuiting and sludge bulking; however, system modifications can address these issues Rich (1996, 1999) has suggested a modified CFID basin design to enhance performance.

Bubble distribution tubes Water movement

The Bio-Dome (Poo-Gloo) Nitrogen Removal Systems utilize aerated lagoon systems for effective nitrification and denitrification, employing in-basin sedimentation to separate solids retention time from hydraulic retention time, while ensuring continuous influent flow Additionally, a proposed nitrification system integrates a modified CFID with a sludge basin, where the modified CFID basin functions as the reactor and the sludge basin stabilizes and stores sludge, offering operational flexibility For further insights into CFID system operations, refer to Rich (1999).

Nitrogen Removal in Ponds Coupled with Wetlands and

WETLANDS AND GRAVEL BED NITRIFICATION FILTERS

The Nitrification Filter Bed (NFB) system is designed as a retrofit for existing FWS and SSF wetland systems struggling to meet ammonia effluent standards Positioned at the front or end of the wetland, the NFB allows effluent to be evenly distributed, promoting denitrification when placed at the head of the system, thereby enhancing nitrogen removal and reducing BOD While locating the NFB at the end requires less pumping capacity, it limits denitrification, allowing nitrates to exit the system Overall, the NFB offers economic benefits and operational simplicity for retrofitting lagoon wetlands and could serve as an effective solution for nitrogen removal in new lagoon wetland designs.

Control of Algae and Design of Settling Basins

Algae control in wastewater stabilization pond effluents has historically been a significant issue, particularly with the implementation of maturation and polishing ponds following various treatment processes State design standards have exacerbated this challenge by mandating extended detention times in lagoon systems Research indicates that the solids present in lagoon effluents are rarely fecal matter, sparking debate over the necessity of algae removal It was noted that algae eventually die, settle, and decay, which contributes to oxygen demand in receiving streams This concern has led to investigations into effective algae removal methods and system designs to limit algae growth in settling basins A study by Toms et al (1975) found that dominant algae species growth rates were consistently below 0.48 d–1, and that growth issues arise only when hydraulic retention times (HRT) exceed 2 d Furthermore, Uhlmann (1971) reported no algae growth in hyperfertilized ponds with detention times under 2.5 d Toms et al also discovered that effluent total suspended solids (TSS) decreased at HRTs below 2.5 d, while significant TSS increases were observed beyond this threshold, particularly in four-cell lagoons after 4 to 5 d.

As lagoon depth increases, light penetration decreases; however, the trapezoidal shape of lagoon cells limits the benefits of depths beyond 3 to 4 meters Without mechanical mixing, thermal stratification in lagoons fosters an ideal environment for algae growth, but disturbing this stratification can hinder algae proliferation Rich (1999) suggests that a certain level of aeration is beneficial for controlling algae in lagoon cells Moreover, the intensity of aeration affects algae growth by suspending more solids, which reduces light transmission and subsequently leads to lower algae levels.

Hydraulic Control of Ponds

Historically, ponds were typically designed to receive wastewater through a single central pipe, but research has demonstrated that this method is inefficient Studies indicate that multiple inlet arrangements are more effective, even for smaller ponds under 0.5 hectares To optimize performance, inlet points should be spaced far apart, with wastewater introduced via a long diffuser Additionally, the placement of inlets and outlets is crucial for ensuring uniform flow throughout the pond.

Single inlets can be effectively utilized when positioned as far as possible from the outlet structure and equipped with baffles to direct flow, minimizing currents and short-circuiting It is essential for outlet structures to facilitate multiple-depth withdrawals, ensuring that all withdrawals occur at least 0.3 meters (1 foot) below the water surface to mitigate the impact of algae and surface debris on effluent quality.

Analysis of performance data from aerated and facultative ponds reveals that utilizing four cells in series optimally enhances BOD and fecal coliform removal in plug-flow systems Additionally, effective results can be achieved with fewer cells by incorporating baffles or dikes to improve the system's hydraulic characteristics.

Carefully guiding the flow through a pond enhances treatment effectiveness, while also considering economic and aesthetic factors in the decision to implement baffling Generally, increased use of baffling improves flow control and treatment efficiency To achieve optimal results, it is essential to specify the lateral spacing and length of the baffles to maintain a nearly constant cross-sectional flow area.

To reduce wind-induced short-circuiting in ponds, it is ideal to align the inlet-outlet axis perpendicular to the prevailing wind direction If this alignment is not feasible, the use of baffles can help manage wind-driven circulation In a pond with a constant depth, surface currents flow with the wind, while the return flow occurs along the bottom in the opposite direction.

Ponds experience distinct behaviors in winter and summer due to temperature stratification between the inflow and pond contents In summer, the colder inflow sinks to the bottom, moving towards the outlet, while in winter, the warmer inflow rises to the surface This stratification can significantly reduce the effective treatment volume of the pond to that of the inflow layer, leading to a drastic decrease in detention time and potentially inadequate treatment levels.

Short et al (2009) investigated pilot-scale duckweed, rock filter, and attached growth media as methods to upgrade lagoon effluent and reduce algae loads before entering a dissolved air flotation unit Their findings revealed that rock filters exhibited the least effective hydraulic patterns, characterized by high dead space Additionally, both pilot-scale duckweed and open-pond reactors showed significant dead space In contrast, the attached growth media suspended in an open pond notably enhanced hydraulic characteristics, achieving less than 9% dead space.

Removal of Phosphorus

Phosphorus removal is typically unnecessary for wastewater treated in lagoons, except in certain cases found in the north central United States and Canada However, if phosphorus removal is mandated, the experiences detailed below can offer valuable guidance.

To achieve a phosphorus discharge requirement of 1 mg/L into the Great Lakes, a chemical treatment approach in controlled-discharge ponds was developed in Canada, utilizing alum, ferric chloride, and lime By employing a motorboat for the distribution and mixing of chemicals, a typical alum dosage of 150 mg/L effectively reduces phosphorus levels to below 1 mg/L, while also maintaining BOD and SS under 20 mg/L The minimal sludge buildup from these chemicals allows for years of operation without the need for cleaning, and the associated costs are significantly lower than conventional phosphorus removal methods This successful technique has been implemented in several midwestern states.

A two-year study in Ontario, Canada, focused on the in-pond precipitation of phosphorus, BOD, and suspended solids (SS) through chemical dosing The research aimed to evaluate the effectiveness of ferric chloride, alum, and lime in phosphorus removal Continuous dosing with ferric chloride at 20 mg/L and alum at 225 mg/L successfully kept pond effluent phosphorus levels below 1 mg/L In contrast, hydrated lime dosages up to 400 mg/L failed to consistently reduce phosphorus below 1 mg/L, achieving levels between 1 to 3 mg/L, and did not decrease BOD while slightly increasing SS concentration Notably, ferric chloride also lowered effluent BOD from 17 to 11 mg/L.

A study on the effectiveness of alum for phosphorus removal in a modified six-cell pond system in Waldorf, Maryland, revealed that alum significantly reduced total phosphorus levels, achieving an average reduction of 81% when added to the third cell compared to 60% in the first cell In contrast, control ponds without alum showed an average phosphorus removal of only 37%, with effluent concentrations of 2.5 mg/L in the alum-treated system versus 8.3 mg/L in the control While alum had minimal impact on biochemical oxygen demand (BOD) and suspended solids (SS), with slight reductions observed, some instances of increased effluent concentrations were noted This indicates that direct chemical addition is primarily effective for phosphorus removal.

Removal of Pharmaceuticals and Personal Care Products

CARE PRODUCTS AND ANTIBIOTIC RESISTANT GENES

Potera (2013) developed a method to analyze 13 pharmaceuticals and personal care products (PPCPs) alongside eight steroid hormones in various water samples The study found ten chemicals in sewage influents and effluents, with concentrations ranging from ng/L to low μg/L Notably, three hormones reached total concentrations of up to 164 ng/L in aerated lagoon influents, although no residues were detected in the effluents While carbamazepine proved challenging to remove, the removal rates for other detected PPCPs varied significantly.

At average air temperatures of 20°C, the removal rates of PPCP contaminants such as caffeine, ibuprofen, triclosan, and hormones reached 88 to 100% In November, when the average temperature dropped to 4.4°C, the removal rates were 1–2 orders of magnitude higher than those observed in September Additionally, PPCP contaminants in the surrounding watershed showed significantly elevated levels in both the receiving creek and the adjacent river.

Potera (2013) highlighted the significant impact of antibiotic-resistant genes (ARGs) from manure pits, referencing laboratory studies by Rysz et al (2013) In the United States, around 10,000 confined animal feeding operations (CAFOs) produce approximately 300 million tons of manure each year, which is more than double the waste generated by humans Notably, nearly half of the antibiotics used in the U.S are administered to livestock, leading to the excretion of ARGs into manure lagoons This issue not only poses a serious medical challenge but also presents significant agricultural and environmental concerns.

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The U.S Environmental Protection Agency (USEPA) has published essential manuals for wastewater management, including the 1993 "Nitrogen Control" guide (EPA/625/R-93/010) aimed at enhancing nitrogen management practices Additionally, the 2011 document titled "Principles of Design and Operations of Wastewater Treatment Pond Systems" (EPA/600/R-11/088) serves as a comprehensive resource for plant operators, engineers, and managers, detailing effective design and operational strategies for wastewater treatment ponds Both publications are vital for improving environmental practices and ensuring compliance with water quality standards.

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Pond Modifications for Polishing Effluents

Solids Removal Methods

Pond systems often face the challenge of high total suspended solids (TSS) concentrations, which can exceed 100 mg/L, primarily due to algae and pond detritus rather than wastewater solids These elevated levels typically occur for 2 to 4 months each year Various solids removal methods are available, including intermittent sand filters, recirculating sand filters, rock filters, coagulation-flocculation, and dissolved-air flotation While the rock filter is not a conventional filter, it is relevant to the discussion of solids removal from lagoon effluents For comprehensive insights into these methods, refer to the references provided at the chapter's end, including the U.S Environmental Protection Agency’s Design Manual on municipal systems, which offers valuable information on the design and performance of intermittent sand and rock filters.

Wastewater Stabilization Ponds (USEPA 1983) Information on both processes can also be found in a document published by the Water Environment Federation (2001)

A literature review revealed a scarcity of information on the use of recirculating sand filters for removing total suspended solids (TSS) from lagoon effluents; however, references were found regarding their effectiveness in treating septic tank effluents This limited data may stem from concerns about algae accumulation in the filter media Nolte and Associates (1992) provided a comprehensive review of both recirculating and intermittent sand filters in their study.

Intermittent sand filters have a proven track record in wastewater treatment, dating back to early studies (Furman et al 1955; Grantham et al 1949; Massachusetts Board of Health 1912) Historical data from Massachusetts around 1900 highlights the design and effectiveness of these systems in treating raw or primary effluent wastewater, resulting in high-quality effluent A typical intermittent sand filter operates similarly to slow sand filtration used in potable water treatment, providing a cost-effective solution for polishing pond effluents through periodic application to a sand filter bed.

As wastewater flows through the filter bed, suspended solids and organic matter are effectively removed via physical straining and biological degradation Over time, particulate matter accumulates in the top 2 to 3 inches (5 to 8 cm) of the filter, leading to surface clogging that hinders further infiltration of effluent When clogging occurs, the filter bed is taken out of service, the top layer of clogged sand is removed, and the system is restored for use The extracted sand can either be washed and reused or disposed of appropriately.

During the 1970s and 1980s, extensive studies on the performance of intermittent sand filters treating lagoon effluents were conducted, as summarized in Tables 5.2 and 5.3 Table 5.2 compiles findings from various literature and EPA documents, while Table 5.3 presents results from field investigations of three full-scale systems combining lagoons and intermittent sand filters These studies demonstrate that it is feasible to achieve effluent quality with total suspended solids (TSS) and biochemical oxygen demand (BOD) levels below 15 mg/L when utilizing anaerobic, facultative, and aerated lagoons followed by intermittent sand filters with effective sizes of 0.3 mm or less.

Design and Performance of Early Massachusetts Intermittent Sand Filters

Ammonia Removal BOD 5 Removal Influent

Source: Data from the Massachusetts Board of Health, Eng Contracting, 37, 271, 1912; and Mancl,

K.M and Peeples, J.A., One hundred years later: reviewing the work of the Massachusetts State Board of Health on the intermittent sand filtration of wastewater from small communities, in

Proceedings of the 6th National Symposium on Individual and Small Community Sewage Systems, Chicago, American Society of Agricultural Engineers (ASAE), December 16–17,

Rich and Wahlberg (1990) assessed the effectiveness of five facultative lagoon–intermittent sand filter systems in South Carolina and Georgia, revealing their superior performance compared to ten aerated lagoon systems The evaluation included a summary of design characteristics and performance metrics, highlighting that six systems featured an aerated cell followed by a polishing pond, while three were dual-power, multicellular systems, and one was a single-cell, dual-power system The findings align with data from Niku et al (1981), further emphasizing the efficiency of facultative lagoon–intermittent sand filter systems.

Filter bed Drain Sand and gravel placement Seal

Liner and filter bed plan view

Seal between liner and pipe

Seal between liner and pipe Seal between drain pipe and liner

Sand media 5” dia rock 75” dia rock 1.5” dia rock

FIGURE 5.1 Cross-sectional and plan views of typical intermittent sand filter (Reprinted by permission from Macmillan Publishers Ltd., Wastewater Stabilization Lagoon Design,

Performance, and Upgrading, Middlebrooks, E.J et al., copyright 1982.)

TABLE 5.2 Intermittent Sand Filter Performance Treating Lagoon Effluents a Pond T ype u b

The study on loading rates (mgd/ac) reveals significant removal efficiencies for Total Suspended Solids (TSS), Volatile Suspended Solids (VSS), and Biochemical Oxygen Demand (BOD) across various influent and effluent concentrations For facultative systems, the removal percentages ranged from 56% to 94% for TSS, 64% to 92% for VSS, and 69% to 90% for BOD, indicating effective treatment processes Notable references include Marshall and Middlebrooks (1974), Harris et al (1978), and Hill et al (1977), each demonstrating varying influent conditions and corresponding removal efficiencies Overall, the data underscores the importance of optimizing loading rates to enhance wastewater treatment performance.

The study presents various performance metrics of different filtration systems, including facultative, aerated, and anaerobic treatments Notably, facultative systems show a mean value of 9.73 with a uniformity coefficient of 0.25, achieving up to 86% efficiency Aerated systems, according to Bishop et al (1977), demonstrate a mean of 9.73 and a notable 52.5% efficiency In contrast, anaerobic systems yield varying results, with Messinger (1976) reporting a mean of 0.1 and efficiencies ranging from 45.1% to 67.6% The findings highlight the effectiveness of 0.17-mm e.s filters in optimizing filtration performance across different wastewater treatment methods.

TABLE 5.3 Mean Performance Values for Three Full-Scale Lagoon–Intermittent Sand Filter Systems Par ameter

The analysis of water quality at various facilities, including Mt Shasta, Moriarty, and Ailey, reveals significant differences in key parameters The influent lagoon at Mt Shasta shows a biochemical oxygen demand (BOD) of 114 mg/L, while the effluent filter exhibits a much lower BOD of 8 mg/L Soluble BOD values also reflect this trend, with influent levels at 41 mg/L decreasing to 5 mg/L in the effluent Total suspended solids (TSS) are notably high in the influent at 83 mg/L, reducing to 16 mg/L after filtration Furthermore, fecal coliform counts are alarmingly high in the Mt Shasta influent at 1.16 × 10^6 number/100 mL, but drop to less than 2 in the effluent pH levels vary across facilities, with Moriarty showing a pH of 8.7, while the effluent from Ailey maintains a pH of 8.0 Dissolved oxygen (DO) levels at the Moriarty facility are optimal at 12.4 mg/L, contrasting with Mt Shasta's lower DO of 4.8 mg/L Chemical oxygen demand (COD) measurements indicate a high influent COD of 244 mg/L at Mt Shasta, which decreases to 68 mg/L in the effluent Lastly, soluble COD levels are also reduced from 159 mg/L in the influent to 50 mg/L in the effluent, demonstrating effective treatment processes across these facilities.

The data presented includes various water quality parameters measured in mg/L, such as alkalinity (Akl), total phosphorus (TP), total Kjeldahl nitrogen (TKN), ammonia (NH3), organic nitrogen, nitrite (NO2), nitrate (NO3), and total algal count (cells/ml), along with flow rates (mgd) Notably, the highest alkalinity recorded was 436 mg/L as CaCO3, while the total phosphorus peaked at 10.3 mg-P/L The TKN levels varied significantly, with a maximum of 60 mg-N/L Ammonia concentrations reached up to 38 mg-N/L, and organic nitrogen levels fluctuated, with a high of 22 mg-N/L Nitrite levels showed a maximum of 159 mg-N/L, while nitrate concentrations were generally low, with a peak of 4.5 mg-N/L The total algal count varied widely, with a notable measurement of 756,681 cells/ml This data is sourced from studies by Russell et al (1980, 1983) and highlights the performance comparison between activated sludge plants and aerated lagoon systems.

Truax and Shindala (1994) conducted a comprehensive evaluation of facultative lagoon-intermittent sand filter systems, utilizing four sand grades with effective sizes ranging from 0.18 to 0.70 mm and uniformity coefficients between 1.4 and 7.0 Their findings indicated a direct correlation between sand effective size and hydraulic loading rate, as shown in Table 5.6 Notably, sands with an effective size of 0.37 mm or smaller, when subjected to hydraulic loading rates of 4.9 gal/ft²·d (0.2 m³/m²·d), achieved effluent quality with BOD5 and TSS levels below 15 mg/L Additionally, total TKN concentrations were successfully reduced from 11.6 mg/L.

Design Characteristics and Performance of Facultative Lagoon–Intermittent Sand Filter Systems

Source: Rich, L.G and Wahlberg, E.J., J Water Pollut Control Fed., 62, 697–699, 1990 With permission. a Based on design flow rate.

Performance Comparison of Lagoon–Intermittent Sand Filters with

Aerated Lagoons and Activated Sludge Plants

Source: Rich, L.G and Wahlberg, E.J., J Water Pollut Control Fed, 62, 697–699, 1990 With permission. a One aerated cell followed by a polishing pond. b Facultative lagoons upgraded to dual-power, multicellular aerated lagoons.

4.3 mg/L at the 4.9-gal/ft 2 ãd (0.2-m 3 /m 2 ãd) loading rate The experiments were con- ducted in a mild climate, and it is unlikely that similar nitrogen removal will occur during cold months of more severe climates.

Melcer et al (1995) examined a full-scale aerated lagoon–intermittent system in New Hamburg, Ontario, operational since 1980 Data from 1990 and January to August 1991 indicated surface loading rates of 79.6 gal/ft²·d (3.24 m³/m²·d), with influent concentrations of BOD₅, TSS, and TKN at 12, 16, and 19 mg/L, respectively The system demonstrated exceptional filter effluent quality, achieving BOD₅, TSS, and TKN concentrations below 2 mg/L.

The length of a filter run is influenced by the effective size of the sand used and the amount of solids that accumulate on the filter's surface, as outlined in the EPA’s Design Manual.

Municipal Wastewater Stabilization Ponds (USEPA 1983) and various studies (Bishop et al 1977; Harris et al 1978; Hill et al 1977; Marshall and Middlebrooks 1974; Messinger 1976; Russell et al 1983; Tupyi et al 1979) provide comprehensive insights into the correlation between the solids accumulated on filter surfaces and the duration of operation Additionally, Truax and Shindala (1994) observed run times that closely align with the findings of these earlier studies, as detailed in Table 5.8.

Modifications and Additions to Typical Designs

Controlled discharge refers to the practice of restricting the release of effluent from lagoon systems to times when the water quality meets regulatory standards Typically, discharges are avoided during winter months, as well as during spring and fall turnover periods and times of algal blooms Many countries have regulations that prohibit lagoon discharges during the winter season.

In a study conducted by Pierce (1974) on 49 lagoon installations in Michigan that utilize controlled discharge, it was found that the majority of these systems feature multiple cells, with 27 having two cells, 19 three cells, 2 four cells, and 1 five cell Discharge typically occurs during late spring and early fall, although some systems operate year-round, with discharge durations ranging from less than 5 days to over 31 days To maintain adequate storage capacity for non-discharge periods, lagoons are emptied to a minimum depth of approximately 0.46 meters (18 inches) during each controlled discharge.

During the discharge period, the lagoon effluent was assessed for BOD5, TSS, and FC levels The concentrations of BOD5 and TSS observed in the study are displayed in Figure 5.10, which illustrates the probability of occurrence The data is organized by magnitude and plotted on normal probability paper, with concentration (mg/L) against the likelihood of exceeding those values under similar conditions This plot facilitates a comparison between the performance of two-cell lagoon systems and those with three or more cells The findings presented in Figure 5.10 are further summarized in Table 5.18.

BOD 3 or more cells BOD - 2 cells

A comparison of effluent quality between two-cell systems and lagoon systems with three or more cells reveals significant differences, particularly in systems with extended storage periods prior to discharge This analysis, conducted in Michigan, highlights the effectiveness of upgrading wastewater stabilization ponds to comply with new discharge standards, as outlined by Pierce in the 1974 report from the Utah Water Research Laboratory at Utah State University.

The study revealed that the expected effluent BOD5 concentration for controlled discharge systems is 17 mg/L for two-cell lagoon systems and 14 mg/L for systems with three or more cells There is a 90% probability that the BOD5 concentration will not surpass 27 mg/L, which is slightly below the U.S Federal Secondary Treatment Standard of 30 mg/L Additionally, the most probable effluent TSS concentration for two-cell lagoon systems is 30 mg/L.

In lagoon systems with three or more cells, the effluent total suspended solids (TSS) concentration averages 27 mg/L, while two-cell systems have a 90% probability level of 46 mg/L, and those with three or more cells show 47 mg/L Additionally, fecal coliform (FC) levels typically remain below 200/100 mL; however, this standard was occasionally surpassed in instances where chlorination was not utilized.

A study conducted in Minnesota on controlled discharge lagoon systems revealed that the discharge practices of 39 installations were comparable to those in Michigan During the fall discharge period, 36 out of 39 installations demonstrated effluent BOD5 concentrations below 25 mg/L and effluent TSS concentrations under 30 mg/L Furthermore, effluent FC concentrations were assessed at 17 of the lagoon installations, with all reporting levels below 200/100 mL.

During the spring discharge period, monitoring of 49 municipal lagoon installations revealed that only three exceeded an effluent BOD5 concentration of 30 mg/L, with a maximum recorded concentration of 39 mg/L Effluent TSS concentrations varied significantly, ranging from 7 to 128 mg/L, and 16 installations reported levels above 30 mg/L Additionally, among the 45 installations assessed for effluent FC concentrations, only three surpassed 200/100 mL.

Effluent Quality Resulting from Controlled Discharge Operations of 49 Michigan Lagoon Installations

Two Cells Three or More Cells Two Cells Three or More Cells

(will not be exceeded 9 out of

Source: Pierce, D.M., in Upgrading Wastewater Stabilization Ponds To Meet New Discharge Standards,

PRWG151, Utah Water Research Laboratory, Utah State University, Logan, 1974.

Controlled discharge of lagoon effluent is an effective and cost-efficient method for achieving high treatment levels Regular monitoring of lagoon effluent is essential to establish appropriate discharge periods, which can last for most of the year Some lagoon systems may require increased storage capacity for controlled discharge, but many already possess sufficient freeboard and storage that can be utilized with minimal modifications.

The hydrograph controlled release (HCR) pond is an innovative variation of the controlled discharge pond, initially developed in the southern United States but applicable globally This system regulates discharge during high-flow periods based on gauging station data from the receiving stream, while storing effluent in the HCR pond during low-flow times Designed with conventional facultative or aerated ponds for initial treatment, the HCR cell's primary function is effluent storage, with residence times ranging from 30 to 120 days The HCR cell typically maintains a maximum water level of 8 ft (2.4 m) and a minimum of 2 ft (0.6 m) A key advantage of HCR systems is the flexibility to meet lower discharge standards during high-flow conditions, contrasting with systems designed for strict low-flow requirements An assessment by Zirschsky and Thomas (1987) highlighted the HCR system's effectiveness, cost-efficiency, and ease of operation, also noting its capability to upgrade existing lagoon systems The article includes illustrations of several simple effluent release structures.

Hydrograph-controlled release pond design in the United States focuses on reducing biochemical oxygen demand (BOD) and suspended solids (SS) loadings during critically low river flows by limiting effluent discharge rates instead of lowering pollutant concentrations These systems must retain wastewater during low flow conditions, defined as the once-in-10-year low flow rate over a 7-day period (Q 10/7), which can involve utilizing existing ponds or constructing new storage ponds Additionally, it is essential to establish the assimilative capacity of the receiving stream through historical data analysis or by employing established estimation techniques (Zirschsky and Thomas 1987).

In areas of the world where the moisture deficit (evaporation minus rainfall) exceeds

A complete retention wastewater pond can be the most cost-effective disposal method when low-cost land is available, requiring sizing to accommodate the annual wastewater volume and precipitation It is crucial to design the pond for the maximum wet year and minimum evaporation year to prevent overflow, although less stringent standards may apply if overflow is acceptable or alternative disposal methods are available Accurate monthly evaporation and precipitation data are essential for proper sizing, and these ponds typically necessitate large land areas that become non-productive The land must be flat or shaped to ensure uniform depth and large surface areas for the ponds Detailed design procedures for complete retention wastewater pond systems can be accessed through the National Technical Information Service, U.S Department of Commerce.

22161 (ask for the EPA-625/1-83-015 document).

Autoflocculation of algae has been observed during some studies (Golueke and Oswald 1965; Hill et al 1977; McGriff and McKinney 1971; McKinney 1971)

Chlorella is the dominant alga found in various cultures, and laboratory experiments have shown that combining activated sludge with algae can create large, well-settling bacteria-algae flocs In certain cases, floating algae blankets have been observed when chemical coagulants are used, potentially due to gas bubbles from metabolic processes or the algae achieving neutral buoyancy A pilot plant operating at 3000 gallons per hour demonstrated this phenomenon with alum doses between 125 to 170 mg/L, resulting in approximately 50% of the algae being skimmed from the surface However, the rare conditions required for autoflocculation make it an ineffective method for algae removal in wastewater stabilization ponds Although phase isolation experiments have had some success in removing algae from lagoon effluents, full-scale implementations have not consistently yielded positive results.

Encouraging microbial growth in oxidation ponds is an effective strategy for sustaining biological populations while achieving desired treatment outcomes Baffles play a crucial role not only in promoting proper mixing and preventing short-circuiting but also in serving as a substrate for the growth of bacteria, algae, and other microorganisms This dual function enhances the overall efficiency of the treatment process.

Attached growth processes generally outperform suspended growth when there is adequate surface area available In anaerobic or facultative ponds equipped with baffling or biological disks, the microbiological community exhibits a gradient ranging from algae to photosynthetic and chromogenic bacteria, ultimately leading to nonphotosynthetic, nonchromogenic bacteria Experiments have shown that the presence of attached growth on baffles significantly enhances treatment efficiency compared to nonbaffled systems Research by Polprasert and Agarwalla highlighted the importance of biofilm biomass developing on the sidewalls and bottoms of ponds, leading to a model for substrate utilization in facultative ponds that incorporates first-order reactions for both suspended and biofilm biomass.

Performance Comparisons with other Removal

Designers and owners of small systems should prioritize simple technology to avoid maintenance issues and ensure compliance with effluent standards Experience indicates that both small and large communities lacking trained personnel and spare parts face significant challenges with complex systems Effective methods discussed include dissolved-air flotation, centrifugation, coagulation-flocculation, and granular media filtration, which require skilled maintenance and operation When adequate support is available, these processes can function effectively.

Effective algae removal or control methods for lagoon effluents are varied, but choosing the right approach depends on several factors For small communities with limited resources and untrained staff, it is advisable to opt for the simplest system that suits their specific conditions.

In rural areas with adequate land, lagoons such as controlled discharge lagoons or hydrograph-controlled release lagoons are an excellent choice In arid areas, the total

List of Chemicals Produced by Decomposing Straw

When considering wastewater treatment options, the use of a containment lagoon is recommended The effectiveness of this treatment method can be optimized by carefully selecting the discharge timing, resulting in high-quality effluent with biochemical oxygen demand (BOD5) and total suspended solids (TSS) levels below 30 mg/L.

In areas with limited land and scarce resources, simple methods for controlling algae in effluents are most effective Techniques such as intermittent sand filters, land application of effluents, overland flow, rapid infiltration, constructed wetlands, and rock filters can be beneficial Specifically, intermittent sand filters operating at low application rates in warm climates facilitate nitrification, while applying effluents to farmland effectively reduces nitrogen and phosphorus levels, resulting in high-quality effluent.

Energy savings from various wastewater treatment processes can be significant, leading to the production of high-quality effluent (Middlebrooks et al., 1981) A comparison of expected effluent quality and energy consumption across different treatment methods is illustrated in Table 5.22, ranging from basic to advanced systems It is essential to account for numerous design variables, with particular emphasis on energy consumption and the complexity of the processes involved.

Total Annual Energy for Typical 1-mgd System Including Electrical and Fuel

Slow rate, ridge, + furrow (facultative lagoon) 1 1 0.1 3 181

Facultative lagoon + intermittent sand filter 15 15 — 10 241

Aerated lagoon + intermittent sand filter 15 15 — 20 506

Extended aeration + intermittent sand filter 15 15 — — 708

Activated sludge + anaerobic digestion + filter 15 10 — — 911

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Lynam, Ettelt, and McAllon (1969) explored the efficacy of rapid sand filtration and microstrainers for tertiary treatment of wastewater in metro Chicago, highlighting its significance in water pollution control Similarly, Mancl and Peeples (1991) reflected on a century of advancements by the Massachusetts State Board of Health regarding the intermittent sand filtration of wastewater from small communities, emphasizing its relevance in contemporary sewage management practices.

Mara, D.D., Pearson, H.W., and Silva, S.A., Eds (1996) Waste stabilization ponds: technol- ogy and applications, Water Sci Technol., 33, 7.

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Mara, D.D and Johnson, M.L (2006) Aerated rock filters for enhanced ammonia and fecal coliform removal from facultative pond effluents, J Environ Eng., 132(4), 574. Marshall, G.R and Middlebrooks, E.J (1974) Intermittent Sand Filtration to Upgrade Existing

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Mota, Soares dos Santos, and Bezerra dos Santos (2009) conducted a study on the quality of water and fish during the reuse of sewage treated in stabilization ponds in Ceara, Brazil Their research was presented at the 8th IWA Specialist Group Conference on Wastewater Stabilization Ponds and the 2nd Latin-American Conference on the same topic in Belo Horizonte, Brazil This study highlights the importance of evaluating water quality and its impact on aquaculture practices in regions utilizing treated sewage.

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In their 2009 study presented at the 8th IWA Specialist Group Conference on Wastewater Stabilization Ponds in Belo Horizonte, Brazil, Short, Fallowfield, and Cromar emphasize the significance of hydraulic characteristics in waste stabilization pond research Their findings from pilot-scale reactors, including algal, duckweed, rock filter, and attached growth media systems, highlight how these characteristics influence the efficiency and effectiveness of wastewater treatment processes This research underscores the critical role of hydraulic design in optimizing the performance of waste stabilization ponds.

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In 1998, K.E Snider communicated insights regarding engineering practices in Fort Collins, Colorado Additionally, a paper by J.B Stamberg and colleagues, presented at the 57th Conference of the Water Pollution Control Federation in New Orleans, Louisiana, discussed how simple rock filter upgrades can enhance lagoon effluent to achieve Advanced Water Treatment (AWT) quality in West Monroe, Louisiana.

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The U.S Environmental Protection Agency's Center for Environmental Research Information in Cincinnati, OH, has highlighted significant studies on algae removal from wastewater Notably, van Vuuren and van Duuren (1965) explored effective methods for removing algae from wastewater maturation pond effluent, as published in the Journal of Water Pollution Control Federation Additionally, van Vuuren and colleagues investigated the flotation of algae in water reclamation processes, contributing valuable insights to the International Journal of Air and Water Pollution.

The Water Environment Federation's "Natural Systems for Wastewater Treatment" (2001) is a comprehensive guide on effective wastewater management practices, emphasizing the importance of natural treatment systems Additionally, Zirschsky and Thomas (1987) discuss the innovative hydrograph controlled release (HCR) lagoons in their article, highlighting advancements in water pollution control techniques Together, these resources provide valuable insights into modern wastewater treatment methodologies and the evolution of sustainable practices in water management.

Free Water Surface Constructed Wetlands

Process Description

Wetlands serve crucial engineering functions, particularly defined by the water surface position, as outlined by Reed and Bastian (1985) FWS wetlands, including natural marshes and swamps, feature a water surface that is open to the atmosphere, while bogs may qualify if water flows over the peat Constructed FWS wetlands typically comprise vegetated shallow basins or channels designed with barriers to prevent seepage, supporting emergent macrophyte vegetation, and equipped with suitable inlet and outlet structures These wetlands generally maintain water depths ranging from 0.2 to 2.6 feet (0.05 to 0.8 meters), with operational design flows varying from less than 1,000 gallons per day (4 m³/d) to over 20 million gallons per day (75,000 m³/d).

Wetlands exhibit biological conditions akin to those found in facultative treatment ponds, characterized by an anoxic/anaerobic environment at the bottom and an aerobic shallow zone near the surface, where atmospheric reaeration provides oxygen Unlike facultative ponds, which benefit from oxygen produced by algae, densely vegetated wetlands lack this additional oxygen source due to shading from the plant canopy A key distinction lies in the presence of a physical substrate in wetlands that supports periphytic attached-growth microorganisms, which play a crucial role in the biological treatment processes within these ecosystems.

In FWS wetlands, the substrate consists of submerged leaves, stems of living plants, standing dead plants, and benthic litter, while SSF wetlands feature submerged media surfaces along with the roots and rhizomes of emergent plants Treatment responses occur more rapidly in wetlands compared to facultative ponds due to the presence of substrates and periphytic organisms Additionally, SSF wetlands exhibit even faster treatment responses than FWS wetlands, attributed to the greater availability of substrate in the gravel media.

The Subsurface Flow (SSF) wetland concept provides several advantages over Free Water Surface (FWS) wetlands, including a higher treatment rate and reduced mosquito issues due to the water being below the gravel surface In colder climates, the SSF design offers better thermal protection thanks to the water's subsurface position and the insulating litter layer Additionally, the SSF system minimizes public exposure to wastewater since the water surface is not easily accessible However, the primary drawback of the SSF design is the high cost of gravel media, which significantly increases project expenses, especially for larger flow rates exceeding 50,000 gallons per day Consequently, SSF wetlands are more suitable for smaller applications where public exposure is a concern, such as individual homes, parks, and schools, while the FWS concept is more economical for larger municipal and industrial systems and offers greater habitat value integration.

The treatment processes in FWS and SSF wetlands involve a complex interplay of biological, chemical, and physical responses Due to shallow water depths and low flow velocities, particulate matter settles quickly or becomes trapped in the wetland's plant and gravel matrix Algae are also hindered from regenerating because of shading in densely vegetated areas This leads to anaerobic decomposition of deposited materials in the benthic layers, releasing dissolved and gaseous substances into the water These dissolved substances are then available for sorption by the wetland's soils and active microbial and plant populations Additionally, oxygen is accessible at the water surface and on plant surfaces, allowing for aerobic reactions to occur within the system.

Wetland Components

Key components that impact the treatment process in constructed wetlands include plants, detritus, soils, bacteria, protozoa, and larger animals The effectiveness of these components is affected by factors such as water depth, temperature, pH levels, redox potential, and dissolved oxygen concentration.

Aquatic plants play a crucial role in wetland systems for wastewater treatment, with larger trees like cypress, ash, and willow often found in natural bogs and strands in Florida However, these species have not been integrated into constructed wetlands, nor has their treatment function been clearly defined The most prevalent species in constructed marsh wetlands are emergent aquatic macrophytes, particularly cattails (Typha), reeds (Phragmites communis), rushes (Juncus spp.), bulrushes (Scirpus), and sedges (Carex), with bulrush and cattails being the dominant plants in many U.S systems While Phragmites is less common in the U.S., it is the preferred species in European constructed wetlands Systems designed for both habitat and treatment typically incorporate a wider variety of plants to enhance food and nesting opportunities for birds and aquatic life.

Free water surface (FWS) wetlands, such as those in Arcata, California, are highlighted in this article For additional information on the characteristics of the plants found in these wetlands, readers can refer to several key references, including Hammer (1992), Lawson (1985), Mitsch and Gosselink (2007), and Thornhurst (1993).

Typha angustifolia (narrow leaf cattail) and Typha latifolia (broad leaf cattail) are common worldwide varieties of cattails, thriving in pH levels between 4 and 10 Narrow leaf cattails can tolerate salinity levels of 15 to 30 ppt, while broad leaf cattails prefer less than 1 ppt These plants exhibit rapid growth through rhizomes, establishing dense cover within a year at a spacing of 2 feet (0.6 meters) Their roots penetrate shallowly, about 1 foot (0.3 meters) in gravel, with an annual yield of 14 tons per acre (30 mt/ha) The tissue composition is 45% carbon, 14% nitrogen, and 2% phosphorus, with 30% solids Cattails serve as a vital food source for water birds, muskrats, nutria, and beavers, while also providing essential nesting cover They can withstand permanent inundation of over 1 foot (0.3 meters) and are adaptable to drought conditions Commonly found in various FWS and SSF wetlands across the United States, their shallow root penetration may require design adjustments for SSF systems.

Bulrush, known scientifically as Scirpus in the U.S and Schoenoplectus globally, encompasses several varieties including Scirpus acutus (hardstem bulrush), Scirpus cyperinus (wool grass), and Scirpus validus (softstem bulrush) This plant is distributed worldwide and thrives in pH levels ranging from 4 to 9 Salinity tolerance varies among species; hardstem, wool grass, river, and softstem bulrushes tolerate salinity levels of 0 to 5 ppt, while alkali and Olney’s bulrush can withstand up to 25 ppt Growth rates differ, with alkali, wool grass, and river bulrush achieving dense cover within one year at 1-ft spacing, while other varieties can reach dense cover at 1- to 2-ft spacing, demonstrating moderate to rapid growth Additionally, deep root penetration occurs in gravel substrates, enhancing stability and water absorption.

Bulrushes, which can grow up to 2 feet (0.6 meters) tall, yield approximately 9 dry tons per acre (20 metric tons per hectare) annually Their tissue composition consists of about 18% nitrogen, 2% phosphorus, and 30% solids These plants serve as a vital food source for various water birds, muskrats, nutria, and fish, while also providing essential nesting areas for fish during inundation Bulrushes can thrive in permanently flooded conditions, with hardstem varieties reaching up to 3 feet (1 meter) and most others growing between 0.5 to 1 foot (0.15 to 0.3 meters); some species are also drought-tolerant They are widely utilized in many constructed wetlands by the U.S Fish and Wildlife Service and other agencies.

Phragmites australis, commonly known as common reed, along with wild reed, are globally distributed varieties that thrive in a pH range of 2 to 8 and can tolerate salinity levels of less than 45 ppt These plants exhibit rapid growth through rhizomes, spreading laterally about 3 feet (1 meter) per year, which allows them to establish dense cover within just one year when spaced 2 feet (0.6 meters) apart Additionally, they demonstrate deep root penetration in gravel, reaching depths of approximately 1.5 feet (0.4 meters), and can yield around 18 dry tons per acre annually.

Reeds, with a tissue composition of approximately 45% carbon, 20% nitrogen, and 2% phosphorus, are characterized by their ability to thrive in various conditions, being able to endure permanent inundation up to 1 meter and exhibiting drought resistance While they provide limited food value for birds and animals, they do offer some nesting cover In the United States, reeds are often viewed as an invasive species in natural wetlands but have been effectively utilized in constructed wastewater treatment wetlands, where they are the predominant species used, particularly in Europe Their low nutritional value makes them less susceptible to damage from pests like muskrats and nutria, which can affect other plant species in similar environments.

Juncus species, including Juncus articulatus, Juncus balticus, and Juncus effusus, are globally distributed and thrive in pH levels ranging from 5 to 7.5 These rushes exhibit slow growth through rhizomes, achieving a lateral spread of less than 0.3 ft (0.1 m) per year, but can create dense cover within a year when planted 0.5 ft (0.15 m) apart With an annual yield of 45 tons per acre (50 mt/ha), their tissue composition is approximately 15% nitrogen, 2% phosphorus, and 50% solids Rushes serve as a vital food source for various bird species and muskrats, and while some can withstand permanent inundation up to 1 ft (0.3 m), they prefer periods of dry-down Although other plants may be more suitable for primary roles in wastewater wetlands, rushes effectively enhance habitats when used as peripheral plantings.

Carex species, including Carex aquatilis (water sedge), Carex lacustris (lake sedge), and Carex stricata (tussock sedge), are globally distributed and thrive in pH levels ranging from 5 to 7.5, with a low salinity tolerance of less than 0.5 ppt These sedges exhibit moderate to slow growth through rhizomes, achieving a lateral spread of less than 0.5 ft (0.15 m) per year, and can provide dense cover within a year when spaced at 0.5 ft (0.15 m) The annual yield is typically below 4 tons (5 mt/ha) on a dry weight basis, with tissue composition around 1% nitrogen, 0.1% phosphorus, and 50% solids Sedges serve as a vital food source for various birds and moose, with some species capable of enduring permanent inundation while others need a drying period While sedges may not be the primary species for wastewater wetlands, they are ideal for peripheral plantings that enhance habitat value.

Submerged plant species play a crucial role in enhancing water quality in deep-water areas of FWS wetlands, contributing to a patented method for improving freshwater ecosystems, including lakes, ponds, and golf course water hazards One notable species utilized in this process is Ceratophyllum demersum, commonly known as coontail or hornwort.

Elodea, Potamogeton pectinatus, Potamogeton perfoliatus, Ruppia maritima, Vallisneria americana, and Myriophyllum spp are aquatic plants found globally, thriving in pH levels of 6 to 10 and salinity tolerances of less than 5 to 15 ppt These species exhibit rapid growth through rhizomes, achieving a lateral spread of over 1 ft/year, creating dense cover within a year when spaced 2 ft apart Annual yields vary, with coontail producing 8.9 tons/ac, Potamogeton yielding 2.7 tons/ac, and watermilfoil at 8 tons/ac Their tissue composition includes approximately 2 to 5% nitrogen and 0.1 to 1% phosphorus, with 5 to 10% solids These plants serve as vital food sources for various birds, fish, and animals, particularly benefiting ducks They can withstand continuous inundation, with their growth dependent on water clarity for sunlight penetration Additionally, some species enhance habitat values in constructed wetlands and are effective for nutrient control in freshwater bodies, with regular harvesting helping to manage plant and nutrient levels.

Floating plants, particularly duckweed (Lemna), are often found in wastewater treatment systems, although they are not typically included in the design of constructed wetlands Duckweed can have both positive and negative effects in FWS wetlands; it suppresses algae growth but also reduces atmospheric oxygen transfer due to its dense mat This fast-growing plant can yield over 18 tons per acre annually, with a tissue composition of approximately 6% nitrogen and 2% phosphorus Duckweed thrives in conditions with salinity levels below 0.5 ppt and serves as a food source for various wildlife, including ducks, muskrats, and beavers Its presence in FWS wetlands is inevitable, as it can tolerate partial shade To maintain optimal conditions, open-water zones should be sufficiently large to allow wind to disrupt the duckweed mat, facilitating necessary reaeration Additionally, the decomposition of duckweed may introduce an unexpected seasonal nitrogen load to the system.

When designing wetlands in arid climates, it's crucial to account for water losses due to evapotranspiration (ET), especially during warm summer months In the western United States, where water use is governed by appropriative laws, compensating for water lost is essential to protect the rights of downstream users Summer evaporative losses can diminish water volume, leading to increased pollutant concentration despite effective treatment For design calculations, the ET rate is typically estimated at 80% of the local pan evaporation rate, which aligns with lake evaporation rates While past debates questioned the impact of plants on evaporation, it is now widely accepted that the shading from emergent or floating plants reduces direct evaporation, although they still transpire, resulting in a net ET rate similar with or without plants Previous studies indicated high ET rates for certain emergent species, but these findings were based on small-scale experiments and may not accurately reflect full-scale wetland systems.

In SSF wetlands, continuous inundation creates anaerobic conditions that are generally unfavorable for most plant species However, emergent plants have adapted to thrive in these environments by absorbing oxygen and essential gases through their leaves and above-water stems They possess large gas vessels that transport these gases to their roots, allowing them to maintain aerobic conditions despite the surrounding anaerobic soil Research indicates that these plants can transfer between 5 and 45 grams of oxygen per square meter daily, influenced by factors such as plant density and oxygen stress levels, with a more typical transfer rate being around 4 grams per square meter.

Performance Expectations

Wetland systems are highly effective in treating elevated levels of BOD, TSS, nitrogen, and various contaminants such as metals, trace organics, and pathogens However, phosphorus removal is limited due to insufficient soil contact The primary treatment mechanisms involve sedimentation, chemical precipitation, adsorption, and microbial interactions with BOD and nitrogen, alongside some uptake by vegetation Even without harvesting, decomposing vegetation contributes to the formation of peat in wetlands, with nutrients and substances in this refractory organic matter being considered permanently removed from the system.

The rapid removal of settleable organics in wetland systems is attributed to the calm conditions found in Free Water Surface (FWS) systems, along with deposition and filtration processes in Subsurface Flow (SSF) systems Comparable outcomes have also been noted in overland flow systems, as discussed in Chapter [insert chapter number].

In constructed wetlands, approximately 50% of the applied Biochemical Oxygen Demand (BOD) is removed within the initial meters of the treatment slope The settled BOD then undergoes either aerobic or anaerobic decomposition, depending on the local oxygen levels As the wastewater interacts with the attached microbial growth, remaining BOD in colloidal and dissolved forms continues to be eliminated Aerobic activity is prominent near the water surface in Free Water Surface (FWS) systems and at aerobic microsites in Subsurface Flow (SSF) systems, while anaerobic decomposition dominates in other areas of the system Data on BOD removals in FWS constructed wetlands can be found in Table 6.1.

The primary mechanisms for removing total suspended solids (TSS) include flocculation and sedimentation in the bulk liquid, as well as filtration through mechanical straining, chance contact, impaction, and interception within detritus interstices Most settleable solids are effectively removed within 50 to 100 feet (15 to 30 meters) of the inlet For optimal TSS removal, a dense vegetation stand is essential to enhance sedimentation and filtration while inhibiting algae regrowth Algal solids typically require a detention time of 6 to 10 days for effective removal The removal rates of TSS in constructed wetlands can be found in Table 6.2.

Biochemical Oxygen Demand (BOD) Removal in Free Water Surface

Arcata, California 26 12 54 Gearheart et al (1989)

Cle Elum, Washington 38 8.9 77 Smith et al (2002)

Iselin, Pennsylvania 140 17 88 Watson et al (1989)

Listowel, Ontario, Canada 56.3 9.6 83 Herskowitz et al (1987)

West Jackson County, Mississippi 25.9 7.4 71 USEPA (1993)

Sacramento County, California 24.2 6.5 73 Nolte Associates (1999)

Total Suspended Solids (TSS) Removal in Free Water Surface Constructed Wetlands

Arcata, California 30 14 53 Gearheart et al (1989)

Cle Elum, Washington 32 4.8 85 Smith et al (2002)

Iselin, Pennsylvania 380 53 86 Watson et al (1989)

9.2 7.1–11.9 23–29 a Nolte Associates (1999) a Effluent collection via surface overflow weir from open water zone contributed to floating solids in the effluent.

Nitrogen removal in constructed wetlands primarily occurs through the processes of nitrification and denitrification, with plant uptake contributing only about 10% to this removal These microbial reactions are influenced by temperature and detention time, and nitrifying organisms thrive in environments with sufficient oxygen and surface area Consequently, they are often scarce in heavily loaded systems or newly constructed wetlands lacking plant cover Field studies indicate that one to two growing seasons are typically required to establish adequate vegetation for effective nitrification Denitrification, on the other hand, necessitates sufficient organic matter, such as plant litter, to facilitate the conversion of nitrate to nitrogen gas Mature free water surface (FWS) wetlands, characterized by reducing conditions from flooding, are particularly conducive to this process When nitrified wastewater is introduced, denitrification of nitrate can occur within a few days, with the efficiency of nitrogen removal largely dependent on nitrate concentration, detention time, and available organic matter In many wetlands treating municipal wastewater, the near-anoxic conditions of the water column promote rapid nitrate reduction Data on nitrogen and ammonia removal can be found in Table 6.3.

Ammonia and Total Nitrogen Removal in Free Water Surface Constructed Wetlands

Source: Adapted from Crites, R.W and Tchobanoglous, G., Small and Decentralized Wastewater Management Systems, McGraw-Hill, New York, 1998. a USEPA (1999). b Nolte Associates (1999). c City of Salem, Oregon (2003).

Phosphorus removal in Free Water Surface (FWS) systems primarily occurs through adsorption, chemical precipitation, and plant uptake While plants rapidly absorb inorganic phosphorus, they release it upon dying, leading to low long-term removal efficiency The effectiveness of phosphorus removal is influenced by soil interactions and detention time; systems with zero discharge or extended detention times can retain phosphorus in the soil or root zone In flow-through wetlands with 5 to 10 days of detention, phosphorus removal typically ranges from 1 to 3 mg/L Seasonal environmental changes, particularly shifts in oxidation-reduction potential (ORP), can cause phosphorus and other constituents to be released The phosphorus removal rate is also affected by the loading rate and the duration of the system's operation, emphasizing the importance of sampling timing and system maturity in evaluating reported removal data.

Heavy metal removal processes closely resemble those used for phosphorus removal, despite the limited data on their specific mechanisms Key methods for removing heavy metals include adsorption, sedimentation, chemical precipitation, and the involvement of aquatic plants.

Phosphorus Removal in Free Water Surface Constructed Wetlands

Average 2.26 4.98 2.36 46 uptake One of the processes that assist in metals removal is burial as metal sulfide precipitates The process is illustrated in Figure 6.2 (USEPA 1999) One metal of concern is mercury Under anaerobic conditions, mercuric ions are biomethylated by microorganisms to methyl mercury, which is the more toxic form of mercury (Kadlec and Knight 1996) A process that may counteract the methylation is precipi- tation with sulfides, as illustrated in Figure 6.2 At Sacramento County, California, the mercury concentrations were reduced by 64% to 4 ng/L (Crites et al 1997) Metals removal depends on detention time, influent metal concentrations, and metal speciation Removal data for heavy metals in the Sacramento County demonstration wetlands, in Brookhaven, New York, and in Prague are presented in Table 6.5 The removal of aluminum, zinc, copper, and manganese with distance down a Prague wetland is shown in Table 6.6.

Constructed wetlands effectively reduce wastewater temperatures when the average daily ambient air temperature is lower than that of the applied wastewater The temperature reduction can be quantified using Equation 6.15 outlined in Section 6.7 Demonstration projects in Sacramento County, California, and Mt Angel, Oregon, illustrate the achieved temperature reductions, as detailed in Table 6.7.

Decreasing redox potential Aerobic zone Anaerobic zone Me 2+

FIGURE 6.2 Metal sulfide burial processes in a wetland (From USEPA, Free Water

Surface Wetlands for Wastewater Treatment: A Technology Assessment, Office of Water

Management, U.S Environmental Protection Agency, Washington, DC, 1999.)

Metals Removal in Free Water Surface Constructed Wetlands

Sources: Data from USEPA, Free Water Surface Wetlands for Wastewater Treatment: A Technology Assessment, Office of Water Management, U.S Environmental Protection Agency, Washington,

DC, 1999; Nolte Associates, Five-Year Summary Report, 1994–1998, Sacramento Constructed

Wetlands Demonstration Project, Sacramento Regional County Sanitation District, Sacramento,

CA, 1999; Hendry, G.R et al., Lowland Recharge Project Operations: Physical, Chemical and

Biological Changes 1975–1978, Final Report to the Town of Brookhaven, Brookhaven National

Laboratory, Brookhaven, NY, 1979. a During the 5 years of monitoring, the influent arsenic dropped from 3.25 to 2.33 μg/L, while the effluent arsenic varied from 2.34 to 3.77 μg/L.

Removal of Metals with Length in a Free Water Surface Constructed

Source: Vymazal, J and Krasa, P., Water Sci Technol, 48(5), 299–305, 2003 With permission.

The removal of trace organic compounds in treatment systems is achieved through volatilization, adsorption, and biodegradation, with adsorption primarily occurring on organic matter In land treatment systems, the removal efficiency exceeds 95%, with few exceptions showing over 90% Constructed wetlands demonstrate even greater effectiveness due to their longer hydraulic retention times (HRT), typically measured in days rather than minutes or hours, allowing for enhanced opportunities for volatilization and adsorption/biodegradation Pilot-scale constructed wetlands with a 24-hour HRT exhibit significant removal rates, which are expected to be even higher and comparable to those in land treatment systems designed for several days of HRT.

Pathogen removal in wetlands is due to the same factors described in Chapter 3 for pond systems, and Equation 3.25 can be used to estimate pathogen removal

Reduction of Temperature through Free Water Surface Constructed Wetlands at Sacramento County, California, and Mt Angel, Oregon

Sacramento County, California Mt Angel, Oregon

In a Out Reduction In b Out Reduction

The average performance data for wetland systems indicates significant effectiveness in removing contaminants, with a five-year average from 1994 to 1998 showing a removal rate of 5.5 for bacteria and 3.0 for viruses (Nolte Associates, 1999) A subsequent four-year average from 1999 to 2002 in Mt Angel, Oregon, supports these findings The additional filtration provided by the plants and litter layer in wetlands enhances the removal efficiency Notably, pilot floating treatment wetland systems (FWS) in Arcata, California, demonstrated impressive results, achieving approximately 95% removal of fecal coliforms and 92% of viruses with a hydraulic retention time (HRT) of around 3.3 days.

A successful wetland treatment system functions as a thriving ecosystem, featuring diverse vegetation and biota The natural life cycles within this ecosystem generate residuals that can be quantified as BOD5, TSS, nitrogen, phosphorus, and fecal coliforms Consequently, achieving a zero effluent concentration of these materials is unattainable, as some residual background concentrations will always be present These typical background concentrations, influenced by system loadings, differ from wastewater constituents and can be higher in wetland systems exposed to nutrient-rich wastewater compared to those receiving clean water Additionally, seasonal variations in plant decomposition and wildlife activity can cause fluctuations in these background concentrations.

Removal of Organic Priority Pollutants in Constructed Wetlands

Compound Initial Concentration (μg/L) Removal in 24 hr (%)

Source: Reed, S.C et al., Natural Systems for Waste Management and Treatment, 2nd ed.,

McGraw-Hill, New York, 1995 With permission.

Potential Applications

This section offers guidance on the diverse applications of constructed wetlands, which can effectively treat various types of wastewater, including municipal and industrial effluents, stormwater runoff, combined sewer overflows (CSO), agricultural runoff, livestock waste, food processing wastewater, landfill leachate, and mine drainage.

Table 6.10 showcases various examples of FWS constructed wetlands The choice between FWS and SSF constructed wetlands for treating municipal wastewater is influenced by the flow volume and specific site conditions of the proposed wetland.

The SSF wetland system necessitates a smaller surface area compared to a FWS wetland for achieving similar effluent goals due to its higher reaction rates for BOD and nitrogen removal However, determining the most cost-effective option for a specific scenario can be complex The ultimate choice hinges on factors such as the availability and cost of suitable land, as well as the expenses associated with acquiring, transporting, and placing the gravel media in the SSF bed.

Background Concentrations of Constituents in Typical

Source: Data from USEPA, Free Water Surface Wetlands for Wastewater

Treatment: A Technology Assessment, Office of Water Management,

U.S Environmental Protection Agency, Washington, DC, 1999;

USEPA, Constructed Wetlands Treatment of Municipal Wastewater, EPA/625/R-99/010, Office of Research and Development, U.S

Environmental Protection Agency, Cincinnati, OH, 2000.

Note: TSS, total suspended solids; BOD, biochemical oxygen demand. a A range from 5 to 12 has been reported for fully covered with emergent vegetation.

Economics is likely to favor the Free Water Surface (FWS) concept for large systems, especially those in remote locations, as the advantages of the Subsurface Flow (SSF) concept may not provide significant benefits Cost trade-offs are expected to favor the FWS concept for design flows exceeding 1 mgd (3785 m³/d), while flows under 0.1 mgd (378 m³/d) may see different considerations Nonetheless, there are instances where the benefits of the SSF system, such as its thermal advantages, justify its use, as demonstrated by a SSF wetland system designed for wastewater treatment in Halifax, Nova Scotia.

Where nitrogen removal to low levels is a project requirement, the use of

In a Subsurface Flow (SSF) system, Phragmites or Scirpus are recommended due to their resilience, particularly against animal damage, making them preferable over Typha Additionally, when facing strict ammonia limits, the nitrifying filter bed (NFB) should be considered as an effective alternative solution.

Incorporating deeper water zones into the FWS concept can significantly increase hydraulic retention time (HRT) in wetlands and improve oxygen transfer from the atmosphere To facilitate this, the deep-water zones should be sufficiently large to allow for the movement of duckweed cover; a permanent duckweed layer can hinder oxygen transfer Additionally, open-water zones, like those illustrated in Cle Elum, Washington, help reduce short-circuiting If deep-water zones exceed 30% of the total area, the system should be designed as a series of wetlands and ponds, following the guidelines outlined in this chapter and Chapter 4 Furthermore, utilizing submerged plant species in these zones can enhance habitat value and potentially improve water quality, provided that water depth aligns with the sunlight needs of the chosen plants and prevents the formation of a duckweed mat.

Municipal Free Water Surface Constructed Wetlands in the United States

Location Pretreatment Flow (mgd) Area (ac) Remarks

Arcata, California, features oxidation ponds with a capacity of 2.3 acres, attracting tourists despite being early research sites In Benton, Kentucky, oxidation ponds have been upgraded with a nitrification filter bed (NFB) for effective ammonia removal, covering 1.0 acres Cle Elum, Washington, utilizes aerated ponds that incorporate alternating vegetated and open-water zones, occupying 0.55 acres Lastly, Gustine, California, boasts aerated ponds with a significant capacity of 1.0 acres, designed to handle high organic loading.

Mt Angel, Oregon Oxidation ponds 2.0 10 Seasonal discharge

Ouray, Colorado Aerated ponds 0.36 2.2 Polishing wetlands

Riverside, California Secondary 10.0 50 Denitrification wetlands Sacramento County,

Secondary 1.0 15 Five-year demonstration project

Water depth 18” Summer 24” Winter Type B fill Type A fill Type A fill Type C fill

NH O 2 4 + N 2 Open water zone Emergent plant zone

NH O 2 4 + N 2 Open water zone Emergent plant zone

FIGURE 6.3 Open-water sketch for free water surface (FWS) wetlands (Courtesy of Brown and Caldwell, Walnut Creek, CA.)

FIGURE 6.4 Free water surface (FWS) wetland at Cle Elum showing bulrush and open water.

Conducting a thorough thermal analysis is essential for systems operating in subfreezing temperatures during winter to ensure optimal performance of temperature-sensitive nitrogen and BOD removal processes This analysis helps identify potential risks of restrictive freezing in extremely cold climates Several FWS systems developed for northwestern Canada have been specifically engineered to mitigate the challenges posed by severe winter freezing, incorporating strategies for winter wastewater storage in lagoon and wetland applications during warmer months.

Integrating habitat and recreational values into the FWS wetland concept is highly effective due to the open water surface, which draws birds and wildlife The inclusion of deep-water zones with nesting islands will greatly improve habitat quality, complemented by the strategic planting of beneficial vegetation like sago pondweed.

SSF and FWS wetlands can effectively treat commercial and industrial wastewater, similar to municipal wastewater, but require careful wastewater characterization These wastewaters may vary in strength, nutrient content, pH levels, and can contain toxic substances that affect biological treatment in wetlands Typically, high-strength waste and priority pollutants undergo anaerobic treatment before entering the wetland system Constructed wetlands, including both SSF and FWS types, are utilized for treating wastewater from various industries such as pulp and paper, oil refining, chemical production, and food processing, often serving as a polishing step following conventional biological treatment Design procedures for these systems are outlined in previous sections, and pilot studies may be essential for addressing unfamiliar toxic substances or optimizing the removal of priority pollutants.

Wetlands designed for urban stormwater treatment primarily focus on sediment removal from runoff originating from parking lots, streets, and landscapes Functioning as stormwater retention basins, these wetlands incorporate vegetation and adhere to fundamental sedimentation basin design principles The inclusion of vegetative fringes, along with deep and shallow water zones and marsh areas, not only improves water treatment efficiency but also enhances habitat quality Research has demonstrated that these wetlands effectively reduce levels of biochemical oxygen demand (BOD), total suspended solids (TSS), pH, nitrates, phosphates, and trace metals.

A stormwater wetland system (SWS) typically incorporates a mix of deep ponds and shallow marshes, along with wet meadows and shrub areas Due to variable flow rates and the risk of inorganic solid accumulation and clogging, the subsurface flow (SSF) wetland approach is not suitable for this application Instead, the marsh component of the SWS is generally designed using free water surface (FWS) constructed wetlands.

Key components of the system include an inlet structure, a sedimentation basin, a spreader swale or weir for lateral flow distribution, a deep pond, and an outlet device to manage overflow during peak storm events while allowing for gradual discharge to the established datum water level This datum level is crucial for maintaining shallow water depths in marsh areas Incorporating drought-resistant plant species in these marsh components enables the system to withstand extended periods of complete dewatering.

Typha, Scirpus, and Phragmites can tolerate temporary inundation levels of up to 3 feet (1 meter), which determines the maximum water level before overflow in the SWS when these species are utilized For grassed wet meadows and shrubs, the ideal maximum storage depth should be around 2 feet (0.6 meters) The wetland's optimum storage capacity, defined as the depth between the datum and the overflow level, should correspond to a volume of 0.5 inches (13 mm) of water from the contributing watershed To ensure effective performance, the minimum storage volume should be at least 0.25 inches (6 mm) of water from the watershed These storage volumes, along with any other depths, can be calculated using Equation 6.1.

V = storage volume in stormwater wetland (ft 3 ; m 3 )

C = coefficient = 360 for U.S units; 10 for metric units y = design depth of water on watershed (mm)

A ws = surface area of watershed (ac; ha)

The minimum surface area of the entire SWS at the overflow elevation is based on the flow occurring during the 5-year storm event and can be calculated with Equation 6.2:

A sws = minimum surface area of SWS at overflow depth (ft 2 ; m 2 )

C = coefficient = 180 for U.S units; 590 for metric units

Q = expected flow from 5-year design storm (ft 3 /d; m 3 /d)

For optimal performance, the aspect ratio of the SWS should ideally be around 2:1, with the inlet positioned as far from the outlet as feasible Utilizing appropriate baffles can also enhance flow management Additionally, the spreader swale and inlet zone need to be adequately wide to effectively decrease the flow velocity.

Planning and Design

The planning and design of wetland treatment systems must consider habitat and recreational potential alongside conventional wastewater treatment factors These systems can serve multiple purposes, from wastewater treatment to environmental enhancement and public recreation Clearly defining the intended functions early in the project development is crucial for feasibility evaluation and cost-effective implementation While wetland systems, including gravel-bed subsurface flow types, will naturally attract wildlife, treatment-focused designs may minimize habitat values To enhance public recreation and attract specific wildlife, special features can be integrated, ensuring that hazardous conditions are avoided An effective strategy combines dedicated treatment wetland units initially, transitioning to units that enhance habitat and recreational value as water quality improves.

Location Wetlands Area, Acres Comments

American Canyon, CA 8.5 Wildlife habitat

Arcata, CA 31 Research/wildlife habitat

Cannon Beach, OR 15 Natural wetlands

Davis, CA 400 Stormwater and effluent/wildlife habitat

Fort Deposit, AL 7.5 Pond upgrade

Grand Strand, SC 702 Carolina Bays

Hillsboro, CA 15 Research at Jackson Bottoms

Houghton Lake, MI 1500 Natural wetlands

Incline Village, NV 390 Zero discharge

Las Gallinas Valley, CA 20 Marsh and ponds

Pacifica, CA 8.7 Stream and riparian wetlands restoration/ endangered species

Petaluma, CA 60 Treatment and wildlife habitat

Showlow, AZ 201 Pinyail Lake/Redhead Marsh

Vermontville, MI 11.5 Former land treatment site

West Jackson Co., MS 56 Former land treatment site

Source: Adapted from USEPA, Constructed Wetlands for Wastewater Treatment and Wildlife Habitat, EPA832-R-93-005, U.S Environmental Protection Agency, Washington, DC,

1993; Crites, R.W., Environmental Enhancement in Bay Area Constructed Wetlands using Recycled Water Proceedings of the 2009 California WateReuse Conference, San Francisco,

Site evaluation criteria for wetlands and other natural systems are given in Chapter

2 The ideal site for a wetland would be within a reasonable distance from the waste- water sources and at an elevation permitting gravity flow to the wetland, between the wetland cells, and to the final discharge point The site would be available at a reasonable cost, would not require extensive clearing or earthwork for construction, would have a deep nonsensitive groundwater table, and would contain subsoils that, when compacted, would provide a suitable liner Any divergence from these ideal characteristics will result in increased project costs The possible future expansion of the system should also be given consideration during the planning and site evaluation effort The 56-ac (23-ha) FWS wetland system in West Jackson County, Mississippi, was constructed in 1990/1991 with a design capacity of 1.6 mgd Because of rapid community growth it was necessary to expand the system capacity to 4 mgd with 50 ac (20.2 ha) of new wetland construction in 1997/1998 This expansion was possible because sufficient land was available adjacent to the original wetland system.

In the United States, all wetland treatment systems require some form of preliminary treatment, ranging from primary to tertiary levels, depending on the wetland's functional intent, public exposure, and habitat protection needs For municipal wastewater, the minimum preliminary treatment is equivalent to primary treatment, often achieved through septic tanks or Imhoff tanks for smaller systems, or a pond unit for larger systems To ensure public safety and promote wildlife habitats, secondary treatment is recommended before allowing public access to wetland components This can be implemented in a first-stage wetland unit with restricted access In cases where the preservation of natural habitats and ecosystems is crucial, tertiary treatment with nutrient removal may be necessary before discharging into natural wetlands Additionally, stormwater wetlands commonly incorporate features like trash racks and forebays to facilitate the settling and removal of large debris from runoff Wetlands designed for treating mine drainage may also require preliminary units for adjusting pH or alkalinity.

Constructed wetland systems function as attached-growth biological reactors, allowing for the estimation of their performance using first-order plug-flow kinetics for the removal of biochemical oxygen demand (BOD) and nitrogen This chapter presents design models specifically for the removal of BOD, total suspended solids (TSS), ammonia nitrogen, nitrate, total nitrogen, and phosphorus in both free water surface (FWS) and subsurface flow (SSF) wetlands Additionally, alternative models from various sources are included for comparative analysis, as there is no universal agreement on the optimal design approach The fundamental relationship governing plug-flow reactors is articulated in Equation 6.8.

K T = temperature-dependent, first-order reaction rate constant (d −1 ) t = hydraulic residence time (d)

The hydraulic residence time in the wetland can be calculated with Equation 6.9: t = LWyn/Q (6.9) where

L = length of the wetland cell (ft; m)

W = width of the wetland cell (ft; m) y = depth of water in the wetland cell (ft; m) n = porosity, or the space available for water to flow through the wetland

In the FWS wetland, vegetation from the litter occupies a significant area, while in the SSF scenario, the media, roots, and other solids also take up space The porosity of these systems is represented as a percentage, expressed in decimal form.

Q = the average flow through the wetland (ft 3 /d; m 3 /d):

To accurately account for water losses or gains from seepage and precipitation in wetlands, the average flow must be determined using Equation 6.10 A conservative design approach may assume no seepage while utilizing local records to estimate monthly evapotranspiration losses and rainfall gains This necessitates an initial assumption about the wetland's surface area to calculate the volume of water lost or added Typically, for preliminary design estimates, it is reasonable to assume that the outflow (Q out) equals the inflow (Q in).

It is then possible to determine the surface area of the wetland by combining Equation 6.8 and Equation 6.9:

( ) ln( 0 ) (6.11) where A s is the surface area of wetland (ft 2 ; m 2 ) The value used for K T in Equation

6.1 or Equation 6.4 depends on the pollutant that must be removed and on the tem- perature; these aspects are presented in later sections of this chapter.

Because the biological reactions involved in treatment are temperature-dependent it is necessary, for a proper design, to estimate the water temperature in the wetland

The performance and feasibility of FWS wetlands in extremely cold climates are significantly affected by ice formation, which can lead to complete freezing of shallow wetlands and halt effective treatment This chapter provides calculation methods for estimating water temperatures within the wetland and determining the thickness of ice formation.

The hydraulic design of wetlands is crucial for effective pollutant removal, as it relies on the critical plug-flow assumption, which requires uniform flow and minimal short-circuiting Early designs of subsurface flow (SSF) and free water surface (FWS) wetlands often overlooked these hydraulic requirements, leading to unexpected flow conditions and negative impacts on performance By following the hydraulic design procedures outlined in this chapter, these issues can be effectively mitigated.

Designing an effective wetland system necessitates careful consideration of hydraulics, thermal dynamics, and pollutant removal kinetics The design process is typically iterative, requiring initial assumptions about water depth and temperature to solve kinetic equations that estimate the necessary wetland area for pollutant removal The pollutant that demands the largest area for effective removal becomes the limiting design parameter (LDP), ultimately dictating the wetland's size Once the wetland area is established, thermal equations can be applied to calculate the theoretical water temperature within the wetland.

If the initially assumed water temperature does not match the calculated temperature, additional iterations of the calculations are necessary until both values align The final step involves applying the correct hydraulic calculations to establish the ultimate aspect ratio (length-to-width) and flow velocity within the wetland.

If these final values differ significantly from those assumed for the thermal calcula- tions, further iterations may be necessary.

Hydraulic Design Procedures

The hydraulic design of constructed wetland systems is essential for their effective performance, relying on uniform flow conditions and optimal contact between wastewater and treatment organisms In subsurface flow (SSF) wetlands, maintaining subsurface flow is crucial throughout the system's lifespan, necessitating meticulous hydraulic design and construction methods Flow must overcome frictional resistance from vegetation and accumulated solids, with energy derived from the head differential between the inlet and outlet While a sloping bottom can contribute to this differential, it is not a sustainable solution alone, as resistance may increase over time Therefore, the design should include a sufficiently sloped bottom for complete drainage and an adjustable outlet to manage water levels, ensuring the necessary water surface slope for effective drainage.

The selected aspect ratio of a wetland significantly impacts its hydraulic regime and flow resistance Early designs of Free Water Surface (FWS) systems recommended high aspect ratios of at least 10:1 to achieve plug-flow conditions and prevent short-circuiting However, this approach can lead to increased flow resistance as the length of the flow path grows For instance, a California FWS system with a 20:1 aspect ratio faced overflow issues due to accumulating vegetative litter Acceptable aspect ratios range from less than 1:1 to about 3:1 or 4:1 To minimize short-circuiting, careful wetland bottom construction, the use of multiple cells, and the inclusion of intermediate open-water zones for flow redistribution are essential strategies, which will be elaborated on in later sections.

A treatment wetland is a shallow body of moving water characterized by a large surface area Its hydraulic design is complex due to significant frictional resistance from plants and litter in free water surface (FWS) systems, as well as gravel media in subsurface flow (SSF) systems Designers assume uniform water movement at a predictable rate across the wetland’s surface However, factors such as precipitation, evaporation, evapotranspiration, and seepage complicate this assumption by influencing the water volume, pollutant concentration, and hydraulic retention time (HRT) within the wetland.

Manning’s equation is a widely recognized model for analyzing water flow in FWS wetland systems According to the equation, flow velocity (v) is influenced by water depth (y), the hydraulic gradient (s), and Manning’s coefficient (n) The relationship is expressed as v = (1/n)(y^(2/3))(s^(1/2)), where v is measured in feet per second (ft/s) or meters per second (m/s), y in feet (ft) or meters (m), and s in feet per foot (ft/ft) or meters per meter (m/m).

In many applications of Manning's equation, flow resistance is primarily influenced by the bottom and submerged sides of an open channel, with numerous published n coefficient values available in technical literature However, in Free Water Surface (FWS) wetlands, flow resistance is affected throughout the entire water depth due to emergent vegetation and litter The relationship between the Manning number (n) and the resistance factor (a) is expressed as n = a/y^1/2, where 'a' represents the resistance factor Reed et al (1995) provided specific values for the resistance factor 'a' in FWS wetlands.

• Sparse, low-standing vegetation — y > 1.2 ft (0.4 m), a = 0.487 sãft 1/6 , (0.4 sãm 1/6 )

• Moderately dense vegetation — y ≥ 1.0 ft (0.3 m), a = 1.949 sãft 1/6 , (1.6 sãm 1/6 )

• Very dense vegetation and litter — y < 1.0 ft (0.3 m), a = 7.795 sãft 1/6 , (6.4 sãm 1/6 )

Dombeck et al (1997) experimentally confirmed the energy needed to overcome resistance in constructed wetlands, which is primarily provided by the head differential between the inlet and outlet water surfaces To optimize this differential, wetlands can be designed with a minimal slope for complete drainage and adjustable outlet structures to manage water levels over time Additionally, the aspect ratio of a Free Water Surface (FWS) wetland significantly affects its hydraulic regime, as increased length leads to greater flow resistance Reed et al (1995) proposed a model to estimate the maximum desirable length for FWS wetland channels.

L = maximum length of wetland cell (m)

A s = design surface area of wetland (m 2 ) y = depth of water in the wetland (m) m = portion of available hydraulic gradient used to provide the necessary head

(% as a decimal) a = resistance factor (sãm 1/6 )

Q A = average flow through the wetland (m 3 /d) = (Q IN + Q OUT )/2

For optimal design, an initial m value of 10 to 20% is recommended to establish a safety reserve Typically, this model yields an aspect ratio of 3:1 or less Equation 6.14 incorporates the average flow (Q A ) to account for the effects of precipitation, evapotranspiration, and seepage on wetland flow Additionally, the design surface area (A s ) in Equation 6.14 represents the bottom area of the wetland, as defined by the pollutant removal models discussed later in this chapter.

Thermal Aspects

Temperature conditions in wetlands significantly influence both physical and biological processes within the ecosystem Prolonged low temperatures and ice formation can lead to the physical degradation of wetlands While biological reactions crucial for biochemical oxygen demand (BOD) removal, nitrification, and denitrification are temperature-dependent, many existing wetland systems in cold climates have not shown a consistent relationship between temperature and BOD removal efficiency.

The long hydraulic residence time in certain systems helps offset lower reaction rates during winter, yet many systems in Canada and the United States still experience reduced nitrogen removal capabilities in colder months This decrease is attributed to the impact of lower temperatures on biological reactions and oxygen depletion due to ice cover forming on the water's surface.

This section outlines reliable methods for estimating water temperature in wetlands, essential for the effective application of biological design models related to BOD and nitrogen removal It details calculation techniques for determining water temperature in Subsurface Flow (SSF) and Free Water Surface (FWS) wetlands, as well as predicting potential ice thickness on FWS wetlands.

In northern regions, FWS wetlands may experience temporary ice formation due to exposure to subfreezing temperatures, which can act as a thermal barrier, slowing the cooling of the water beneath Unlike ponds and lakes where ice floats freely, ice in FWS wetlands may be anchored by vegetation, potentially reducing flow volume as it thickens In extreme cases, this can lead to ice cracking and surface flow, resulting in freezing and halting biological treatment activities until warmer weather returns To mitigate these risks, constructed wetlands in areas with prolonged low temperatures may utilize seasonal components, such as storing wastewater in lagoons during winter Successful operations have been documented in regions like Ontario and Iowa, highlighting the importance of conducting thermal analyses for projects in northern climates to ensure winter stability and maintain suitable water temperatures for biological processes.

(1986) with the assistance of Darryl Calkins (USA CRREL; Hanover, NH) The procedure has three parts:

1 Calculate water temperatures in the wetland until conditions that allow ice for- mation (3°C water temperature) commence Separate calculations are required for densely vegetated wetland segments and for large area open-water zones.

2 Water temperature calculations are then continued for the ice-covered case.

3 An estimate is made of the total depth of ice that may form over the period of concern.

The temperatures assessed in the initial steps are crucial for evaluating the feasibility of a FWS wetland at the proposed site and for validating temperature assumptions used in sizing the wetland according to BOD or nitrogen removal models These models are foundational in the design process, as they inform the necessary wetland dimensions, hydraulic retention time (HRT), and flow velocity for subsequent thermal calculations Additionally, the estimated total ice depth from the third step offers insights into the wetland's viability and helps establish the required operating water depth during winter months.

6.7.1 c ase 1 f ree W ater s urface W etLand p rior to i ce f ormation

Equation 6.15 is used to calculate the water temperature at the point of interest in the wetland Experience has shown that ice formation commences when the bulk temperature in the liquid approaches 3°C (37°F) because of density differences and convection losses at the water surface (Ashton 1986; Calkins 1995) Equation 6.15 is therefore repeated until a temperature of 3°C is reached or until the end of the wetland cell is reached, whichever comes first If a temperature of 3°C is reached prior to the end of the wetland, then Equation 6.17 is used to calculate the temperatures under an ice cover If the wetland is composed of vegetated zones interspersed with deeper open water zones, Equation 6.15 must be used sequen- tially, with the appropriate heat-transfer coefficient (U s ) to calculate the water temperatures:

T w = T air + (T 0 − T air ) exp [−U s (x − x 0 )/δyvc p ] (6.15) where

T w = water temperature (°C; °F) at distance x (m; ft)

T air = average air temperature during period of interest (°C; °F)

T 0 = water temperature (°C; °F) at distance x 0 (m; ft) the entry point for the wet- land segment of interest

U s = heat-transfer coefficient at the wetland surface (W/m 2 ã°C; Btu/ft 2 ãhrã°F)

= 1.5 W/m 2 ã°C (0.264 Btu/ft 2 ãhrã°F) for dense marsh vegetation

= 10–25 W/m 2 ã°C (1.761–4.403 Btu/ft 2 ãhrã°F) for open water; high value used for windly conditions with no snow cover c p = specific heat = 1.007 Btu/lb °F (4215 J/kg °C)

In the first iteration, if the final effluent temperature from the wetland is below 37°F (3°C), Equation 6.15 can be manipulated to determine the distance x at which the temperature reaches 37°F This is expressed by the formula: x − x₀ = −[δyvcₚ /Uₛ][ln (37 − Tₐᵢʳ)/(T₀ − Tₐᵢʳ)].

Calculate the water temperature in a three-stage FWS constructed wetland:

Stage 1: Length 300 ft, depth 1 ft, densely vegetated, flow velocity 0.2 ft/hr Stage 2: Deep open-water zone, length 100 ft, depth 4 ft, flow velocity 0.1 ft/hr Stage 3: Same as stage 1

Air temperature, 49°F; influent wastewater temperature, 70°F

1 Use Equation 6.15 to calculate the temperature at the end of stage 1:

2 Calculate the water temperature at the end of stage 2:

3 Calculate the water temperature at the end of the wetland:

6.7.2 c ase 2 f LoW under an i ce c over

When ice forms on water, the heat transfer from the water to the ice occurs at a constant rate, unaffected by air temperature or snow cover The ice surface remains at 32°F (0°C) until all water beneath it is frozen While air temperature and snow presence influence the rate of ice formation, they do not affect the cooling rate of the underlying water The temperature of wetland water under an ice cover can be estimated using a specific equation, which is similar to another equation but adjusted for the presence of ice.

T 0 = water temperature at distance x 0, assuming 37.4°F (3°C) where an ice cover commences

U i = heat-transfer coefficient at ice/water interface (Btu/ft 2 hr °F; W/m 2 °C)Other terms are as defined previously.

The U i value in Equation 6.17 depends on the depth of water beneath the ice and the flow velocity:

U i = heat-transfer coefficient at ice/water interface (Btu/ft 2 hr °F; W/m 2 °C) Φ = proportionality coefficient = 0.0022 Btu/ft 2.6 hr 0.2 °F (1622 J/m 2.6 s 0.2 °C) v = flow velocity (ft/hr; m/s), assuming no ice conditions y = depth of water (ft; m)

6.7.3 c ase 3 f ree W ater s urface W etLand and t Hickness of i ce f ormation

Ice begins to form on the surface of the FWS wetland when the bulk water temperature drops to 37.4°F (3°C) and continues to develop as long as temperatures are at or below 32°F (0°C) In northern regions with prolonged low air temperatures, the FWS wetland may not be suitable for year-round treatment due to the risk of extensive ice formation leading to system failure The thickness of ice that forms over a 24-hour period can be estimated using a specific equation, which takes into account factors such as time, temperature differences, and the properties of ice.

The average air temperature during the specified time period is denoted as T air (°F or °C) The snow cover depth is represented by y s (in feet or meters), while the depth of daily ice formation is indicated by y i (also in feet or meters) Additionally, the conductivity of snow and ice is represented by k s and k i, respectively, with values sourced from Table 6.18.

U s = heat transfer coefficient at the wetland surface, W/m 2 ã°C; (Btu/ft 2 ãhrã°F)

= 1.5 W/m 2 ã°C (0.264 Btu/ft 2 ãhrã°F) for dense marsh vegetation

= 10–25 W/m 2 ã°C (1.761–4.403 Btu/ft 2 ãhrã°F) For open water; high value used for windy conditions with no snow cover

U i = heat transfer coefficient water to ice (from Equation 6.18)

T w = average water temperature during period of concern (from Equation 6.17)

To accurately assess ice formation, it is essential to recalculate for each day of interest, adjusting the depth of ice and snow accordingly The relevant time frame for previous FWS thermal models corresponds to the design Hydraulic Retention Time (HRT) for the wetland, which may extend to the entire winter season if prolonged subfreezing temperatures occur A practical initial estimate of potential ice formation can be obtained by utilizing the average monthly air temperatures from the coldest recorded winter during the period in question This model is based on research by Ashton (1986) with contributions from Darryl Calkins of USA CRREL in Hanover, NH.

The highest rate of ice formation occurs on the first day of freezing, as the absence of an ice cover or snow layer allows for maximum heat loss For estimation purposes, the final term in Equation 6.19 can typically be disregarded, simplifying it to the Stefan formulation (Stefan 1891): y = m [T m − T air (t)] 0.5 In this equation, y represents the depth of ice formed over time period t, while m is the proportionality coefficient.

= 0.066 ft/°F 0.5 ãd 0.5 (0.027 m/°C 0.5 ãd 0.5 ) for open-water zones with no snow

= 0.024 ft/°F 0.5 ãd 0.5 (0.010 m/°C 0.5 ãd 0.5 ) for wetland with dense vegetation and litter

= 0.044 ft/°F 0.5 ãd 0.5 (0.018 m/°C 0.5 ãd 0.5 ) for open-water zones with snow

T air = average air temperature during time period t (°F; °C) t = number of days in the period of interest (d)

Thermal Conductivities for Wetland Components

Source: Reed, S.C et al., Natural Systems for Waste Management and Treatment,

The native soil beneath the wetland bed plays a crucial role in thermal dynamics, facilitating heat transfer into the wetland during winter months and releasing heat during the summer.

3-ft (1-m) depth for this soil layer.

Equation 6.20 can be used to estimate total ice formation on FWS wetlands over the entire winter season or for shorter time periods if desired This equation can be used to determine the feasibility of winter operations for a wetland in locations with very low winter temperatures For example, a site with persistent air temperatures at –13°F (–25°C) would result in a wetland that is 1.5 ft (0.45 m) deep freezing to the bottom in about 84 days.

Design Models and Effluent Quality Prediction

Recent advancements in constituent removal design procedures have led to the comparison of three models in the Water Environment Federation’s Manual of Practice (WEF 2010) These models, based on wetland input/output data and mass balance relationships, utilize a first-order plug-flow model but do not fully capture the complex reactions in wetlands, relying instead on a lumped apparent rate constant to represent concentration changes (Tchobanoglous et al 2003) While fundamentally equivalent and expected to yield similar results, discrepancies arise due to variations in data sets and model structures The models are categorized into volumetric models, as developed by Reed et al (1995) and Crites and Tchobanoglous (1998), and areal loading models by Kadlec and Knight (1996), each with distinct advantages and limitations.

• The design is based on average flow through the system This allows compen- sation for water losses and gains due to precipitation and evapotranspiration.

• The safety factors and irreducible background concentrations are treated as external boundary conditions and have no limiting impact on the math- ematical results of design models.

Understanding water depth is crucial for the procedure, but managing this aspect can be challenging during the construction of large systems Additionally, water depth may vary over time, complicating long-term control.

Understanding the porosity of vegetation and accumulated litter is crucial, as the assumed design values are derived from a limited database It is important to note that these values may fluctuate over time, highlighting the need for ongoing assessment and adaptation in design considerations.

The removal of Biochemical Oxygen Demand (BOD) in wastewater treatment is often thought to be influenced by temperature, drawing on insights from various treatment processes However, evidence from numerous operational wetland systems indicates that BOD removal does not consistently show temperature dependence.

The design calculations for wetland models focus on the mass loading of the wetland surface area, making water depth a non-factor, even in large systems where measuring depth can be challenging.

• These models are more flexible mathematically It is possible to produce a better fit of existing data with this two-variable (K, C*) model as compared to the single-variable (K) volumetric models.

• The models deal only with input wastewater volume (Q) This does not allow for compensation for water gains and losses in the design calculations.

The FWS database utilized for model development comprises numerous lightly loaded polishing wetland systems Relying on this data may lead to the calculation of low-valued rate constants (K), potentially resulting in overly large designs for wetland systems.

• The internal position of the background concentration (C*) and safety factor (z) terms in the models for determining wetland area may result in excessive wetland sizes to achieve low concentrations.

The volumetric and detention time models outlined in Table 6.19 can be utilized to predict the necessary land area and effluent quality Additionally, Table 6.20 provides design equations based on the areal loading rate approach.

K T = rate constant at temperature T (d −1 ) θ = temperature coefficient at 20°C

T w = average water temperature in wetland during period of concern (°C)

A s = treatment area (bottom area) of wetland (m 2 )

Q A = average flow of the wetland (m 3 /d) = (Q IN + Q OUT )/2 y = average depth of water in the wetland (m) n = porosity of the wetland (% as a decimal)

Note: (1) The effluent concentration (C e ) cannot be less than the background concentrations listed below (2) The average flow (Q A ) accounts for water gains and losses from precipitation, evapo- transpiration, seepage, etc.

C e /C 0 = [0.1139 + 0.00213 (HLR)] where HLR = hydraulic loading rate (mm/d × 0.1), and TSS removal is not dependent on temperature Background concentration (mg/L) = 6

K NH is a rate constant at 20°C for FWS wetlands (d –1 ).

Note: It is prudent to assume that all TKN (from municipal wastewater) entering the wetland can appear as ammonia, so assume C 0 for ammonia is equal to influent TKN.

In designing for nitrate removal, it is conservative to assume that all ammonia eliminated in the preceding step can be converted to nitrate Therefore, the concentration for nitrate removal (C0) should be calculated as the sum of the concentration of ammonia removed (Ce) and any existing nitrate in the influent.

Effluent TN = C e ( NO 3 ) + ( C e ( NH 4 ) − C e ( NO 3 ) )

Note: A specific model for total nitrogen removal is not available in this set The effluent total nitrogen

(TN) can be estimated as the sum of residual ammonia and remaining nitrate (C 0 – C e ).

C e /C 0 = exp (−K p /HLR) (6.24) where HLR is the average hydraulic loading rate (cm/d), and total phosphorus removal is not dependent on temperature.

C e /C 0 (MPN/100 mL) = [1/(1 + K T (t/d))] x (6.25) where t, d = HRT in the system, and x = number of wetland cells in series.

Note: This model was developed for facultative ponds and is believed to give a conservative estimate for fecal coliform removal in both FWS and SF wetlands.

The background concentration values for each parameter serve as essential external boundary conditions for the design models in this set It is crucial that none of the models are solved for concentrations below these established background levels.

The key design factor for determining the necessary treatment area is the parameter with the highest removal requirement, such as BOD 5 It is essential to select this area based on the specific needs of the project, ensuring that the wetland effectively addresses all other relevant parameters.

In engineering design projects, applying a safety factor is standard practice, typically implemented after preliminary calculations are finalized For wetland size determination, this safety factor ranges from 15% to 25%, influenced by data uncertainty and performance expectations The choice of safety factor involves engineering judgment and reflects a proportional increase in the calculated treatment area.

Source: Adapted from Reed, S.C et al., Natural Systems for Waste Management and Treatment, 2nd ed.,

McGraw-Hill, New York, 1995 With permission.

Areal-Based Process Design Models

HLR A = annual hydraulic loading rate (m/yr)

K T = rate constant at temperature T (m/yr)

K 20 = rate constant at 20°C (m/yr) θ = temperature coefficient

Q 0 = annual influent wastewater flow rate (m 3 /yr) z = safety factor

The article discusses areal-based models that express hydraulic loading per unit area, contrasting with detention time-based models While detention time (HRT) and hydraulic loading rate (HLR) are related under specific wetland conditions, discrepancies in model results arise from different data sets and the treatment of safety factors The influent flow rate (Q0) and HLR in these models do not account for water gains or losses from precipitation, evapotranspiration, or seepage Additionally, factors such as porosity, water depth, and detention time are excluded from these design models The safety factor (z) reflects the ratio of annual average to maximum monthly pollutant concentrations based on Kadlec and Knight's database Due to the internal positioning of C* and z, designing a system to achieve effluent concentrations matching background levels is impractical, as the required wetland area approaches infinity as the effluent concentration (Ce) nears background concentration (C*).

Note: The treatment area model cannot be solved for BOD 5 effluent values approaching background levels For example

Areal-Based Process Design Models

The treatment area model is ineffective for low effluent concentration values (C e) For instance, to achieve an effluent total suspended solids (TSS) concentration of 15 mg/L, the influent concentration (C 0) must not exceed 17 mg/L to ensure the model can be solved.

The background concentration (C*) is included internally in each design model.

To effectively design a FWS wetland aimed at reducing the biochemical oxygen demand (BOD) from 100 mg/L to 15 mg/L, with a flow rate of 0.9 mgd and a wastewater temperature of 68°F, it is essential to compare the land areas required using two distinct methodologies: the volumetric/detention time approach and the areal loading rate approach This comparison will provide insights into the spatial requirements for optimal wastewater treatment performance in the wetland system.

1 Using Equation 6.23, solve for the land area for the volumetric/detention time approach Use a depth of 2 ft and a porosity of 0.8 Use K T = 0.68:

Use a safety factor of 20%, A = 5.77 ac.

2 Calculate the land area using the hydraulic loading rate method Using Equation 6.24, calculate the land area Convert flow into m 3 /yr; use K T = 34; and set z = 1, C* = 8.8:

The primary distinctions between these models lie in the C* values and the placement of the safety factor (z) within the logarithmic section of the area model For instance, if the safety factor z is set at 0.59 rather than 1, the calculated area in step 2 would amount to 67 acres.

Areal-Based Process Design Models

Physical Design and Construction

Wetland design and construction share fundamental civil engineering principles with shallow lagoons, primarily involving earthen berms for lateral water containment and seepage control at the wetland cell's bottom A distinctive characteristic of wetland systems is the incorporation of vegetation along with inlet and outlet structures that facilitate uniform water flow throughout the wetland.

Berms in wetland cells are designed with a 3:1 slope on the interior and must have at least 2 feet of freeboard above the average water level in a Free Water Surface (FWS) wetland For municipal wetlands, external berms should be a minimum of 10 feet wide at the top to allow service vehicle access Each wetland cell needs an access ramp for maintenance equipment, and it is preferable to balance cut and fill on-site to eliminate the need for remote borrow pits or spoil disposal.

Typical Design Criteria and Expected Effluent Quality for

Free Water Surface Constructed Wetlands

BOD loading rate lb/acãd 1 at the 95th Percentile of the HQ Distribution by Pathway for the Agricultural Land Application Scenario

CASRN Chemical Pathway Receptor HQ

14797-65-0 Nitrite Irrigation of surface water Child 1.1

7440-22-4 Silver Ingestion of milk Adult 3.8

Source: USEPA, Technical Background Document for the Sewage Sludge Exposure and Hazard Screening Assessment, Document No 822-B-03-001, Office of Water, U.S Environmental

On-Site Wastewater Systems

Types of On-Site Systems

While many types of on-site systems exist, most involve some variation of subsur- face disposal of septic tank effluent The four major categories of on-site systems are:

• Modified conventional on-site systems

• On-site systems with additional treatment

The conventional on-site system, comprising a septic tank and a soil absorption system, is the most prevalent method for wastewater treatment The septic tank serves as a pretreatment unit before on-site disposal Modified systems, such as shallow trenches and pressure-dosed systems, enhance this process Alternative disposal methods include mounds, evapotranspiration systems, and constructed wetlands, which may require additional treatment for septic tank effluent Intermittent and recirculating granular-medium filters are often cost-effective solutions for this purpose For areas needing further nitrogen removal, various alternatives can be explored, as detailed in Section 10.4 A summary of disposal and reuse systems for individual on-site setups is provided in Table 10.1.

Native soil backfill Fabric or building paper

12 in minimum 4-in distribution pipe

Side wall absorption area (both sides) 18–24 in min

2-in minimum rock over pipe

6-in minimum rock under pipe

0.75- to 2.5-in.-diameter washed drainrock

FIGURE 10.1 Typical cross section through conventional soil absorption system.

Types of On-Site Wastewater Disposal/Reuse Systems

Gravity leachfields/conventional trench Most common system

Deep trench To get below restrictive layers

Shallow trench Enhanced soil treatment

Conventional trench To reach uphill fields

Shallow trench Uphill and shallow sites

Drip application Following additional treatment of septic tank effluent; to optimize use of available land area

Sand-filled trenches Added treatment

At-grade systems Less expensive than mounds

Constructed wetlands Requires a discharge or subsequent infiltration

Effluent Disposal and Reuse Options

Alternative infiltration systems (presented in Table 10.2) have been developed to overcome restrictive conditions such as:

On-site system effluent can be reused through various methods, including drip irrigation, spray irrigation, groundwater recharge, and toilet flushing Drip irrigation is increasingly favored for water reuse, while spray irrigation is ideal for handling larger flows from commercial, industrial, and small community sources Groundwater recharge is applicable in regions with deep permeable soils, both of which are explored in detail in Chapter 8.

Site Evaluation and Assessment

Selecting an appropriate on-site location for disposal involves several key steps, including identification, reconnaissance, and assessment It starts with a comprehensive analysis of soil characteristics such as permeability, depth, texture, structure, and pore sizes, which are crucial for evaluating the site Additionally, factors like groundwater depth, site slope, existing landscape and vegetation, and surface drainage features play a significant role in the assessment process Once a potential site is identified, the evaluation continues in two phases: a preliminary site evaluation followed by a detailed site assessment.

Types of On-Site Wastewater Disposal/Reuse Systems

Drip irrigation Usually follows added treatment

Holding tanks Seasonal use alternative

Surface water discharge Allowed in some states following added treatment

When considering on-site disposal methods to address site constraints, various factors such as soil permeability, bedrock, groundwater levels, and slope must be evaluated For small lot sizes with varying soil drainage rates, options include trenches, beds, pits, mounds, fill systems, and sand-lined trenches and beds These methods can be categorized based on their effectiveness in shallow or deep applications and their adaptability to slopes of 0%–5% or greater than 5%.

• • • • • • • • Drained systems • • • • • Ev aporation ponds • • • • • • • • ET beds • • • • • • • • ET A beds • • • • • • • • • Spray irrig ation • • • • • • • • • • • Drip irrig ation • • • • • • • • • • • Note: The symbol • indicates appropriate system; ET , e vapotranspiration; ET A, e vapotranspiration–absorption.

The first step in a preliminary site evaluation involves assessing the current and proposed land use, along with understanding the anticipated flow and characteristics of wastewater Additionally, it is essential to observe the site's characteristics Following this, gather information on various relevant features.

• Soil permeability (general or qualitative)

• Existence of streams, drainage courses, or wetlands

When the pertinent data have been collected, the local regulatory agency should be contacted to determine the regulatory requirements The tests required for the phase

During the wet-test period, investigations can assess groundwater depth and conduct permeability tests to evaluate water absorption rates Additionally, Table 10.3 outlines common regulatory factors for on-site disposal.

Typical Regulatory Factors in On-Site Systems

Setback distances (horizontal, separation from wells, springs, surface waters, escarpments, site boundaries, buildings) ft (See Table 10.12)

Maximum slope for on-site disposal field % 25–30

Percolation rate min/in >1 to