United States

Environmental Protection

Agency

Office of Research and

Development

Cincinnati, Ohio 45268

EPA/625/R-99/010

September 2000

http://www.epa.gov/ORD/NRMRL

Manual

Constructed Wetlands

Treatment of Municipal

Wastewaters
i

EPA/625/R-99/010

September 1999

National Risk Management Research Laboratory

Office of Research and Development

U.S. Environmental Protection Agency

Cincinnati, Ohio 45268

Manual

Constructed Wetlands Treatment of

Municipal Wastewaters
ii

Notice

This document has been reviewed in accordance with the U.S. Environmental Protection

Agency’s peer and administrative review policies and approved for publication. Mention of trade

names or commercial products does not constitute endorsement or recommendation for use.
iii

Foreword

The U.S. Environmental Protection Agency is charged by Congress with protecting the

Nation’s land, air, and water resources. Under a mandate of national environmental laws, the

Agency strives to formulate and implement actions leading to a compatible balance between

human activities and the ability of natural systems to support and nurture life. To meet this

mandate, EPA’s research program is providing data and technical support for solving environmental prob-lems today and building a science knowledge base necessary to manage our ecological re-sources wisely, understand how pollutants affect our health, and prevent or reduce

environmen-tal risks in the future.

The National Risk Management Research Laboratory is the Agency’s center for investigation of technicological and management approaches for reducing risks from threats to human

health and the environment. The focus of the Laboratory’s research program is on methods for

the prevention and control of pollution to air, land, water and subsurface resources; protection

of water quality in public water systems; remediation of contaminated sites and ground water;

and prevention and control of indoor air pollution. The goal of this research effort is to catalyze

development and implementation of innovative, cost-effective environmental technologies; develop scientific and engineering information needed by EPA to support regulatory and policy

decisions; and provide technical support and information transfer to ensure effective implementation of environmental regulations and strategies.

This publication has been produced as part of the Laboratory’s strategic long-term research

plan. It is published and made available by EPA’s Office of Research and Development to assist

the user community and to link researchers with their clients.

E. Timothy Oppelt, Director

National Risk Management Research Laboratory
iv

Abstract

This manual discusses the capabilities of constructed wetlands, a functional design approach, and the management requirements to achieve the designed purpose. The manual also

attempts to put the proper perspective on the appropriate use, design and performance of constructed wetlands. For some applications, they are an excellent option because they are low in

cost and in maintenance requirements, offer good performance, and provide a natural appearance, if not more beneficial ecological benefits. In other applications, such as large urban areas

with large wastewater flows, they may not be at all appropriate owing to their land requirements.

Constructed wetlands are especially well suited for wastewater treatment in small communities

where inexpensive land is available and skilled operators hard to find and keep.

Primary customers will be engineers who service small communities, state regulators, and

planning professionals. Secondary users will be environmental groups and the academics.
v

Contents

Chapter 1 Introduction................................................................................................................................................... 1

1.1. Scope ........................................................................................................................................................ 1

1.2. Terminology ............................................................................................................................................... 1

1.3. Relationship to Previous EPA Documents ................................................................................................. 2

1.4. Wetlands Treatment Database .................................................................................................................. 2

1.5. History........................................................................................................................................................ 4

1.6. Common Misperceptions ........................................................................................................................... 4

1.7. When to Use Constructed Wetlands.......................................................................................................... 5

1.8 Use of This Manual .................................................................................................................................... 8

1.9 References ................................................................................................................................................ 8

Chapter 2 Introduction to Constructed Wetlands ........................................................................................................ 10

2.1 Understanding Constructed Wetlands ..................................................................................................... 10

2.2 Ecology of Constructed Wetlands............................................................................................................ 12

2.3 Botany of Constructed Wetlands ............................................................................................................. 12

2.4 Fauna of Constructed Wetlands .............................................................................................................. 16

2.5 Ecological Concerns for Constructed Wetland Designers ....................................................................... 16

2.6 Human Health Concerns ......................................................................................................................... 18

2.7 Onsite System Applications ..................................................................................................................... 19

2.8 Related Aquatic Treatment Systems ....................................................................................................... 19

2.9 Frequently Asked Questions .................................................................................................................... 20

2.10 Glossary................................................................................................................................................... 23

2.11 References .............................................................................................................................................. 27

Chapter 3 Removal Mechanisms and Modeling Performance of Constructed Wetlands ........................................... 30

3.1 Introduction .............................................................................................................................................. 30

3.2 Mechanisms of Suspended Solids Separations and Transformations .................................................... 30

3.3 Mechanisms for Organic Matter Separations and Transformations ........................................................ 35

3.4 Mechanisms of Nitrogen Separations and Transformations .................................................................... 42

3.5 Mechanisms of Phosphorus Separations and Transformations .............................................................. 46

3.6 Mechanisms of Pathogen Separations and Transformations .................................................................. 48

3.7 Mechanisms of Other Contaminant Separations and Transformations ................................................... 49

3.8 Constructed Wetland Modeling................................................................................................................ 50

3.9 References .............................................................................................................................................. 52

Chapter 4 Free Water Surface Wetlands .................................................................................................................... 55

4.1 Performance Expectations ...................................................................................................................... 55

4.2 Wetland Hydrology .................................................................................................................................. 64

4.3 Wetland Hydraulics .................................................................................................................................. 65

4.4 Wetland System Design and Sizing Rationale ........................................................................................ 68

4.5 Design ..................................................................................................................................................... 69

4.6 Design Issues .......................................................................................................................................... 78

4.7 Construction/Civil Engineering Issues ..................................................................................................... 81

4.8 Summary of Design Recommendations .................................................................................................. 83

4.9 References .............................................................................................................................................. 83
vi

Contents (cont.)

Chapter 5 Vegetated Submerged Beds ...................................................................................................................... 86

5.1 Introduction .............................................................................................................................................. 86

5.2 Theoretical Considerations ...................................................................................................................... 86

5.3 Hydrology................................................................................................................................................. 91

5.4 Basis of Design ........................................................................................................................................ 93

5.5 Design Considerations .......................................................................................................................... 101

5.6 Design Example for a VSB Treating Septic Tank or Primary Effluent .................................................... 103

5.7 On-site Applications ............................................................................................................................... 106

5.8 Alternative VSB Systems ....................................................................................................................... 106

5.9 References ............................................................................................................................................ 107

Chapter 6 Construction, Start-Up, Operation, and Maintenance ............................................................................... 111

6.1 Introduction ............................................................................................................................................. 111

6.2 Construction ........................................................................................................................................... 111

6.3 Start-Up ..................................................................................................................................................117

6.4 Operation and Maintenance ...................................................................................................................118

6.5 Monitoring ...............................................................................................................................................119

6.6 References .............................................................................................................................................119

Chapter 7 Capital and Recurring Costs of Constructed Wetlands ............................................................................ 120

7.1 Introduction ............................................................................................................................................ 120

7.2 Construction Costs ................................................................................................................................ 120

7.3 Operation and Maintenance Costs ........................................................................................................ 125

7.4 References ............................................................................................................................................ 127

Chapter 8 Case Studies ............................................................................................................................................ 128

8.1 Free Water Surface (FWS) Constructed Wetlands ................................................................................ 128

8.2 Vegetated Submerged Bed (VSB) Systems .......................................................................................... 141

8.3 Lessons Learned ................................................................................................................................... 152
vii

List of Figures

2-1 Constructed wetlands in wastewater treatment train ...........................................................................................11

2-2 Elements of a free water surface (FWS) constructed wetland .............................................................................11

2-3 Elements of a vegetated submerged bed (VSB) system .....................................................................................11

2-4 Profile of a 3-zone FWS constructed wetland cell .............................................................................................. 18

3-1 Mechanisms which dominate FWS systems....................................................................................................... 32

3-2 Weekly transect TSS concentration for Arcata cell 8 pilot receiving oxidation pond effluent .............................. 34

3-3 Variation in effluent BOD at the Arcata enhancement marsh .............................................................................. 36

3-4 Carbon transformations in an FWS wetland ....................................................................................................... 37

3-5 Dissolved oxygen distribution in emergent and submergent zones of a tertiary FWS ........................................ 40

3-6 Nitrogen transformations in FWS wetlands......................................................................................................... 43

3-7 Phosphorus cycling in an FWS wetland.............................................................................................................. 47

3-8 Phosphorus pulsing in pilot cells in Arcata .......................................................................................................... 48

3-9 Influent versus effluent FC for the TADB systems .............................................................................................. 49

3-10 Adaptive model building ...................................................................................................................................... 51

4-1 Effluent BOD vs areal loading ............................................................................................................................. 57

4-2 Internal release of soluble BOD during treatment ............................................................................................... 57

4-3 Annual detritus BOD load from Scirpus & Typha ................................................................................................ 58

4-4 TSS loading vs TSS in effluent ........................................................................................................................... 58

4-5 Effluent TKN vs TKN loading .............................................................................................................................. 59

4-6 Effluent TP vs TP areal loading ........................................................................................................................... 61

4-7 Total phosphorus loading versus effluent concentration for TADB systems ....................................................... 61

4-8 Hydraulic retention time vs orthophosphate removal .......................................................................................... 62

4-9 Influent versus effluent FC concentration for TADB systems .............................................................................. 63

4-10 TSS, BOD and FC removals for Arcata Pilot Cell 8 ............................................................................................ 63

4-11 Tracer response curve for Sacramento Regional Wastewater Treatment Plant Demonstration

Wetlands Project Cell 7 ....................................................................................................................................... 67

4-12 Transect BOD data for Arcata Pilot Cell 8 ........................................................................................................... 71

4-13 Elements of a free water surface (FWS) constructed wetland ............................................................................ 71

4-14 Generic removal of pollutants in a 3-zone FWS system ..................................................................................... 72

5-1 Seasonal cycle in a VSB ..................................................................................................................................... 90

5-2 Preferential flow in a VSB ................................................................................................................................... 93

5-3 Lithium chloride tracer studies in a VSB system ................................................................................................. 94

5-4 Effluent TSS vs areal loading rate....................................................................................................................... 95

5-5 Effluent TSS vs volumetric loading rate .............................................................................................................. 95

5-6 Effluent BOD vs areal loading rate ...................................................................................................................... 96

5-7 Effluent BOD vs volumetric loading rate ............................................................................................................. 96

5-8 Effluent TKN vs areal loading rate ...................................................................................................................... 98

5-9 Effluent TP vs areal loading rate ......................................................................................................................... 99

5-10 NADB VSBs treating pond effluent ................................................................................................................... 100

5-11 Proposed Zones in a VSB................................................................................................................................. 102

6-1 Examples of constructed wetland berm construction.........................................................................................112

6-2 Examples of constructed wetland inlet designs .................................................................................................114
viii

6-3 Outlet devices ....................................................................................................................................................115

8-1 Schematic diagram of wetland system at Arcata, CA ....................................................................................... 129

8-2 Schematic diagram of Phase 1&2 wetland systems at West Jackson County, MS .......................................... 132

8-3 Schematic diagram of Phase 3 wetland expansion at West Jackson County, MS ........................................... 132

8-4 Schematic diagrams of the wetland system at Gustine, CA ............................................................................. 136

8-5 Schematic diagram of the wetland system at Ouray, CO ................................................................................. 140

8-6 Schematic of Minoa, NY, VSB system .............................................................................................................. 142

8-7 Schematic diagram of typical VSB (one of three) at Mesquite, NV................................................................... 145

8-8 Schematic of VSB system at Mandeville, LA .................................................................................................... 147

8-9 Schematic of VSB system at Sorrento, LA ....................................................................................................... 151

List of Figures (cont.)
ix

List of Tables

1-1. Types of Wetlands in the NADB ............................................................................................................................ 3

1-2. Types of Wastewater Treated and Level of Pretreatment for NADB Wetlands ..................................................... 3

1-3. Size Distribution of Wetlands in the NADB ........................................................................................................... 4

1-4. Distribution of Wetlands in the NADB by State/Province ...................................................................................... 4

1-5. Start Date of Treatment Wetlands in the NADB .................................................................................................... 4

2-1 Characteristics of Plants for Constructed Wetlands ............................................................................................ 14

2-2 Factors to Consider in Plant Selection ................................................................................................................ 15

2-3 Characteristics of Animals Found in Constructed Wetlands ............................................................................... 16

3-1 Typical Constructed Wetland Influent Wastewater.............................................................................................. 30

3-2 Size Distributions for Solids in Municipal Wastewater ........................................................................................ 31

3-3 Size Distribution for Organic and Phosphorus Solids in Municipal Wastewater ................................................. 31

3-4 Fractional Distribution of BOD, COD, Turbidity and TSS in the Oxidation Pond Effluent and

Effluent from Marsh Cell 5................................................................................................................................... 34

3-5 Background Concentrations of Contaminants of Concern in FWS Wetland Treatment System Effluents .......... 35

3-6 Wetland Oxygen Sources and Sinks .................................................................................................................. 41

4-1 Loading and Performance Data for Systems Analyzed in This Document ......................................................... 56

4-2 Trace Metal Concentrations and Removal Rates, Sacramento Regional Wastewater Treatment Plant ............ 63

4-3 Fractional Distribution of BOD, COD and TSS in the Oxidation Pond Effluent and Effluent from

Marsh Cell 5 ........................................................................................................................................................ 64

4-4 Background Concentrations of Water Quality Constituents of Concern in FWS Constructed Wetlands ............ 70

4-5 Examples of Change in Wetland Volume Due to Deposition of Non-Degradable TSS (Vss) and

Plant Detritus (Vd) Based on 100 Percent Emergent Plant Coverage ................................................................ 74

4-6 Lagoon Influent and Effluent Quality Assumptions ............................................................................................. 77

4-7 Recommended Design Criteria for FWS Constructed Wetlands ........................................................................ 83

5-1 Hydraulic Conductivity Values Reported in the Literature ................................................................................... 92

5-2 Comparison of VSB Areas Required for BOD Removal Using Common Design Approaches ........................... 97

5-3 Data from Las Animas, CO VSB Treating Pond Effluent .................................................................................. 100

5-4 Summary of VSB Design Guidance .................................................................................................................. 106

7-1 Cost Comparison of 4,645m2 Free Water Surface Constructed Wetland and Vegetated Submerged Bed ...... 121

7-2 Technical and Cost Data for Wetland Systems Included in 1997 Case Study Visitations................................. 121

7-3 Clearing and Grubbing Costs for EPA Survey Sites ......................................................................................... 122

7-4 Excavation and Earthwork Costs for EPA Survey Sites .................................................................................... 122

7-5 Liner Costs for EPA Survey Sites ...................................................................................................................... 123

7-6 Typical Installed Liner Costs for 9,300m2 Minimum Area .................................................................................. 123

7-7 Media Costs for VSBs from EPA Survey Sites .................................................................................................. 124

7-8 Costs for Wetland Vegetation and Planting from EPA Survey Sites ................................................................. 124

7-9 Costs for Inlet and Outlet Structures from EPA Sites ........................................................................................ 124

7-10 Range of Capital Costs for a 0.4 ha Membrane-Lined VSB and FWS Wetland ............................................... 126

7-11 Annual O&M Costs at Carville, LA (570m3

/d) Vegetated Submerged Bed ....................................................... 127

7-12 Annual O&M Costs for Constructed Wetlands, Including All Treatment Costs ................................................. 127
x

8-1 Summary of Results, Phase 1 Pilot Testing, Arcata, CA ................................................................................... 130

8-2 Long-Term Average Performance, Arcata WWTP ............................................................................................ 131

8-3 Wetland Water Quality, West Jackson County, MS ........................................................................................... 134

8-4 Performance Results in Mature Vegetated vs Immature Vegetated FWS Cells, Gustine, CA .......................... 138

8-5 Wetland Effluent Characteristics, Gustine, CA.................................................................................................. 139

8-6 BOD & TSS Removal for Ouray, CO................................................................................................................. 141

8-7 Village of Minoa VSB Construction Costs ......................................................................................................... 143

8-8 Summary Performance, Mesquite, NV, VSB Component ................................................................................. 146

8-9 Monthly Effluent Characteristics, Mesquite, NV, VSB Component ................................................................... 146

8-10 Water Quality Performance, Mandeville, LA, Treatment System (June - Sept., 1991) ..................................... 149

8-11 Water Quality Performance, Mandeville, LA, Treatment System (Jan. 1996 - July 1997) ................................ 150

8-12 VSB Effluent Water Quality, Sorrento, LA ......................................................................................................... 152

List of Tables (cont.)
xi

Acknowledgments

Many people participated in the creation of this manual. Technical direction throughout the

multi-year production process was provided by USEPA’s National Risk Management Research

Laboratory (NRMRL). Technical writing was carried out in several stages, but culminated into a

final product as a cooperative effort between the NRMRL and the contractors named below.

Significant technical reviews and contributions based on extensive experience with constructed

wetlands were made by a number of prominent practitioners. Technical review was provided by

a group of professionals with extensive experience with the problems specific to small community wastewater treatment systems. The production of the document was also a joint effort by

NRMRL and contractual personnel. All of these people are listed below:

Primary Authors and Oversight Committee

Donald S. Brown, Water Supply and Water Resources Division, NRMRL, Cincinnati, OH

James F. Kreissl, Technology Transfer and Support Division, NRMRL, Cincinnati, OH

Robert A. Gearheart, Humboldt University, Arcata, CA

Andrew P. Kruzic, University of Texas at Arlington, Arlington, TX

William C. Boyle, University of Wisconsin, Madison, WI

Richard J. Otis, Ayres Associates, Madison, WI

Major Contributors/Authors

Sherwood C. Reed, Environmental Engineering Consultants, Norwich, VT

Richard Moen, Ayres Associates, Madison, WI

Robert Knight, Consultant, Gainesville, FL

Dennis George, Tennessee Technological University, Cookeville, TN

Brad Finney, Humboldt University, Arcata, CA

Jeff Anderson, Humboldt University, Arcata, CA

Margaret Lang, Humboldt University, Arcata, CA

Michael Ogden, Southwest Wetlands Group, Inc., Santa Fe, NM

Ronald Crites, Brown and Caldwell, Sacramento, CA

George Tchobanoglons, Consultant, Davis, CA

Contributing Writers/Production Specialists

Ian Clavey, CEP Inc., Cincinnati, OH

Vince Iacobucci, CEP Inc., Cincinnati, OH

Julie Hotchkiss, CEP Inc., Cincinnati, OH

Peggy Heimbrock, TTSD - NRMRL, Cincinnati, OH

Stephen E. Wilson, TTSD - NRMRL, Cincinnati, OH

Denise Ratliff, TTSD - NRMRL, Cincinnati, OH

Betty Kampsen, STD - NRMRL, Cincinnati, OH

Technical Reviewers

Arthur H. Benedict, EES Consulting, Inc., Bellevue, WA

Pio Lombardo, Lombardo Associates, Inc., Newton, MA

Rao Surampalli, USEPA - Region VII, Kansas City, KS

Robert K. Bastian, USEPA - Office of Wastewater Management, Washington, DC
12

1.1. Scope

Constructed wetlands are artificial wastewater treatment

systems consisting of shallow (usually less than 1 m deep)

ponds or channels which have been planted with aquatic

plants, and which rely upon natural microbial, biological,

physical and chemical processes to treat wastewater. They

typically have impervious clay or synthetic liners, and engineered structures to control the flow direction, liquid detention time and water level. Depending on the type of system, they may or may not contain an inert porous media

such as rock, gravel or sand.

Constructed wetlands have been used to treat a variety

of wastewaters including urban runoff, municipal, industrial, agricultural and acid mine drainage. However, the

scope of this manual is limited to constructed wetlands

that are the major unit process in a system to treat municipal wastewater. While some degree of pre- or post- treatment will be required in conjunction with the wetland to

treat wastewater to meet stream discharge or reuse requirements, the wetland will be the central treatment component.

This manual discusses the capabilities of constructed

wetlands, a functional design approach, and the management requirements to achieve the designed purpose. This

manual also attempts to put the proper perspective on the

appropriate use of constructed wetlands. For some applications, they are an excellent option because they are low

in cost and in maintenance requirements, offer good performance, and provide a natural appearance, if not more

beneficial ecological benefits. However, because they require large land areas, 4 to 25 acres per million gallons of

flow per day, they are not appropriate for some applications. Constructed wetlands are especially well suited for

wastewater treatment in small communities where inexpensive land is available and skilled operators are hard to

find.

1.2 Terminology

A brief discussion of terminology will help the reader differentiate between the constructed wetlands discussed in

this manual and other types of wetlands. Wetlands are

defined in Federal regulations as “those areas that are inundated or saturated by surface or ground water at a frequency and duration sufficient to support, and that under

Chapter 1

Introduction to the Manual

normal circumstances do support, a prevalence of vegetation typically adapted for life in saturated soil conditions. Wetlands generally include swamps, marshes, bogs

and similar areas.” (40 CFR 230.3(t)) Artificial wetlands

are wetlands that have been built or extensively modified

by humans, as opposed to natural wetlands which are

existing wetlands that have had little or no modification by

humans, such as filling, draining, or altering the flow patterns or physical properties of the wetland. The modification or direct use of natural wetlands for wastewater treatment is discouraged and natural wetlands are not discussed in this manual (see discussion of policy issues in

Section 1.7.2).

As previously defined, constructed wetlands are artificial wetlands built to provide wastewater treatment. They

are typically constructed with uniform depths and regular

shapes near the source of the wastewater and often in

upland areas where no wetlands have historically existed.

Constructed wetlands are almost always regulated as

wastewater treatment facilities and cannot be used for

compensatory mitigation (see Section 1.7.2). Some EPA

documents refer to constructed wetlands as constructed

treatment wetlands to avoid any confusion about their primary use as a wastewater treatment facility (USEPA,

1999). Constructed wetlands which provide advanced

treatment to wastewater that has been pretreated to secondary levels, and also provide other benefits such as

wildlife habitat, research laboratories, or recreational uses

are sometimes called enhancement wetlands.

Constructed wetlands have been classified by the literature and practitioners into two types. Free water surface (FWS) wetlands (also known as surface flow wetlands) closely resemble natural wetlands in appearance

because they contain aquatic plants that are rooted in a

soil layer on the bottom of the wetland and water flows

through the leaves and stems of plants. Vegetated submerged bed (VSB) systems (also known as subsurface

flow wetlands) do not resemble natural wetlands because

they have no standing water. They contain a bed of media

(such as crushed rock, small stones, gravel, sand or soil)

which has been planted with aquatic plants. When properly designed and operated, wastewater stays beneath

the surface of the media, flows in contact with the roots

and rhizomes of the plants, and is not visible or available

to wildlife.
13

The term “vegetated submerged bed” is used in this manual

instead of subsurface flow wetland because it is a more accurate and descriptive term. The term has been used previously to describe these units (WPCF, 1990; USEPA, 1994).

Some VSBs may meet the strict definition of a wetland, but a

VSB does not support aquatic wildlife because the water

level stays below the surface of the media, and is not conducive to many of the biological and chemical interactions that

occur in the water and sediments of a wetland with an open

water column. VSBs have historically been characterized as

constructed wetlands in the literature, and so they are included in this manual.

Constructed wetlands should not be confused with created or restored wetlands, which have the primary function

of wildlife habitat. In an effort to mimic natural wetlands, the

latter often have a combination of features such as varying

water depths, open water and dense vegetation zones, vegetation types ranging from submerged aquatic plants to

shrubs and trees, nesting islands, and irregular shorelines.

They are frequently built in or near places that have historically had wetlands, and are often built as compensatory mitigation. Created and restored wetlands for habitat or compensatory mitigation are not discussed in this manual.

Finally, the term vertical flow wetland is used to describe a

typical vertical flow sand or gravel filter which has been

planted with aquatic plants. Because successful operation

of this type of system depends on its operation as a filter (i.e.

frequent dosing and draining cycles), this manual does not

discuss this type of system.

1.3 Relationship to Previous EPA

Documents

Several Offices or Programs within USEPA have published

documents in recent years on the subject of constructed

wetlands. Some examples of publications and their USEPA

sponsors are:

• Design Manual: Constructed Wetlands and Aquatic Plant

Systems for Municipal Wastewater Treatment (1988)

(Office of Research and Development, Cincinnati, OH,

EPA 625-1-88-022)

• Subsurface Flow Constructed Wetlands for Wastewater Treatment: A Technology Assessment (1993) (Office

of Wastewater Management, Washington, DC, EPA 832-

R-93-008)

• Habitat Quality Assessment of Wetland Treatment Systems (3 studies in 1992 and 1993) (Environmental Research Lab, Corvallis, OR, EPA 600-R-92-229, EPA 600-

R-93-117, EPA 600-R-93-222)

• Constructed Wetlands for Wastewater Treatment and

Wildlife Habitat: 17 Case Studies (1993) (Office of Wastewater Management, Washington, DC, EPA 832-R-93-

005)

• Guidance for Design and Construction of a Subsurface Flow Constructed Wetland (August 1993) (USEPA

Region VI, Municipal Facilities Branch)

• A Handbook of Constructed Wetlands (5 volumes,

1995) (USEPA Region III with USDA, NRCS, ISBN 0-

16-052999-9)

• Constructed Wetlands for Animal Waste Treatment: A

Manual on Performance, Design, and Operation With

Cases Histories (1997) (USEPA Gulf of Mexico Program)

• Free Water Surface Wetlands for Wastewater Treatment: A Technology Assessment (1999) (Office of

Wastewater Management, Washington, DC, EPA /832/

R-99/002)

Some information presented in this manual may contradict information presented in these other documents. Some

contradictions are the result of new information and understanding developed since the publication of earlier documents; some contradictions are the result of earlier misconceptions about the mechanisms at work within constructed wetlands; and some contradictions are the result

of differing opinions among experts when insufficient information exists to present a clear answer to issues surrounded by disagreement. As stated previously, this manual

attempts to put an environmental engineering perspective

on the use, design and performance of constructed wetlands as reflected by the highest quality data available at

this time. In areas where there is some disagreement

among experts, this manual assumes a conservative approach based on known treatment mechanisms which fit

existing valid data.

1.4 Wetlands Treatment Database

Through a series of efforts funded by the USEPA, a

Wetlands Treatment Database, “North American Wetlands

for Water Quality Treatment Database or NADB” (USEPA,

1994) has been compiled which provides information about

natural and constructed wetlands used for wastewater

treatment in North America. Version 1 of the NADB was

released in 1994 and contains information for treatment

wetlands at 174 locations in over 30 US states and Canadian provinces. Information includes general site information, system specific information (e.g., flow, dimensions,

plant species), contact people with addresses and phone

numbers, literature references, and permit information. It

also contains some water quality data (BOD, TSS, N-series, P, DO, and fecal coliforms), but the data is not of uniform quantity and quality, which makes it inappropriate for

design or modeling purposes.

Version 2 of the NADB is currently undergoing Agency

review and contains information on treatment wetlands at

245 locations in the US and Canada. Because each location may have multiple wetland cells, there are over 800

individual wetland cells identified in Version 2. Besides

expanding the number of wetland locations from Version

1, Version 2 also contains information regarding vegetation, wildlife, human use, biomonitoring and additional water

quality data. As with Version 1, the data is not adequate

for design or modeling.
14

Data did not exist or were incomplete for many of the

wetlands included in the NADB. Only existing information was collected for the NADB; no new measurements

were made. Therefore, the NADB is very useful for obtaining general information about the status of constructed wetlands usage, as well as the locations of

operating systems and people to contact. However, it is

not useful as a source of water quality data for wetland

design or prediction of treatment performance.

Tables 1.1 through 1.4 give an overview of Version 2 of

the NADB. The size range and median size are shown in

several tables to give the reader a feel for the size of each

type of wetland. The median size is shown because there

are a few very large wetlands in some of the groups, which

makes the median size more characteristic of the group

than the mean size.

Tables 1.1 and 1.2 group the wetlands by type of wetland and type of wastewater being treated, respectively. In

general FWSs are larger than VSBs, with the median size

of FWS wetlands being twice that of VSBs. The summary

statistics for “other water” wetlands in Table 1.2 are somewhat misleading because they are influenced by the large

Everglades Nutrient Removal project in Florida.

Table 1-1. Types of Wetlands in the NADB

Size (hectares)

Type of Wetland Qty. Min. Median Max.

Constructed Wetlands 205 0.0004 0.8 1406

Free Water Surface 138 0.0004 1 1406

Marsh* 125 0.0004 1 1406

Other 13 0.08 3 188

Vegetated Submerged Bed (all Marsh) 49 0.004 0.5 498

Combined FWS & VSB (all Marsh) 8 0.1 0.4 17

Other or Not Classified 10 0.01 1 14

Natural Wetlands (all Free Water Surface) 38 0.2 40 1093

Forest 18 1 40 204

Marsh 16 0.2 33 1093

Other or Not Classified 4 6 64 494

Not Classified 2

*Marshes are characterized by soft-stemmed herbaceous plants, including emergent species, such as cattails, floating species, such as water lilies,

and submerged species, such as pondweeds. (Niering, 1985)

Table 1-2. Types of Wastewater Treated and Level of Pretreatment for NADB Wetlands

Size (hectares)

Wastewater Type Pretreatment Qty. Min. Median Max.

Agricultural 58 0.0004 0.1 47

None 8

Primary 35

Facultative 14

Not classified 1

Industrial 13 0.03 10 1093

Primary 1

Facultative 2

Secondary 6

Advanced Secondary 1

Not classified 3

Municipal 159 0.004 2 500

Primary 9

Facultative 78

Secondary 49

Advanced Secondary 9

Tertiary 4

Other 4

Not classified 6

Stormwater 6 0.2 8 42

None 4

Secondary 1

Other 1

Other Water 7 3 376 1406

None 4

Primary 1

Facultative 1

Secondary 1

Not classified 2
15

Table 1.3 groups all the wetlands, regardless of type of

wetland or wastewater being treated, by size. In terms of

area, the majority of the wetlands are less than 10 hectares (25 acres), and almost 90% are less than 100 hectares (250 acres). In terms of design flow rate, the majority

are less than 1000 m3/d (about 0.25 mgd), and 82% are

less than 4060 m3/d (1 mgd).

Table 1.4 groups all the wetlands, regardless of type of

wetland or wastewater being treated, by location. Treatment wetlands are located in 34 US states and 6 Canadian provinces. The number of wetlands per state is probably more a function of having an advocate for treatment

wetlands in the state than climate or some other favorable

condition.

Table 1-3. Size Distribution of Wetlands in the NADB

Area (hectares) Design Flow (m3/d)

Size Cumulative Size Cumulative

Range Percentage Range Percentage

less than 1 46 less than 10 19

less than 10 75 less than 100 31

less than 100 93 less than 1000 62

less than 1000 99 less than 10,000 93

Table 1-4. Distribution of Wetlands in the NADB by State/Province

 Number of Size (hectares)

State or Province* Wetlands Min. Median Max.

SD 42 0.3 2 134

FL 24 0.2 44.5 1406

AR 21 0.3 0.8 4

KY 19 0.01 0.1 5

LA 15 0.02 0.3 17

MS 11 0.02 0.9 101

CA 9 0.1 14 59

AL 8 0.04 0.2 6

ONT 8 0.02 0.09 0.4

MD 5 0.1 0.2 2

OR 5 0.1 4 36

SC 5 20 36 185

IN 4 0.002 0.12 1

MI 4 5 56.5 110

MO 4 0.04 0.25 37

NY 4 0.03 0.25 2

PA 4 0.01 0.055 0.2

TX 4 0.1 0.2 0.5

AZ 3 2 38 54

GA 3 0.01 0.3 0.4

ND 3 14 17 33

NVS 3 0.1 0.1 0.4

TN 3 0.1 0.2 0.3

WI 3 0.01 6 156

ALB, IA, ME, MN,

NC, NM, NV, QUE 2

CT, IL, MA, NJ, NWT,

PEI, VA, WA 1

*Two-letter abbreviations are states; three-letter abbreviations are

provinces.

1.5 History

Kadlec and Knight (1996) give a good historical account

of the use of natural and constructed wetlands for wastewater treatment and disposal. As they point out, natural

wetlands have probably been used for wastewater disposal

for as long as wastewater has been collected, with documented discharges dating back to 1912. Some early constructed wetlands researchers probably began their efforts

based on observations of the apparent treatment capacity

of natural wetlands. Others saw wastewater as a source

of water and nutrients for wetland restoration or creation.

Research studies on the use of constructed wetlands for

wastewater treatment began in Europe in the 1950's, and

in the US in the late 1960's. Research efforts in the US

increased throughout the 1970's and 1980's, with significant Federal involvement by the Tennessee Valley Authority

(TVA) and the U.S. Department of Agriculture in the late

1980's and early 1990's. USEPA has had a limited role in

constructed wetlands research which might explain the

dearth of useful, quality-assured data.

Start dates for constructed wetlands in the NADB are

shown in Table 1.5, with the start dates for natural wetlands used for treatment included for comparison. The table

shows that the use of FWS wetlands and VSBs in North

America really began in the early and late1980's, respectively, and the number continues to increase. No new natural wetland treatment systems have begun since 1990, and

at least one-third of the natural wetland treatment systems

included in the NADB are no longer operating.

1.6 Common Misconceptions

Many texts and design guidelines for constructed wetlands, in addition to those listed above sponsored by the

various offices of USEPA, have been published since

USEPA’s 1988 design manual (EC/EWPCA (1990); WPCF

(1990); Tennessee Valley Authority (1993); USDA (1993);

Reed, et al (1995); Kadlec and Knight (1996); Campbell

and Ogden (1999)). Also, a number of international conferences have been convened to present the findings of

constructed wetlands research from almost every continent (Hammer (1989); Cooper and Findlater (1990); Moshiri

Table 1-5. Start Date of Treatment Wetlands in the NADB

before 1950's & 1990's

Type 1950 60's 1970's ‘80-‘84 ‘85-‘89 (latest*)

Constructed, 1 0 3 8 33 85

FWS (‘96)

Constructed, 0 0 0 0 21 31

VSB (‘94)

Constructed, 0 0 0 1 4 6

hybrid (‘94)

Natural, 4 3 9 5 8 1

FWS (‘90)

*Year of last wetland included in database for this type of wetland - other

wetlands may have started after this date, but are not in the database.
16

(1993); IAWQ (1994)(1995) (1997)). However, in spite of

the great amount of resources devoted to constructed wetlands, questions and misconceptions remain about their application, design, and performance. This section briefly describes four common misconceptions; further discussion of

these items is found in other chapters.

Misconception #1: Wetland design has been well-characterized by published design equations. Constructed wetlands

are complex systems in terms of biology, hydraulics and water

chemistry. Furthermore, there is a lack of quality data of sufficient detail, both temporally and spatially, on full-scale constructed wetlands. Due to the lack of data, designers have

been forced to derive design parameters by aggregating performance data from a variety of wetlands, which leads to

uncertainties about the validity of the parameters. Data from

wetlands with detailed research studies with rigorous quality

control (QC) might be combined with data from wetlands

with randomly collected data with little QC. Data from small

wetlands with minimal pretreatment might be combined with

data from large wetlands used for polishing secondary effluent. Additional problems with constructed wetlands data include: lack of paired influent-effluent samples; grab samples

instead of composited samples; lack of reliable flow or detention time information; and lack of important incidental information such as temperature and precipitation. The resulting data combinations, completed to obtain larger data sets,

have sometimes been used to create regression equations

of questionable value for use in design. Finally, data from

constructed wetlands treating relatively high quality (but inadequately characterized) wastewater has sometimes been

used to derive design parameters for more concentrated

municipal treatment applications, which is less than assuring for any designer.

Misconception #2: Constructed wetlands have aerobic as

well as anaerobic treatment zones. Probably the most common misconception concerns the ability of emergent wetland plants to transfer oxygen to their roots. Emergent aquatic

plants are uniquely suited to the anaerobic environment of

wetlands because they can move oxygen from the atmosphere to their roots. Research has shown that some oxygen “leaks” from the roots into the surrounding soils (Brix,

1997). This phenomenon, and early work with natural and

constructed wetlands that treated wastewater with a low

oxygen demand, has led to the assumption that significant

aerobic micro-sites exist in all wetland systems. Some constructed wetlands literature states or implies that aerobic biodegradation is a significant treatment mechanism in fully

vegetated systems, which has led some practitioners to believe that wetlands with dense vegetation, or many sources

of “leaking” oxygen, are in fact aerobic systems. However,

the early work with tertiary or polishing wetlands is not directly applicable to wetlands treating higher strength wastewater because it fails to account for the impacts of the wastewater on the characteristics of the wetland. Treatment mechanisms that function under light loads are impaired or overwhelmed due to changes imparted by the large oxygen demand of more contaminated municipal wastewater. Field

experience and research have shown that the small amount

of oxygen leaked from plant roots is insignificant compared

to the oxygen demand of municipal wastewater applied at

practical loading rates.

Misconception #3: Constructed wetlands can remove significant amounts of nitrogen. Related to the misconception

about the availability of oxygen in constructed wetlands is

the misconception about the ability of constructed wetlands

to remove significant amounts of nitrogen. Harvesting removes less than 20% of influent nitrogen (Reed, et al,1995)

at conventional loading rates. This leaves nitrification and

denitrification as the primary removal mechanisms. If it is

assumed that wetlands have aerobic zones, it then follows

that nitrification of ammonia to nitrate should occur. Furthermore, if the aerobic zone surrounds only the roots of the

plants, it then follows that anaerobic zones dominate, and

denitrification of nitrate to nitrogen gas should also occur.

Unfortunately, the nitrogen-related misconceptions have been

responsible for the failure of several constructed wetlands

that were built to remove or oxidize nitrogen. Because anaerobic processes dominate in both VSBs and fully vegetated

FWS wetlands, nitrification of ammonia is unlikely to occur

in the former and will occur only if open water zones are

introduced to the latter. Constructed wetlands can be designed to remove nitrogen, if sufficient aerobic (open water)

and anaerobic (vegetated) zones are provided. Otherwise,

constructed wetlands should be used in conjunction with other

aerobic treatment processes that can nitrify to remove nitrogen.

Misconception #4: Constructed wetlands can remove significant amounts of phosphorus. Phosphorus removal in constructed wetlands is limited to seasonal uptake by the plants,

which is not only minor compared to the phosphorus load in

municipal wastewater, but is negated during the plants’ senescence, and to sorption to influent solids which are captured, soils or plant detritus, all of which have a limited capacity. Two problems have been associated with phosphorus data in the literature. First, some phosphorus removal

data has been reported in terms of percent removal. However, many of the early phosphorus studies were for natural

wetlands or constructed wetlands that received wastewater

with a low phosphorus concentration. Because of low influent concentrations, removal of only a single mg/L of phosphorus was reported as a large percent removal. Second,

for studies evaluating the performance of newly constructed

wetlands, phosphorus removal data will be uncharacteristic

of long-term performance. New plants growing in a freshly

planted wetland will uptake more phosphorus than a mature

wetland, which will have phosphorus leaching from dying

(senescent) plants as well as uptake by growing plants. Also,

newly placed soils or media will have a greater phosphorus

sorption capacity than a mature system which will have most

sorption sites already saturated.

1.7 When to Use Constructed Wetlands

1.7.1 Appropriate Technology for Small

Communities

Appropriate technology is defined as a treatment system which meets the following key criteria:
17

Affordable - Total annual costs, including capital, operation, maintenance and depreciation are within the

user’s ability to pay.

Operable - Operation of the system is possible with

locally available labor and support.

Reliable - Effluent quality requirements can be consistently meet.

Unfortunately, many rural areas of the U.S. with small

treatment plants (usually defined as treating less than 3,800

m3/d (1 mgd)) have failed to consider this appropriate technology definition, and have often adopted inappropriate

technologies such as activated sludge. In 1980, small,

activated sludge systems constituted 39% of the small

publicly owned treatment facilities (GAO, 1980). Recent

information from one state showed that 73% of all treatment plants of less than 3,800m3/d capacity used some

form of the activated sludge process. Unfortunately, the

activated sludge process is considered by almost all U.S.

and international experts to be the most difficult to operate

and maintain of the various wastewater treatment concepts.

Presently, small treatment plants constitute more than 90%

of the violations of U.S. discharge standards. At least one

U.S. state, Tennessee, has required justification for the use

of activated sludge package plants for very small treatment plant applications (Tennessee Department of Public

Health, 1977).

Small community budgets become severely strained by

the costs of their wastewater collection and treatment facilities. Inadequate budgets and poor access to equipment,

supplies and repair facilities preclude proper operation and

maintenance (O&M). Unaffordable capital costs and the

inability to reliably meet effluent quality requirements add

up to a prime example of violating the prior criteria for appropriate technology. Unfortunately, no consideration for

reuse, groundwater recharge, or other alternatives to

stream discharge has heretofore been common, except in

a few states where water shortages exist.

Presently there are a limited number of appropriate technologies for small communities which should be immediately considered by a community and their designer. These

include stabilization ponds or lagoons, slow sand filters,

land treatment systems, and constructed wetlands. All of

these technologies fit the operability criterion, and to varying degrees, are affordable to build and reliable in their

treatment performance. Because each of these technologies has certain characteristics dNd requirements for preand post- treatment to meet a certain effluent quality, they

may be used alone or in series with others depending on

the treatment goals.

For example, the designer may wish to supplement stabilization ponds with a tertiary system to meet reuse or

discharge criteria consistently. Appropriate stabilization

pond upgrading methods to meet effluent standards include FWS wetlands, which can provide the conditions for

enhanced settling to attain further reduction of fecal

coliforms and removal of the excess algal growth which

characterizes pond system effluents. FWS wetlands are

normally used after ponds because of their ability to handle

the excess algal solids generated in the ponds. Although

VSBs have been employed after ponds, excess algal solids have caused problems at some locations, thus defeating the operability factor in the appropriate technology definition. VSBs are more appropriately applied behind a process designed to minimize suspended and settleable solids, such as a septic or Imhoff tank or anaerobic lagoon.

Constructed wetlands may also require post-treatment

processes, depending on the ultimate goals of the treatment system. More demanding effluent requirements may

require additional processes in the treatment train or may

dictate the use of other processes altogether. For example,

the ability of constructed wetlands to remove nitrogen and

phosphorus has frequently been overestimated. Two appropriate technologies that readily accomplish ammonium

oxidation are intermittent and recirculating sand filters.

There is at least one case study of the successful use of a

recirculating gravel filter in conjunction with a VSB (Reed,

et al, 1995 ). FWS systems can both nitrify and denitrify,

thus removing significant portions of nitrogen from the

wastewater, by alternating fully vegetated and open water

zones in proper proportions. If the municipal facility is required to have significant phosphorus removal (e.g., to attain 1 mg/L from a typical influent value of 6 to 7 mg/L),

constructed wetlands will need to be accompanied by some

process or processes that can remove the phosphorus.

In conclusion, constructed wetlands are an appropriate

technology for areas where inexpensive land is generally

available and skilled labor is less available. Whether they

can be used essentially alone or in series with other appropriate technologies depends on the required treatment

goals. Additionally, they can be appropriate for onsite systems where local regulators call for and allow systems other

than conventional septic tank - soil absorption systems.

1.7.2 Policy and Permitting Issues

An interagency workgroup, including representatives

from several Federal agencies, is presently developing

Guiding Principles for Constructed Treatment Wetlands:

Providing Water Quality and Wildlife Habitat (USEPA,

1999). The essence of the current draft of the guidelines is

that constructed treatment wetlands will:

— receive no credit as mitigation wetlands;

— be subject to the same rules as treatment lagoons

regarding liner requirements;

— be subject to the same monitoring requirements as

treatment lagoons;

— should not be constructed in the waters of the United

States, including existing natural wetlands; and

— will not be considered Waters of the United States

upon abandonment if the first and the fourth conditions are met.
18

The guidance encourages use of local plant species and

expresses concern about permit compliance during lengthy

startup periods and vector attraction and control issues.

To avoid additional permitting and regulatory requirements, constructed wetlands should be designed as a treatment process and built in uplands as opposed to wetlands

or flood plains, i.e., outside of waters of the U.S.. Consider

the following from the draft guidelines.

If your constructed treatment wetland is constructed

in an existing water of the U.S., it will remain a water

of the U.S. unless an individual CWA section 404 permit is issued which explicitly authorizes it as an excluded waste treatment system designed to meet the

requirements of the CWA. ... Once constructed, if your

treatment wetland is a water of the U.S., you will need

a NPDES permit for the discharge of pollutants ... into

the wetland. ... [Additionally,] if you wish to use a degraded wetland for wastewater treatment and plan to

construct water control structures, such as berms or

levees, this construction will ... require a Section 404

permit. Subsequent maintenance may also require a

permit.

As stated in the guidelines:

If the constructed wetland is abandoned or is no longer

being used as a treatment system, it may revert to a

water of the U.S. if ... the following conditions exist:

the system has wetland characteristics (i.e., hydrology, soils, vegetation) and it is either (1) an interstate

wetland, (2) is adjacent to another water of the U.S.

(other than waters which are themselves wetlands),

or (3) if it is an isolated intrastate water which has a

nexus to interstate commerce (e.g., it provides habitat

for migratory birds).

None of preceding discussion precludes designing and

building a wetland which provides water reuse, habitat or

public use benefits in addition to wastewater treatment.

Constructed wetlands built primarily for treatment will generally not be given credit as compensatory mitigation to

replace wetland losses. However, in limited cases, some

parts of a constructed wetland system may be given credit,

especially if additional wetland area is created beyond that

needed for treatment purposes. Also, current policy encourages the use of properly treated wastewater to restore

degraded wetlands. For example, restoration might be

possible if:

1 the source water meets all applicable water qual

ity standards and criteria, (2) its use would result in a net

environmental benefit to the aquatic system’s natural

functions and values, and if applicable, (3) it would help

restore the aquatic system to its historical condition.

Prime candidates for restoration may include wetlands

that were degraded or destroyed through the diversion

of water supplies, ... For example, in the arid west, there

are often historic wetlands that no longer have a reliable

water source due to upstream water allocations or sinking groundwater tables. Pre-treated effluent may be the

only source of water available for these areas and their

dependent ecosystems. ... EPA has developed regional

guidance to assist dischargers and regulators in demonstrating a net ecological benefit from maintenance of

a wastewater discharge to a waterbody.

This discussion of policy and permitting issues is very

general and regulatory decisions regarding these issues

are made on a case-by-case basis. Planners and designers should seek guidance from State and Regional regulators about site specific constructed wetland criteria including location, discharge requirements, and possible

long-term monitoring requirements.

1.7.3 Other Factors

Probably the most important factor which impacts all

aspects of constructed wetlands is their inherent aesthetic

appeal to the general public. The desire of people to have

such an attractive landscape enhancement treat their

wastewater and become a valuable addition to the community is a powerful argument when the need for wastewater treatment upgrading becomes a matter of public

debate. The appeal of constructed wetlands makes the

need to accurately assess the capability of the technology

so important and so difficult. The engineering community

often fails to appreciate this inherent appeal, while the

environmental community often lacks the understanding

of treatment mechanisms to appreciate the limitations of

the technology. The natural attraction of constructed wetlands and the potential for other aesthetic benefits may

sometimes offset the treatment or cost advantages of other

treatment options, and public opinion may dictate that a

constructed wetland is the preferred option. In other situations, constructed wetlands will be too costly or unable to

produce the required effluent water quality, and the designer will have to convince the public that wetlands are

not a viable option, in spite of their inherent appeal.

The use of constructed wetlands as a treatment technology carries some degree of risk for several reasons.

First, as noted in a review of constructed wetlands for

wastewater treatment by Cole (1998), constructed wetlands

are not uniformly accepted by all state regulators or EPA

regions. Some authorities encourage the use of constructed

wetlands as a proven treatment technology, due in part to

the misconceptions noted in Section 1.6. Others still consider them to be an emerging technology. As with any new

treatment technology, uniform acceptance of constructed

wetlands will take some time. Other natural treatment processes which are now generally accepted, such as slow

rate or overland flow land treatment systems, went through

a similar course of variable acceptance.

Second, although there is no evidence of harm to wildlife using constructed wetlands, some regulators have expressed concern about constructing a system which will

treat wastewater while it attracts wildlife. Unfortunately,

there has not been any significant research conducted on

the risks to wildlife using constructed wetlands. Although
19

they are a distinctly different type of habitat, the lack of

evidence of risks to wildlife using treatment lagoon systems for many years suggests that there may not be a

serious risk for wetlands treating municipal wastewater.

Of course, if a wetland is going to treat wastewater with

high concentrations of known toxic compounds, the designer will need to use a VSB system or incorporate features in a FWS wetland which restrict access by wildlife.

Finally, as noted earlier, due to the lack of a large body

of scientifically valid data, the design process is still empirical, that is, based upon observational data rather than

scientific theories. Due to the variability of many factors at

constructed wetlands being observed by researchers (e.g.,

climatic effects, influent wastewater characteristics, design

configurations, construction techniques, and O&M practices), there will continue to be disagreement about some

design and performance issues for some period of time.

1.8 Use of This Manual

Chapters 1, 2, 7 and 8 provide information for non-technical readers, such as decision-makers and stakeholders,

to understand the capabilities and limitations of constructed

wetlands. These chapters provide the type of information

required to question designers and regulators in the process of determining how constructed wetlands may be used

to expand, upgrade or develop wastewater treatment infrastructure.

Chapters 3 through 6 provide information for technical

readers, such as design engineers, regulators and planners, to plan, design, build and manage constructed wetlands as part of a comprehensive plan for local and regional management of municipal wastewater collection,

treatment, and reuse.

Chapter 2 describes constructed wetland treatment systems and their identifiable features. It answers the most

frequently asked questions about these systems and includes a glossary of terms which are used in this manual

and generally in discussion of constructed wetland systems. There are brief discussions of other aquatic treatment systems that are in use or are commercially available and an annotated introduction to specific uses for

constructed wetlands outside the purview of this manual.

Chapter 3 discusses the treatment mechanisms occurring in a constructed wetland to help the reader understand the most important processes and what climatic conditions and other physical phenomena most affect these

processes. A basic understanding of the mechanisms involved will allow the reader to more intelligently interpret

information from other literature sources as well as information in chapters 4 and 5 of this manual.

Chapters 4, 5, and 6 describe the design, construction,

startup and operational issues of constructed wetlands in

some detail. It will be apparent to the reader that there are

presently insufficient data to create treatment models in

which there can be great confidence. Most data in the literature has been generated with inadequate quality assurance and control (Qa/Qc), and most research studies

have not measured or focused on documentation of key

variables which could explain certain performance characteristics. Chapters 4 and 5 use the existing data of sufficient quality to create a viable approach to applicability

and design of both FWS and VSB systems and sets practical limits on their performance capabilities. Chapter 6

deals with the practical issues of construction and start-up

of these systems which have been experienced to date.

Chapter 7 contains cost information for constructed wetlands. Subsequent to standardizing the costs to a specific

time, it becomes clear that local conditions and requirements can dominate the costs. However, the chapter does

provide a reasonable range of expected costs which can

be used to evaluate constructed wetlands against other

alternatives in the facility planning stage. Also, there is

sufficient information presented to provide the user with a

range of unit costs for certain components and to indicate

those components that dominate system costs and those

that are relatively inconsequential.

Chapter 8 presents eight case studies to allow readers

to become familiar with sites that have used constructed

wetlands and their experiences. The systems in this chapter are not ones which are superior to other existing facilities, but they are those which have been observed and

from which lessons can be learned by the reader about

either successful or unsuccessful design practices.

1.9 References

Brix, H. 1997. Do Macrophytes Play a Role in Constructed

Treatment Wetlands?. Water Science & Technology,

Vol 35, No. 5, pp.11-17.

Campbell, C.S. and M.H. Ogden. 1999. Constructed Wetlands in the Sustainable Landscape. John Wiley and

Sons, New York, New York.

Cole, Stephen. 1998. The Emergence of Treatment Wetlands. Environmental Science & Technology, Vol. 3,

No.5, pp 218A-223A.

Cooper, P.F., and B.C. Findlater, eds. 1990. Constructed

Wetlands in Water Pollution Control. Pergamon Press,

New York, New York.

EC/EWPCA. 1990. European Design and Operations

Guidelines for Reed Bed Treatment Systems. Prepared

for the EC/EWPCA Expert Contact Group on Emergent Hydrophyte Treatment Systems. P.F. Cooper, ed.,

European Community/European Water Pollution Control Association.

Government Accounting Office. 1980. Costly wastewater

treatment plants fail to perform as expected. CED-81-

9. Washington, D.C.

Hammer, D.A., ed. 1989. Constructed Wetlands for Wastewater Treatment. Lewis Publishers, Inc. Chelsea,

Michigan.
20

IAWQ. 1992. Proceedings of international conference on

treatment wetlands, Sydney, Australia. Water Science

& Technology. Vol. 29, No. 4.

IAWQ. 1995. Proceedings of international conference on

treatment wetlands, Guangzhou, China. Water Science

& Technology. Vol. 32, No. 3.

IAWQ. 1997. Proceedings of international conference on

treatment wetlands, Vienna, Austria. Water Science &

Technology. Vol. 35, No. 5.

Kadlec, R.H. and R.L. Knight. 1996. Treatment Wetlands.

CRC Press LLC. Boca Raton, FL.

Moshiri, L., ed. 1993. Constructed Wetlands for Water Quality Improvement. Lewis Publishers, Inc., Chelsea, MI.

Niering, W.A. 1985. Wetlands. Alfred A. Knopf, Inc., New

York, NY.

Reed, S.C., R.W. Crites, and J.E. Middlebrooks. 1995.

Natural Systems for Waste Management and Treatment. Second edition. McGraw-Hill, Inc., New York,

NY.

Tennessee Department of Public Health. 1977. Regulations for plans, submittal, and approval; Control of construction; Control of operation. Chapter 1200-4-2, State

of Tennessee Administrative Rules. Knoxville, TN.

Tennessee Valley Authority. 1993. General Design, Construction, and Operation Guidelines: Constructed Wetlands Wastewater Treatment Systems for Small Users Including Individual Residences. G.R. Steiner and

J.T. Watson, eds. 2nd edition. TVA Water Management

Resources Group. TVA/WM--93/10. Chattanooga, TN.

U.S. Department of Agriculture. 1995. Handbook of Constructed Wetlands. 5volumes. USDA-Natural Resources Conservation Service/US EPA-Region III/

Pennsylvania Department of Natural Resources.

Washington, D.C.

U.S. Environmental Protection Agency. 1988. Design

Manual: Constructed Wetlands and Aquatic Plant Systems for Municipal Wastewater Treatment. EPA/625/

1-88/022. US EPA Office of Research and Development, Cincinnati, OH.

U.S. Environmental Protection Agency. 1993. Subsurface

Flow Constructed Wetlands for Wastewater Treatment:

A Technology Assessment. S.C. Reed, ed., EPA/ 832/

R-93/008. US EPA Office of Water, Washington, D.C.

U.S. Environmental Protection Agency. 1994. Wetlands

Treatment Database (North American Wetlands for

Water Quality Treatment Database). R.H. Kadlec, R.L.

Knight, S.C. Reed, and R.W. Ruble eds., EPA/ 600/C94/002. US EPA Office of Research and Development,

Cincinnati, OH.

U.S. Environmental Protection Agency. 1999. Final Draft --

Guiding Principles for Constructed Treatment Wetlands:

Providing Water Quality and Wildlife Habitat. Developed

by the Interagency Workgroup on Constructed Wetlands

(U.S. Environmental Protection Agency, Army Corps of

Engineers, Fish and Wildlife Service, Natural Resources

Conservation Services, National Marine Fisheries Service, and Bureau of Reclamation). Final Draft 6/8/1999.

http://www.epa.gov/owow/wetlands/constructed/

guide.html

Water Pollution Control Federation. 1990. Natural Systems

for Wastewater Treatment. Manual of Practice FD-16,

S.C. Reed, ed., Water Pollution Control Federation,

Alexandria, VA.
21

2.1 Understanding Constructed Wetlands

Constructed wetlands are wastewater treatment systems

composed of one or more treatment cells in a built and

partially controlled environment designed and constructed

to provide wastewater treatment. While constructed wetlands have been used to treat many types of wastewater

at various levels of treatment, the constructed wetlands

described in this manual provide secondary treatment to

municipal wastewater. These are treatment systems that

receive primary effluent and treat it to secondary effluent

standards and better, in contrast to enhancement systems

or polishing wetlands, which receive secondary effluent

and treat it further prior to discharge to the environment.

This distinction emphasizes the degree of treatment more

than the means of treatment, because the constructed

wetlands described in this manual receive higher-strength

wastewater than the polishing wetlands that have been

widely used as wastewater treatment systems for the last

20 years.

While constructed wetlands discussed in this manual

provide secondary treatment in a community’s wastewater treatment system, this technology also can be used in

combination with other secondary treatment technologies.

For example, a constructed wetland could be placed upstream in the treatment train from an infiltration system to

optimize the cost of secondary treatment. In other uses,

constructed wetlands could discharge secondary effluent

to enhancement wetlands for polishing. Constructed wetlands are not recommended for treatment of raw wastewater. Figure 2-1 portrays a hypothetical wastewater treatment train utilizing constructed wetlands in series.

The distinction between constructed wetlands for secondary treatment and enhancement systems for tertiary

treatment is critical in understanding the limitations of earlier accounts of wetland-based treatment systems and

databases of system performance. Most of the commonly

available information on constructed wetland treatment

systems is derived from data gathered at many larger polishing wetlands and a relatively few smaller constructed

wetlands for secondary treatment. In the past, largely unverified data from these disparate sources has been aggregated, statistically rendered, and then applied as guidance for constructed wetland systems, with predictably

Chapter 2

Introduction to Constructed Wetlands

inconsistent results. In contrast, guidance offered in this

manual is drawn from reliable research data and practical

application in constructed wetlands for secondary treatment of higher-strength municipal wastewater.

Constructed wetlands comprise two types of systems

that share many characteristics but are distinguished by

the location of the hydraulic grade line. Design variations

for both types principally affect shapes and sizes to fit sitespecific characteristics and optimize construction, operation, and performance. Both types of constructed wetlands

typically may be fitted with liners to prevent infiltration,

depending on local soil conditions and regulatory requirements.

Free water surface (FWS) constructed wetlands closely

resemble natural wetlands in appearance and function, with

a combination of open-water areas, emergent vegetation,

varying water depths, and other typical wetland features.

Figure 2-2 illustrates the main components of a FWS constructed wetland. A typical FWS constructed wetland consists of several components that may be modified among

various applications but retain essentially the same features. These components include berms to enclose the

treatment cells, inlet structures that regulate and distribute

influent wastewater evenly for optimum treatment, various

combinations of open-water areas and fully vegetated surface areas, and outlet structures that complement the even

distribution provided by inlet structures and allow adjustment of water levels within the treatment cell. Shape, size,

and complexity of design often are functions of site characteristics rather than preconceived design criteria.

Vegetated submerged bed (VSB) wetlands consist of

gravel beds that may be planted with wetland vegetation.

Figure 2-3 provides a schematic drawing of a VSB system. A typical VSB system, like the FWS systems described

above, contains berms and inlet and outlet structures for

regulation and distribution of wastewater flow. In addition

to shape and size, other variable factors are choice of treatment media (gravel shape and size, for example) as an

economic factor, and selection of vegetation as an optional

feature that affects wetland aesthetics more than performance.

The apparent simplicity and natural function of constructed wetlands may obscure the complexity of interac-
22

Figure 2-1. Constructed wetlands in wastewater treatment train

Figure 2-2. Elements of a free water surface (FWS) constructed wetland

Figure 2-3. Elements of a vegetated submerged bed (VSB) system

Raw

Wastewater

Primary

Treatment

Primary

Effluent

Secondary

Effluent

Final

Discharge

Constructed

Wetlands

Secondary

Treatment

Disinfection

or

Tertiary

Treatment

Pretreated

(Lagoon)

Influent

Inlet Settling Zone

Floating and

Emergent Plants

Submerged

Growth Plants

Floating and Emergent

 Plants

Zone 3

Fully Vegetated

D.O. (–)

H ≤ 0.75 m

Zone 2

Open-Water Surface

D.O. (+)

H ≥ 1.2 m

Zone 1

Fully Vegetated

D.O. (–)

H ≤ 0.75 m

Outlet

Zone

Length (L)

Depth (H)

Width (W)

Variable

Level

Outlet

Length

(L)

Liner Bottom Slope (≤ 1%) Depth (H)

Width

(W)

Effluent

Outlet Zone

Treatment Zone

Top Slope

(Media)

( 0% - flat)

Inlet Zone

Variable-Level

Outlet

Pretreated

(Settled)

Influent
23

tions required for effective wastewater treatment. Unlike

natural wetlands, constructed wetlands are designed and

operated to meet certain performance standards. Once a

constructed wetland is designed and becomes operational,

the system requires regular monitoring to ensure proper

operation. Based on monitoring results, these systems may

need minor modifications, in addition to routine management, to maintain optimum performance.

In this chapter, a basic understanding of constructed

wetland ecology is presented for planners, policy makers,

local government officials, and others involved in the application of constructed wetlands for wastewater treatment.

Basic ecological components and functions of wetlands

are briefly described to bring readers to a common level of

understanding, but detailed descriptions are purposely

omitted for the sake of focus and relative correlation to

treatment performance. To enhance one’s knowledge of

wetland ecology, many publications are commonly available. For designers and operators, general knowledge of

wetland ecology is assumed, and detailed information on

constructed wetlands is offered in succeeding chapters.

While municipal wastewater treatment systems utilizing

constructed wetlands modeled on the functions of natural

wetlands systems are the focus of this manual, related

systems utilizing components of natural wetland systems

also are briefly described. In addition, constructed wetlands

for on-site domestic wastewater systems and non-municipal wastewater treatment are introduced.

Because VSB wetlands are not dependent on wetland

vegetation for treatment performance and do not require

open-water areas, portions of this chapter describe design and management considerations that pertain only to

FWS wetlands. For reference purposes, important terms

are highlighted in bold type and are explained in a glossary at the end of the chapter.

2.2 Ecology of Constructed Wetlands

Constructed wetlands are ecological systems that combine physical, chemical, and biological processes in an

engineered and managed system. Successful construction and operation of an ecological system for wastewater

treatment requires a basic knowledge and understanding

of the components and the interrelationships that compose

the system.

The treatment systems of constructed wetlands are

based on ecological systems found in natural wetlands. A

main distinction between constructed wetlands and natural wetlands is the degree of control over natural processes.

For example, a constructed wetland operates with a relatively stable flow of water through the system, in contrast

to the highly variable water balance of natural wetlands,

mostly due to the effects of variable precipitation. As a result, wetland ecology in constructed wetlands is affected

by continuous flooding and concentrations of total suspended solids (TSS), biochemical oxygen demand (BOD),

and other wastewater constituents at consistently higher

levels than would otherwise occur in nature.

In a constructed wetland, most of the inflow is a predictable volume of wastewater discharged through sewers.

Lesser volumes of precipitation and surface runoff are subject to seasonal and annual variations. Losses from these

systems can be calculated by measuring outflow and estimating evapotranspiration as well as by accounting for

seepage in unlined systems. Even with predictable inflow

rates, however, modeling the water balance of constructed

wetlands must comprehend weekly and monthly variations

in precipitation and runoff and the effects of these variables on wetland hydraulics, especially detention time required for treatment. See Chapter 3 for a more thorough

discussion of modeling concerns.

Temperature variations also affect the treatment performance of constructed wetlands, although not consistently

for all wastewater constituents. Treatment performance for

some constituents tends to decrease with colder temperatures, but BOD and TSS removal through flocculation, sedimentation, and other physical mechanisms is less affected.

In colder months, the absence of plant cover would allow

atmospheric reaeration and solar insolation to occur without the shading and surface covering that plant cover provides during the growing season. Ice cover is another seasonal variable that affects constructed wetlands by altering wetland hydraulics and restricting solar insolation, atmospheric reaeration, and biological activity; however, the

insulating layer provided by ice cover would slow down

the rate and degree of cooling in the water column but

would not affect physical processes such as settling, filtration, and flocculation. Plant senescence and decay also

decreases under ice cover, with a corresponding decrease

in effluent BOD.

2.3 Botany of Constructed Wetlands

Successful performance of constructed wetlands depends on ecological functions that are similar to those of

natural wetlands, which are based largely on interactions

within plant communities. Research has confirmed that

treatment of typical wastewater pollutants (TSS and BOD)

in FWS constructed wetlands generally is better in cells

with plants than in adjoining cells without plants (Bavor et

al., 1989; Burgoon et al., 1989; Gearheart et al., 1989;

Thut, 1989). However, the mechanisms by which plant

populations enhance treatment performance have yet to

be determined fully. Some authors have hypothesized a

relationship between plant surface area and the density

and functional performance of attached microbial populations (EPA, 1988; Reed et al., 1995), but demonstrations

of this relationship have yet to be proven.

Plant communities in constructed wetlands undergo significant changes following initial planting. Very few constructed wetlands maintain the species composition and

density distributions envisioned by their designers. Many

of these changes are foreseeable, and many have little

apparent effect on treatment performance. Other changes,

however, may result in poor performance and the consequent need for increased management. The following sections summarize basic principles of plant ecology that may

aid in understanding of constructed wetlands.
24

2.3.1 Wetland Microbial Ecology

In any wetland, the ecological food web requires microscopic bacteria, or microbes, to function in all of its complex transformations of energy. In a constructed wetland,

the food web is fueled by influent wastewater, which provides energy stored in organic molecules. Microbial activity is particularly important in the transformations of nitrogen into varying biologically useful forms. In the various

phases of the nitrogen cycle, for example, different forms

of nitrogen are made available for plant metabolism, and

oxygen may be either released or consumed. Phosphorus

uptake by plants also is dependent in part on microbial

activity, which converts insoluble forms of phosphorus into

soluble forms that are available to plants. Microbes also

process the organic (carbon) compounds, and release

carbon dioxide in the aerobic areas of a constructed wetland and a variety of gases (carbon dioxide, hydrogen sulfide, and methane) in the anaerobic areas. Plants, plant

litter, and sediments provide solid surfaces where microbial activity may be concentrated.

Microbial activity varies seasonally in cold regions, with

lesser activity in colder months, although the performance

differential in warm versus cold climates is less in full-scale

constructed wetlands than in small-scale, controlled experiments (Wittgren and Maehlum, 1996), apparently because of the multiplicity of physical, chemical, and biological transformations taking place simultaneously over a

larger contiguous area.

2.3.2 Algae

Algae are ubiquitous in wet habitats, and they inevitably

become components of FWS systems. While algae are a

major component in certain treatment systems (for example, lagoons), algae can affect treatment performance

of FWS constructed wetlands significantly. As a result, the

presence of algae must be anticipated in the design stage.

Algae in open areas, especially in areas of submergent

vegetation, can form a living canopy that blocks sunlight

from penetrating the water column to that vegetation, which

results in reduced dissolved oxygen (DO) levels. The presence of open, unshaded water near the outlet of a constructed wetland typically promotes seasonal blooms of

phytoplanktonic algal species, which results in elevated

concentrations of suspended solids and particulate nutrient forms in the effluent.

Several floating aquatic plant species, especially duckweed, have very high rates of primary production, which

result in large quantities of biomass and trapped nonliving

elements accumulating within the fully vegetated portion

of the FWS wetland and pond systems (Table 2-1). Water

hyacinth can also perform well in pond systems in tropical

climates to enhance TSS and algal removal. However, both

species block sunlight and lower DO levels by eliminating

atmospheric reaeration at the water/air interface.

High growth rates of these plants have led to specialized wastewater treatment systems that use these plants

for harvesting nutrients from wastewater. The disadvantages of harvesting these plants arise from their low %

solids (typically less than 5% on a wet-weight basis) and

the consequent need for drying prior to disposal, which

simultaneously creates secondary odor and water-quality

problems. For disposal, harvested duckweed, which has a

high protein content, typically has been incorporated into

agricultural soils as green manure, and water hyacinths

have been partially dried and landfilled or allowed to decompose in a controlled environment to produce methane

as a useful by-product. However, numerous attempts to

demonstrate beneficial and cost-effective by-product recovery have been mostly unsuccessful under North American social and economic conditions.

2.3.3 Emergent Herbaceous Plants

Emergent herbaceous wetland plants are very important structural components of wetlands. Their various adaptations allow competitive growth in saturated or flooded

soils. These adaptations include one or more of the following traits: lenticels (small openings through leaves and

stems) that allow air to flow into the plants; aerenchymous

tissues that allow gaseous convection throughout the length

of the plant, which provides air to plant roots; special morphological growth structures, such as buttresses, knees,

or pneumatophores, that provide additional root aeration;

adventitious roots for absorption of gases and plant nutrients directly from the water column; and extra physiological tolerance to chemical by-products resulting from growth

in anaerobic soil conditions.

The primary role of emergent vegetation in FWS systems is providing structure for enhancing flocculation, sedimentation, and filtration of suspended solids through idealized hydrodynamic conditions. Emergent wetland plant

species also play a role in winter performance of FWS

constructed wetlands by insulating the water surface from

cold temperatures, trapping falling and drifting snow, and

reducing the heat-loss effects of wind (Wittgren and

Maehlum, 1996).

Limited information is available to demonstrate significant or consistent effects of plant species selection on

constructed wetland performance. For example, in two similar FWS treatment cells at the Iron Bridge Wetland in

Florida, the major difference between the cells was the

dominant plant species. Bulrush appeared to perform

nearly the same as cattail in treatment of BOD, TSS, total

nitrogen (TN), and total phosphorus (TP). As research and

application of constructed wetlands have expanded, documentation of actual performance differences between

emergent marsh plant species in constructed wetlands has

become increasingly less valuable to constructed wetland

designers.

The wetland designer is strongly encouraged to seek

information from experienced local wetland practitioners

when selecting emergent herbaceous species to ensure

selection of locally successful species. Table 2-2 provides

guidelines for initial selection and establishment of plant

species adapted to wetland environments.
25

Table 2.1 Characteristics of plants for constructed wetlands

General Types General Characteristics Function or Importance Function or Importance Design & Operational

of Plants and Common Examples to Treatment Process for Habitat Considerations

Free-Floating

Aquatic Roots or root-like structures Primary purposes are nutrient Dense floating mats limit Duckwood is a natural

suspended from floating uptake and shading to retard oxygen diffusion from the invasive species in

leaves. Will move about with algal growth. Dense floating mats atmosphere and block North America. No

water currents. Will not stand limit oxygen diffusion from the sunlight from submerged specific design is

erect out of the water. atmosphere. Duckweed will be plants. Plants provide shelter required.

Common duckweed (Lemna), present as an invasive species. and food for animals.

Big duckweed (Spirodela).

Rooted Floating Usually with floating leaves, Primary purposes are providing Dense floating mats limit Water depth must be

Aquatic but may have submerged structure for microbial attachment oxygen diffusion from the designed to promote

leaves. Rooted to bottom. and releasing oxygen to the water atmosphere and block the type of plant (i.e.

Will not stand erect out of the column during daylight hours. sunlight from submerged floating, submerged,

water. Water lily (Nymphea), Dense floating mats limit oxygen plants. Plants provide emergent) desired

Pennywort (Hydrocotyle). diffusion from the atmosphere. shelter and food for animals. while hindering other

types of plants.

Submerged Usually totally submerged; Primary purposes are providing Plants provide shelter and Retention time in open

Aquatic may have floating leaves. structure for microbial attachment. food for animals (especially water zone should be

Rooted to bottom. Will not and providing oxygen to the water fish). less than necessary

stand erect in air. Pondweed column during daylight hours. to promote algal

growth which can

(Potamogeton), Water weed destroy these plants

(Elodea). through sunlight

blockage.

Emergent Herbaceous (i.e. non-woody). Primary purpose is providing Plants provide shelter and Water depths must be

Aquatic Rooted to the bottom. Stand structure to induce enhanced food for animals. Plants in the range that is

erect out of the water. Tolerate flocculation and sedimentation. provide aesthetic beauty for optimum for the

flooded or saturated conditions. Secondary purposes are shading humans. specific species

Cattail (Typha), Bulrush to retard algal growth, windbreak chosen (planted).

(Scirpus), Common Reed to promote quiescent conditions for

(Phragmites). settling, and insulation during winter

months.

Shrubs Woody, less than 6 m tall. Treatment function is not defined: Plants provide shelter and Possible perforation of

Tolerate flooded or saturated it is not known if treatment data food for animals (especially liners by roots.

soil conditions. Dogwood from unsaturated or occasionally birds). Plants provide aesthetic

(Cornus), Holly (Ilex). saturated phytoremediation sites beauty for humans.

in upland areas is applicable to

continuously saturated wetland

sites.

Trees Woody, greater than 6 m tall. (same as for shrubs) (same as for shrubs) (same as for shrubs)

Tolerate flooded or saturated

soil conditions. Maple (Acer).

Willow (Salix).

2.3.4 Plant Nutrition and Growth Cycles

Wetland plants require optimum environmental conditions in each phase of their life cycles, including germination and initial plant growth, adequate nutrition, normal

seasonal growth patterns, and rates of plant senescence

and decay. For more detailed information on wetland plant

ecology, the nonbiologist is referred to the wetland ecology text by Mitsch and Gosselink (1993) and portions of

the constructed wetland text by Kadlec and Knight (1996).

A wide variety of references describe growth cycles, timing of seed release, overwintering ability, energy cycling,

and other characteristics and processes that provide wetland plant species with a competitive advantage in their

natural habitats; the reader is referred to other sources for

detailed information. An overview of important characteristics follows.

Emergent herbaceous wetland species planted early in

the growing season in temperate climates generally multiply by vegetative reproduction to a maximum total standing biomass in late summer or early fall within a single

growing season. This biomass may represent multiple

growth and death periods for individual plants during the

course of the growing season, or it may represent a single

emergence of plant structures, depending on the species.

For many species, seeds are produced along with maximum standing crop and released with maturation in the

fall for early germination in the spring.
26

For some species with high lignin content, particularly

cattail, bulrush, and common reed, much of the plant remains standing as dead biomass that slowly decays during the winter season. In FWS systems, this standing dead

biomass provides additional structure for enhanced flocculation and sedimentation that is important in wetland

treatment performance throughout the annual cycle. Dead

biomass, both standing and fallen, also is important to root

viability under flooded, winter conditions because of the

insulating layer it provides, in addition to its contribution to

the internal load on the system.

Like all plants, wetland plants require many macro- and

micronutrients in proper proportions for healthy growth.

While municipal wastewater can supply adequate quantities of these limiting nutrients, other types of wastewater,

including industrial wastewater, acid mine drainage, and

stormwater, may not.

Nitrogen and phosphorus are key nutrients in the life

cycles of wetland plants. However, plant uptake of nitrogen and phosphorus is not a significant mechanism for

removal of these elements in most wetlands receiving partially treated municipal wastewater because nitrogen and

phosphorus are taken up and released in the cycle of plant

growth and death. Nonetheless, undecomposed litter from

dead biomass provides storage for phosphorus, metals,

and other relatively conservative elements (Kadlec and

Knight, 1996).

While uptake rates of nitrogen and phosphorus are potentially high, harvesting plant biomass to remove these

nutrients has been limited to floating aquatic plant communities, in which the plants can be harvested with only

brief altering of system performance. Although common

reed is harvested annually from certain European constructed wetlands as a by-product (and not for nutrient reduction), full-scale constructed wetlands where plants are

Table 2.2 Factors to Consider in Plant Selection (adapted from Thunhorst, 1993)

Factors Comments

Consult local experts The number of professional wetland scientists, practitioners, and plant nurseries has increased dramatically in the

past 10 years. Help from an experienced, local person should be available from a variety of sources, including

government agencies and private companies.

Native species Using plants that grow locally increases the likelihood of plant survival and acceptance by local officials.

Invasive or aggressive Plants that have extremely rapid growth, lack natural competitors, or are allelopathic* can crowd out all other spespecies cies and destroy species diversity. State or local agencies may ban the use of some species.

Tolerant of high Unlike natural wetlands, constructed wetlands will receive a continuous inflow of wastewater with high nutrient

nutrient load concentrations. Plants that can not tolerate this condition will not survive.

Tolerant of continuous Unlike natural wetlands, which may experience periodic or occasional dry periods, constructed wetlands will

flooding receive a continuous inflow of wastewater. Plants that require periodic or occasional drying as part of their

reproductive cycle will not survive.

Growth characteristics Perennial plants are generally preferred over annual plants because plants will continue growing in the same area

and there is no concern about seeds being washed or carried away. For emergent species, persistent plants are

generally preferred over semi- or non-persistent plants because the standing plant material provides added

shelter and insulation during the winter season.†

Available form for planting Costs of obtaining and planting the plants will vary depending on the form of planting material, which may be

available in a variety of forms depending on the plant species. Entire plant forms (e.g. bare root plants or plugs) will

usually cost more than partial plant material (e.g. seeds or rootstock), but the plant supplier may guarantee a

higher survival rate.‡

Rate of growth Slower growing plants will require a greater number of plants, planted closer together, at start-up to obtain the

same density of plant coverage in the initial growing season.

Wildlife benefits If the wetland is to be used for habitat, plants that provide food, shelter/cover and nesting/nursery for the desired

animals should be chosen.

Plant diversity Mono-cultures of plants are more susceptible to decimation by insect or disease infestations; catastrophic

infestations will temporarily affect treatment performance. Greater plant diversity will also tend to encourage a

greater diversity of animals.

* Allelopathic - plants that have harmful effects on other plants by secreting toxic chemicals

† Perennial - aboveground portion dies, but below-ground portion remains dormant and sprouts in the next growing season.

Annual - entire plant dies and reproduction is only by seed produced before the plant dies.

Persistent - aboveground dead portions remain upright through the dormant season.

Semi-persistent - aboveground dead portions may remain standing for some part of the dormant season before falling into clumps.

Non-persistent - aboveground dead portions decay and wash away at the end of the growing season.

‡ Bare root plant - seedling with soil washed from roots. Plug - seedling with soil still on roots. Rootstock - piece of underground stem (rhizome).
27

routinely harvested have not been documented in the

United States.

2.4 Fauna of Constructed Wetlands

The role that animal species may play in constructed

wetlands is a consideration for management of FWS wetlands. Animals typically compose less biomass than do

wetland plants, but animals are able to alter energy and

mass flows disproportionately to their biomass contribution. During outbreaks of insect pests in constructed wetlands, for example, entire marshes and floating aquatic

plant systems can be defoliated, which interrupts mineral

cycles and upsets water-quality treatment performance.

In another example, the rooting action of bottom-feeding

fish (primarily carp) causes sediment resuspension, which

affects performance of constructed wetlands in removing

suspended solids and associated pollutants. The presence

of large seasonal waterfowl populations has had similar

results in constructed wetlands at Columbia, Missouri, and

elsewhere. In VSB wetlands, only avian species play a

significant role.

While wildlife species play generally positive, secondary roles in constructed wetlands, their presence also may

generate unintended consequences. Bird species common

to wetland environments, for example, typically attract

birdwatchers, who may provide public support for municipalities and industries employing this treatment technology. The presence of the public at constructed wetlands

for secondary treatment, however, necessitates management efforts to ensure adequate protection from human

health and safety risks presented by exposure to primary

effluent (see also section 2.6). Conversely, regulatory concern for potentially vulnerable wildlife species has impeded

plans for constructed wetlands at certain sites and for certain wastewaters with toxic constituents.

Free water surface wetlands closely resemble the ecology of natural wetlands and aquatic habitats, and they inevitably attract animal species that rely on wet environments during some or all of their life history. All animal

groups are represented in constructed wetlands: protozoans, insects, mollusks, fish, amphibians, reptiles, birds, and

mammals. Table 2-3 summarizes animal species that may

be found in constructed wetlands.

2.5 Ecological Concerns for Constructed

Wetland Designers

Wetland ecology is integral to the success of constructed

wetlands because of their complexity and their accessibility to wildlife. While the ecology of VSB systems relates

more to its subsurface than its surface environment, wetland plants and other surface features that are characteristic of VSB wetlands also require consideration.

Table 2.3 Characteristics of Animals Found in Constructed Wetlands

Members of Group Commonly Found in Function or Importance to Treatment

Animal Group Treatment Wetlands Process Design & Operational Considerations

Invertebrates, A wide variety will be present, but diversity Undoubtedly play a role in Mosquito control must be considered;

including and populations will vary seasonally and chemical and biological cycling and mono-cultures of plants are more

protozoa, spatially. transformations and in supporting food susceptible to decimation by insect

insects, spiders, web for higher organisms, but exact infestations.

and crustaceans functions have not been defined

Fish Species adapted to living at or near the Consumers of insects and decaying Anaerobic conditions will limit

surface (mosquitofish, mudminnow); material (e.g. mosquitofish eat populations; nesting areas required;

species adapted to living in polluted mosquito larvae). bottom-feeders can uproot plants and

waters (bowfin, catfish, killifish, carp). resuspend sediments.

Amphibians and Frogs, alligators, snakes, turtles Consumers of lower organisms Turtles have an uncanny ability to fall

Reptiles into water control structures and to

get caught in pipes, so turtle exclusion

devices are needed; monitoring of

control structures and levees for

damage or obstruction is needed.

Birds A wide variety (35-63 species*) are Consumers of lower organisms Heavy use, especially by migratory

present, including forest and prairie waterfowl, can contribute to pollutant

species as well as waterfowl, but load on a seasonal basis.

diversity and populations vary

seasonally and spatially.

Mammals Small rodents (shrews, mice, voles); Consumers of plants and lower Nutria and muskrat populations can

large rodents (rabbits, nutria, muskrats, organisms reach nuisance levels, removing

beaver); large grazers (deer); large vegetation and destroying levees;

carnivores (opossums, raccoons, foxes). structural controls and animal

removal may be required.

* McAllister, 1992, 1993a, 1993b
28

Constructed wetlands invariably attract wildlife, a factor

that must be considered in the design and management of

constructed wetlands. As components of an ecological

community, animals in general perform vital ecological functions in constructed wetlands. Specific roles of animals in

the development and operation of constructed wetlands,

however, are not well researched. Experience has shown

that many animals are beneficial elements in constructed

wetlands, but many other are nuisance species. Proper

attention to desirable and undesirable wildlife species, as

well as primary and ancillary functions of constructed wetlands, will aid the success of a constructed wetland.

2.5.1 Primary and Ancillary Functions of

Constructed Wetlands

Primary functions of most constructed wetlands include

water storage and water-quality improvement. Some of

these constructed wetlands are designed intentionally for

ground water recharge. Numerous other functions attributed to natural wetlands are important in constructed wetlands and are described in succeeding chapters.

Ancillary functions include primary production of organic

carbon by plants; oxygen production through photosynthesis; production of wetland herbivores, as well as predator species that range beyond the wetland boundaries;

reduction of export of organic matter and nutrients to downstream ecosystems; and creation of cultural values in terms

of educational and recreational resources. One or more of

these ancillary functions may be an important goal in some

constructed wetland projects. For detailed descriptions of

ancillary functions, the reader is referred to information

presented elsewhere (Feierabend, 1989; Sather, 1989;

Knight, 1992).

2.5.2 Wildlife Access Controls

Successful wildlife management in FWS wetlands requires maintaining a balance between attracting beneficial species and controlling pest species (EPA, 1993a).

While most wildlife species in wetlands are attractive but

often unnoticed, many species are attractive for aesthetic

reasons but are impediments to the success of constructed

wetlands. Nuisance species in constructed wetlands include burrowing rodents, especially beavers, nutria, and

muskrats, which burrow through berms and levees and

consume beneficial emergent vegetation; mosquitoes,

which cause annoyance and health concerns; and certain

bottom-feeding fish, such as carp, which uproot aquatic

vegetation and cause increases in effluent TSS and associated pollutants by stirring up sediments and resuspending them in the water column. Waterfowl in large numbers

also may be undesirable because they cause similar problems, and their nutrient-rich droppings place additional

demands on the water-quality performance of constructed

wetlands.

Control of wildlife access in constructed wetlands is highly

site-specific; as a result, control measures must be based

on geographic location, nuisance species, wetland design,

and preferred levels of management. Control methods are

applied throughout the planning, construction, and operation of constructed wetland projects. Control of carp, for

example, can be anticipated during design and managed

with winter drawdown of water levels and subsequent indepth freezing in northern climates. Also effective is drawdown and physical removal of stranded individuals, but

this method is more labor intensive and less effective in

eradicating carp populations. Large rodents can be

screened out of culverts to limit access and prevent damming; however, trapping and physical removal may be

needed to prevent burrowing and subsequent undermining of banks and other damage. For waterfowl control, limited open-water areas will discourage many species, but

treatment requirements will dictate the size and use of these

zones. Netting suspended over unavoidable open-water

areas can prevent their use for feeding, but this method

deviates from the intent to incorporate natural methods of

wildlife control.

Wetland wildlife species frequently have home ranges

well outside the borders of an individual constructed wetland cell; consequently, they can become a public resource

that may need to be protected and promoted for reasons

unrelated to their perceived value to constructed wetlands.

Although the values of constructed wetlands for wildlife

habitat may be subject to public and scientific debate, this

topic nonetheless must be considered in all project phases

to determine optimum design and management features

to promote or discourage the presence of wildlife (Knight,

1997; Worrall et al., 1996).

2.5.3 Mosquito Habitat Controls

Mosquitoes may be integral components of the ecological food web, but mosquitoes generally are considered a

pest species. While a constructed wetland’s attractiveness

to wildlife may be regarded as a benefit to the human community, the potential for breeding mosquitoes can be an

obstacle to permitting, funding, and other steps essential

to the siting of a constructed wetland.

Several methods of mosquito control can be employed

in the planning, construction, and operation of constructed

wetlands. Predation is one means. Mosquito fish have been

found to be effective in reducing mosquito populations when

habitat conditions are optimized by manipulating water levels and when channels are kept free of dead vegetation.

Drawdown of water levels aids mosquito fish spawning in

spring and provides the fish with better access to mosquito larvae during mosquito breeding season (Dill, 1989).

In warm climates, mosquito fish habitat must be monitored

for excessive water temperatures and fluctuations in effluent strength and content. Bats and several avian species

also are effective predators, but planning and managing

optimum conditions have yet to be standardized.

In the planning and construction stages, management

of mosquito habitat can be enabled with steep slopes on

water channels that reduce standing water area in shallow

areas. In contrast to this design is the use of more natural,

undulating banks that have been popular in polishing wet-
29

lands. This natural appearance is more visually appealing

but is ineffective for mosquito-control purposes. A channel

profile that has been effective in mosquito control is a steepsided channel flanked by relatively flat aprons leading outward to steep-sided banks (Dill, 1989). This profile allows

the facility operator to draw down water levels to the lower

channel during the mosquito-breeding season. Figure 2-4

illustrates this design. With standing water eliminated from

emergent vegetation in the shallow flanks of the channel,

deeper water in the lower channel provides an environment more conducive to mosquito predation by fish species. Flexible drainage capability is essential to this means

of control.

Water spray systems also have been used for mosquito

control, but such mechanical systems are inconsistent with

the passive nature of constructed wetlands, which utilize

natural systems to accomplish wastewater treatment and

manage ancillary concerns.

Vegetation management is another approach to mosquito control, especially in the absence of water-level control features (Dill, 1989). Taller vegetation especially needs

management. Cattails and bulrushes, for example, tend to

fall over late in the growing season, which creates conditions favorable for mosquito reproduction in the following

growing season, as well as unfavorable conditions for predation by mosquito fish (Martin and Eldridge, 1989). Channels planted with lower-growing vegetation and cleared

annually of dead standing stock can reduce mosquito populations and optimize predation, providing that this vegetation imparts the same structural role beneath the water

surface.

Larvicide is a proven means of active mosquito control

when employed in conjunction with other management

techniques. A bacterium (Bacillus sphaericus) has been

found effective in reducing culex mosquitoes, one of the

most common species in the United States. Tests have

indicated that a commercial larvicide containing the bacteria may be capable of eliminating most of the populations

of culex in treatment lagoons (WaterWorld, 1996). The

concentrated bacteria in powdered form is applied to standing water as a coating on granulated corncobs, which

quickly releases protein crystals and bacteria spores to

the water surface. Upon ingestion, the bacteria enter mosquito larvae tissues through pores in the gut wall and multiply rapidly, and the infected larvae typically die within two

days. However, fully vegetated zones are more difficult to

treat than open water zones or lagoons.

2.6 Human Health Concerns

Many studies of constructed wetlands’ biological effectiveness and attractiveness to humans for aesthetic and

cultural reasons have focused on polishing wetlands that

receive secondary effluent, which are outside the focus of

this manual. At many of these successful polishing wetlands for tertiary treatment, interpretive centers and signage

invite visitors, and boardwalks and naturalists guide them

through the outdoor experience. Constructed wetlands that

receive primary effluent for secondary treatment, on the

other hand, may not be visitor-friendly places, and human

visitors may best enjoy them from the periphery for several reasons.

Partially treated wastewater in a constructed wetland for

secondary treatment, despite the proven effectiveness of

this ecological approach to treatment, presents essentially

the same risks to human health as wastewater in primary

treatment and lagoons. Risk of dermal contact and possible transmission of disease is equally unappealing in FWS

wetlands for secondary treatment as it is in open lagoons.

This concern is distinguished from human interaction with

Inlet

Distribution

Pipe Vegetated

Zone

(1)

Vegetated

Zone

(3)

Open Water (2)

Emergent Plants

VariableLevel

Outlet

Floating Plants

Outlet

Collector

Pipe

Submerged Plants

Figure 2-4. Profile of a three-zone FWS constructed wetland cell
30

polishing systems, where influent wastewater has already

met effluent quality requirements which are set by regulatory authorities.

In constructed wetlands receiving primary effluent, human exposure to wastewater is a greater concern at the

inlet end of the system, where influent has achieved primary treatment only. Lesser concern for human exposure

is warranted at the outlet end, where wastewater has been

treated to the quality of secondary treatment or better, which

is the quality of wastewater entering the polishing wetlands

that have been popular for environmental awareness and

education activities.

As a result, humans must be considered an unwanted

species in most areas of FWS wetlands treating municipal

wastewater to meet secondary treatment (defined as 30

mg/L of BOD and TSS). Nonetheless, constructed wetlands can serve as recreational areas and outdoor laboratories, especially at the outlet end where wastewater has

been treated to secondary effluent standards. Management

considerations may include the public’s access, perceptions, and exposure to health threats (Knight, 1997). To

effectively address these concerns, fencing, signage, and

other controls must be considered in the proposal stage

as well as in design and operation of the system.

Mosquito populations may represent merely an annoyance factor to be managed, as described above, but some

species of mosquitoes also carry a health risk that must

be addressed. In warmer climates, including the southern

United States, the encephalitis mosquito (Culex tarsalis)

thrives in the extended breeding season provided by constructed wetlands, but water-level manipulation and mosquito fish predation in the two-tiered pond design described

previously have been effective in controlling these mosquito populations (Dill, 1989). The two-tiered design allows

water levels to be drawn down to concentrate prey species (mosquitoes) in smaller areas for more efficient consumption by predators (mosquito fish).

Most of the health concerns described above do not apply

to VSB systems, in which wastewater typically is not exposed at the land surface.

2.7 On-site System Applications

On-site constructed wetland systems may also be applied to wastewater treatment and disposal at individual

properties. On-site constructed wetlands generally utilize

the same technologies as the municipal VSB systems described in this manual, and they share with municipal systems the advantages of cost-effectiveness and low-maintenance requirements. However, on-site constructed wetlands are distinguished typically by final effluent discharge

to soils instead of surface water. For purposes of this discussion, on-site constructed wetland systems treat septic

tank effluent, or primary effluent, in small-scale VSB systems for subsurface disposal to soils.

On-site constructed wetlands also differ from municipal

systems in scale. On-site constructed wetlands typically

occupy only a few hundred square feet. Municipal VSB

systems may serve hundreds of residential, commercial,

and industrial properties, while on-site systems would serve

a single home or several residences in a cluster.

An on-site VSB system typically consists of a lined VSB

that receives primary effluent from a septic tank, and in

some designs, a second VSB that receives effluent from

the upstream VSB system. The second VSB can be unlined to allow treated wastewater to infiltrate to soil for disposal. Variations of this treatment train include use of

supplemental absorption trenches to facilitate soil absorption and direct surface discharge with or without subsequent disinfection. Each VSB typically is planted with wetland vegetation.

Applied studies and research experiments of on-site

constructed wetland systems have shown adequate treatment performance for most wastewater constituents, including BOD, TSS, and fecal coliforms, with variations in

performance for removal of ammonia nitrogen (Burgan and

Sievers, 1994; Huang et al., 1994; Johns et al., 1998;

Mankin and Powell, 1998; Neralla et al., 1998; White and

Shirk, 1998).

2.8 Related Aquatic Treatment Systems

Several types of aquatic treatment systems have been

constructed to treat municipal and other wastewaters, and

most of these systems fall outside the definition of constructed wetlands discussed in this manual. These other

types of systems are briefly described to provide the reader

with additional background and references to source material.

Polishing wetlands have been used also to remove trace

metals, including cadmium, chromium, iron, lead, manganese, selenium, and zinc in a variety of situations. The

primary removal mechanism for metals in wastewater appears to be sedimentation. Plant uptake results in deposition of metals to soil via plant roots and requires harvest of

plants to partially remove metals from the system. In some

cases, however, effluent concentrations of metals have

exceeded influent levels, apparently due to evaporation of

wastewater.

One proprietary treatment system, which among its many

manifestations has used both FWS-like and VSB-like treatment units as part of its treatment train, is known as the

Advanced Ecologically Engineered System (AEES), or

“Living Machine.” This system incorporates conventional

treatment system components, including sedimentation/

anaerobic bioreactors, extended aeration, clarifiers, fixedfilm reactors, and a final clarifier, sometimes with a VSB

for polishing, in a greenhouse setting. The AEES was applied to four demonstration projects funded with federal

grants. The four projects underwent evaluation of treatment performance for various wastewater types and settings (e.g., raw wastewater in a moderate climate, raw

wastewater at higher flow rates in a colder climate, in situ

water-quality improvements to pond water, and polishing
31

of secondary effluent). One of the demonstration projects

also was evaluated by an independent firm under contract

to the U.S. EPA (EPA, 1997b). Results of performance

evaluations indicated that wastewater treatment met performance goals for certain wastewater constituents; other

goals were unmet. Although this technology is presented

by its developers as a type of natural system, the use of

wetland plants appears to influence aesthetics more than

treatment performance. The reader is directed to other

sources for further information (EPA, 1993b; EPA, 1997b;

Living Technologies, 1996; Reed et al., 1995; Todd and

Josephson, 1994).

Floating macrophyte systems rely only partially on treatment processes provided by wetlands and require mechanized components to achieve the intended treatment performance. Larger duckweed systems and water hyacinth

systems utilize mechanical systems to remove floating

macrophytes. Both have been employed to treat wastewater by removing some of the wastewater constituents,

primarily BOD and TSS. In both systems, removal of plants

usually requires additional mechanical systems for drying,

disposal, and other residuals handling (Zirschky and Reed,

1988).

2.9 Frequently Asked Questions

1. What are constructed wetlands?

The term “constructed wetlands” refers to a technology designed to employ ecological processes found

in natural wetland ecosystems. These systems utilize wetland plants, soils, and associated microorganisms to remove contaminants from wastewater. As

with other natural biological treatment technologies,

wetland treatment systems are capable of providing

additional benefits. They are generally reliable systems with no anthropogenic energy sources or chemical requirements, a minimum of operational requirements, and large land requirements. The treatment

of wastewater using constructed wetland technology

also provides an opportunity to create or restore wetlands for environmental enhancement, such as wildlife habitat, greenbelts, passive recreation associated

with ponds, and other environmental amenities.

2. What are wetland treatment systems?

The term “wetland treatment system” generally refers to two types of passive treatment systems. One

type of system is a free water surface (FWS) constructed wetland, which is a shallow wetland with a

combination of emergent aquatic plants (cattail, bulrush, reeds, and others), floating plants (duckweed,

water hyacinth, and others), and submergent aquatic

plants (sago pondweed, widgeon grass, and others).

A FWS wetland may have open-water areas dominated by submergent and floating plants, or it may

contain islands for habitat purposes. It may be lined

or unlined, depending on regulatory and/or performance requirements. These systems exhibit complex aquatic ecology, including habitat for aquatic

and wetland birds.

A second type of system is termed “vegetated submerged bed (VSB)” and is known to many as a subsurface flow wetland. A VSB is not an actual wetland because it does not have hydric soils. Emergent wetland plants are rooted in gravel, but wastewater flows through the gravel and not over the

surface. This system is also shallow and contains

sufficiently large gravel to permit long-term subsurface flow without clogging. Roots and tubers (rhizomes) of the plants grow into pore spaces in the

gravel. Most current data indicate that these systems perform as well without plants as with plants;

as a result, wetland ecology is not a critical factor

in VSB systems.

3. Are constructed wetlands reliable? What do they

treat?

Constructed wetlands are an effective and reliable

water reclamation technology if they are properly

designed, constructed, operated, and maintained.

They can remove most pollutants associated with

municipal and industrial wastewater and stormwater

and are usually designed to remove contaminants

such as biochemical oxygen demand (BOD) and

suspended solids. Constructed wetlands also have

been used to remove metals, including cadmium,

chromium, iron, lead, manganese, selenium, zinc,

and toxic organics from wastewater.

4. How does a constructed wetland treat wastewater?

A natural wetland acts as a watershed filter, a sink

for sediments and precipitates, and a biogeochemical engine that recycles and transforms some of

the nutrients. A constructed wetland performs the

same functions for wastewater, and a constructed

wetland can perform many of the functions of conventional wastewater treatment trains (sedimentation, filtration, digestion, oxidation, reduction, adsorption, and precipitation). These processes occur sequentially as wastewater moves through the

wetland, with wastewater constituents becoming

comingled with detritus of marsh plants.

5. What is the difference between treatment and enhancement wetlands?

Constructed wetlands generally are designed to

treat municipal or industrial effluents as well as

stormwater runoff. Enhancement marshes, or polishing wetlands, are designed to benefit the community with multiple uses, such as water reclamation, wildlife habitat, water storage, mitigation banks,

and opportunities for passive recreation and environmental education. Both types of wetland systems can be designed as separate systems, or
32

important attributes of each can be integrated into

a single design with multiple treatment and enhancement objectives.

6. Can a constructed wetland be used to meet a secondary effluent standard?

Both FWS and VSB constructed wetlands can be

used to meet a 30/30 mg/L BOD and TSS discharge

standard. It is not advisable to put raw wastewater

into a constructed wetland.

7. Can a constructed wetland be used to meet an

advanced secondary/tertiary discharge standard?

With sufficient pretreatment and wetland area, FWS

constructed wetlands can meet discharge standards

of less than 10 mg/L BOD, TSS, and TN on a

monthly average basis. Many examples of FWS

wetland systems meeting these standards on a

monthly average basis can be found in the United

States (EPA, 1999). VSB systems have been used

extensively in England for polishing secondary effluents and treating effluent from combined sanitary and storm sewers. In the U.S., they have generally not performed well in consistently reaching

advanced treatment goals with primary treatment

influent.

8. How much area is required for constructed wetlands?

As a general rule, a constructed wetland receiving

wastewater with greater degrees of pretreatment

(for example, primary clarification, oxidation pond,

trickling filter, etc.) requires less area than a constructed wetland receiving higher-strength wastewater. Historically, constructed wetlands designers

have employed from <2 to over 200 acres/MGD (4

to 530 L/m2

-d). However, there is no generic answer to the question since it depends on the effluent criteria to be met and buffer areas required.

9. Do these systems have to be lined?

The requirement for liners in constructed wetlands

depends on each state’s regulatory requirements

and/or the characteristics of surface and subsurface soils. In a general sense, if soils are porous

(e.g., sands), well-drained, and contain small

amounts of loams, clays, and silts, lining is likely to

be a requirement for constructed wetlands. On the

other hand, if soils are poorly drained and composed

mostly of clays, then lining might not be required.

These systems would tend to produce a layer of

peat on the bottom that would reduce infiltration with

time. The concept of a “leaky wetland,” which may

take advantage of natural processes to purify wastewater as it moves downward through soil to recharge the ground water, may be considered a potential benefit in certain areas.

10. What is the role of the plants in constructed wetlands?

In FWS constructed wetlands, plants play several

essential roles. The most important function of

emergent and floating aquatic plants is providing a

canopy over the water column, which limits production of phytoplankton and increases the potential

for accumulation of free-floating aquatic plants (e.g.,

duckweed) that restrict atmospheric reaeration.

These conditions also enhance reduction of suspended solids within the FWS constructed wetland.

Emergent plants play a minor role in taking up nitrogen and phosphorus. The effect of litter fall from

previous growing seasons as it moves through the

water column and eventually decomposes into humic soil and lignin particles may be significant in

terms of effluent quality.

The role of plants in VSB systems is not clear. Initially it was believed that translocation of oxygen

by plants was a major source of oxygen to microbes

growing in the VSB media, and therefore plants

were critical components in the process. However,

side-by-side comparisons of planted and unplanted

systems have not confirmed this belief. Nevertheless, planted VSB systems are more desirable aesthetically than unplanted horizontal rock-filter systems, and plants do not appear to hinder performance of VSB systems.

11. How much time is needed for a constructed wetland to become fully operational and meet discharge

requirements?

For FWS wetland systems, several growing seasons may be needed to obtain the optimum vegetative density necessary for treatment processes.

The length of this period is somewhat dependent

on the original planting density and the season of

the initial planting. Effluent quality has been observed to improve with time, suggesting that vegetation density and accumulated plant litter play an

important role in treatment effectiveness.

VSB systems also require more than one growing

season to achieve normal wetland plant densities.

However, the time required for VSB systems to

become fully operational is considerably less than

FWS systems because of the minor role of plants

in the treatment process. Development of the microbial biomass in the media of a VSB system typically requires from three to six months.

12. How long can a FWS wetland operate before accumulated plant material and settled solids need to

be removed?

FWS wetland systems receiving oxidation pond

effluent may operate for 10 to 15 years without the

need to remove accumulated litter and settled
33

nondegradable influent solids. Treatment capacities of these wetlands have not shown a decrease

in treatment effectiveness with time. However, it is

assumed that further experience will reveal that

there is a finite period of accumulation that will result in the need to remove solids. In both types of

systems, the bulk of the solids accumulation occurs at the influent end of the system. As a result,

solids may need to be removed from only a portion

of the system that may be as small as 10 to 25% of

the surface area.

13. How much effort is required to operate and maintain a constructed wetland?

These systems require a minimum of operational

control. Monthly or weekly inspection of weirs and

weekly sampling typically are required at the effluent end, and periodic sampling between multiple

cells is recommended.

Maintenance of constructed wetlands generally is

limited to the control of unwanted aquatic plants

and control of disease vectors, especially mosquitoes. Harvesting of plants generally is not required,

but annual removal or thinning of vegetation or replanting of vegetation may be needed to maintain

flow patterns and treatment functions.

Effective vector control can be achieved by appropriately applying integrated pest management practices, such as introducing mosquito fish or providing habitat for mosquito-eating birds and bats. Bimonthly monitoring of mosquito larvae and pupae

and applications of larvacides may be required on

an as-needed basis.

Sediment accumulation typically is not a problem

in a well-designed and properly operated constructed wetland, thus partial dredging is required

only rarely.

These tasks would require approximately one day

per week of labor for a wetland system treating a

flow of one million gallons per day (MGD) (3,880

m3

/d) or less, and monitoring may be the most demanding task.

14. Do constructed wetlands produce odors?

Conventional wastewater treatment processes produce odors mostly associated with anaerobic decomposition of human waste and food waste found

in sewage. These odors usually are concentrated

in areas of small confinement and point discharges,

like influent pump stations, anaerobic digesters, and

sludge-handling processes. Wetlands, in contrast,

incorporate normal processes of decomposition

over a relatively large area, potentially diluting odors

associated with the natural decomposition of plant

material, algae, and other biological solids. However, wetland treatment systems receiving septic

tank and primary effluents can release anaerobic

odors around the inlet piping, and both types are

generally anaerobic, which makes odor generation

a major operational concern.

15. Are mosquitoes a potential problem with constructed wetlands? If so, how are they managed?

Mosquitoes generally are not a problem in properly

designed and operated VSB systems. However,

mosquitoes can be a problem in FWS constructed

wetlands. If a FWS wetland is designed with sufficient open water (40 to 60% of the surface area) to

permit effective control with mosquito fish, and inlet and outlet weirs are placed to allow level control

and drainage of wetland cells, the potential for

mosquito populations to thrive is reduced. This latter concept provides for isolation of various wetland cells to allow them to be drained and/or to allow predators and mosquitoes to become concentrated in pools and channels.

Along with these physical factors, the development

of a balanced ecosystem that includes other aquatic

invertebrates (beetles), aquatic insects (dragon flies

and damsel flies), fish (top-feeding minnows, sticklebacks, gobis, and others), birds (swallows, ducks,

and others), and mammals (bats) will help maintain acceptable levels of mosquitoes. Under these

conditions, the mosquito is simply a component in

a balanced food web. If an imbalance develops,

then intervention with certain biological and chemical agents may be required.

A successful intervention method has been the use

of Bti, a bacterium spore that interferes with development of the adult. In essence, Bti kills the larva

via physical actions. Several applications over the

mosquito season are needed to interfere with the

mosquito’s natural growth cycle, which may be three

to four months in length. Other larvacides, such as

methoprene, are chemicals that are not selective

for certain stages of mosquitoes’ life cycle.

Adulticides also are not selective for life cycles but

could be used at critical times.

In general, proper design that supports a healthy

wetland ecosystem produces conditions that maintain sufficiently low mosquito populations.

16. What is the present level of application of this technology?

As of late 1999, more than 200 communities in the

United States were reported to be utilizing constructed wetlands for wastewater treatment. Most

of these communities use wetlands for polishing

lagoon effluent. In addition, communities in a wide

range of sizes use this technology, including large

cities such as Phoenix, Arizona, and Orange
34

County, Florida. For the most part, however, FWS

technology has been utilized by small- to mediumsized communities ranging from 5,000 to 50,000 in

population.

Even though constructed wetlands for municipal

wastewater treatment have been around for as long

as 40 years, there have been widespread problems

in their performance with respect to nitrogen transformations and removal as well as phosphorus removal (WRC, 2000). This manual has been created to help future owners and designers avoid

unrealistic expectations from these systems.

17. Can these systems operate at elevations other than

sea level?

FWS and VSB systems are found in a wide range

of elevations extending, for example, from the

desert Southwest to New England, and from the

southeast United States to the Rocky Mountains

and Pacific coast regions. The common wetland

plants used in these systems are found in all areas

of the United States and Canada. There is no inherent biological or ecological basis for these types

of systems to not work in the normal range of physiographic conditions in the United States, Canada,

and Mexico.

18. Can constructed wetlands work in cold temperatures?

Constructed wetlands are found in a wide range of

climatological settings, including cold climates

where ice forms on the surface for four to six months

of the year. For example, these systems are found

in Canada, North Dakota, Montana, Vermont, Colorado, and other cold-climate areas. Special considerations must be included in the design of these

systems for the formation of an ice layer and the

effect of cold temperatures on mechanical systems,

such as the influent and effluent works. The absence of living plants that have died back for the

winter and the presence of a layer of ice approximately 0.5 to 1.0 ft. thick have not been shown to

severely affect the secondary treatment capabilities of these systems. Nitrogen transformation and

removal is, however, impaired during very cold periods.

19. Can you receive full treatment benefits from a constructed wetland that also provides ancillary benefits such as wildlife habitat?

Multiple benefits can accrue from a FWS constructed wetland if it is properly sited and designed.

For example, FWS wetlands that have a significant

portion of surface area occupied by submergent

aquatic plants and deeper water have been found

to produce higher-quality effluent and provide

greater habitat value than other configurations. This

open space is used by aquatic fowl for feeding,

access to refugia, and as a source of fresh water.

The same submerged aquatic plants that provide

wastewater treatment also serve as a food source

for aquatic birds and mammals.

Because reduced human health risk is associated

with tertiary treatment or “polishing” wetlands, they

have commonly enjoyed full recreational access to

the FWS systems, but they provide minimal removal

of several key pollutants in comparison to the treatment wetlands that are the focus of this manual.

Therefore, human access to these systems entails

greater health risks because the wastewater is actively being treated. The wildlife and other natural

ecological populations may be equally abundant in

these systems as in the polishing systems, but human access may be restricted, at least in the inlet

environs.

The potential for ancillary benefits is reduced with

VSB systems. Depending on its size and degree of

vegetation, a VSB system could provide wildlife

habitat. VSB wetlands also can be used for environmental education and awareness activities.

2.10 Glossary

Abiotic Nonbiological processes or treatment mechanisms

in a constructed wetland.

Adsorption Adherence by chemical or physical bonding

of a pollutant to a solid surface.

Adventitious roots provide a competitive advantage to a

plant by growing from stems into the surrounding air

(around terrestrial plants) or water (around aquatic plants)

before entering the soil substrate to provide additional uptake or absorption directly from the surrounding medium.

Aerenchymous tissues in aquatic plants provide for transfer of gases within a plant. In wastewater treatment systems, emergent aquatic plants rely on aerenchymous tissues for transfer of oxygen to their roots.

Aerobic processes in wastewater treatment systems take

place in the presence of dissolved oxygen.

Algae are single-celled to multicelled organisms that rely

on photosynthesis for growth. Most algae are classified as

plants.

Anaerobic processes in wastewater treatment systems

take place in the absence of dissolved oxygen and instead

rely on molecular oxygen available in decomposing compounds.

Aspect ratio The length of a constructed wetland divided

by its width (L/W).

Atmospheric reaeration introduces atmospheric oxygen

into the water at the water’s surface, which provides dissolved oxygen to the aquatic environment.
35

Autotrophic Types of reactions that generally require only

inorganic reactants; for example, nitrification.

Biochemical oxygen demand (BOD) is the demand for

dissolved oxygen that decomposition of organic matter

places on a wastewater treatment process. BOD as expressed in milligrams per liter (mg/L) is used as a measure of wastewater organic strength and as a measure of

treatment performance. This constituent is represented

throughout the text of this manual as “BOD,” which stands

for the U.S. standard 5-day BOD test result.

Biomass is the total amount of living material, including

plants and animals, in a unit volume.

Biotic is a term which implies microbiological or biological

mechanisms of treatment.

BOD removal is the lowering of demand for dissolved oxygen required for biological decomposition processes in the

water column; hence, BOD removal can be accomplished

by biological decomposition in open-water zones and by

flocculation and sedimentation in fully vegetated zones and

in VSBs.

Bulrush is the common name for a number of plants of

the genus Scirpus found in wetlands. Several species of

bulrush commonly used in constructed wetlands thrive in

the wide range of environmental conditions in constructed

wetlands, including varying levels of water depth and quality. The large, terete bulrush species include S. validus, S.

californicus, and S. acutus, all of which form dense stands

with large numbers of round-sectioned stems that maintain an upright posture for one or more years. Other species of Scirpus include the three-square varieties, such as

S. americanus (olynei), S. fluviatilis, and S. robustus, which

offer tolerance to salinity, a variety of color shades, and

attractiveness to various animal species.

Canopy Uppermost or tallest vegetation in a plant community.

Cattail is the common name for a number of plants of the

genus Typha that are common in constructed wetlands in

the United States, with at least three species predominant:

T. latifolia, T. domingensis, and T. angustifolia. Along with

their hybridized forms, these species occupy numerous

water-depth and water-quality niches within constructed

wetlands. The wetland designer is advised to consult local

botanists and geographic references to determine which

local cattail species or hybrid is best adapted to the specific water quality, water depth, and substrate planned for

a constructed wetland.

Common reed (Phragmites) probably is the most widely

used plant in constructed wetlands on a worldwide basis,

but it typically is not used in the United States. Although

this plant has excellent growth characteristics in very shallow constructed wetlands, it is an invasive species in some

natural wetlands, and its transport and intentional introduction to some localities are discouraged. Common reed

is considered to offer little value as food or habitat for wetland wildlife species (Thunhorst, 1993).

Constructed wetlands are wastewater treatment systems

that rely on physical, chemical, and biological processes

typically found in natural wetlands to treat a relatively constant flow of pretreated wastewater.

Deciduous Woody plants that shed their leaves in cold

seasons.

Dentrification Biotic conversion of nitrate-nitrogen to nitrogen gases.

Detritus Loose, dead leaves and stems from dead vegetation.

Dike A wall of mounded soil that contains or separates

constructed wetlands from surrounding areas. Dissolved

oxygen (DO) is required in the water column of a wastewater treatment system for aerobic biochemical processes

that take place in constructed wetlands.

Dominant plant species The plant species that exerts a

controlling influence on the function of the entire plant community.

Duckweed Duckweed naturally moves on a large water

surface by movement induced by wind action unless it is

protected from the wind and held in place by dense stands

of emergent plants (e.g., macrophytes) or artificial barriers. In FWS systems, this results in dense growths of duckweed within the fully vegetated zones. Duckweed effectively seals the water surface and prevents atmospheric

reaeration. This action combined with the inherent oxygen

demand of the incompletely treated municipal wastewater

results in anaerobic conditions in these fully vegetated

zones.

Emergent herbaceous wetland plants grow rooted in the

soil, with plant structures extending above the surface of

the water. These plants are herbaceous by virtue of their

relatively decomposable (leafy) plant structures, but they

also have sufficient internal structure to maintain their upright growth, even without the support of surrounding waters. Most emergent wetland plants grow with or without

the presence of surface water; however, they generally

grow in shallow water near the banks of a water body.

Emergent vegetation (see macrophytes).

Evapotranspiration Loss of water to the atmosphere

through water surface and vegetation.

Exotic species A plant not indigenous to the region.

Fecal coliform A common measure for pathogenicity of

wastewater. This analytical test reveals the number of these

types of organisms in counts/100 milliliters (#/100mL)

Filtration is the process of filtering influent solids from the

wastewater and typically is provided by plant stems and

leaves and other vegetation in the water column.
36

Floating aquatic plants are commonly found in FWS systems, including water hyacinth (Eichhornia crassipes),

duckweed (Lemna spp., Spirodela spp., and Wolffia spp.),

water fern (Azolla caroliniana and Salvinia rotundifolia),

and water lettuce (Pistia stratiotes). Also common are

rooted plants growing in a floating form, including pennywort (Hydrocotlyle spp.), water lily (Nymphaea spp.), frog's

bit (Limnobium spongia), spatterdock (Nuphar spp.), and

pondweed (Potemogeton spp.).

Floating aquatic systems are in essence shallow basins

covered with floating aquatic plants. One type of plant that

can remain in place is the water hyacinth, but it is very

sensitive to other than tropical temperatures and is considered to be an invasive species. Duckweed has been

held in place with artificial barriers in these types of systems.

Flocculation is the process of very small particles of matter clumping together to reach a collectively larger size. In

wastewater treatment processes, flocculation typically agglomerates colloidal particulates into larger, settleable solids that are then removed by sedimentation processes.

Free water surface (FWS) wetlands are constructed wetlands that provide wastewater treatment through flocculation and sedimentation during the flow of wastewater

through stands of aquatic plants growing in shallow water.

In some FWS wetlands, there are also open areas where

aerobic bio-oxidation complements the physical removal

processes. FWS systems resemble natural wetlands in

function and appearance. FWS systems have also been

termed “surface flow systems.”

Function refers to the purpose, role, or actions expected

of constructed wetlands in the process of wastewater treatment. Function is expressed in terms of expected results,

such as nutrient uptake, removal of TSS and BOD, maintenance of dissolved oxygen in open water zones, and

reduction of wastewater constituents to acceptable levels,

waterfowl habitat, and water storage.

Habitat value The suitability of an area to support a given

species.

Herbaceous Plant material that has no woody parts.

Herbivores are members of the animal kingdom that consume plant matter.

Hydric soils, or wetland soils, exhibit distinct chemical and

physical changes that result from periodic inundation and

saturation. Flooding and subsequent decomposition and

oxidation of soil chemicals typically result in anaerobic soil

conditions.

Hydrophyte Any plant growing in a soil that is deficient in

oxygen.

Indigenous species Species of plants that are native to

an area.

Inorganics Compounds that do not contain organic carbon.

Lagoons are also called stabilization ponds, oxidation

ponds, etc. In conventional wastewater treatment systems,

they typically are used to provide intermediate treatment

of wastewater through a variety of physical, chemical, and

biological processes.

Limiting nutrient is the nutrient that controls a particular

plant’s growth. When present in insufficient amounts relative to a given plant’s needs, a limiting nutrient limits that

plant’s growth.

Limnetic The open water zone of a FWS system where

light can penetrate to induce photosynthesis.

Macrophytes are plants that are readily visible to the unaided eye and include vascular or higher plants. Vascular

plants include mosses, ferns, conifers, monocots, and dicots. Macrophytes also may be categorized by a variety of

ecological growth forms.

Marsh A common term applied to treeless wetlands.

Microbes or microorganisms are microscopic organisms

(only viewed with a microscope), such as bacteria, protozoans, and certain species of algae, which are responsible for many of the biochemical transformations necessary in wastewater treatment processes.

Nitrification Biotic conversion of ammonium nitrogen to

nitrite and nitrate-nitrogen.

Nuisance species Plants that detract from or interfere with

the designated purpose(s) of constructed wetlands.

On-site constructed wetland systems are wastewater

systems for treatment and disposal at the site where wastewater is generated. For example, a residential septic system is an on-site system.

Organics Compounds that contain organic carbon (also

volatile solids).

Oxygen demand Generally expressed through relatively

high BOD concentrations, the property of municipal wastewater that removes dissolved oxygen from the water column.

Photosynthesis is the conversion of sunlight into organic

matter by plants through a process of combining carbon

dioxide and water in the presence of chlorophyll and light,

which releases oxygen as a by-product.

Phytoplankton are algae that are microscopic in size

which float or drift in the upper layer of the water column

and depend on photosynthesis and the presence of phosphorus and nitrogen in the water.

Pneumatophores are structures that provide air channels

for emergent plants growing in water environments.

Polishing wetlands are designed to provide tertiary treatment to secondary effluent to meet performance standards
37

required by National Pollutant Discharge Elimination System (NPDES) permits. Design considerations for polishing wetlands are outside the scope of this manual.

Primary effluent is the product of primary treatment of

wastewater that typically involves settling of solids in a

containment structure, such as a septic tank, settling pond,

or lagoon.

Primary production is the production of biomass (organic

carbon) by plants and microscopic algae, typically through

photosynthesis, as the first link in the food chain.

Primary treatment of wastewater is a settling process for

removal of settleable solids from wastewaters.

Rhizome Root-like stem that produces roots and stems to

propogate itself in a surrounding zone.

Secondary effluent is wastewater that has undergone secondary treatment and is discharged to the environment or

receives further treatment in tertiary treatment processes.

Secondary treatment continues the process begun in primary treatment by removing certain constituents, such as

biochemical oxygen demand (BOD) and total suspended

solids (TSS) from primary effluent to prescribed treatment

levels; typically, 30 mg/L in the United States.

Sediment Organic and mineral particulates that have

settled from the overlying water column (also sludge).

Seepage Loss of water from a constructed wetland to the

soil through infiltration below the system. Senescence is

the phase at the end of a plant’s life that leads to death

and, finally, decay.

Solar insolation refers to the amount of solar radiation

that reaches the constructed wetland. Solar radiation in

the summer months may play an important role in photosynthesis in open-water zones of a FWS system.

Standing biomass in a constructed wetland is the total

amount of plant material that stands erect. This term typically is used as “dead standing biomass” to refer to dead,

standing plants, in contrast to green plants and plant litter

composed of broken and fallen dead plant parts.

Structure refers to the form and amount of living and

nonliving components of an ecosystem. For example,

emergent vegetation provides the structure to perform

wetland functions. Wetland structure is expressed in

qualitative terms such as species of flora and fauna, or

type of wetland such as marsh, bog, or bottomland forest.

Submerged aquatic plants or submergent vegetation

are rooted plants that grow in open water zones within

the water column of an aquatic environment (compare to

emergent aquatic plants) and provide dissolved oxygen

for aerobic biochemical reactions. They lie below the water surface, except for flowering parts in some species.

Subsurface flow (SF) wetlands (see vegetated submerged bed (VSB) systems).

Tertiary treatment (see polishing wetlands).

Total nitrogen (TN) is the sum of all the forms of nitrogen,

including nitrate, nitrite, ammonia, and organic nitrogen in

wastewater, and is typically expressed in milligrams per

liter (mg/L).

Total phosphorus (TP) is a measure of all forms of phosphorus in wastewater, typically expressed in milligrams per

liter (mg/L).

Total suspended solids (TSS) are particulate matter in

wastewater consisting of organic and inorganic matter that

is suspended in the water column. The numeric value is

provided by specific analytical test. Typically, municipal

wastewaters include the settleable solids and some portion of the colloidal fraction.

Vascular plant Plant that readily conducts water, minerals and foods throughout its boundaries.

Vegetated submerged bed (VSB) systems provide wastewater treatment in filter media that is not directly exposed

to the atmosphere but may be slightly influenced by the

roots of surface vegetation. VSB systems also have been

termed subsurface flow (SF) wetlands, rock reed filters,

submerged filters, root zone method, reed bed treatment

systems, and microbial rock plant filters. In this manual,

the term “vegetated submerged bed systems” is used because gravel beds rather than hydric soils are the support

media for wetland plants; as a result, the systems are not

truly wetlands.

Vegetative reproduction is the process of asexual reproduction, in which new plants develop from roots, stems,

and leaves of the parent plant.

Wastewater treatment is the process of improving the

quality of wastewater. The term can refer to any parts

or all parts of the process by which raw wastewater is

transformed through biological, biochemical, and physical means to reduce contaminant concentrations to prescribed levels prior to release to the environment. A

wastewater treatment process typically consists of primary, secondary, and tertiary treatment.

Wetland hydraulics refers to movement of water

through constructed wetlands, including volumes,

forces, velocities, rates, flow patterns, and other characteristics.

Woody plants are plants that produce bark and vascular structures that are not leafy in nature. Woody plants

have trunks, stems, branches, and twigs that allow them

to occupy a greater variety of available niches than herbaceous plants can occupy. General terms that describe

categories of woody plants found in wetlands are shrubs,

trees (canopy or subcanopy), and woody vines.
38

Wrack Plant debris carried by water.

Zooplankton Microscopic and small animals that live in

the water column.

2.11 References

Arizona Department of Environmental Quality (ADEQ).

1995. Arizona guidance manual for constructed wetlands for water quality improvement. Prepared by

R.L. Knight, R. Randall, M. Girts, J.A. Tress, M.

Wilhelm, and R.H. Kadlec. ADEQ TM 95-1.

Armstrong, W. 1978. Root aeration in the wetland environment. In: D.D. Hook and R.M.M. Crawford (eds.)

Plant life in anaerobic environments. Ann Arbor, MI:

Ann Arbor Science, Chapter 9, pp. 269–297.

Bavor, H.J., D.J. Roser, P.J. Fisher, and I.C. Smalls.

1989. Performance of solid-matrix wetland systems

viewed as fixed-film bioreactors. In: D.A. Hammer

(ed.) Constructed wetlands for wastewater treatment, municipal, industrial, and agricultural.

Chelsea, MI: Lewis Publishers, Chapter 39k, pp.

646–656.

Burgan, M.A. and D.M. Sievers. 1994. On-site treatment

of household sewage via septic tank and two- stage

submerged bed wetland. In: Proceedings of the Seventh International Symposium on Individual and

Small Community Sewage Systems. American Society of Agricultural Engineers, Atlanta, GA: pp. 77–

84.

Burgoon, P.S., K.R. Reddy, and T.A. DeBusk. 1989. Domestic wastewater treatment using emergent plants

cultured in gravel and plastic substrates. In: D.A.

Hammer (ed.) Constructed wetlands for wastewater treatment, municipal, industrial, and agricultural.

Chelsea, MI: Lewis Publishers, Chapter 38f, pp.

536–541.

Dennison, M.S. and J.F. Berry. 1993. Wetlands: guide

to science, law, and technology. Park Ridge, NJ:

Noyes Publications, 439 pp.

Dill, C.H. 1989. Wastewater wetlands: user friendly mosquito habitats. In: D.A. Hammer (ed.) Constructed

wetlands for wastewater treatment, municipal, industrial, and agricultural. Chelsea, MI: Lewis Publishers, Chapter 39m, pp. 664–667.

Feierabend, J.S. 1989. Wetlands: the lifeblood of wildlife. In: D.A. Hammer (ed.) Constructed wetlands

for wastewater treatment, municipal, industrial, and

agricultural. Chelsea, MI: Lewis Publishers, Chapter 7, pp. 107–118.

Friend, M. 1985. Wildlife health implications of sewage

disposal in wetlands. In: P.J. Godfrey, E.R. Kaynor,

S. Pelczarski, and J. Benforado (eds.) Ecological

considerations in wetlands treatment of municipal

wastewaters. New York, NY: Van Nostrand Reinhold,

Chapter 17, pp. 262–269.

Gearheart, R.A., F. Klopp, and G. Allen. 1989. Constructed free surface wetlands to treat and receive

wastewater: pilot project to full scale. In: D.A. Hammer (ed.) Constructed wetlands for wastewater treatment, municipal, industrial, and agricultural.

Chelsea, MI: Lewis Publishers, Chapter 8, pp. 121–

137.

Hammer, D.A. 1992. Creating freshwater wetlands. Boca

Raton, FL: Lewis Publishers, 298 pp.

Huang, J., R.B. Renau, Jr., and C. Hagedorn. 1994. Constructed wetlands for domestic wastewater treatment. In: Proceedings of the Seventh International

Symposium on Individual and Small Community

Sewage Systems. American Society of Agricultural

Engineers, Atlanta, GA, pp. 66–76.

Johns, M.J., B.J. Lisidar, A.L. Kenimer, and R.W. Weaver.

1998. Nitrogen fate in a subsurface flow constructed

wetland for on-site wastewater treatment. In: Proceedings of the Eighth National Symposium on Individual

and Small Community Sewage Systems. American

Society for Agricultural Engineers, Orlando, FL, pp.

237–246.

Kadlec, R.H. and R.L. Knight. 1996. Constructed wetlands.

Boca Raton, FL: Lewis Publishers, 893 pp.

Kadlec, R.H., D.A. Hammer, and M.A. Girts. 1990. A total

evaporative constructed wetland treatment system. In:

P.F. Cooper and B.C. Findlater (eds.) Constructed wetlands in water pollution control. Oxford, UK: Pergamon

Press, pp. 127–138.

Kent, D.M. 1994. Applied wetlands science and technology. Boca Raton, FL: Lewis Publishers, 436 pp.

Knight, R.L. 1992. Ancillary benefits and potential problems with the use of wetlands for nonpoint source pollution control. Ecological Engineering, 1:97–113.

Knight, R.L. 1997. Wildlife habitat and public use benefits

of constructed wetlands. In: R. Haberl, R. Perfler, J.

Laber, and P. Cooper (eds.) Wetland systems for water pollution control 1996. Water Science & Technology 35(5). Oxford, UK: Elsevier Science Ltd., pp. 35–

43.

Kroodsma, D.E. 1978. Habitat values for nongame wetland birds. In: P.E. Greeson, J.R. Clark, and J.E. Clark

(eds.) Wetland functions and values: The state of our

understanding. Minneapolis, MN: American Water

Resources Association, pp. 320–326.

Living Technologies, Inc. 1996. Interim performance report for the South Burlington, Vermont “Living Machine,” January–August 1996. Burlington, VT: Living

Technologies, Inc., September 1996.
39

Majumdar, S.K., R.P. Brooks, F.J. Brenner, and R.W.

Tiner (eds.) 1989. Wetlands ecology and conservation: Emphasis in Pennsylvania. Easton, PA: The

Pennsylvania Academy of Science, 394 pp.

Mankin, K.R. and G. M. Powell. 1998. Onsite rock-plant

filter monitoring and evaluation in Kansas. In: Proceedings of the Eighth National Symposium on Individual and Small Community Sewage Systems.

American Society for Agricultural Engineers, Orlando, FL, pp. 228–236.

Marble, A.D. 1992. A guide to wetland functional design. Boca Raton, FL: Lewis Publishers, 222 pp.

Martin, C.V. and B.F. Eldridge. 1989. California’s experience with mosquitoes in aquatic wastewater treatment systems. In: D.A. Hammer (ed.) Constructed

wetlands for wastewater treatment, municipal, industrial, and agricultural. Chelsea, MI: Lewis Publishers, pp. 393–398.

McAllister, L.S. 1992. Habitat quality assessment of two

wetland treatment systems in Mississippi—A pilot

study. EPA/600/R-92/229. U.S. Environmental Protection Agency, Environmental Research Laboratory,

Corvallis, OR. November 1992.

McAllister, L.S. 1993a. Habitat quality assessment of

two wetland treatment systems in the arid west—A

pilot study. EPA/600/R-93/117. U.S. Environmental

Protection Agency, Environmental Research Laboratory, Corvallis, OR. July 1993.

McAllister, L.S. 1993b. Habitat quality assessment of

two wetland treatment systems in Florida—A pilot

study. EPA/600/R-93/222. U.S. Environmental Protection Agency, Environmental Research Laboratory,

Corvallis, OR. November 1993.

Merritt, A. 1994. Wetlands, industry & wildlife: A manual

of principles and practices. Gloucester, UK: The

Wildfowl Trust, 182 pp.

Mitsch, W.J. and J.G. Gosselink. 1993. Wetlands. 2d

ed. New York, NY: Van Nostrand Reinhold, 722 pp.

Neralla, S., R.W. Weaver, and B.J. Lesikar. 1998. Plant

selection for treatment of septic effluent in subsurface wetlands. In: Proceedings of the Eighth National Symposium on Individual and Small Community Sewage Systems. American Society for Agricultural Engineers, Orlando, FL, pp. 247–253.

Niering, W.A. 1985. Wetlands. New York, NY: Alfred A.

Knopf, 638 pp.

Payne, N.F. 1992. Techniques for wildlife habitat management of wetlands. New York, NY: McGraw-Hill.

Post, Buckley, Schuh & Jernigan, Inc. 1993. Compliance and performance review for the City of

Orlando’s easterly wetland treatment system. Prepared for the City of Orlando, FL.

Reed, S.C., R.W. Crites, and E.J. Middlebrooks. 1995.

Natural systems for waste management and treatment. 2d ed. New York, NY: McGraw-Hill, 433 pp.

Reed, S.C., J. Salisbury, L. Fillmore, and R. Bastian.

1996. An evaluation of the “Living Machine” wastewater treatment concept. In: Proceedings, WEFTEC

‘96, 69th Annual Conference and Exhibition, Dallas, TX. Water Environment Federation, Alexandria,

VA, October 1996.

Sather, J.H. 1989. Ancillary benefits of wetlands constructed primarily for wastewater treatment. In: D.A.

Hammer (ed.) Constructed wetlands for wastewater treatment, municipal, industrial, and agricultural.

Chelsea, MI: Lewis Publishers, Chapter 28a, pp.

353–358.

Schueler, T.R. 1992. Design of stormwater wetland systems: Guidelines for creating diverse and effective

stormwater systems in the mid-Atlantic region.

Anacostia Restoration Team, Metropolitan Washington Council of Governments, Washington, DC. 133

pp.

Snoddy, E.L., G.A. Brodie, D.A. Hammer, and D.A.

Tomljanovich. 1989. Control of the armyworm,

Simyra henrici (Lepidoptera: Noctuidae), on cattail

plantings in acid drainage constructed wetlands at

Widows Creek Steam-Electric Plant. In: D.A. Hammer (ed.) Constructed wetlands for wastewater treatment, municipal, industrial, and agricultural.

Chelsea, MI: Lewis Publishers, Chapter 42I, pp.

808–811.

Stowell, R., S. Weber, G. Tchobanoglous, B.A. Wilson,

and K.R. Townzen. 1985. Mosquito considerations

in the design of wetland systems for the treatment

of wastewater. In: P.J. Godfrey, E.R. Kaynor, S.

Pelczarski, and J. Benforado (eds.) Ecological considerations in wetlands treatment of municipal

wastewaters. New York, NY: Van Nostrand Reinhold,

Chapter 3, pp. 38–47.

Thunhorst, G.A. 1993. Wetland planting guide for the

northeastern United States: Plants for wetland creation, restoration, and enhancement. St. Michaels,

MD: Environmental Concern, Inc., 179 pp.

Thut, R.N. 1989. Utilization of artificial marshes for treatment of pulp mill effluents. In: D.A. Hammer (ed.)

Constructed wetlands for wastewater treatment, municipal, industrial, and agricultural. Chelsea, MI:

Lewis Publishers, Chapter 19, pp. 239–244.

Todd, J. and B. Josephson. 1994. Living machines:

Theoretical foundations and design precepts.

Falmouth, MA: Ocean Arks International.
40

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manual: Constructed wetlands and aquatic plant systems for municipal wastewater treatment. Center for

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habitat . EPA 832-R-005. Office of Water, Washington,

DC.

U.S. Environmental Protection Agency (EPA). 1993b. Subsurface flow constructed wetlands for wastewater treatment: A technology assessment. EPA 832-R-93-008.

Office of Water, Washington, DC.

U.S. Environmental Protection Agency (EPA). 1997a. Response to Congress on use of decentralized wastewater treatment systems. EPA 832-R-97-001b. Office of

Water, Washington, DC.

U.S. Environmental Protection Agency (EPA). 1997b. Response to Congress on the AEES “Living Machine”

Wastewater Treatment Technology. EPA 832-R-97-002.

Office of Water, Washington, DC.

U.S. Environmental Protection Agency (EPA). 1999. Free

water surface wetlands for wastewater treatment: A technology assessment. EPA 832/R-99/002. Office of Water, Washington, DC.

Visher, S.S. 1954. Climatic atlas of the United States. Cambridge, MA: Harvard University Press.

Vymazal, J. 1995. Algae and element cycling in wetlands.

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wildlife. In: R.E. Good, D.F. Whigham, and R.L. Simpson

(eds.) Freshwater wetlands: Ecological processes and

management potential. New York, NY: Academic Press,

pp. 267–284.

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recommendations for on-site wastewater treatment using constructed wetlands. In: Proceedings of the Eighth

National Symposium on Individual and Small Community Sewage Systems. American Society for Agricultural

Engineers, Orlando, FL, pp. 195–201.

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Control Foundation, 60:1253–1258.
41

3.1 Introduction

Constructed wetlands are highly complex systems that

separate and transform contaminants by physical, chemical, and biological mechanisms that may occur simultaneously or sequentially as the wastewater flows through

the system. In a qualitative sense, the processes that occur are known, but in only a few cases have they been

adequately measured to provide a more quantitative assessment. The predominant mechanisms and their sequence of reaction are dependent on the external input

parameters to the system, the internal interactions, and

the characteristics of the wetland. The external input parameters most often of concern include the wastewater

quality and quantity and the system hydrological cycle.

Typical characteristics of municipal wastewaters most

often treated in constructed wetlands are described in Table

3-1. The emphasis of this manual is on the treatment of

municipal wastewater with the objectives of achieving target levels of suspended solids, organic matter, pathogens,

and in some instances, nutrients (specifically total nitrogen) and heavy metals. Wastewaters that will be considered include septic tank effluent, primary effluent, pond

effluents, and some secondary effluents from overloaded

or poorly controlled systems. Table 3-1 shows that the character of the wastewater is dependent on pretreatment and

Chapter 3

Removal Mechanisms and Modeling Performance of

Constructed Wetlands

may contain both soluble and particulate fractions of organic and inorganic constituents in reduced or oxidized

forms. As will be seen later, these characteristics play an

important role in the major mechanisms of removal.

The two major mechanisms at work in most treatment

systems are liquid/solid separations and constituent transformations. Separations typically include gravity separation, filtration, absorption, adsorption, ion exchange, stripping, and leaching. Transformations may be chemical, including oxidation/reduction reactions, flocculation, acid/

base reactions, precipitation, or a host of biochemical reactions occurring under aerobic, anoxic, or anaerobic conditions. Both separations and transformations may lead to

contaminant removal in wetlands but often only result in

the detention of the contaminant in the wetland for a period of time. There may be changes in the contaminant

composition that will effectively achieve treatment objectives, such as the biochemical transformation of organic

compounds to gases such as CO2

 or CH4. A biochemical

transformation, however, may produce biomass or organic

acids that may not achieve the treatment objective if these

materials escape in the effluent. In the case of biomass, it

may escape as volatile suspended solids, or it may undergo further bacterial reaction, which may result in the

leaching of a soluble carbon compound back into the water column.

The remainder of this chapter will review potential mechanisms that may be at work in constructed wetlands. These

reactions may occur in the water column, on the surfaces

of plants, within the litter and detritus accumulating at the

wetland surface or on the bottom, or within the root zone

of the system. The reactions unique to the wetland type

will also be delineated.

3.2 Mechanisms of Suspended Solids

Separations and Transformations

3.2.1 Description and Measurement

Suspended solids in waters are defined by the method

of analysis. Standard Methods (1998) defines total suspended solids as those solids retained on a standard glass

fiber filter that typically has a nominal pore size of 1.2μm.

The type of filter holder, the pore size, porosity, area and

thickness of the filter, and the amount of material deposTable 3-1. Typical Constructed Wetland Influents

Constituent Septic Tank Primary Pond

(mg/L) Effluent1 Effluent2 Effluent3

BOD 129-147 40-200 11-35

Sol. BOD 100-118 35-160 7-17

COD 310-344 90-400 60-100

TSS 44-54 55-230 20-80

VSS 32-39 45-180 25-65

TN 41-49 20-85 8-22

NH3 28-34 15-40 0.6-16

NO3 0-0.9 0 0.1-0.8

TP 12-14 4-15 3-4

OrthoP 10-12 3-10 2-3

Fecal coli (log/100ml) 5.4-6.0 5.0-7.0 0.8-5.6

1

EPA (1978), 95% confidence interval. Prior to major detergent

reformulations which reduce P species by ~50%.

2Adapted from Metcalf and Eddy, (1991) assuming typical removal by

primary sedimentation-soluble BOD = 35 to 45% total.

3

EPA (1980).
42

ited on the filter are the principal factors affecting the separation of suspended from dissolved solids. As a result, the

measurement reported for total suspended solids may include particle sizes ranging from greater than 100μm to about

1μm. Soluble (dissolved) solids would therefore include colloidal solids smaller than 1μm and molecules in true solution. A classical method of solids classification by size would

include the following:

Settleable Solids >100μm

Supracolloidal Solids 1–100μm

Colloidal Solids 10-3–1μm

Soluble Solids <10-3μm

Solids are also classified as volatile or fixed, again based

on the method of analysis. Standard Methods (1998) defines a volatile solid as one that ignites at 550°C. Although

the method is intended to distinguish between organic solids and inorganic solids, it is not precise since volatile solids

will include losses due to the decomposition or volatilization

of some mineral salts depending on the time of exposure to

the ignition temperature.

Wastewater influents to wetlands may contain significant

quantities of suspended solids (Table 3-1). The composition

of these solids is quite different, however. Septic tank and

primary effluents will normally contain neutral density colloidal and supracolloidal solids emanating from food wastes,

fecal materials, and paper products. Pond effluent suspended

solids are likely to be predominantly algal cells. All three will

be high in organic content. Size distribution is also different

among the waste streams. Tables 3-2 and 3-3 present information on size distributions of suspended solids, organic

matter, and phosphorus in domestic wastewaters with various levels of pretreatment. It should be noted that methods

differed between investigators on estimating size ranges.

High settleable fractions are not surprising for raw wastewater samples or for the pond effluent containing algal cells. It

is important to note the association of organic matter and

phosphorus with the various solid fractions.

Table 3-2. Size Distributions for Solids in Municipal Wastewater

Type of Sample

(% by Weight)

Size Range Primary Eff.1 Primary Eff.2 Primary Eff.3 Raw Sewage4 Raw Sewage5

 (μm)

<10-3 - - - 31 48

10-3-1.0 20 22(10-30) - 14 9

1.0-12 54 35(24-51) - - -

>12 26 43(30-60) - - -

1.0-100 - - 81 24 18

>100 - - 19 31 23

1 Levine et al. (1984).

2 Tchobanoglous et al. (1983).

3 Gearheart, et al. (1993).

4 Heukelekian and Balmat (1959).

5 Rickert and Hunter ((1972).

Table 3-3. Size Distribution for Organic and Phosphorus Solids in Municipal Wastewater

Type of Solids

(% by Weight)

Size Range Organic Organic Organic Organic Organic Organic Total Total

 (μm) Solids1 Solids2 Solids2 Solids2 Solids2 Solids3 Phos.4 Phos.4

(primary (primary (primary (primary (primary (raw (primary (primary

effluent) effluent) effluent) effluent) effluent) sewage) effluent) effluent)

<10-3 9 - - - - 42 - -

<0.1 - 51 50 25 35 - 17.2 15.7

10-3-1.0 ? - - - - 11 - -

0.1-1.0 - 8 19 2 1 - 54.6 67.0

1.0-12 - 34 26 13 13 - 7.15 6.7*

>12 - 7 5 60 41 - 21.06 8.5**

1.0-100 15 - - - - 20 - -

>100 28 - - - - 27 - -

1 Munch et al. (1980).

2 Levine et al. (1991).

3 Rickert and Hunter (1972).

4 Levine et al. (1984).

5 Range: 1-5 μm 6

Range: >5 μm
43

3.2.2 Suspended Solids in Free Water

Surface Wetlands

Total suspended solids are both removed and produced

by natural wetland processes. The predominant physical

mechanisms for suspended solids removal are flocculation/

sedimentation and filtration/interception, whereas suspended

solids production within the wetland may occur due to death

of invertebrates, fragmentation of detritus from plants, production of plankton and microbes within the water column or

attached to plant surfaces, and formation of chemical precipitates such as iron sulfide. Figure 3-1 illustrates the most

important of these processes as they occur in a FWS system. Resuspension of solids may occur due primarily to turbulence created by animals, high inflows, or winds. A brief

discussion of some of these processes and how they may

affect free water surface systems follows.

3.2.2.1 Discrete and Flocculant Settling

Typically, particulate settling produced by gravity may

be categorized as discrete or flocculant settling. Both separation processes exploit the properties of particle size, specific gravity, shape, and fluid specific gravity and viscosity.

Discrete settling implies that the particle settles independently and is not influenced by other particles or changes

in particle size or density. A mathematical expression for

the terminal settling velocity of the discrete particle may

be derived from Newton’s Law. Under laminar flow conditions, which exist in fully vegetated zones of a FWS and in

VSBs, the velocity of a spherical particle can be estimated

by Stokes’ Law, which states that the settling velocity is

directly proportional to the square of the nominal diameter

and the difference in particle and fluid densities and is inversely proportional to fluid viscosity. Drag on the particle

that influences settling velocity is affected by particle shape,

fluid/particle turbulence, and fluid viscosity.

Whereas discrete settling can be estimated given the

independent variables discussed previously, flocculant

settling cannot be so easily determined, requiring experimental measurement. It occurs as the result of particle

growth and, perhaps, change in characteristics over time.

As a result, particle settling velocity typically increases with

time. Flocculent settling is promoted by the relative movement of target particles in such a fashion as to cause an

effective collision. This relative velocity (velocity gradient)

is often calculated as G, the mean velocity gradient, which

is a function of power input, dynamic viscosity, and system

volume (Camp and Stein, 1943). Effective velocity gradiFigure 3-1. Mechanisms that dominate FWS systems

TSS

Toxics

Metals

Particulate BOD

Organic P&N

Settled TSS & Detritus from Plants

Emergent Plants

Duckweed, Floating Litter & Detritus

Atmospheric

Reaeration

Solar

 Radiation

BOD Oxidation

Submerged Vegetation

NH4-N NO3-N

O2

O2

Dissolved Oxygen O (Anoxic)

Fully Vegetated Zone

Removals due to Flocculation,

Sedimentation, Adsorption and

Anaerobic Reactions, Primarily

Dissolved Oxygen +++(Aerobic)

Open Water Zone

Transformations by Aerobic Biological

Treatment, Primarily

Pathogen Kill by Sunlight & Time
44

ents for flocculation range from 10 to 75 sec-1. Flocculation may occur naturally, as when fresh water flows into

saline water forming a delta, or it may require chemical

(coagulant) addition. It may affect large particles (100μm

to 1000μm) of low to moderate specific gravity (1.001 to

1.01) and small particles (1.0μm to 10μm) with high specific gravity (1.5 to 2.5). The formation of larger flocculant

particles is dependent on the electric charge on the particle surface. Like electrical charges on the double layer

surrounding particles may produce particle stability that

hinders attachment even if collisions take place. This

charge is sensitive to the composition of the fluid. Adsorption of solutes to the surface occurs as a result of a variety

of binding mechanisms, which may eventually result in the

destabilization of the particles and result in particle adhesion. There has been little work done on the evaluation of

natural flocculation phenomena with primary effluent or

algal cells. The existence of emergent plant stems in FWS

wetlands will promote effective velocity gradients for particle collisions, but the adhesion of these particles would

be dependent on surface properties that would be influenced by water column quality.

In wetland systems treating primary or septic tank effluents (or secondary effluents), particle sizes are mostly in

the colloidal to low supracolloidal range (Table 3-2). Typically sedimentation processes will remove material larger

than about 50μm with specific gravity of about 1.20. The

remaining particles are normally the lower density materials. Using Stokes’ Law to approximate discrete settling

velocity, particles ranging from 1.0μm to 10μm with a specific gravity ranging from 1.01 to 1.10 will settle at a rate of

from 0.3 to 4 x 10-4m/d. Typical hydraulic loads to FWS

wetlands are in the range of 0.01 to 0.5 m/d (note that the

hydraulic load is equivalent to the mean settling velocity of

a particle that will be removed exactly at that loading). Assuming the higher settling velocity of 0.3 m/d and a typical

FWS system velocity of 50 m/d and depth of 0.8m, the

larger particles would settle by gravity in approximately

2.7 days, or 133 m along the wetland longitudinal axis.

The smaller, less dense particles would require over 200

days and a length of over 11,000 m. Therefore it can be

concluded that the larger, denser particles could be removed in the primary zone of a wetland based on simple

discrete settling theory (see Chapter 4 for more details on

design). The smaller, neutral-density particles, which make

up a significant fraction of septic tank and primary effluent,

are not likely removed in this primary zone by simple discrete sedimentation, but may be flocculated due to the

velocity gradients imposed by emergent plant stems in the

water column. It is also possible that some particles may

be intercepted by angular emergent plant tissue as would

occur in settling basins equipped with plate or tube settlers. Clearly, removal of TSS by a FWS wetland is more

complex than predicted by discrete settling theory. There

is currently insufficient transect data available on wastewater influents of interest to develop a rational separation

model, either qualitatively or quantitatively, for TSS removal

from primary or septic tank effluents in FWS systems.

For wetland systems receiving pond effluents, the primary source of suspended solids for much of the season

is algal cells. These cells include green algae, pigmented

flagellates, blue-greens, and diatoms. Sizes range from

1μ to 100μm, and shapes may range from coccoid to filamentous. Specific gravity of actively growing algal cells

may be close to that of water insofar as they must remain

suspended high up within the water column in order to

survive. Flotation may be accomplished by gas vacuoles

(blue-green algae), gelatinous sheathes, or shapes that

increase particle drag. It is believed that wind-induced turbulence and vertical water motion greatly influence algae

distribution in ponds (Bella, 1970). Motile algae are not

typically predominant in wastewater pond systems. Once

algal cells die for lack of nutrients and/or sunlight, they

lose this flotation characteristic and will settle. Settling velocities range from 0.0 to 1.0 m/s (typically, 0.1 to 0.3 m/s)

depending on species and physiological condition

(Hutchinson, 1967). It is likely that many of these cells will

be removed by sedimentation in wetlands covered by

emergent vegetation providing shading and reducing wind

action. Flocculation of the cells within the wetland is also

possible although little experimental evidence has been

presented to date. Table 3-4 was generated by Gearheart

and Finney (1996), and it represents the only known application of the particle-size theory to show that colloidal fractions are flocculated in FWS systems (see Chapter 4 for

further explanation). Figure 3-2 illustrates the removal of

TSS observed for a fully vegetated FWS wetland treating

pond effluent. Attempts to settle algae from ponds in open

settling basins have not been successful, however, likely

due to the presence of light and wind action.

3.2.2.2 Filtration/Interception

Filtration, in the usual sense of this unit process, is not

likely to be significant in surface wetlands. Stems from

emergent plants are too far apart to effect significant entrapment of the particle sizes found in influent to these

wetlands. Furthermore, plant litter and detritus at the surface and bottom of the wetland are high in void fraction

such that filtration is not likely an important mechanism.

On the other hand, interception and adhesion of particles

on plant surfaces could be significant mechanisms for removal. The efficiency of particle collection would be dependent on particle size, velocity, and characteristics of

the particle and the plant surfaces that are impacted. In

wetlands, plant surfaces in the water column are coated

with an active biofilm of periphyton. This biofilm can adsorb colloidal and supracolloidal particles as well as absorb soluble molecules. Depending on the nature of the

suspended solids, they may be metabolized and converted

to soluble compounds, gases, and biomass or may physically adhere to the biofilm surfaces to eventually be

sloughed off into the surrounding water column. Similar

reactions may occur in the surface detritus or at the surficial

bottom sediment. To date, there have been no definitive

studies reported on the importance of this mechanism in

suspended solids removal in free surface wetlands.

3.2.2.3 Resuspension

In FWS wetlands, velocity induced resuspension is minimal. Water velocities are too low to resuspend settled par-
45

ticles from bottom sediments or from plant surfaces. Furthermore, fully vegetated wetlands provide excellent stabilization of sediments by virtue of sediment detritus and

root mats. The reintroduction of settled solids in wetlands

is most likely due to gas-lift in vegetated areas or

bioturbation or wind-induced turbulence in open water areas. Wetland sediments and microdetritus are typically near

neutral buoyancy, flocculant, and easily disturbed.

Bioturbation by fish, mammals, and birds can resuspend

these materials and lead to increases in wetland suspended

solids. The oxygen generated by algae and submerged

plants, nitrogen oxides and nitrogen gas from denitrification, or methane formed in anaerobic process may cause

flotation of particulates (Kadlec and Knight, 1996).

As discussed previously, the generation of new biomass by primary production or through the metabolism of

influent wastewater constituents will eventually result in

the return of some suspended materials back into the water column. The magnitude of wetland particulate cycling

is large, with high internal levels of gross sedimentation

and resuspension, and almost always overshadows influent TSS loading in natural or tertiary treatment wetlands.

The effluent TSS from a wetland rarely results directly from

nonremovable TSS in the influent wastewater and is often

dictated by the wetland processes that generate TSS in

the wetland. Typical background TSS concentrations expected in FWS wetlands appear in Table 3-5. It should be

noted that large expanses of open wetland prior to disFigure 3-2. Weekly transect TSS concentration for Arcata cell 8 pilot receiving oxidation pond effluent (EPA, 1999)

Table 3-4. Fractional Distribution of Bod, COD, Turbidity, and SS in the Oxidation Pond Effluent and Effluent from Marsh Cell 5 (Gearheart and

Finney, 1996)

BOD COD Turbidity SS

mg/l % mg/l % NTU % MG/L %

Oxidation Pond

Fraction

Total 27.5 100 80 100 11.0 100 31.0 100

Settleable 3.7 13 5 6 2.5 23 5.8 19

Supracolloidal 13.7 50 23 29 5.3 48 25.2 81

Colloidal, soluble 10.1 37 52 65 3.2 29 - -

Marsh Fraction

Total 4.8 100 50 100 3.9 100 2.3 100

Settleable 0 0 0 0 0.6 15 0.3 13

Supracolloidal 1.2 24 4 8 1.6 42 2.0 87

Colloidal, soluble 3.6 76 46 92 1.7 43 - -

Detention Time

TSS (mg/L)

0

20

40

60

80

100

120

140

160

Influent

0.75 days

1.5 days

2.25 days

3.0 days

3.75 days

4.5 days

5.25 days

6 days - Effluent

Average
46

charge structures could result in unusually high effluent

TSS concentrations due to the production of excessive

amounts of algae and induced high levels of wildlife activities that could produce effluent variations as typified in Figure 3-3.

High incoming TSS or organic loading will result in a

measurable increase in bottom sediments near the inlet

structure (Van Oostrom and Cooper, 1990; EPA, 1999).

However, no FWS treatment wetland has yet required

maintenance because of sediment accumulation, including some that have been in service for over 20 years.

3.2.3 Suspended Solids in Vegetated

Submerged Beds

One of the primary intermediate mechanisms in the removal of suspended solids by VSB systems is the flocculation and settling of colloidal and supracolloidal particulates. These systems are relatively effective in TSS removal

because of the relatively low velocity and high surface area

in the VSB media. VSBs act like horizontal gravel filters

and thereby provide opportunities for TSS separations by

gravity sedimentation (discrete and flocculant), straining

and physical capture, and adsorption on biomass film attached to gravel and root systems. Clogging of the filter

media has been of some concern especially with high TSS

loading, but documentation of this phenomenon has not

been forthcoming. The accumulation of recalcitrant or

slowly degradable solids may eventually lead to increased

headlosses near the influent end of the system. Design

features to overcome this are described in Chapter 5.

The importance of vegetation in VSB systems has been

debated for some time. Several recent studies have compared pollutant removal performance of planted and

unplanted VSB systems and have shown no significant difference in performance (Liehr, 2000; Young et al., 2000).

The importance of plant type has also been evaluated

(Gersberg et al., 1986; Young et al., 2000). Maximum root

length and growth rates have been reported. Although some

investigators claimed that certain treatment goals are likely

to benefit from certain plants, these claims have not been

sustained by others (Young et al., 2000). The extension of

root system within the gravel bed is dependent on system

loading, plant type, climate, and wastewater characteristics,

among other variables. It appears that a dominant fraction

of the flow passes below the root system in VSB facilities.

The role of root surfaces in TSS removal has not been proven

experimentally.

The contributions of internal biological processes to effluent TSS is likely similar to that found in FWS systems,

although algal contributions should be negligible.

Resuspension of separated solids is not likely since system velocities are low and scouring should not be significant. Furthermore, bioturbation in these systems should

be minimal. Background concentrations for VSB systems

have not yet been definitively documented with reliable

information.

3.3 Mechanisms for Organic Matter

Separations and Transformations

3.3.1 Description and Measurement

Organic matter in wastewater has been measured in a

number of ways over the years. Because the organic fraction in wastewater is often complex and the concentrations of the individual components relatively low, analyses

Table 3-5. Background Concentrations of Contaminants of Concern in FWS Wetland Treatment System Effluents

Constituent Range (mg/L) Typical (mg/L) Factors Governing Value Reference

TSS 2-5 3 Plant types, coverage, Reed et al.,1995;

Climate, wildlife Kadlec and Knight, 1996

BOD 5 

1 2-8 5 Plant types, coverage, Reed et al., 1995

Climate, plant density Gearheart, 1992

BOD 5

2 5-12 10 Plant types, coverage, Kadlec and Knight, 1996

Climate, plant density

TN 1-3 2 Plant types, coverage, Kadlec and Knight, 1996;

Climate, oxic/anoxic Reed et al., 1995

NH4

-

 N 0.2-1.5 1.0 Plant types, coverage, Kadlec and Knight, 1996;

Climate, oxic/anoxic Reed et al., 1995

TP 0.1-0.5 0.3 Plant types, coverage, Kadlec and Knight, 1996;

Climate, soil type Reed et al., 1995

Fecal Coli CFU/100 ml 50-5,000 200 Plant types, coverage, Watson et al, 1987;

Climate, wildlife Gearheart et al., 1989

1

Wetland system with significant open water and submergent vegetation

2

Wetland system fully covered by emergent vegetation
47

are often performed on the aggregate amount of organic

matter comprising organic constituents with common characteristics. Methods for total organic carbon (TOC) and

volatile solids (VS) measure the total amount of organic

matter present. Chemically oxidizable organic matter is

often measured as Chemical Oxygen Demand (COD),

expressed in units of oxygen, and biodegradable organic

matter is determined by the Biochemical Oxygen Demand

(BOD) procedure. All of these aggregate methods have a

place in assessing pollutant levels in water, but none will

provide information on specific organic molecules or their

fate in treatment processes. As a result, any qualitative or

quantitative model attempting to express mechanistic behavior of organic matter in a system is an empirical model

based on observation of the parameter of interest. In wetlands, physical, chemical, and biochemical reactions will

transform and/or separate organic matter, often leading to

different species of organic molecules. Thus the BOD (or

COD, or TOC) of the influent to a wetland does not measure the same organic constituents that appear in the effluent. This phenomenon is no different than what has already been discussed for TSS.

Most regulatory agencies today establish wastewater

discharge permit limits based on BOD values; thus much

of the data available is expressed as that parameter. It is

difficult to use BOD in mass balance calculations insofar

as it is a dynamic measure conducted over a finite time

(usually five days) and at a specified temperature (usually

20°C). One can use it to estimate roughly the oxygen requirements for aerobic systems if the rate of oxygen demand exertion (expressed by the biodegradation constant,

k) is known. To further complicate the use of this parameter, the analysis of BOD may or may not include the nitrogenous oxygen demand (NOD), which may be expressed concurrently or simultaneously with the carbonaceous oxygen demand. Thus primary influent BOD5

 may

measure carbonaceous organic matter, whereas effluent

from the wetland may include both carbonaceous and nitrogenous (ammonia) matter depending on the system.

Table 3-1 presents typical values of total and soluble BOD

for primary and septic tank effluents as well as lagoon effluents. Values of COD and VSS (volatile suspended solids) are also provided. There is no simple way to relate

BOD and COD values for wastewater insofar as the differences that exist between degradable and chemically oxidizable fractions.

3.3.2 Organic Matter in Free Water

Surface Wetlands

3.3.2.1 Physical Separations of Organic Matter

Table 3-4 illustrates that influents typically received by

FWS systems contain some particulate organic matter. A

significant amount of the influent from septic and primary

effluents is in the dissolved and colloidal fraction as would

be expected, whereas pond effluents may contain a sizeable fraction of supracolloidal material represented by algal cells. Separation of particulate organic matter would

occur by the same mechanisms as those described for

TSS. It is not uncommon to find organic matter removal in

the influent end of a FWS system that parallels TSS removal. Figure 3-4 illustrates predominant organic matter

separations and transformations that occur in FWS systems. As noted in section 3.3.2.2, biochemical transformations of the entrapped and settled organic matter will greatly

influence the apparent removal of total organic matter within

the water column.

Soluble organic matter may also be removed by a number of separation processes. Adsorption/absorption (the

movement of contaminants from one phase to another) is

an important process affecting some organic molecules.

The process is often referred to as sorption to cover both

adsorption and absorption processes in natural systems

because the exact manner in which partitioning to the solids occurs is often not known. Partitioning of organic matter between solids can be understood and predicted to

Figure 3-3. Variation in effluent BOD at the Arcata enhancement marsh (EPA, 1999) Effluent BOD (mg/L)

Date

0

2

4

6

8

10

12

14

16

July-91 July-92 July-93 July-94 July-95 July-96 July-97
48

some degree using physiochemical properties of the organic (e.g., water solubility, octanol-water partition coefficient). Sorption is often described by an isotherm

(Freundlich) relationship. The degree of sorption and its

rate are dependent on the characteristics of both the organic and the solid. In wetlands, the important solid surfaces would include the plant litter and detritus occurring

at the wetland surface or on the bottom and plant stems

and leaves often covered by a periphyton biofilm. The sorption process may be reversible or irreversible depending

on the organic and solid. Sorbents may eventually become

“saturated” with sorbate although often the sorbed organic

is biochemically transformed, renewing the sorbent capacity. Many of the wetland solid surfaces are also renewed

by continuous turnover of biomass that makes up the major component of the sorbent. It is believed that sorption

processes play an important role in organic separation in

wetlands, but the processes have not been adequately

quantified at this time.

Volatilization may also account for loss of certain organics. In this reaction, the organic partitions to the air. The

propensity for a given compound to move from liquid to

gaseous phase is measured by Henry’s constant. The

higher the value for a compound, the more likely it will partition to the gaseous phase. Generally, organic matter entering a wetland receiving pretreatment will not contain significant quantities of volatile compounds (VOCs). Facultative lagoons have been shown to remove 80 to 96% of

volatile organics from municipal wastewater (Hannah et

al., 1986). However, some of these organic materials may

be produced by biological transformations. Precipitation

reactions with organic matter found in wetland influents

have not been documented for municipal wastewater.

3.3.2.2 Biological Conversions of Organic

Matter

The Biochemistry

Biochemical conversions are important mechanisms

accounting for changes in concentration and composition

of biodegradable organic matter in wetlands. They may

account for removal of some organic constituents by virtue of mineralization or gasification and the production of

organic matter through synthesis of new biomass. Organisms will consume organic matter (and inorganic matter

as well) in order to sustain life and to reproduce. The organic matter in wastewater serves as an energy source as

well as a source of molecular building blocks for biomass

synthesis. The reactions occurring therefore are ones that

prepare the molecule for use by the organism or are directly involved in the extraction of energy or incorporation

of building blocks in the synthesis reaction. End products

of the reaction are waste products of the system. These

Figure 3-4. Carbon transformations in a FWS wetland

Air

Water

Floating Plants

Macrophytes

Emergent

Submergent

Roots

Epiphytes

Epiphytes

Influent Effluent

Sediment POC - Particulate Organic Carbon

DOC - Dissolved Organic Carbon

DIC - Dissolved Inorganic Carbon

VOC - Volatile Organic Carbon

PIC - Particulate Inorganic Carbon

Phytoplankton

DIC

DOC

POC

PIC

PIC

DIC

DIC POC

POC

DOC

POC

DOC

DIC

VOC

VOC

VOC CH4

H2

CO2

O2

CO2

CO2 O2

O2

Root Zone/Pore Water

Litter/Detritus

Peat

SO4 S

S

N2

POC

DOC

DIC

DIC

NO3 N2

CO2

CH4

H2

N2
49

reactions include oxidation/reduction processes, hydrolysis, and photolysis.

Since energy is a key element in the biochemical system, the reaction is classified as chemotrophic or phototrophic depending on whether the reaction utilizes a

chemical source of energy or light energy. The environment greatly influences the energetics of the reaction that

is driven by electron transfers. If the element at the end of

this set of transfers is oxygen, the reaction is referred to as

an aerobic reaction. Aerobic metabolism requires dissolved

oxygen and results in the most efficient conversion of biodegradable materials to mineralized end products, gases,

and biomass. Anoxic (anaerobic respiration) reactions use

nitrates, carbonates, or sulfates as terminal electron acceptors (in place of oxygen). Terminal electron acceptors

are subsequently reduced in these reactions, producing

end products such as nitrogen oxides, free nitrogen, sulfur, thiosulfate, and so forth. These reactions are typically

less efficient (biomass produced per unit substrate utilized)

than aerobic reactions and produce mineralized end products and gases, but less biomass per unit of substrate converted. Anaerobic metabolism takes place in the absence

of dissolved oxygen and uses organic matter as the terminal electron acceptor as well as the electron donor. These

transformations are the least efficient of the biochemical

reactions and will not result in the reduction in BOD (or

COD) unless hydrogen or methane is produced, since electrons released in the oxidation of the organic are passed

to electron acceptors (reduced products such as alcohols

and organic acids) that remain in the medium. The only

energy loss in the system is that due to microbial inefficiencies.

The transformations described previously rely on chemical energy to drive them. A very important transformation

that takes place in open water areas of FWS wetlands uses

sunlight energy to drive the reaction (photosynthesis). In

these reactions, inorganic carbon (CO2

 and carbonates) is

synthesized to form biomass (primary production), releasing oxygen. Thus both oxygen and organic matter are produced within the system.

The previous reactions and their stoichiometry, equilibrium, and rates are dependent on environmental variables

including dissolved oxygen, temperature, oxidation/reduction potential, and chemical characteristics, to name a few.

An excellent reference on microbial metabolism in wastewater systems can be found in Grady and Lim (1980).

The Organisms

Biochemical reactions of importance in FWS constructed

wetlands are carried out by a large number of organism

classes. They may be classified based on their position in

the energy or food chain (producers, consumers, or decomposers) or on their life form or habitat. A classical description relative to a wetland habitat would include

• Benthos—organisms attached or resting on the bottom within the plant litter and detritus or living within

the sediments

• Periphyton—both plants and animals attached to stems

and leaves of rooted plants

• Plankton—floating plants or animals whose movements are generally dictated by currents

• Neuston—plants or animals resting or swimming on

the wetland surface

• Nekton—swimming organisms able to navigate at will

Although no quantitative assessments have been reported, it is believed that the decomposers (bacteria, actinomycetes, and fungi) play the most important role in wetlands relative to the removal of organic matter by way of

mineralization and gasification. They are associated with

all habitats listed. They are also responsible for the synthesis of biomass and the production of organic metabolic

end products that may leach into the water column. These

organisms are often classified as aerobic, facultative, or

anaerobic, depending on the type of reactions that they

perform. They may also be classified with respect to their

source of energy and the type of chemical compound that

they use as a source of carbon.

The primary producers generate organic residues within

the wetland and add dissolved oxygen to the system. They

also play an important role in the removal and recycling of

nutrients in the wetland. The macrophytes will also prevent incoming radiation from entering the water column.

This shading will affect the growth of plant periphyton, phytoplankton, and submerged macrophytes, will interfere with

surface mass transfer of oxygen, and will moderate wetland water temperatures. These rooted plants provide significant submerged surfaces for growth of periphyton. The

submerged surface area provided by selected FWS vegetation varies from 2.2 to 6.5 m2

/m2 for a wetland 0.5 m

deep, depending on plant species and coverage (EPA,

1999).

Oxygen Transfer to FWS Wetlands

Oxygen is a critical element in the biochemical transformation of organic matter (as well as other compounds to

be described later). As briefly discussed previously, biochemical reactions that use dissolved oxygen (DO) are

efficient processes and yield mineralized end products. An

aerobic environment is highly desirable in wastewater treatment systems in which the target is effective removal of

BOD. There are three possible sources of DO in wetland

systems—surface aeration, photosynthesis, and plant oxygen transfer.

When a gas is dissolved in water, the process is generally treated as a mass transfer occurring over four steps

(two-film theory). It presumes a thin but finite gaseous and

liquid film at the air-water interface. The gas must pass

through the bulk gas phase, then through the gas film, to

the liquid film, and then into the bulk liquid. For oxygen,

the liquid film is the rate-limiting step and controls the rate

of mass transfer, although for a quiescent water body, the

bulk liquid transport may control the process. By use of

Fick’s Law, the diffusional process can be modeled pro-
50

ducing the commonly used oxygen mass transfer relationship:

dC/dt = KL

a (Cs-Co) (3-1)

Where C = concentration of oxygen in water (M/L3

)

t = time (T)

KLa = overall mass transfer coefficient for oxygen in water (1/T)

Cs

 = steady-state oxygen saturation concentration in water (M/L3

)

Co

 = the dissolved oxygen concentration in the

bulk water (M/L3)

This equation indicates that the rate of oxygenation is

dependent on the overall mass transfer coefficient, the

steady-state DO saturation concentration for oxygen in

water, and the bulk water DO. Generic symbols for mass

(M), time (T), and length (L) are employed for units. The

value of KL

a is a function of the physical characteristics of

the wetland. For typical FWS systems, isotropic turbulence

would occur whereby there is neither a significant velocity

gradient nor shearing stress. For this system, the value of

KLa would be a direct function of wetland velocity and the

oxygen molecular diffusion coefficient and an indirect function of system depth (O’Connor and Dobbins, 1958) according to the following equation:

K2

 = [DLU]1/2/ H3/2 (3-2)

Where K2 = surface reaeration coefficient (1/T)

DL = oxygen molecular diffusion coefficient (L2/T)

U = velocity of flow (L/T)

H = depth (L)

An estimate of surface aeration in an open water zone

of a FWS system can be made assuming a typical wetland water velocity of 30 m/d and a depth of 0.3 m. At

20°C, the mass transfer coefficient would be approximately

0.43/d. Assuming a bulk wetland water DO of 0 mg/L, the

mass transfer of oxygen would be about 3.9 mg/L/d or 1.2

g/m2

/d. For more realistic values of the water DO, the transfer would range from 0.5 to 0.9 g/m2

/d. These numbers

compare favorably with values reported by the TVA, which

range from 0.50 to 1.0 g/m2

/d (Watson et al., 1987). These

open water-zone values will be higher than would be observed in wetlands with plant cover because plant debris,

floating plants, and emergent plant surfaces impede mass

transfer across the surface. In the open zones, the bulk

water DO (which might approach 0 mg/L) would not be the

same as the DO near the water surface (which might approach saturation during the daylight hours). See Figure

3-5 for a submergent plant (open water) zone, which illustrates the inaccuracy that would result in assuming a DO

of 0 mg/L in the bulk water column. The actual DO values

in Figure 3-5 are from polishing FWS applications and are

therefore higher than would be expected in a treatment

FWS system. Actually, a value near zero DO is normal in

fully vegetated or emergent zones (Figure 3-5b) of these

latter systems.

Oxygen will also be transferred to the wetland by virtue

of photosynthesis carried out by phytoplankton, periphyton, and submerged plants. The general relationship for

photosynthesis of green plants implies that approximately

2.5 g of oxygen would be evolved per g of carbon fixed as

cell mass. In the absence of sunlight energy, oxygen would

be consumed by these plants in respiration. Presuming

highly productive conditions, about 1.0 g C/m2

/d may be

produced during daylight hours resulting in a generation

of 2.5 g O2/m2/d (Lewin, 1962). One may deduce a somewhat larger value presuming typical net primary production of a wetland to be about 4.0 g total biomass/m2

/d

(Mitsch and Gosselink, 1993) and about 1.0 g O2

/g net

biomass produced by photosynthesis. These numbers

imply that in open areas where sufficient photosynthesis

might occur, oxygen should be available for aerobic oxidation within the water column, at least during part of the

day. Active plant respiration during the evening may completely negate the oxygenation estimated during daylight

hours, however. The DO concentrations have been documented for both emergent and submergent plant zones in

the Arcata tertiary/polishing/enhancement marsh system

(Figure 3-5). Clearly, the submergent plant zone provided

more DO by virtue of greater photosynthesis and surface

aeration. This figure also clearly demonstrates that the

bulk water column is not mixed with respect to DO, being

highest at the surface where atmospheric reaeration and

photosynthesis are predominant. Rose and Crumpton

(1996) have demonstrated the effects of emergent macrophyte stands and open water in a prairie wetland. They

found that the emergent sites had low DO concentrations

and were almost always anoxic, whereas sites at the edge

of the stand had higher DO concentrations, with significant increases in DO during the daylight hours. The open

water areas consistently exhibited higher values of DO

with diurnal changes up to 10 mg/L. These results are

consistent with the presumption that emergent stands provided a heavy canopy cover, along with small floating

plants and plant litter that obscured light penetration and

subsequent photsynthesis and also affected surface aeration. The plant litter stands at the margin were open and

allowed light penetration, encouraging photosynthesis by

periphyton, phytoplankton, and submerged plants. Open

areas allowed even more opportunity for photosynthesis

as well as surface aeration.

The role of rooted, emergent plants on oxygen transfer

to the wetland system is subject to controversy. Because

these wetland plants are typically rooted in soils that are

anaerobic, they have evolved special airways to allow the

efficient movement of atmospheric oxygen to the root system. Dead and broken plant shoots may also allow for transport of some oxygen to the root zone. There is a sufficient

body of evidence to show that significant amounts of oxygen are transported down these passages and that other

gases, such as carbon dioxide, hydrogen, and methane,

may pass upward through these same channels. Gas transfer is the result of humidity-, thermally-, and/or Venturi-induced convective flow as well as diffusion. Gas exchange

through the root surface is by diffusion. Although photo-
51

synthesis does not seem to enhance the concentration of

oxygen in the shoots, sunlight can cause convection to

increase because light amplifies the degree of stomatal

openings in the leaves, and higher temperatures cause

steeper diffusion gradients for gases (Armstrong and

Armstrong, 1990). A large proportion of the oxygen entering the plant is vented to the atmosphere instead of being

used in plant respiration or effluxed to the root system.

A number of studies have been made on the transport of

oxygen by plants. Only one field study has been reported

on an operating wetland (Brix and Schierup, 1990). They

found that for Phragmites in both FWS and VSB systems,

the oxygen transported almost exactly balanced respiration by the plants. The range of values reported by other

investigators is 0 to 28.6 g O2/m2/d, whereas rates of 0 to 3

g O2/m2/d were found in 15 of the 23 studies. Kadlec and

Knight (1996) pointed out that studies to date infer that

oxygen transfer measurements made from BOD and ammonium losses are not accurate. At present, it seems reasonable to presume that plant oxygen transfer is not an

important source of oxygen in most wetland systems.

The DO concentration found in FWS systems is dependent on the rate of oxygen transfer and the rate of oxygen

uptake. If one equals the other, system DO will remain

constant. If respiration exceeds transfer, DO values will

fall to zero and anaerobic conditions will prevail. The principle consumers of oxygen (respiration) in wetlands include

microorganisms that consume oxygen during normal exogenous and endogenous respiration and plants that carry

out respiration when sunlight energy is unavailable as an

energy source. Sources of oxygen requirements will include those from influent organic matter, stored organic

matter in living biomass (endogenous respiration), dead

plant litter at the surface and bottom of the wetland, dead

periphyton and plankton suspended in the water column

or residing at the wetland surface or within the benthal

deposits, and influent ammonium nitrogen. One factor that

should be considered in wetland design is the determination of an oxygen balance. A mass balance on oxygen is

difficult to perform in FWS systems owing to the highly

complex and dynamic characteristics of the system. An

estimate of influent oxygen demand (carbonaceous and

nitrogenous) is possible, but uptake due to plant respiration and decomposition of plant carbon is more difficult to

predict (Table 3-6). The sources of oxygen are also difficult to quantify as discussed previously. Total coverage of

wetlands by emergent plants results in little reliable oxygenation (some oxygen will likely be produced by periphyton near the water surface, but quantification of this is unreliable at the present time). Use of open areas to promote photosynthesis by submerged plant and plankton

appears to be a logical approach, but again, quantitative

estimates of transfer are difficult to assess based on current data. It should be emphasized that DO is not a steady

value in wetlands but will vary in a diurnal fashion with

photosynthesis. It is not unrealistic to presume that wide

fluctuations in DO will occur, especially in active systems.

As will be described later, the DO concentration will also

change along the wetland length as demands and supplies change.

FWS Biological Reactions

Influent particulate organic matter may be entrapped

within biofilms attached to emergent plant surfaces or accumulated on the wetland floor within the plant litter and

sediments (Figure 3-4). Experience suggests that much of

this material is accumulated very close to the influent strucFigure 3-5. Dissolved oxygen distribution in emergent and submergent zones of a tertiary FWS (EPA, 1999)

top middle bottom

Gearheart-1

Gearheart-3

Gearheart-5

0

2

4

6

8

10

12

14

16

DO (mg/L)

a) Vertical distribution of DO in a submergent

 plant zone of the Arcata Enhancement Marsh

top middle bottom

b) Vertical distribution of DO in an emergent

 plant zone of the Arcata Enhancement Marsh

Gearheart-2

Gearheart-4

Gearheart-6

0

0.5

1

1.5

2

2.5

3

DO (mg/L)
52

ture. In addition, particulate organic matter deriving from

dead plant litter accumulates on the floor of the wetland as

well as in surface mats. The quantity of this material is

dependent on the species of plants growing in the area

and their coverage. Emergent plant communities have

higher potential production rates than submerged communities (Wetzel, 1983). The accumulated organic debris at

the wetland floor degrades at different rates depending on

the composition of the organic matter. Influent particulate

organic matter from primary settling or septic tank effluents is easily degradable in most environments. Algal cells

from pond effluents are biologically less available. Emergent macrophytes produce much more structural material

(lignin, cellulose, and hemicellulose) than do submerged

and floating leafed plants, and this material degrades relatively slowly (Godschalk and Wetzel, 1978).

Much of the particulate organic matter would be hydrolyzed, producing lower molecular weight organic compounds that are more soluble in water, which leach back

into the water column and contribute to the soluble BOD

downstream. In the presence of oxygen, these compounds

would be oxidized by microbes to CO2

, oxidized forms of

nitrogen and sulfur, and water. Under anaerobic conditions,

these compounds may be converted to low molecular

weight organic acids and alcohols. Under strict anaerobic

conditions, methanogenesis may occur whereby these

compounds are converted to gaseous end products of CH4,

CO2, and H2. In the presence of sulfates, sulfur-reducing

microbes will convert these low molecular weight organic

compounds to CO2 and sulfide. Either of these anaerobic

reactions will essentially remove organic matter from the

system. As the degradability of the material decreases,

the decomposition rates slow and the nature of the metabolic end products change. It has been suggested that

soluble organic matter has a half-life of about three days

while organic sediment may exhibit a half-life on the order

of four months (EPA, 1999). The rates of degradation are

also temperature dependent. Thus sediment organic matter may accumulate during the colder months and be more

rapidly degraded in the spring when water temperatures

rise. This increase in degradation will result in an increase

in soluble organic matter released to the water column and

a concomitant increase in oxygen demand, a phenomenon

that has also been observed in facultative ponds for nearly

a century.

The result of these metabolic activities is that (1) organic

matter concentrations (and oxygen demand) may increase

downstream as particulate organic material is solubilized,

(2) the effect of temperature on observed organic matter

removal may not be as significant in wetlands as would be

predicted by typical temperature correction relationships

for biochemical reactions, and (3) the organic matter residual downstream is a combination of recalcitrant organic

matter in the influent, likely very small, soluble organic

compounds released from plant decomposition and particulate organic matter released from dead plant and microbial materials.

The source of the soluble fraction of organic matter in

the wetland is, as has been stated, influent to the system

and soluble material released from the decomposition of

influent particulate matter and dead plant and microbial

tissue. This soluble material is most likely sorbed onto plant

surface biofilms, although a fraction may be sorbed onto

biomass suspended in the water column, and still another

fraction will diffuse to the debris at the wetland floor or

surface. The sorbed organic matter will be metabolized by

organisms associated with the biofilms. The metabolic

pathway and the end products of this metabolism will be

dependent on the presence or absence of oxygen. As previously described, areas of the wetland populated with

dense emergent macrophytes can supply only a small fraction of the oxygen necessary to satisfy demand. Typical

wetland profiles in these areas suggest that most of the

water column is anoxic, as are the sediments below (small

microsites containing oxygen may be found adjacent to

active plant roots). The degree of anoxia would be dependent on the organic and nutrient load to the wetland area.

Open areas of wetland (containing submerged plants) and

the margins adjacent to emergent macrophyte coverage

do demonstrate aerobic conditions throughout the wetland

depth (again, dependent on organic and nutrient load and

type of vegetation).

As described previously, the mechanisms that regulate

dissolved organic matter removal in wetlands include biodegradation, sorption, and photolysis. Different operative

mechanisms may act on different types of organic matter.

As a result, fundamental mechanisms should be considered in wetland designs and operation to enhance removal

processes. Wetlands receiving municipal wastewater pond

effluents may, for example, produce a net increase in dissolved organic matter (Barber et al., 1999). Details of design strategies for FWS systems based on these observations are found in Chapter 4.

3.3.3 Organic Matter in Vegetated

Submerged Beds

Vegetative submerged beds act as fixed-film bioreactors.

As described in section 3.2.3, the actual role of plants in

these beds is controversial. Some research (Gersberg et

al., 1986) has claimed effective results with selected species, but these claims have not been substantiated by others. The presence of a root structure would provide addiTable 3-6. Wetland Oxygen Sources and Sinks

FWS VSB

Process Source Sink Source Sink

Reaeration + (*) +

Photosynthesis +

Plant O2 Transpiration + +

Influent BODult; (oxid./endog.) + (†) + (†)

Influent NH4 + (†) + (†)

Plant Respiration + + (†)

Plant Decomposition + + (†)

(*) may be estimated in some instances

(†) May be calculated for wetland system
53

tional surface for biofilm attachment. Macrophytes may also

contribute some oxygen to the granular bed as previously

described. Based on that review, rates of oxygen transport by macrophytes range from 0 to 3 g 02

/m2/d. However, it has been found that root penetration in the bed is

only partial, and there is a significant amount of flow under

the root zone. Furthermore, plant oxygen transfer would

be unreliable for a significant portion of the year due to

plant senescence.

Particulate organic matter is removed in VSB systems

by mechanisms similar to suspended solids separation in

horizontal gravel beds with the same media size. The separated solids from the influent wastewater and plant litter

would undergo decomposition much the same way as occurs in FWS systems. Hydrolysis will generate soluble

compounds. Those compounds and soluble organic matter from the influent to the system or cycled from solids

decomposition will most likely sorb to biofilm surfaces attached to the media, plant roots, and plant litter accumulated at the bed surface or within the interstices of the

media. Oxygen sources to the VSB would be limited to

some small amount of surface aeration and plant-mediated transport. The thatch that accumulates at the bed

surface would inhibit or at least slow down surface transport. It is possible that some aerobic metabolism would

occur in these beds, but the predominant biological mechanism is likely to be facultative/anaerobic. Typical values of

DO in VSB systems are very low (<0.1 mg/L). In VSB systems where ORP has been measured, values were typically quite low, indicating strong reducing conditions

(Lienard, 1987). Thus the predominant metabolic pathways

are most likely anaerobic. As described previously, the

anaerobic pathways leading to BOD removal from the system would be methanogenesis, sulfate reduction, or denitrification, all yielding gaseous end products. These reactions are temperature dependent and therefore are likely

to slow down or cease in winter months. As discussed previously, the processes will resume as the water warms up,

and high releases of gaseous end products and soluble

organic matter may occur. Some clogging of the bed may

occur due to accumulation of slowly degradable and recalcitrant solids. Low-loaded systems may exhibit some

aerobic reactions, especially near the effluent end of the

process. Residual effluent BOD from VSB systems is likely

to be somewhat more consistent than that from FWS systems because of the presence of less plant matter in the

water column.

3.4 Mechanisms of Nitrogen Separations

and Transformations

3.4.1 Description and Measurement

In waters and wastewater, the forms of nitrogen of greatest interest are, in order of decreasing oxidation state, nitrate, nitrite, ammonia, and organic nitrogen. All nitrogen

forms are reported in wastewater as nitrogen, N. All of these

forms, including nitrogen gas (N2), are biochemically

interconvertible and are components of the nitrogen cycle.

Analytically, organic nitrogen and ammonia can be determined together, and are termed “Total Kjeldahl nitrogen”

(TKN). Organic nitrogen in wastewater includes proteins,

peptides, nucleic acids, and urea. Organic nitrogen may

be found in both soluble and particulate forms. The other

nitrogen species are water soluble. Ammonia nitrogen may

be found in the un-ionized form, NH3

, or the ionized form,

NH4

+, depending on water temperature and pH. The ionized form is predominant in wetlands. At 25°C and a pH of

7.0, the % un-ionized ammonia is approximately 0.6%.

The discharge of nitrogen to receiving surface and

ground water sources is of concern for a number of reasons. Excessive accumulation of nitrogen in surface waters can lead to ecological imbalances that may cause

overgrowth of plants and animals, leading to water quality

degradation (eutrophication). High concentration of the unionized ammonia species are toxic to fish and other aquatic

life. Nitrate and nitrite nitrogen constitute a public health

concern, primarily related to methemoglobinemia and carcinogenesis. Ammonia nitrogen may deplete dissolved

oxygen in natural waters by way of microbial nitrification

reactions. As a result, discharge permits may be written to

control any or all species of nitrogen. Most commonly,

ammonia or total nitrogen are the target pollutants specified depending on the receiving stream. Typical concentrations of the nitrogen species found in primary, septic

tank, and treatment pond effluents are shown in Table 3-1.

It should be noted that whereas primary and septic tank

effluents would contain organic nitrogen and ammonia,

treatment ponds may contain either reduced or oxidized

forms of nitrogen depending on loading and season of the

year. Organic nitrogen in the latter systems would be primarily associated with algal cells. It is important to note

that when evaluating the performance of wetlands relative

to nitrogen, both total nitrogen and the species of nitrogen

are important. Mass balances must be conducted on total

nitrogen species, not on just one or two forms, to generate

meaningful data.

3.4.2 Nitrogen in Free Water Surface

Systems

3.4.2.1 Physical Separation of Nitrogen

Species

There are a number of separation processes that will

affect nitrogen species in wetlands. Nitrogen associated

with suspended solids (organic nitrogen) may be removed

by many of the processes described earlier for the removal

of TSS including flocculation, sedimentation, filtration, and

interception processes (Figure 3-6). Sorption of both particulate and soluble organic nitrogen may occur on biofilms

associated with emergent macrophytes, plant litter, or other

detritus at the FWS surface or on the bottom. Ion exchange

of ammonium (NH4

+) by clay minerals in the wetland soils

may play a role in nitrogen separation if this species either

diffuses into the soil layer or is biologically produced by

the ammonification of organic nitrogen solids located in

the benthal layer. The exchange capacity of the clay minerals would be important in this assessment. It must be

emphasized that the exchange mineral would have limited
54

long-term capacity unless regeneration by chemical or biochemical action takes place. It is not likely that ion exchange

would play an important or dependable role in nitrogen

removal after a start-up period in most wetland systems.

Furthermore, the native soil layer is buried under detritus

and plant litter within a period of time, thereby isolating

these clay minerals from the wetland system. Ammonia

(NH3

) gas may also be removed from the system by stripping. As discussed previously, the quantity of un-ionized

ammonia at neutral pH values is low, but during active

photosynthesis in open water zones, pH values may rise

to values as high as 8.0 to 8.5 depending on water alkalinity. At that pH, the fraction of NH3

 (NH3/NH4

+) may increase

to 20 to 25% at 20°C. If surface turbulence is high due to

wind action, significant losses of nitrogen may occur in

these open water areas if ammonia concentrations are

high. Few, if any, well-controlled studies have been conducted to examine the importance of volatilization losses

in FWS systems with open areas.

3.4.2.2 Biologically Mediated Transformations

of Nitrogen Species

Ammonification

Almost half of the municipal wastewater nitrogen content, as received at the treatment facility, is in the organic

nitrogen form. The rest has usually already been converted

to ammonium-nitrogen (EPA, 1993) in the sewer. The biological transformation of organically combined nitrogen to

ammonium nitrogen during organic matter degradation is

referred to as ammonification, hydrolysis, or mineralization. The process occurs under aerobic and anaerobic

conditions, but has been described as slower for the latter

by Mitsch and Gosselink (1993). Other environmental engineering literature, however, does not support any difference in rates under varying oxygen states because of their

primary dependence on enzymatic pathways. The rate of

this process is primarily dependent on pH and temperature, increasing with increased temperatures. Municipal

wastewaters have been demonstrated to be fully hydrolyzed in 19 hours at temperatures of 11° to 14°C (Bayley

et al., 1973). Once the ammonium is formed, it can be

absorbed by plants through their root systems, immobilized by ion exchange in the sediments, solubilized and

returned to the water column, volatilized as gaseous ammonia, anaerobically converted back to organic matter by

microbes, absorbed by phytoplankton/floating aquatic

macrophytes in the water column, or aerobically nitrified

by aerobic microorganisms.

Nitrification

In the presence of dissolved oxygen, microbes in the

water column or within the biofilms may convert ammonium to nitrite and nitrate nitrogen in a two-step process.

In this process about 4.3g of 02

 are required per g ammonium nitrogen oxidized to nitrate and 7.14 g of alkalinity as

CaCO3 are consumed. The process is temperature and

pH dependent (EPA, 1993). The reaction may take place

in an aerobic water column by suspended bacteria and

Figure 3-6. Nitrogen transformations in FWS wetlands

Air

Water

Floating Plants

Macrophytes

Emergent

Submergent

Roots

Epiphytes

Epiphytes

Influent Effluent

Sediment - Particulate Organic Nitrogen

Phytoplankton

PON

DON

O2

O2

N2

O2

Root Zone/Pore Water

Litter/Detritus

N2

NOx

NO2

-

NO3

-

NH4

+

PON

DON - Dissolved Organic Nitrogen

NO2

-

NO3

-

PON

DON

NO2

-

NO3

-

NH4

+

PON

PON

PON NOx NH4 , N2

DON NH4

+

NH4

+

PON

PON

DON

NH4

+

NH3

NH3

NO2

-

NO3

-
55

within any aerobic biofilms. Nitrate is not immobilized by

soil minerals and remains in the water column or pore water

of the sediments. It may be absorbed by plants or microbes

in assimilatory nitrate reduction (converted to biomass via

ammonium) or may undergo dissimilatory nitrogenous oxide reduction (nitrate reduction pathways referred to as

denitrification).

Denitrification

Dissimilatory nitrate reduction or denitrification is carried out by microorganisms under anaerobic (anoxic) conditions, with nitrate as the terminal electron acceptor and

organic carbon as the electron donor (EPA, 1993). That is

to say, the reaction occurs in the absence of oxygen and

requires an organic carbon source. The products of denitrification are N2

 and N2O gases that will readily exit the

wetlands. The denitrification reaction occurs primarily in

the wetland sediments and in the periphyton films in the

water column below fully vegetated growth where DO is

low and available carbon is high. The minimum carbon to

nitrate-nitrogen ratio for denitrification would be about 1 g

C/g NO3

-N. Decomposing wetland plants and plant root

exudates are potential sources of biodegradable organic

carbon for this purpose. These sources are most readily

available at the beginning of senescence. Organic carbon

will be consumed (satisfying some oxygen demand: approximately 2.86 g O2

 per g nitrate nitrogen reduced) and

alkalinity is produced (approximately 3.0 g CaCO3

 alkalinity per g nitrate nitrogen reduced). The process is temperature and pH dependent. Denitrification in the sediments

may supply N2 for fixation by bacteria and for plant uptake

in the root zone (nitrogen fixation) if the system is nitrogen-poor. The remainder will remain in equilibrium with N2

in the water column. The nitrogen gas in the water phase

may be available for nitrogen fixation by some periphyton

and phytoplankton. There will also be an air-water interchange of N2

. The losses to the air represent nitrogen

losses from the system, and these are greater in open water

than in fully vegetated zones.

Nitrogen Fixation

Nitrogen gas may be converted to organic nitrogen by

way of selected organisms that contain the enzyme nitrogenase. The reaction may be carried out aerobically or

anaerobically by bacteria and blue-green algae. Nitrogen

fixation occurs in the overlying water in FWS open water

zones, in the sediment, in the oxidized rhizosphere of the

plants, and on the leaf and stem surfaces of plants (Reddy

and Graetz, 1988). It may be a significant source of nitrogen in natural wetlands but is not important in systems

treating wastewater in which nitrogen is plentiful.

Plant Uptake (Assimilation)

Wetland plants will assimilate nitrogen as an important

part of their metabolism. Inorganic nitrogen forms are reduced by the plant to organic nitrogen compounds used

for plant structure. During the growing season, there is a

high rate of uptake of nitrogen by emergent and submerged

vegetation from the water and sediments. Increased immobilization of nutrients by microbes and uptake by algae

and epiphytes also lead to a retention of inorganic forms

of nitrogen in the wetland. Estimates of net annual nitrogen uptake by emergent wetland plant species vary from

0.5 to 3.3 gN/m2

/yr (Burgoon et al., 1991). Reeds and bulrushes are at the lower end of this range, whereas cattails

are at the higher end. Estimates for epiphytes and microbes

in wetland systems have not been found. During the active growth period, a significant amount of the total plant

nitrogen is in the stems and leaves above the sediments.

During senescence, the nitrogen translocates back to the

roots and rhizomes for storage. However, a substantial

portion of the nitrogen is lost to the water column through

litter fall and subsequent leaching. This generally leads to

a net export of nitrogen in the fall and early spring. The

extent of recycling of nitrogen within the wetland is dependent on nitrogen loading to the system. Treatment wetland

systems are considered as eutrophic wetlands with excessive nutrient levels. As a result, intrasystem cycling is

less important to the treatment process than it would be if

the hydrologic regime were more varied (Twinch and

Ashton, 1983).

3.4.2.3 Nitrogen in Free Water Surface

Wetlands

Figure 3-6 illustrates the highly complex series of reactions for nitrogen in FWS systems. Particulate organic nitrogen entering the wetland as wastewater influent or produced in the wetland by plants is separated. It may be

found associated with biofilms attached to plant structures

in the water column, the wetland sediment, or in floating

litter and detritus. The biodegradable compounds will be

ammonified by aerobic or anaerobic organisms associated

with the biofilms and sediment surfaces. The recalcitrant

organic nitrogen will accumulate and eventually become a

part of the deep sediments.

Ammonium released from the particulate organic nitrogen in the sediments is available to the emergent and submerged macrophytes as an important nutrient. Uptake

occurs during the growing season, which increases the

concentration gradient and the release of more ammonium.

Excess ammonia may remain in reserve in the sediment

and leach from the sediment into the water column, where

it may undergo biological oxidation (nitrification) under

aerobic conditions. Release of ammonium into the water

column in the fall and early spring is not an unusual phenomenon in wetlands. The leached ammonium may be

taken up by epiphytes or plankton found in the water column or be attached to emergent and floating plants. Nitrification of ammonium requires DO and is therefore limited

to areas of the wetland where oxygen is available. There

may be some nitrification occurring adjacent to plant rhizomes where oxygen leaks from the plant. The relative

importance of this pathway in wetlands is minor in treatment wetlands because most sediments below emergent

canopies are anaerobic. Little nitrification would be expected in these regions of the wetland. In the water col-
56

umn near the surface in open areas, oxygenation may be

sufficient to ensure significant nitrification. The important

variable in sizing an open water zone of a FWS system for

nitrification is organic plus nitrogenous loading. More precisely, it is the oxygen demand loading (carbonaceous plus

nitrogenous oxygen demand [CBOD + NOD]) to each wetland zone that will dictate whether oxygen is present. Typically, nitrification will not be initiated until a majority of the

organic compounds have been removed. Thus, nitrification in the water column would not be expected in the initial settling zone, but could occur in subsequent open water/aerobic zones.

The nitrate produced by nitrification or introduced with

the system influent (e.g., from oxidation ponds achieving

nitrification) may be taken up by periphyton or plankton.

Small amounts produced within any aerobic sediments

would be taken up by plant roots or may diffuse upward

into the water phase. Under anaerobic conditions and in

the presence of organic matter, microbes associated with

attached biofilms or suspended in the water column may

convert nitrate to nitrogen gases (NO2, N2) via denitrification. Some nitrate will also diffuse into the sediments where

it is available for plant uptake or can be denitrified as well.

Moving downstream from the wetland influent, the reactions of nitrogen could be expected to occur sequentially

such that total nitrogen levels should drop in the settling

zone (owing to separation of organic nitrogen), followed

by ammonia release, nitrification, and denitrification. Plants

will attenuate this sequence by releases and uptake

throughout the annual growth/senescence cycle. Systems

loaded at a level such that oxygen demand exceeds oxygen supply will not exhibit significant nitrification. Seasonal

change as well as influent variability will greatly impact

system performance. Highly nitrified influent from pretreatment systems may provide excellent nitrogen removal

during warm seasons when sufficient organic matter is

available (primarily from plant decomposition) for denitrification. Nitrogen release may be significant in fall and early

spring seasons during plant senescence and death. The

background nitrogen values found in Table 3-5 reflect contributions by internal recycling of nitrogen within the system.

In open water zones of FWS systems, elevated pH and

water temperature may enhance NH3-N volatization to the

degree that it becomes a significant removal mechanism.

This mechanism has been shown to reach levels as high

as 50% or more under optimum conditions in stabilization

ponds, but FWS open water zones are smaller in size,

therefore minimizing this pathway under normal conditions

(EPA, 1983).

Wastewater discharge permits are normally written so

as to limit effluent ammonia concentrations (either seasonally or all year) or total nitrogen concentrations. For

ammonia removal, the processes that will achieve effective removal include plant uptake, nitrification, volatilization, and ion exchange. The latter two typically have only

minor impact in most FWS systems. Plant uptake is seasonal and requires harvesting prior to plant senescence.

Fortunately, seasonal nitrogen uptakes may parallel ammonia restrictions in those cases in which these restrictions are based on summer oxygen demands and/or certain game fish maximization. Generally, the designer must

be concerned with achieving reliable ammonia removal by

means of nitrification. This may be achieved at low loading (oxygen demand) with sequencing closed and open

wetland areas (See Chapter 4). It should be noted that

nitrification is temperature dependent, so that rates will significantly slow in winter months, especially in colder climates.

Nitrogen removal may be achieved by plant uptake/harvesting, nitrification/denitrification, volatilization, and ion

exchange. Again, the latter two are considered to be of

minor consequence in most FWS systems. Plant uptake

and harvesting requires careful system management and

can be costly. The requirement for denitrification requires

nitrification as explained previously plus adequate decaying plant organic carbon in a region free of dissolved oxygen. Sequential designs are also best to achieve this goal

(See Chapter 4). Note that temperature affects both nitrification and denitrification so that rates can be significantly

reduced during the colder months, which may control design requirements.

3.4.3 Nitrogen in Vegetated Submerged

Beds

As described in section 3.3.3, VSB systems incorporate

anaerobic fixed-film biological reactions. Organic nitrogen

trapped within the bed will undergo ammonification. The

released ammonia may be available for plant uptake depending on the location of plant roots. Flow below the plant

roots will carry ammonium downstream. Plant uptake (0.03

to 0.3 g/m2/d) of nitrogen is low compared to typical nitrogen loading to VSB systems. As described previously, dependence on plant uptake for nitrogen removal requires

harvesting and is not effective during plant senescence

and death.

Oxygen sources in VSB systems are negligible, and it is

most likely there will be insufficient oxygen to promote reliable nitrification in all but the most lowly loaded systems.

Any nitrification occurring may be found in the root zone

adjacent to rhizomes or near the bed surface where some

surface oxygen transfer might occur. If nitrification occurs,

it would occur downstream where oxygen demand is lowest.

Conventional VSB systems would seem to be well suited

for denitrification of nitrified influents. These beds are

anaerobic. However, they require a supply of organic carbon from decomposing plant residue entrapped within the

bed or aerobically decomposed products of plant biomass

at the bed surface, which may also leach into the anaerobic zones during rainfall events. The supply of carbon is

seasonal, however, since it would be highest after plant

senescence. Low temperature will slow the process during the winter months.
57

From this discussion, it is clear that conventional VSB

systems do not represent reliable, cost-effective systems

for ammonia removal. To improve, the loading to the system would necessarily have to be low. Nitrogen removal

from well-nitrified influents to the system may be possible,

but seasonal availability of carbon may create carbon limitations that should be considered in the design (See Chapter 5).

3.5 Mechanisms of Phosphorus

Separations and Transformations

3.5.1 Description and Measurement

Phosphorus occurs in natural waters and wastewater

primarily as phosphates. They are classified as orthophosphates, condensed (pyro-, meta-, and poly-) phosphates,

and organically bound phosphates. They may be in solution or particulate form. Organic phosphates are formed

primarily by biological processes and are found in raw

wastewater as food residues and body wastes and in

treated wastewater as living or nonliving biota (e.g., algae

and bacteria from treatment ponds). Inorganic phosphorus found in wastewater most often comes from various

forms of personal and commercial cleaning solutions or

from the treatment of boiler waters. Storm waters carry

inorganic forms of phosphorus from fertilizers into combined sewers. Classification of phosphorus is based on a

variety of analytical methods. The typical concentrations

of phosphates in influent wastewaters to a treatment wetland appear in Table 3-1. Modern P concentrations are

~50% of those shown for septic tank effluent in this table.

Table 3-3 indicates that the major fraction of these phosphate forms are in the colloidal range, and Munch et al.

(1980) indicated that 80% of the phosphate was split almost equally between colloidal and supracolloidal fractions.

Phosphorus is one of the most important elements in

ecosystems. It is often the major limiting nutrient in freshwater systems. Since there is no important gaseous component in the biogeochemical cycle, phosphorus tends to

move to the sediment sink in natural systems and become

scarce in the ecosystem. In fact, it is the accretion of mineral phosphates and biomass in the sediment that is the

primary mechanism for phosphorus removal in the wetland environment.

3.5.2 Phosphorus in Free Water Surface

Systems

3.5.2.1 Physical/Chemical Separations of

Phosphorus

Particulate phosphate may be deposited onto the FWS

system sediment by sedimentation or entrapped within the

emergent macrophyte stem matrix and attached (sorbed)

onto biofilms (Figure 3-7). Soluble phosphate may be

sorbed onto plant biofilms in the water column, onto biofilms

in the floating plant litter, or onto the wetland sediments.

The exchange of soluble phosphate between sediment

pore water and the overlying water column by diffusion

and sorption/desorption processes is a major pathway for

soluble phosphates in wetlands. In the sediment pore water, these phosphates may be precipitated as the insoluble

ferric, calcium, and aluminum phosphates or adsorbed onto

clay particles, organic peat, and ferric and aluminum oxides and hydroxides. The precipitation as calcium phosphates occurs at pH values above 7 and may occur within

the sediment pore water or in the water column near active phytoplankton growth where pH values may rise well

above 7. The sorption of phosphorus on clays involves

both the chemical bonding of the negatively charged phosphates with positively charged clay and the substitution of

phosphates for silicates in the clay matrix (Stumm and

Morgan, 1970). Phosphate can be released (desorbed)

from the metal complexes depending on the redox potential of the sediment. Under anoxic conditions, for example,

the ferric compound is reduced to the more soluble ferrous compound and phosphate is released. Phosphates

may also be released from ferric and aluminum phosphates

under anoxic conditions by hydrolysis. Phosphate sorbed

to clays and hydrous oxides may also be resolubilized

through the exchange of anions. The release of phosphate

from insoluble salts will also occur if the pH decreases as

a result of the biological formation of organic acids, nitrates,

or sulfates. Over time, however, a significant fraction of

the initially removed phosphate will become bound within

the sediments and lost to the system. At the start-up of a

FWS system (possibly for more than one year), the phosphorus removal will be abnormally high owing to the initial

reactions with the soils of the wetland. This removal mechanism is finite and essentially disappears after this period.

3.5.2.2 Biological Transformations of

Phosphates

Dissolved organic phosphate and insoluble inorganic and

organic phosphate are not usually available to plants until

transformed to a soluble inorganic form. These transformations may take place in the water column by way of

suspended microbes and in the biofilms on the emergent

plant surfaces and in the sediments. Uptake of phosphates

by microorganisms, including bacteria, algae, and duckweed, acts as a short-term, rapid-cycling mechanism for

soluble and insoluble forms. Cycling through the growth,

death, and decomposition process returns most of the

phosphate back into the water column. Some phosphate

is lost in the process due to long-term accretion in newly

formed sediments. Uptake by the macrophytes occurs in

the sediment pore water by the plant root system. The

estimate of net annual phosphorus uptake by emergent

wetland species varies from 1.8 to 18 g P/m2

/y (Burgoon

et al., 1991). The cycle of uptake and release is similar to

that of the microbes, but these reactions operate over a

longer time scale of months to years. Uptake occurs during the growth phase of the plant and release occurs during plant senescence and death in the late summer and

fall, followed by decomposition in the plant litter. Again,

some phosphate is lost to the system through accretion

processes within the sediments.

3.5.2.3 FWS System Reactions—Phosphate

Sustainable phosphate removal in a FWS system involves the accretion and burial of phosphate within the
58

wetland sediments (Figure 3-7). Phosphate cycling and

storage involves a complex set of processes with a number of forms of phosphate. Insoluble organic and inorganic

phosphates are settled or captured on solid surfaces within

the water column or in the wetland sediments or floating

litter. Many of these insoluble forms may be chemically

and biologically transformed to available forms of phosphate for uptake by macrophytes, epiphytes, and floating

plants. The phosphate taken up by these biological systems is recycled back into the system and is subsequently

available for other organisms or may leave the system

through the water column. The undecomposed portion of

the biological growths may accumulate and thereby be

removed from the system as accreted sediment. This material along with any recalcitrant phosphate separated from

the water column and accumulated within the sediment

represents the total net phosphate removal from the system.

Phosphate removal in FWS systems follows a seasonal

pattern in most temperate climate conditions. The form of

phosphate, the type and density of the aquatic plants, the

phosphate loading rate, and the climate determine the

pattern and amount of phosphate removed from the wetland over any given time period. Aquatic plants serve as

the seasonal reservoir for phosphate as they take up

soluble reactive phosphate (SRP) during the growing season. There is a finite amount of SRP that can be incorporated in the aquatic plants, epiphytes, and plankton in the

water column. In climates where senescing of plants occurs in the fall, the majority of the phosphorus taken up will

be released back into the water column (Figure 3-8). In

this figure, during the second year Marsh 1 received tap

water without phosphorus, whereas Marsh 3 received

wastewater at a loading rate of 0.15 kg P/ha/d throughout.

Release in excess of influent phosphorus is noted in the

Marsh 3 effluent in the fall. The effluent phosphorus exhibited for Marsh 1 is the result of release from the standing

crop developed the year before. Note a background residual phosphorus concentration of about 0.5 mg/L SRP

for this system. Maximum removal of phosphate was found

to be about 1.5 mg/L at a loading of less than 1.5 kg PO4/

ha/d and reduced to negligible (<0.2 mg/L) at loadings

above 5 (Gearheart, 1993). This maximum is consistent

with the theory of Stumm (1975), which suggested approximately 20% of influent P could be removed under an equilibrium condition in a lake.

3.5.3 Phosphorus in Vegetative

Submerged Beds

The removal of phosphate from VSB systems relies on

accretion of phosphorus from decomposing plants and from

separated, recalcitrant phosphate from the influent to the

process. Phosphate loading to these systems is large relative to plant uptake, and reliable sustained removal by harvesting of plants prior to senescence would not provide

significant removal. Accretion of partially decomposed plant

tissue may provide some additional removal of phosphorus. Cycling of phosphorus will produce seasonal effluent

variations similar to those seen for FWS systems. Some

minerals associated with the media can provide tempoFigure 3-7. Phosphorus cycling in a FWS wetland (adapted from Twinch and Ashton, 1983)

PO4

PO4

PO4

Particulate

P

Soluble

Organic P

Particulate

Mobile P

Air

Water

Floating

Plants

Macrophytes

Emergent

Submerged

Roots

Epiphytes

Epiphytes

Flow Outflow

Bound P

Sediment

Litter/Detritus/

Microbial Compartment
59

rary removal by means of precipitation/exchange/sorption

mechanisms, but these effects would be short term (<1

year) and dependent on the source of the granular materials.

3.6 Mechanisms of Pathogen Separations

and Transformations

3.6.1 Description and Measurement

Waterborne pathogens including helminthes, protozoans,

fungi, bacteria, and viruses are of great concern in assessing water quality. Since routine examination for pathogenic

organisms is not recommended because of cost and the

low numbers of a specific pathogen present at any given

time, indicator organisms are used. The most common indicators of the level of waterborne pathogen contamination in water are the coliform group. Today, the fecal coliform

test is considered to be a better indicator of human fecal

contamination than the more general total coliform procedure. Even so, the fecal coliform test is not specific and

can produce false positive results for human contamination, since these organisms are excreted by a number of

warm-blooded animals including those residing in wetland

environments. Fecal streptococci analysis may also be

used as an additional indicator of fecal pollution. Together

with the fecal coliform test, these two procedures are sometimes used to discriminate between human and other warmblooded animals. Table 3-1 indicates typical ranges of the

indicator organisms in typical influents to treatment wetland systems.

Separation of pathogens (and indicators) from the water

column does not in itself mean that the organisms are no

longer viable. They may be released from the matrix to

which they are attached and become available again in

the water column as infectious agents. The true removal

of pathogens is only achieved by rendering them nonviable.

3.6.2 Fate in Constructed Wetlands

Pathogens (and indicators) entering wetlands may be

incorporated within the TSS or may be found as suspensions in the influent wastewater. Those associated with

TSS would be separated from the water column by the

same mechanisms as discussed for TSS (sedimentation,

interception, and sorption). Once separated, the viable

organisms may be released from the solid matrix and be

retained within the biofilm or sediment pore-water, or they

may be reentrained into the water column. Regardless of

their location, they must compete with the consortium of

organisms surrounding them. As intestinal organisms, they

will normally require a rich substrate and high temperature

to favorably compete. Most will not survive in this competition. They will also be destroyed by predation or, if near

the open water surface, by UV irradiation.

Removal of pathogens (indicators) in wetlands appears

to be correlated with TSS removal and hydraulic residence

times (Gearheart et al., 1999; Gersberg et al., 1989). Few

studies have been performed on the effect of wetlands on

specific pathogens, but Gearheart has found similar removals to FC with Salmonella and MS2 coliphage. Many

pathogens are more sensitive to the wetland environment

than indicators, but some viruses and protozoans (spores)

may be more resistant. Erratic results have been reported

on viruses, and mechanisms affecting their removal may

be different than those that destroy indicators.

Figure 3-8. Phosphorus pulsing in pilot cells in Arcata; Marsh 1 received tap water until June 1982 (no phosphorus load), while Marsh 3 received

oxidation pond effluent (Gearheart, 1993)

12/1/81 3/2/82 6/2/82 9/1/82

Date

Phosphorus (mg/L)

0

1

2

3

4

5

6

7

8

9

10

Influent

Marsh 1

Marsh 3
60

It is significant to note that indicator organisms and perhaps pathogens may be generated within the wetland. Thus

background levels of indicators will be found even in natural wetland systems (see Table 3-5). These background

levels are variable, influenced by season and other operational parameters of the system (Figure 3-9). It should be

noted that in general these indicator organisms are not

from human sources. However, constructed wetlands are

unlikely to consistently meet stringent effluent fecal coliform

permit levels. Therefore, regulators may require disinfection of treatment wetland effluents prior to discharge.

Gearheart (1998) consistently attained an FC count of less

than 2/100 mL with UV disinfection of FWS effluent.

3.7 Mechanisms of Other Contaminant

Separations and Transformations

3.7.1 Metals

While some metals are required for plant and animal

growth in trace quantities (barium, beryllium, boron, chromium, cobalt, copper, iodine, iron, magnesium, manganese, molybdenum, nickel, selenium, sulfur, and zinc),

these same metals may be toxic at higher concentrations.

Other metals have no known biological role and may be

toxic at even very low concentrations (e.g., arsenic, cadmium, lead, mercury, and silver) (Gersberg et al., 1984;

Crites et al., 1997). Influent wastewater to wetlands may

carry metals as soluble or insoluble species.

Metals entering wetlands as insoluble suspended solids

are separated from the water column in a manner similar

to TSS. Depending on pH and redox potential, these insoluble species may be resolubilized and returned to the

liquid phase. Important removal mechanisms for metals

include cation exchange and chelation with wetland soils

and sediments, binding with humic materials, precipitation

as insoluble salts of sulfides, carbonates, and

oxyhydroxides, and uptake by plants, algae, and bacteria.

The chemically bound metals may eventually become buried in the anoxic sediments where sulfides occur. These

bound metals are often not bio-available and remain removed from the system. If sediments are disturbed or resuspended and moved to oxic regions of the wetland, sequestered metals may resolubilize.

Metals may be incorporated into the wetland biomass

by way of the primary production process. For macrophytes, metals are taken up through the root system and

distributed through the plant. The extent of uptake is dependent on the metal species and plant type. Gersberg et

al. (1984), found only minor uptake by plants in VSB systems, while others claim that metals can be found on root

surfaces due to precipitation and adsorption. The accumulation of heavy metals was found to be variable in a

marsh in New Jersey receiving wastewater (Simpson et

al., 1981; 1983). Cadmium, copper, lead, nickel, and zinc

had accumulated in the litter at the end of the growing season in much higher concentrations than in the live vegetation. Other studies have shown that metals like cadmium,

chromium, copper, lead, mercury, nickel, and zinc can be

sequestered by wetland soils and biota or both (Mitsch

and Gosselink, 1993). The high uptake of selenium by biota

in a wetland marsh receiving irrigation waters was discussed in Hammer (1992), but some could have been volatilized. Studies have shown that some algae will sequester selected metals (Kadlec and Knight, 1996). Floating

plants such as duckweed have been shown to be excellent accumulators of cadmium, copper, and selenium, but

only moderate accumulators for chromium and poor accumulators for nickel and lead (Zayed et al., 1998). A review

of metal removal in wetlands is found in Kadlec and Knight

(1996).

Figure 3-9. Influent versus effluent FC for the TADB systems (EPA, 1999)

10

1 10 100 1,000 10,000 100,000 1,000,000

100

1,000

10,000

Influent FC (number/100 mL)

Effluent FC (number/100 mL)
61

At the present time there is insufficient long-term data

on full-scale constructed wetlands to provide a reliable

estimate of performance on the removal of metals from

wastewater. However, in VSBs and in fully vegetated FWS

systems, the anaerobic conditions are conducive to retaining most metals with the settled TSS and minimizing

resolubilization. Similarly, the actual removals will be affected by the speciation of the influent metals.

3.7.2 Other Organic Compounds

There is concern about the fate of many trace organic

compounds in the environment. These include pesticides,

fertilizers, process chemicals, and others that fall under

the category of priority pollutants. The fate of these compounds in wetlands is dependent on the properties of the

compound, the characteristics of the wetland, the species

of plants, and other environmental factors. The most important separation and transformation mechanisms involved include volatilization, sedimentation/interception,

biodegradation, adsorption, and uptake. These mechanisms have been discussed previously. Recalcitrant organics that have been separated may accumulate in the wetland sediments. Some may be taken up by plants and be

returned to the system upon plant decomposition. Biodegradation of some organic compounds may result in completely mineralized end products, or the process may produce end products that may be more toxic than the parent

compound. At this time, there is insufficient data available

on full-scale wetland systems to evaluate how effective

they are in the long-term removal and destruction of most

priority pollutants. Based on pretreatment performance,

oxidation or facultative lagoons remove a high %age of

volatile and semivolatile organic compounds (Hannah et

al., 1986), resulting in low influent concentrations to the

FWS system that follows, while primary sedimentation is

less effective and results in higher influent concentrations

of both to subsequent VSB systems.

3.8 Constructed Wetland Modeling

3.8.1 Modeling Concepts

The modeling of wastewater treatment operations and

processes has long been of interest to environmental engineers. This interest stems primarily from a need to quantify the performance of the process and a desire to optimize the design and operation of the treatment facility.

Modeling of many of the treatment processes used today

has met with only partial success primarily because of the

lack of rigor in most models. This is due to the enormous

complexity of the reaction mechanisms that may take place

within many of these systems and with the difficulty in characterizing the constituents within the wastewater. Constructed wetlands fall in this category of a highly complex

system in which a multiplicity of reactions and reaction

mechanisms occur, even in the simplest of systems. Adsorption, sedimentation, flocculation, biological catalysis,

precipitation, exchange, and diffusional processes are but

a few of the important functional mechanisms that may

control the removal of a given constituent. Furthermore,

these mechanisms are dependent on a number of physical, chemical, and biological variables within the system

(e.g., temperature, redox potential, pH, plant density, etc.).

It is apparent that in a highly complex system such as

the constructed wetland, the quantification of all specific

rate-controlling mechanisms seems unlikely. The transient

nature of the influent wastewater characteristics and the

lack of substantial control over the process undoubtedly

will result in frequent changes of the rate-controlling mechanisms of the process.

In developing a model for any process, the first question

that should be asked is, “What is the value of modeling

this particular system?” Currently, the design engineer may

be using empirical models that assist in interpolating information from lab-scale or pot-scale studies. The major problem is scale-up, and most of the designer’s time is spent

on concerns about operating conditions within rather narrow ranges in order to meet permit requirements. Added

to that is the ever-present shortcoming of analytical and

sampling methods. The answer to the question would appear to be that there is little value in developing more empirical curve-fitting models for the process. The design

engineer might be well advised at the present time to develop performance parameters, curves, and operational

charts that could be used for the plants under investigation. If a deterministic model of mechanistic proportions is

to be developed, it needs to be developed through a series of rigorous processes ensuring that it is, in fact, a true

model of the process.

Writing a mechanistic, mathematical model is generally

easier than verifying it. That is not to say that conceiving

and writing the model is trivial, but rather to emphasize the

difficulty in giving the model a fair chance to fail the experimental test. Note that the emphasis is on a chance for

verification failure rather than a chance to pass. Thus an

investigator’s problem of ownership of a particular model

presents itself, which is not easily subjugated to the scientific issue of fairly and rigorously testing the model.

The experimental learning process is basically iterative

and consists of successive and repeated use of the sequence. Box and Hunter (1965), although not the first to

note this important underlying pattern in experimentation,

have best exploited the pattern and developed strategies

for experimentation that efficiently lead one through the

cycle and advise one in determining the details of subsequent cycles. One such cycle is illustrated in Figure 3-10.

In environmental engineering, a field that relies heavily on

empirical methods, the unavoidable long-term character

of the experiments required for a system, such as a constructed wetland, has resulted in many different individuals conducting rather inefficient iterations of different parts

of the problem. As a result, the development of a rigorous

model of the process, or parts of it, has not been achieved

as is true with many of the wastewater treatment processes

designed today.

Of particular importance in the iterative sequence are

the steps of design and analysis. In this context, design is
62

devising experiments suggested by the current status of

the process that will most likely improve an understanding

of it. Analysis would be the examination of the data in such

a manner as to discover how far conjectures are born out

and spark modified or new conjectures when the old ones

are found wanting. Both engineering and statistics play an

important role in the process.

3.8.2 Status of Wetland Modeling

What is the status of model building for constructed wetlands at this time? There are a number of ways one might

classify models. One classification would include linguistic models, stochastic models, and deterministic models.

Linguistic, or word, models are qualitative and have as their

main advantage the capability of representing incomplete

states of knowledge not explicitly represented by the other

two types of models. Stochastic models of processes are

the result of data-driven, top-down approaches to modeling and can reflect information contained in the data used

to prepare them. These models are useful but are data

intensive and require that the data be representative of

the behavior of the process that they are designed to model.

Deterministic models, both empirical and mechanistic, are

either built from the bottom up based on first principles or

result from careful observation of phenomena. They are

robust and general, and are effective provided that the data

are available to support the theory and to calibrate the

model.

An examination of the literature on constructed wetlands

suggests that all three of these models have been proposed or used. Yet it is worth noting that none have really

met the test of the iterative process. There are very few

full-scale studies of constructed wetlands that have provided a database sufficient to demonstrate success or failure of a given model. Most studies lack sufficient spatial

and temporal sampling to even identify whether a model

fails the experimental test of verification. In an effort to provide more data, investigators use databases from a number of sites. Site-to-site variability makes this process suspect. Quality control on the data collected is also of concern. For the most part, the data fit to constructed wetlands is regressed from empirical models, and fit is expressed in terms of the coefficient of determination, R2

. It

is important to note that a high value of R2 does not assure

a valid relation nor does a low value mean that the model

is useless. It is often stated that R2 explains a certain proportion of the variability in the observed response. If the

data were from a well-designed controlled experiment, with

proper replication and randomization, it is reasonable to

infer that a significant association of the variation in y with

variation in the level of x is a causal effect of x. If, however,

as in the case of most wetland data, the data are observational, there is a high risk of a causal relationship being

wrong. With observational data, there can be many reasons for associations among variables, only one of which

exhibits causality. Totally spurious correlations, often with

Figure 3-10. Adaptive model building

Model

Reliable and Pertinent

Data

Parametric Estimates

Quality of Fit

Adequacy of Model

Yes No

Experimentation
63

high R2 values, can arise when unrelated variables are

combined.

Often rival models may be tentatively considered for a

constructed wetland process. It is not uncommon to find

that more than one model can be calibrated to fit the data

and give residual errors that are acceptable. In some cases,

the selection of the seemingly acceptable model is of little

practical consequence over the range of interest. In other

cases, knowing which model is better may throw light on

fundamental questions about reaction mechanisms and

other phenomena under investigation. A fundamental concept in model discrimination is that rival models often diverge noticeably only at extreme conditions. Thus extremes

must be evaluated to provide useful statistical information.

Models must be put into jeopardy of failing. An excellent

discussion of this may be found in Berthouex and Brown

(1994).

In summary, it appears premature, given the lack of quality-assured data, for designers of constructed wetlands to

rely on regressed empirical models. The problem is one of

extrapolation of these mostly “black box” models from site

to site and extension of the model outside of the database.

Examination of the wide variation in parameter values (reaction coefficients, background concentrations, etc.) suggests that fitting data to simplistic models may be insufficient at this time to provide reliable design information in

most cases. Currently the design of these systems should

be based on design parameters (e.g., hydraulic loading,

nitrogen loading, detention time, etc.) and operating criteria that are required to meet a specific effluent limitation.

This is not a novel approach, but has been used for many

decades by environmental engineers in the design of highly

complex unit processes including the activated sludge process and waste stabilization ponds. Only when a carefully

designed series of iterative studies have been conducted,

and data based on quality-controlled specifications have

been analyzed, can rigorous models be provided for use

in wetland system design.

Currently there is a North American Database (NADB)

on wetlands that has been relied on for purposes of wetland design. Efforts have been made to refine that database to improve reliability. What is urgently needed at this

time is an effective plan for the design of studies that will

provide a comprehensive understanding of the processes

that occur within wastewater wetlands. Efforts have already

been made to mathematically model some of the wetland

processes based on first principles. These models should

serve as the starting point for the adaptive iterative process as described in Figure 3-10. The experimental design should include extensive, quality-assured, transect

data at numerous selected sites (spatial variations) over

an extended period of time (temporal variations). Characterization of the wastewater must include both particulate

and soluble fractions of contaminants and must ensure

quality control of both sampling and analyses. Characterization of the wetland is also important and should include

data on residence time distribution of flow, geometry, plant

species and distribution, monolithic zone coverage and

distribution, and so forth. Once this database is developed,

the iterative modeling of this very complex system can

begin in earnest. Given the number of years of constructed

wetland experience, such efforts are generally overdue.

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Reddy, K.R. and D.A. Graetz. 1988. Carbon and nitrogen

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The ecology and management of wetlands. Vol. I. Portland OR: Timber Press, p. 307.

Reed, S.C., R.W. Crites, and E.J. Middlebrook. 1995. Natural systems for waste management and treatment. 2d

ed. New York, NY: McGraw-Hill.

Rickert, D.A. and J.V. Hunter. 1972. Colloidal matter in

wastewater secondary effluents. J. Water Pollution

Control Federation, 44(1):134.

Rose, C. and W.G. Crumpton. 1996. Effects of emergent

macrophytes and dissolved oxygen dynamics in a prairie pothole wetland. Wetlands, 16(4):495.

Simpson R. L., et al. 1981. Dynamics of nitrogen, phosphorus, and heavy metals in Delaware River freshwater tidal wetland. U.S. Environmental Protection

Agency, Final Report, Corvallis Environmental Research Lab, Corvallis, OR.
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Simpson, R.L., et al. 1983. The ecology of freshwater

tidal wetlands. Bioscience, 33:225.

Standard methods for the examination of water and

wastewater. 1998. 20th ed. APHA, AWWA, WEF.

Stumm, W. 1975. Man’s acceleration of

hydrogeochemical cycling of phosphorus: Eutrophication of inland and coastal waters. Water Pollution

Control (GB), pp. 124–133.

Stumm, W. and J.J. Morgan. 1970. Aquatic chemistry:

An introduction emphasizing chemical equilibria in

natural waters. New York, NY: John Wiley & Sons,

p. 583.

Twinch, A.J. and P.J. Ashton. 1983. Wetlands as

accreting systems: Nutrient cycling in wetlands.

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U.S. Environmental Protection Agency (EPA). 1978. Design manual—Onsite wastewater treatment and disposal systems. EPA 625/1-80-012. Office of Research and Development, Municipal Environmental

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Wastewater stabilization lagoons intermittent sand

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U.S. Environmental Protection Agency (EPA). 1983.

Wastewater stabilization ponds—Nitrogen removal.

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DC.

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Van Oostrom, A.J. and R.N. Cooper. 1990. Meat processing effluent treatment in surface flow and gravel

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U.S. EPA.

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66

There are many design factors which effect the effluent

quality from a free water surface (FWS) constructed wetland. Consideration of some of these factors can significantly reduce the effluent variation. Site specific factors

which need to be considered in the design of wetlands

include topography, climate (growing season, temperature

variation, evapotranspiration and precipitation), wastewater characteristics, flows and loads, and wildlife activity

(Gearheart and Finney, 1996). Design variables include

total area, the number, size, depth, and shape of wetland

cells, hydraulic retention time, vegetation types and coverage, inlet and outlet type and location, and internal flow

patterns.

This chapter provides a brief summary of expected wetland treatment performance based on the most reliable

data from existing operating wetlands, describes issues

that are important in the design and layout of a FWS constructed wetland, and provides an overall strategy for sizing a constructed wetland system. Several design examples are included, and construction issues unique to

FWS wetland systems are discussed.

4.1 Performance Expectations

4.1.1 Introduction

The performance of operating FWS constructed wetland

treatment systems reveals the range of effluent quality and

the variability of performance possible with these types of

systems. Evident in this analysis is the great variety in the

range of treatment capacities, and thus the loadings, to

which FWS treatment wetlands are subjected. Given the

wide variety of settings, design criteria, configuration, and

inlet/outlet (I/O) placement, considerable variation in performance should be expected. The largest single compilation of FWS wetland operating data is the North American

Database (NADB, 1993). Unfortunately, much of the data

within the NADB is difficult to use for performance evaluation because of the absence of hydraulic description of the

system’s influent and effluent flows. Another major constraint is the lack of intra-cellular and individual cell influent and effluent water quality data. The 40 systems included in the detailed portion of this performance evaluation are known as the Technology Assessment (TA) sites

(USEPA, 1999). For purposes of this design manual only

those sites treating oxidation pond effluent and primary

treated effluent are considered. There are 22 sites in this

Chapter 4

Free Water Surface Wetlands

category (designated DM sites), which are used to assess

loading versus effluent concentration. The data sets employed in this analysis are from Arcata, Gustine and Manila, CA, Beaumont, TX, Brookhaven, NY, Columbia, MO,

Fort Deposit, AL, Mount Angel, OR, Ouray, CO, West Jackson County, MS, and Listowel, Ontario, Canada.

The brief performance summary included here offers an

overview of loading, influent and effluent ranges for a number of water quality constituents, and presents the observed

relationship between loading rates and effluent quality for

all systems. A more broad performance review, including

cumulative effluent probability plots for the 40 TA sites is

provided in FWS Wetlands for Wastewater Treatment: A

Technology Assessment (USEPA, 1999). The focus of this

manual for FWS application in a strictly non-tertiary treatment mode resulted in the need to cull out those tertiary

applications from the TADB to produce the (DMDB) database which is used herein. In all cases, after the DMDB is

used to analyze the information, the TA database is evaluated along with other mechanistically viable, controlled

studies and/or mechanistic analyses to produce the most

reasonable engineering interpretations of the data. The

additional quality-assured data which are needed to advance the understanding of the free water surface constructed wetlands are described at several locations in the

text of this chapter.

4.1.2 Range of Operating Conditions and

Performance

Table 4-1 shows the mean and range of loadings for

each water quality constituent for the 22 DM systems included in this analysis. Preliminary treatment for the 22

DM systems includes 21 lagoon systems and one primary

effluent system. This first level of assessment is useful only

in the context of summarizing the range of performance

and operating conditions of these FWS constructed wetlands and their range of response in terms of effluent concentration. This level of analysis summarizes the wide

range of applications and expected performance for operating FWS treatment wetlands. No accounting for geometric

configuration, planting strategy, inlet/outlet works, climate,

etc., is made at this time.

The data in Table 4-1 are a summary of the treatment

effectiveness, and are not meant to suggest a mass bal-
67

ance of constituents, e.g., not all systems reported all constituents.

4.1.3 BOD Performance

The relationship between average biochemical oxygen

demand (BOD) loading and average BOD effluent concentration for the studied systems is shown in Figure 4-1.

. There is a general trend between increased BOD loading

and increased effluent concentration up to the highest loading of 183 kg/ha-day if only the fully vegetated designs are

taken into account. The figure reveals considerable effluent variation for a given BOD loading and shows considerable variation in effluent quality at the lower BOD loading

rates. The effect of the background BOD, due to release

from previously settled influent TSS and plant decomposition is especially evident in systems with low loading rates.

Figure 4-2 illustrates the internal BOD load which occurs

from the partial anaerobic digestion of previously settled

organic solids and plant exudates and other byproducts of

anaerobic biodegradation. The portion of that internal loading which is due to the plant detritus has been measured

and is illustrated in Figure 4-3. Since anaerobic processes

are extremely sensitive to temperature, these internal loadings begin in the spring as water temperature rises and

continue until the backlog of settleable organics and plant

detritus accumulated over the winter is exhausted. Plant

exudates and senescent byproducts occur in their own

cycles which provide internal loading to the system in a

dynamic manner. Background BOD concentrations can

range up to almost 10 mg/L.

A more conservative analysis of Figure 4-1 indicates that

a fully vegetated FWS should not be loaded above 40 kg

BOD/ha-d if a secondary effluent BOD standard (30 mg/L)

is to be met. Restricting the analysis to open water FWS

systems indicates that loadings up to 45 kg/ha-d yielded

effluent BOD of less than 20 mg/L, and loadings up to 130

kg/ha-d always met this quality. However, there is only one

data point above 45 kg/ha-d and it is at 130 kg/ha-d. The

open water FWS systems which permit reaeration and

aerobic oxidation should be designed for an areal loading

of no more than 60 kg/ha-d to consistently attain an effluTable 4-1. Loading and Performance Data for Systems Analyzed in this Document (DMDB).

Pollutant Loading

Rate (kg/ha-day) Influent (mg/L) Effluent (mg/L)

Constituent Min Mean Max Min Mean Max Min Mean Max

BOD5 2.3 51 183 6.2 113 438 5.8 22 70

TSS 5 41 180 12.7 112 587 5.3 20 39

NH4

-N 0.3 5.8 16 3.2 13.4 30 0.7 12 23

TKN 1.0 9.5 20 8.7 28.3 51 3.9 19 32

TP ——— 0.56 1.39 2.41 0.68 2.42 3.60

FC ——— 42000 73000 250000 112 403 713

BOD = Biochemical Oxygen Demand (5 day)

TSS = Total Suspended Solids

NH4-N = Ammonia Nitrogen

TKN = Total Kjeldahl Nitrogen

TP = Total Phosphorus

FC = Fecal Coliform, cfu/100mL

ent BOD of <30 mg/L until more performance data are obtained. Coincidently, Stowell (1988) recommended an upper limit of 60 to 70 kg BOD/ha-d to prevent odors from an

FWS system. Open water FWS systems loaded below 45

kg/ha-d can be expected to attain effluent BODs of 20 mg/

L or less. Analysis of the TADB yields a similar maximum

areal BOD loading rate of 50 kg/ha-d without differentiation between fully vegetated and open water FW systems

(EPA, 1999).

4.1.4 TSS Performance

The effectiveness of FWS treatment wetlands to remove

total suspended solids (TSS) is recognized as one of their

principal advantages. Over a range of loadings from 5 to

180 kg/ha-day, there are several relationships between

loading and effluent TSS quality with the DM data, as shown

in Figure 4-4. Under a fairly narrow range of solids loadings, (up to 30 kg/ha-d) secondary effluent TSS concentrations (≤ 30 mg/L) can be attained with fully vegetated

systems. Since physical processes dominate the removal

of TSS, similar designs should produce similar effluent

qualities. Analysis of the TADB (EPA, 1999) yields similar

maximum loading.

TSS removal is most pronounced in the inlet region of a

FWS constructed wetland. Generally, the influent TSS from

oxidation pond systems are removed in the first 2 to 3 days

of the nominal hydraulic retention time in fully vegetated

zones near the inlet (Gearheart,et al, 1989; Reed, et al,

1995; Kadlec and Knight, 1996). Enhanced settling and

flocculation processes account for most of this removal,

and the overall removal efficiency is a function of the terminal settling velocity of the influent and flocculated solids. Long-term removal of detrital material will likely be required 10-15 years into operation. The separated solids

undergo anaerobic decomposition, releasing soluble dissolved organic compounds and gaseous by-products, carbon dioxide and methane gas, to the water column. Figure

4-2 shows the reduction of total and soluble BOD and TSS

through a pilot project wetland cell. Approximately 80% of

the TSS is removed in the first two days of theoretical HRT

primarily due to enhanced sedimentation and flocculation.
68

80

70

60

50

40

30

20

10

0

0 50 100 150 200

BOD Load (kg/ha•day)

Effluent BOD (mg/L)

V = Fully Vegetated

O = Significant Open Area O V

O

VV

V

V

V

O O

O OO O

V

V

V

V

V

V

Figure 4-1. BOD effluent vs. areal loading (DMDB)

60

50

40

30

20

10

0

01 4 2

HRT (days)

0 3 5

BOD and TSS (mg/L)

Total BOD (mg/L)

TSS (mg/L)

Soluble BOD (mg/L)

Figure 4-2. Release of soluble BOD during early stage
69

5

4

3

2

1

0

0 20 40 60 80 100

BOD (mg/L)

Typha-annual BOD (mg/L)

Scirpus-annual-BOD (mg/L)

Percent Dry Weight of Standing Crop to

Water Column

Figure 4-3. Annual detritus BOD load Scripus and Typha - assuming 15,000 kg/Ha standing crop (Gearheart & Finney, 1996)

50

40

30

20

10

0

0 50 100 150 200

TSS Load (kg/ha•day)

Effluent TSS (mg/L)

V = Fully Vegetated

O = Significant Open Area

O

V

O

V

V V V

O

O

OO

O

V

V

V V

V

V

V

Figure 4-4. TSS loading vs. TSS in effluent (DMDB)
70

Subsequently, the removal essentially ceases without subsequent open zones which can provide conditioning and

transformation processes which may improve overall removal of TSS attainable by the system.

A closer analysis of the DMDB again shows that TSS

loadings can be higher for FWS systems with open-water

zones. Only one very small site with such zones exceeded

secondary effluent TSS standards (30 mg/L), and it was

loaded at more than 90 kg/ha-d. Below a loading of 30 kg/

ha-d an effluent of 20 mg/L of TSS was consistently achievable. It is therefore recommended that in addition to that

areal loading limitation a maximum loading of 50 kg/ha-d

be employed to attain an effluent of 30 mg/L of TSS until

more performance data can be obtained.

4.1.5 Nitrogen Performance

Any discussion of nitrogen species in FWS constructed

wetlands must be predicated by a return to first principles,

as described in Chapter 3. Given the numerous transformation possibilities and the dearth of removal mechanisms,

there are only a few meaningful explanations. Influent nitrogen for the typical applications of this manual will be

primarily in the form of ammonia-nitrogen (NH4-N) with a

significant organic nitrogen (ON) contribution. Approximately 10 to 15% of the oxidation pond effluent TSS due

to algae is organic nitrogen (Balmer and Vik, 1978). Since

both are measured by the total Kjeldahl nitrogen (TKN)

test, it becomes the likely standard of areal loading analysis. Any discussion of just one species (typically, NH4-N)

is of little value and often misleading to readers.

Another key discussion point is the inherent inability of

fully vegetated FWS systems to nitrify a typical FWS influent, as described in the preceding paragraph, within a practical number of days of HRT. During periods of senescence

when fully vegetated zones become partially open-water

zones, different mechanisms of treatment can replace

these which dominate during the normal growing season,

if the climate can sustain them. This is further reinforcement for the fact that there are few absolutes in natural

systems. Sadly, these conditions are rarely recognized or

adequately described and measured, making use of most

existing data sets open to some question. However, in this

analysis the fact that a system is classified as either fully

vegetated or as having significent open-water zones aids

in explaining many anomolies.

4.1.5.1 TKN Performance

Figure 4-5 illustrates that for a fully vegetated FWS which

receives 30 to 50 mg/L of TKN the effluent exceeds 24

mg/L since the only removal mechanism is sedimentation

35

30

25

20

10

5

15

0

0 5 10 15 20 25

TKN Load (kg/ha•day)

Effluent TKN (mg/L)

V = Fully Vegetated

O = Significant Open Area

V

OOO

V

V

V V

V

V

Figure 4-5. Effluent TKN vs. TKN loading
71

of organic nitrogen (ON). One more lightly loaded (3.3 kg/

ha-d) fully vegetated system did produce an effluent TKN

of about 9 mg/L. The three open-water zone systems were

all lightly loaded (< 2.8 kg/ha-d) and produced effluent TKN

of about 4 mg/L. Since these latter designs offer a mechanism to transform the TKN to nitrate-nitrogen (NO3 -N),

which could subsequently be removed through denitrification, they could conceivably be loaded more heavily and

still meet a stringent total nitrogen standard (e.g., 10 mg/

L). No more heavily loaded systems were indicated in the

DMDB. The TADB analysis (EPA, 1999) did indicate that

any FWS system which received a TKN loading of less

than 3.3 kg/ha-d could meet an effluent TKN of less than

10 mg/L, but does not subdivide results into the two

subcatagories used herein. Arcata’s open-water-zone systems have been able to maintain effluent total nitrogen (TN)

below 5 mg/L at loadings of up to 3 kg TN/ha-d (Gearheart,

1995) through the denitrification provided near the FWS

system outlet which is fully vegetated. Maximum TKN loadings to sustain an effluent TKN of less than 10 mg/L can

conservatively be set at 5 kg/ha-d until more data are made

available. This only applies to open water FWS systems,

while fully vegetated systems are limited to only a small

percentage of TKN removal due to the settling of organic

nitrogen particulate matter.

4.1.5.2 Denitrification

The extent of nitrate removal via denitrification is dependent on the extent of the prior conversion of TKN to

NO3-N, a labile carbon or other energy source and anaerobic/anoxic conditions in the water column. Therefore, denitrification, which converts NO3-N to gaseous end products which can leave the constructed wetland, is best suited

to a fully vegetated condition. Further, any NO3-N which

enters a FWS wetland is likely to be quickly removed, while

any nitrate formation (nitrification), which occurs in the

open-water zones of the FWS can be removed near the

fully vegetated outlet zone if the conditions noted above

are met. The DMDB offers no assistance in that all the

systems which had nitrate-nitrogen in their influent had it

at very low concentrations (average = 2.47 mg/L), even

though effluent concentrations were lower (average = 2.22

mg/L).

Gearheart (1995) reports that the carbon produced from

decaying macrophytes is sufficient to denitrify 100 mg NO3-

N/L in an FWS and that the reaction rate is temperature

dependent, signifying its biological nature. Reed, et al

(1995) and Crites and Tchobanoglous (1998) also indicate

that FWS systems have the capablility to denitrify, but they

offer no specific examples. Therefore, denitrification should

be feasible in FWS systems as long as there is sufficient

detention time in fully vegetated zones with anaerobic/anoxic conditions.

4.1.5.3 Ammonia Nitrogen Performance

Also, as noted previously, ammonia-nitrogen (NH4-N)

limits are often specified for treatment facilities in their

NPDES permit. However, the level of effluent ammonia in

an FWS constructed wetland effluent bears only a tenuous relationship to its influent NH4-N concentration. The

normal case will find that all influent nitrogen is measured

as TKN, and that this total will be divided between organic

nitrogen and NH4-N. It is likely that this total nitrogen will

be reduced in any FWS owing to the loss of organic nitrogen due to enhanced settling. In the DMDB the average

TKN removal was 32%, while the fully vegetated systems

reported 28%. This difference is not larger because the

open-water FWS systems were few in number and were

all loaded more lightly. Although one can plot the effluent

NH4-N vs NH4-N loading from the DMDB, the data demonstrate no useful relationship. Therefore, unless a FWS

is designed for very low TKN loadings with an ample openwater zone to nitrify the influent, there is no meaningful

chance to meet any NH4-N effluent standard which is significantly lower (>30%) than the influent concentration.

4.1.5.4 Other Nitrogen Performance

Since total nitrogen is the sum of all forms of nitrogen, it

will be reduced through nitrification/denitrification, the loss

of organic nitrogen due to flocculation and sedimentation

and plant uptake of NH4-N. Since some of this settled fraction will return to the mainstream as the settled organics

partially digest and the plant uptake will return due to senescence, the total nitrogen budget should be evaluated

on an annual basis. The returned nitrogen will likely be in

the same two forms (organic and ammonia-nitrogen) as

the normal influent TN load, making this internal load very

compatible with the external load. Given the transformability

of individual nitrogen components between each other

based on the conditions existing at different locations in

the FWS wetland, the designer needs to provide passive

controls(e.g.,depth and vegetation patterns) if he or she

wishes to remove a substantial portion of the incoming nitrogen load.

4.1.6 Total Phosphorus Performance

Only 4 of the DMDB systems measured TP loadings and

effluent quality, as shown in Figure 4-6. While some approximate comparisons can be made, the need to separate the inorganic phosphorus performance from the organic particulate phosphorus performance makes the lack

of DM data impossible to utilize effectively, therefore, the

TA database (EPA,1999) is used for an approximate analysis.

Figure 4-7 shows that over a range of loading up to 4.5

kg/ha-day at the TADB sites, total phosphorus effluent concentration generally increased with loading (USEPA, 1999).

At the lower loading rates (<0.55 kg/ha-d), however, the

effluent phosphorus concentration was less than 1.5 mg/

L. The fractional distribution of TP in municipal wastewater previously treated by lagoons is variable and has not

been well documented. Balmer and Vik (1978) found filtered P/total P to be 20-25%, but the flocculation and sedimentation superiority of the initial FWS fully vegetated zone

will remove a significant % of those forms which enter as

supracolloidal and settleable solids. Gearheart (1993) has

performed extensive studies on the Arcata FWS systems

and found a relationship between areal loading and phos-
72

4

3.5

3

2.5

1.5

1

2

0.5

0 0.5 1 1.5 2 2.5 3

TP Load (kg/ha•day)

Effluent TP (mg/L)

Figure 4-6. Effluent TP vs. TP areal loading

5.0

4.5

4.0

3.5

3.0

2.5

2.0

1.5

1.0

0.5

0.0

0.0 1.0

Total P Loading (kg/ha•d)

2.0 3.0 4.0 5.0

Effluent Total P (mg/L)

Figure 4-7. Average total phosphorus loading rate vs. total phosphorus effluent concentration for TADB wetland systems
73

phorus removal. An upper limit of 1.5 mg/L removal of orthophosphates was found at loadings of less than 1.5 kg/

ha-d and a hydraulic retention time (HRT) of at least 15

days. HRTs of less than 7 days yielded a maximum removal of 0.7 mg/L of orthophosphate, as shown in Figure

4-8.

The phosphorus cycle of uptake during the growing season and release during senescence and the initial substantial uptake of phosphorous by the soil of the FWS further confounds short-time studies of phosphorus transformations and removal. As a final issue in this conundrum,

the particulate fractionation and chemical speciation of the

influent phosphorus will also have some impact on transformations and fate of P. Therefore, long-term studies which

differentiate between P forms and document vegetation

condition, climate, temperature and other pertinent water

quality parameters are necessary to provide meaningful

information for future FWS system designers, but the annual P removal by these systems is quite limited based on

mechanistic evaluations and controlled studies.

4.1.7 Fecal Coliform Performance

Only four systems in the DMDB reported fecal coliform

(FC) data, and three of those were fully vegetated systems. The average removal was just over 2 logs, from

72,700/100mL to 400/100ml. The TADB (EPA, 1999) results are shown in Figure 4-9 which offers no obvious relationship between inlet and outlet concentrations. Figure 4-

10 (Gearheart, 1989) shows results from the Arcata pilot

study which demonstrate that flocculation/sedimentation

is the primary FC removal mechanism for fully vegetated

cells or zones. This figure is especially valuable in light of

the world literature on FC dieoff in lagoons which identify

solar radiation as the primary disinfection mechanism

(Mara, 1975). Such a mechanism can be effective in openwater zones of the FWS, but the same mechanisms which

remove settleable and colloidal solids in fully vegetated

zones are responsible for FC removal in those zones.

Estimates of the internal addition of background fecal

coliform by wildlife in treatment wetlands are provided by

those systems that receive disinfected influent. For example, the Arcata Enhancement Wetland received chlorinated effluent, and during the period 1990-1997

(Gearheart,et al,1998), the effluent FC was less than 500

MPN/100mL about 80% of the time. This is a system with

large open-water zones that supports a wide variety and

high population of aquatic birds and mammals. Higher levels of background FC levels are found during the fall and

winter bird migration period. A similar study on the same

system during 1995-1996 showed that the effluent mean

was 40 CFU/100mL, was less than 300 CFU/100mL over

90 percent of the time, and on no occasion exceeded 500

CFU/100mL. Studies of MS-2 coliphage showed similar (2

logs) removal to that which was obtained with FC

(Gearheart, 1995).

The considerable temporal variability in the effluent microorganism counts produced by treatment wetlands and

conventional treatment technologies suggests the use of

geometric averaging to determine monthly mean values

from daily or weekly measurements. Even with geometric

means, individual monthly values are frequently 10 times

larger or smaller than the long-term mean for many treatment wetlands, possibly due to wildlife habitat features.

This implies that at sites which have strict FC restrictions,

the ability to disinfect the FWS effluent is required.

4.1.8 Metals & Other Particulate-Oriented

Pollutants

While some metals are required for plant and animal

growth in trace quantities (barium, boron, chromium, cobalt, copper, iodine, iron, magnesium, manganese, molybdenum, nickel, selenium, sulfur, and zinc), these same

metals may be toxic at higher concentrations. Other metals may be toxic at even very low concentrations (e.g. arsenic, cadmium, lead, mercury, and silver) (Gearheart,

1993).

Information from FWS treatment wetlands indicates that

a fraction of the incoming metal load will be trapped and

effectively removed through sequestration with settleable

suspended solids and soils. For many metals, the limited

data indicate that concentration reduction efficiency and

mass reduction efficiency correlate with TSS reduction.

Wetland background metal concentrations and internal

profiles are not well known. It has been shown that chromium levels higher than 0.1 mg/L and copper levels higher

than 1.0 mg/L have detrimental effects on a floating duckweed species (Lemna gibba). Table 4-2 shows metal concentration data obtained from a constructed wetland demonstration project in Sacramento, where disinfected activated sludge effluent was applied to parallel 1.44 acre cells,

each with a hydraulic loading of 65 m3/ha-d and similar

plant density.

The valence and form of each metal was not determined,

but nickel and arsenic appeared to be the most resistent

to removal (SCRSD, 1998). Several researchers have stud1.5

1

0.5

0

0 100 200 300 400 500 600 700

HRT (hr)

Orthophosphate Removal

(mg/L)

y = 0.13655 + 0.0021878x R = 0.86206

Figure 4-8. Retention Time vs. Orthophosphate Removal (mg/l)
74

10,000

1,000

100

10

1 100 1,000 10,000 100,000 1,000,000 10

Influent FC (number/100 mL)

Effluent FC (number/100 mL)

Figure 4-9. Average influent FC concentration vs. FC effluent concentration for TADB wetland systems

100

10

1

11 2345 67

Days

100

1,000

105

104

BOD and TSS (mg/L)

Fecal Coliform (cfu/100mL)

TSS

BOD

(HRT = 6 days)

Fecal Coliform

Figure 4-10. Arcata pilot Cell 8, TSS, BOD, FC

Table 4-2. Trace Metal Concentrations and Removal Rates, Sacramento Regional Wastewater

Treatment Plant (SCRSD, 1998).

Removal (%)

Metal Influent (mg/L) Effluent (mg/L) Rate

Min Mean Max Min Mean Max

Silver 0.25 0.29 0.32 0.02 0.03 0.03 90

Arsenic 2.00 2.23 2.60 1.50 2.20 3.10 1.3

Cadmium 0.040 0.077 0.140 0.005 0.009 0.019 88

Chromium 0.50 1.05 1.40 0.50 0.77 3.10 27

Copper 4.60 8.62 17.00 1.60 4.04 7.00 19

Mercury 0.0084 0.0105 0.0144 0.0021 0.0031 0.0041 71

Nickel 4.30 8.23 23.00 4.10 8.96 20.00 —

Lead 0.25 0.58 1.20 0.05 0.14 0.26 55

Antimony 0.40 0.41 0.42 0.12 0.15 0.20 63

Selenium 0.50 0.50 0.50 0.50 0.50 0.50 —

Zinc 6.4 26.2 34.0 1.30 3.53 8.70 70
75

ied particle sizes vs removal rates. Most have concentrated

on urban mechanical wastewater treatment systems.

Odegaard (1987) noted that except for nickel 50 to 75% of

the incoming metals (zinc, copper, chromium, lead, and

cadmium) in the wastewater were associated with the TSS.

Hannah, et al, (1986) showed that facultative lagoons removed 40 to 80% of metals, including nickel. The summary of these and other studies is that most metals, with

the exception of nickel, boron, selenium, and arsenic, tend

to associate with removable solids fractions. Gearheart and

Finney (1996) evaluated particle size removals of oxidation pond effluent in a FWS wetland (see Table 4-3). Their

results show that the settleable solids (> 100 &m) portion

of BOD, COD and TSS are essentially completely removed,

the supracolloidal 1 to 100 &m) fraction is 80 to 90% removed, while the remaining (<1&m) fractions are less impacted.

4.1.9 Stochastic Variability

Free water surface (FWS) treatment wetlands demonstrate the same type of water quality variability typical of

other complex biological treatment processes. While inlet

concentration pulses are frequently dampened through the

long hydraulic and solids retention times of the treatment

wetland, there is always significant spatial and temporal

variability in wetland water pollutant concentrations. The

stochastic character of rainfall and the periodicity and seasonal fluctuation in ET contribute to much of this variability

in the concentrations in wetland effluents. Better design

and operational factors could reduce some of the variation

seen in systems to date. Each site and its unique climatology require the designer to consider the effect these variables will have on the sizing, depth, and configuration of

the system.

4.2 Wetland Hydrology

The hydrology of FWS wetlands is considered by many

to be the most important factor in maintaining wetland structure and function, determining species composition, and

developing a successful wetlands project (Mitsch and

Gosselink, 1993). The following section describes the most

common method for characterizing wetland hydrology: the

development of a wetland water balance.

4.2.1 Wetland Water Balance

The wetland water balance quantifies the hydrologic

balance between inflows, outflows, and internal gains and

losses of water to a wetland, in relation to the wetland volume or storage capacity. The sources of water to a FWS

constructed wetland are wastewater inflow and precipitation, snowmelt and direct runoff from the wetland catchment (i.e. berms and roads). Water losses from a FWS

constructed wetland occur through the outlet, evapotranspiration, infiltration, and bank storage (wicking). A thorough

understanding of the dynamic nature of the wetland water

balance, and how this balance affects pollutants, is useful

in the planning and design of FWS constructed wetlands.

An overall wetland water balance is the first step in designing a FWS constructed wetland, and should be completed prior to the actual design steps described later in

this chapter. At a minimum, a detailed monthly or seasonal

water balance, which considers all potential water losses

and gains, should be conducted for any proposed FWS

constructed wetland. An annual water balance may miss

important seasonal wetland water gains or losses, such

as heavy periods of winter precipitation or high summer

evapotranspiration rates, which can affect FWS constructed

wetland pollutant effluent concentrations. Water balances

performed over shorter time periods than monthly will capture additional information about the dynamics of a wetlands hydrology, but the increased cost of data acquisition

will not generally be justified.

The wetland water balance for a FWS constructed wetland can be expressed in generic units (L=length; T=time)

as:

 (4-1)

where:

Aw = wetland water surface area (L2

),

ET = evapotranspiration rate (L/T),

I = infiltration to groundwater (L/T),

P = precipitation rate (L/T),

Qb = berm loss rate (L3/T),

Qc = catchment runoff rate (L3/T),

Qo = wastewater inflow rate (L3/T),

Qe = wetland outflow rate (L3

/T),

Qsm = snowmelt rate (L3/T),

t = time (T), and

Vw = water volume or storage in wetland (L3).

The impact of wet weather and snowmelt on the wastewater inflow (Qo

) is external to the water balance.

Some of the terms in Equation 4-1 may be deemed insignificant and can be neglected, simplifying calculations

and data collection requirements. For example, groundTable 4-3. Fractional Distribution of BOD, COD and TSS in the

Oxidation Pond Effluent and Effluent from Marsh Cell 5

(Gearheart and Finney, 1996)

BOD COD TSS

mg/L % mg/L % mg/L %

Oxidation Pond

Fraction

Total 27.5 100 80 100 31.0 100

Settleable 3.7 13 5 6 5.8 19

Supracolloidal 13.7 50 23 29 25.2 81

Colloidal-soluble 10.1 37 52 65 — —

Marsh Fraction

Total 4.8 100 50 100 2.3 100

Settleable 0 0 0 0 0.3 13

Supracolloidal 1.2 24 4 8 2.0 87

Colloidal-soluble 3.6 76 46 92 — —

dVw

dt = + + − − ++ + Q Q Q Q Q (P ET I)A o c sm b e w
76

water infiltration (I) and berm losses (Qb) can be neglected

if the wetland is lined with an impermeable barrier, and the

snowmelt (Qsm) is only important in certain locations.

4.2.2 Wastewater Inflow

The daily wastewater inflow flow rate (Qo) will almost

always be the primary inflow into a FWS constructed wetland. If the FWS wetland is being added to an existing

wastewater treatment process, wastewater flow rates may

already be measured. If the wastewater flow rates and

variability are not known, they can be estimated using conventional engineering methods. Examples of variable

wastewater flows include seasonal peaks from vacation

communities and seasonal high infiltration and inflow rates

into collection systems.

4.2.3 Precipitation, Snowmelt, and

Catchment Runoff

Depending on the time period of the water balance, daily,

monthly or seasonal precipitation and snowmelt data may

be required. Precipitation inflows into a wetland come from

direct precipitation (P) onto the wetland surface area and

runoff from the wetland catchment (Qc). Snowmelt

(Qsm)accounts for the amount of water entering the wetland from melting snow from the wetland catchment. The

effects of precipitation on the wetland water balance are

normally significant, while snowmelt can be seasonably

significant only in certain climates. Qc

 is a significant factor

only in the rarest of circumstances.

4.2.4 Wastewater Outflow

The wastewater outflow (Qe) corresponds to the amount

of treated wastewater leaving the FWS constructed wetland over a specified time period. Wastewater outflow reflects the balance between inflows, additional water gains

and losses, and the change in storage of the FWS constructed wetland.

4.2.5 Evapotranspiration

Wetland evapotranspiration (ET) is the combined water

loss due to evaporation from the water surface and transpiration from wetland vegetation. The loss of water from

ET affects the wetland in two ways. It increases the hydraulic retention time by removing water, and can concentrate certain pollutants, especially conservative dissolved

constituents. For non-conservative constituents, such as

BOD, an increase in the hydraulic retention time may provide a modified removal rate which can either partially offset or enhance the concentrating effects of ET.

Specific ET rates have proven difficult to accurately

measure in FWS wetlands. As a consequence, it is common practice in wetland design to assume that wetland

ET rates are equivalent to some percentage of open water

or pan evaporation rates. Kadlec and Knight (1996) recommend that ET be assumed equal to 70 to 80% of Class

A pan evaporation in fully vegetated FWS systems. Reed,

et al, (1995) suggest 80% of the pan rate. Since ET rates

in FWS systems may vary from those in open waters to

those in fully vegetated zones, an overall average rate may

be useful. A rate of 70 to 75% of the pan rate is a reasonable assumption since the two are not significantly different. Maximum ET rates have been found in smaller wetlands or wetland test cells with small area to perimeter

ratios (Gearheart, et al, 1993). ET rates of up to 5 mm/d

are found in the southern U.S., and Qe may approach zero

during these periods (EPA, 1999).

4.2.6 Infiltration and Berm Losses

Infiltration (I) is the loss of water that occurs into the bottom soils or berms of a FWS constructed wetland. If

present, infiltration decreases the outlet flow rate, effectively increasing the water retention time and increasing

the potential for constituent removal. Constituent reduction may be further improved by the loss of soluble pollutants into the soil as the water infiltrates. Infiltration tends

to reduce with time as clogging of soil pores progresses

(Middlebrooks, et al, 1982). If the FWS constructed wetland is lined with some type of impermeable barrier, infiltration can be neglected in the water balance. If not, siting

on more permeable soils might endanger ground water

quality.

4.2.7 Wetland Volume

The outlet in a FWS constructed wetland generally consists of a control structure that can regulate water depths

in the wetland. Increasing or decreasing water levels

changes the wetland volume, which influences the water

balance by providing more or less storage capacity. The

wetland volume (Vw) or storage capacity directly influences

the time required for the wastewater to pass through the

wetland. Water storage capacity can be increased to offset the effects of high seasonal precipitation or evapotranspiration. Since FWS constructed wetlands have a continuous wastewater inflow and some form of outlet water level

control, water surface elevations do not change significantly, unless the wetland operation and maintenance

schedule dictates water level fluctuations. The type of

emergent aquatic plants in each region of an FWS is primarily determined by the depth of the FWS in that zone.

For the most part 1.2m (4 feet) is considered the maximum seasonal water depth for fully vegetated sections of

the wetland. Normal operating depths vary from 0.5 to

0.75m ( 1.7 to 2.5 feet) depending on the types of plants

and types of physical substrate.

4.3 Wetland Hydraulics

From a design perspective, wetland hydraulics defines

the movement of water through a FWS constructed wetland. A FWS constructed wetland with poor hydraulic design can be problematic in terms of effluent water quality,

odors, and vector nuisances. This section first defines some

basic wetland hydraulics terms, and then briefly summarizes basic wetland hydraulic principles.
77

4.3.1 Wetland Hydraulics Terminology

and Definitions

4.3.1.1 Water Depth

Water depth is an important physical measure for the

design, analysis, and operation and maintenance of FWS

constructed wetlands. The ability to vary water depth in a

FWS constructed wetland is one operational control available to operators to manipulate wetland performance.

The actual water depth at all locations in a FWS constructed wetland will generally not be known with a high

degree of accuracy due to basin bottom irregularities. In

addition, the water depth in the wetland may decrease due

to the buildup of peat from the deposition of detritus and

settled solids buildup. Increasing the water depth by changing the outlet weir elevation can help offset decreases in

water depth. This slow-rate change progresses with time

from the inlet toward the outlet. Detrital plant material builds

up on and under the water surface of any vegetated zone,

especially the initial vegetated zone. Wetlands operating

for 15 years have documented a 0.08 - 0.12m (3 to 5 inch)

depth change due to plant detritus along the initial vegetated zone in an FWS wetland. The largest accumulation

was in the inlet region (Kadlecik, 1996).

Estimated operating water depths for FWS constructed

wetlands in the NADB (1993) have ranged from approximately 0.1 to over 2.0m (0.3 to 6+ feet) with typical depths

of 0.15 - 0.60m (0.5 to 2 feet). Operating depths are generally different for the area with emergent plants (0.6 m)

than those areas with submergent plants (1.2m). Since

most of the NADB systems were designed to be fully vegetated, these depths are less than one would expect in the

future. For some calculations the average water depth (h)

can be used, as it represents the average water depth over

the total wetland surface area (Aw).

4.3.1.2 Volume

The volume (Vw) of a FWS constructed wetland is the

potential quantity of water (neglecting vegetation, litter and

peat) that could be stored in the wetland basin. The wetland water volume can be determined by multiplying average water depth (h) by area (Aw):

Vw = (Aw)(h) (4-2)

4.3.1.3 Wetland Porosity or Void Fraction

In a FWS wetland, the vegetation, settled solids, litter

and peat occupy a portion of the water column, thereby

reducing the space available for water. The porosity of a

wetland (0), or void fraction, is the fraction of the total volume available through which water can flow. Wetland porosity has proven difficult to accurately measure in the field.

As a result, wetland porosity values listed in the literature

are highly variable. For example, Reed, et al (1995) and

Crites and Tchobanoglous (1996) suggest wetland porosity values ranging from 0.65 to 0.75 for fully vegetated

wetlands, for dense to less-mature wetlands, respectively.

Kadlec and Knight (1996) report that average wetland porosity values are usually greater than 0.95, and 0 = 1.0

can be used as a good approximation. Gearheart (1997)

found porosity values in the range of 0.75 in dense mature

portions of the Arcata wetland. For hydrological design, an

average porosity value should be used which is based on

the areal percent of open water zones (non-emergent vegetation) to vegetated zones. For example, a wetland with

50% open water (0 = 1.0) and 50% emergent vegetation

(0 = 0.75), would have an average 0 = 0.875.

The overall effects of decreasing porosity are to reduce

the wetland volume available for water, which reduces the

retention time of water within the wetland, and to increase

flocculation of colloidal material which improves removal

by sedimentation. It is recommended that a porosity value

of 0.65 to 0.75 for fully vegetated zones be used in FWS

constructed wetland design calculations, with lower values for the most densely vegetated areas. A value of 0 =

1.0 should be used for wetland open water zones. The

use of conservative average porosity values provides a

factor of safety, and results in a more conservative design.

4.3.1.4 Average Wastewater Flow

The average wastewater flow accounts for the effects of

water gains and losses (precipitation, evapotranspiration

and infiltration) that occur in a FWS constructed wetland.

Defining Qo as the FWS influent flow rate and Qe as the

FWS effluent flow rate, the average wastewater flow rate

is expressed as:

 (4-3)

If actual wastewater inflow and outflow are known, these

values can be used in Equation 4-3. If only one of these

flows has been measured, a water balance can be conducted to determine the other. If neither are known, a water balance can be useful to show the relationship between

the two under the extreme circumstances of operation.

4.3.1.5 Hydraulic Retention Time

The nominal hydraulic retention time (HRT) is defined

as the ratio of the useable wetland water volume to the

average flow rate (Qave). The theoretical hydraulic retention time as t can be calculated as:

t = (Vw)(0) / Qave (4-4)

The flow rate used in the hydraulic retention time calculation can be the average wetland flow (Qave) or the maximum or minimum flows, depending on the purpose of the

calculation.

4.3.1.6 Hydraulic Loading Rate

The hydraulic loading rate (q) is the volumetric flow rate

divided by the wetland surface area and represents the

depth of water distributed to the wetland surface over a

specified time interval. The hydraulic loading rate can be

written as:

q = Qo

 / Aw (4-5)

Q Q Q

2 ave o e = +
78

 where q has units of L/T. Generally the hydraulic loading rate is determined using the wastewater inflow (Qo

).

4.3.2 Water Conveyance

Water conveyance in FWS wetlands is hydraulically complex, varying in both space and time due to wetland vegetation and litter, changing inflow conditions, and the stochastic nature of hydrologic events.

When designing a constructed wetland, it is necessary

to understand how water moves through the wetland, and

how this water movement influences various design considerations.

4.3.2.1 Ideal versus Actual Flow in a FWS

Constructed Wetland

Though plug flow is generally assumed for the purposes

of FWS constructed wetland design, actual wetland flow

hydraulics do not follow an ideal plug flow model. The deviation from plug flow of an existing FWS constructed wetland can be determined through the use of tracer tests.

One result of a tracer test is the determination of the average tracer retention time, which is defined as the centroid

of the response curve, as shown in Figure 4-11. The average tracer retention time is equal to the active water volume (Vw) (0) divided by the average volumetric flow rate

(Qave), and thus represents a direct measure of actual retention time. Results from some tracer studies have shown

that the hydraulic characteristics of a FWS constructed

wetland can be approximated by a series of 4 to 6 equally

sized complete mix reactors (Kadlec and Knight, 1996; and

Crites and Tchobanoglous, 1998). In other studies, the

complete mix reactor model has resulted in a poor fit to

the data, and other models have been more successful.

Figure 4-11 shows the observed tracer concentration from

one wetland cell of the Sacramento Regional Wastewater

Treatment Plant Demonstration Wetlands Project compared to the predicted tracer concentrations using the finite state model first suggested by Hovorka (1961). The

finite stage model integrates components of completely

mixed, plug flow, and off line storage addition/feedback

into one hydraulic model. Coefficients are unique for each

geometry, planting pattern, etc. For any given site with

appropriate data the finite stage model gives the best fit of

the tracer data. This method was first applied to FWS systems at the Arcata pilot studies (Gearheart, et al, 1983)

and has subsequently been applied to the Sacramento

system (Dombeck, 1998). The value of multiple cells and

periodic open-water zones have been recognized for minimizing short-circuiting by numerous authors.

4.3.2.2 Hydraulic Gradient in a FWS

Constructed Wetland

For FWS constructed wetlands, some assessment of the

energy loss or head loss from inlet to outlet is necessary

to ensure that the wetland is designed to handle all potential flows without creating significant backwater problems,

such as flooding the inlet structures or overtopping berms.

It has historically been assumed that Manning's equation,

which defines flow in open channels, can be adapted to

0.90

0.80

0.70

0.60

0.50

0.40

0.30

0.20

0.10

0.00

0 2 4 6 8 10 12 14 16

Time (days)

Li Concentration (mg/L)

Finite Stage Model

Observed

Figure 4-11. Tracer response curve for Sacramento Regional Wastewater Treatment Plant Demonstration Wetlands Project Cell 7 (SERSD, 1998).
79

estimate head loss in FWS wetlands. By assuming that

the submerged wetland vegetation, peat and litter provides

more frictional resistance to flow than the wetland bottom

and sides, Manning's equation has been adapted as follows:

 (4-6)

where:

v = average flow velocity (L/T),

n = Manning’s resistance coefficient (T/L1/3),

h = average wetland depth (L), and

S = hydraulic gradient or slope of water surface (L/L).

In the above equation, the average wetland depth and

water surface slope is fairly easily estimated, and the average velocity (v) is defined as the average flow (Qave) divided by the available average cross sectional area (Av

)(0),

or (Width)(depth)(0). The determination of Manning's resistance coefficient (n) is not as straightforward. In wetlands, the vegetation and litter providing resistance to water flow is distributed throughout the water column, with

settled particles and detritus on the bottom and a thicker

thatch level at the top. Thus, n should be a function of the

water depth as well as the resistance of specific surfaces.

Measurements of n in operating wetlands range from approximately 0.3 to 1.1 s/m1/3 with the higher numbers corresponding to water depths less than 0.2 m (Kadlec and

Knight, 1996). Reed, et al, (1995) use an equation to estimate n at 0.2m depth to vary from 1 to 14 s/m1/3. Linsley,

et.al., (1982) offer a series of values for n from0.024 to

0.112. A default value of 0.1to 0.5 is suggested for those

wishing to pursue this issue. A typical solution provided in

Crites and Tchobanoglous (1996) is a slope of 1 in 10,000,

or 1cm in 100 meters. Since multiple cells are recommended as good design practice to minimize short-circuiting and to maximize treatment performance, the above

analysis is superfluous for most applications where aspect

ratios (length/width) are within suggested limits of 3:1 to

5:1, or even larger.

4.4 Wetland System Design and Sizing

Rationale

4.4.1 Introduction

As FWS constructed wetlands became recognized as a

viable wastewater treatment process, FWS design models soon followed. These models were intended to aid engineers/designers in the process of FWS wetland design

and performance assessment. To date, a number of wetland design methods have been proposed for predicting

constituent removals in FWS wetlands. These may be

found with explanation in Reed, et al, (1995), Kadlec and

Knight, (1996) and Crites and Tchobanoglous, (1998). The

design models and methods have been used to attempt to

predict the fate of BOD, TSS, TN, NH4- , NO3, TP and fecal

coliforms in a FWS system.

Free water surface constructed wetlands have usually

been modeled as attached growth biological reactors, in

which the plants and detrital material uniformly occupy the

entire volume of the wetland.

The current trend in wetland design modeling is the development of simple mass balance or input/output models. These simplified models do not explicitly account for

the many complex reactions that occur in a wetland, either

in the water column or at interfaces such as the water/

sediment interface. Instead, all reactions are lumped into

one overall biological reaction rate parameter that can be

estimated from collected FWS wetland performance data.

At this stage of wetland understanding, more complex and

theoretical wetland models which explicitly describe the

kinetics of known wetland processes may not be possible

due to severe limitations in almost all of the existing wetlands data.

4.4.2 Existing Models

In essence the types of models that have been used in

FWS constructed wetland design are known as plug-flowreactor (PFR) models.. One assumes horizontally based (linear) kinetics (Reed, et al,1995; Crites and Tchobanoglous,

1998), while the other assumes vertical (areal) kinetics

(Kadlec and Knight, 1996). Several varieties of these models exist. Some assume average kinetic rate constants, while

others assume retarded kinetic rate constants. All provide a

list of effluent background concentrations below which an

FWS cannot dependably attain and specific default values

for temperature adjustments to correct kinetic rate constants.

Some suggest monthly multipliers for average computed

design performance. Some include safety factors within the

equation while others apply them to the model result. All assume first-order biological kinetics, despite the fact that the

initial fully vegetated treatment zone is anaerobic, and none

of these models can account for a sequencing, i.e., fully vegetated and open-water zones in sequence, design which is

recommended herein for better performance. Recently, one

of the primary model creators has also noted the inadequacy

of these models (Kadlec, 2000). Readers are referred to

Kadlec and Knight (1996), Reed, et al, (1995), and Crites

and Tchobanoglous (1998) for details. For the purpose of

this manual, i.e., providing secondary (BOD=SS =30mg/l)

and advanced secondary treatment of municipal wastewaters, none of these equations alone are able to accurately

predict the performance of a multi-zone FWS constructed

wetland. Even if they could be calibrated “to fit” a specific set

of data their non-deterministic basis belies their ability to fit

other circumstances of operation.

4.4.3 Areal Loading Rates

The areal loading rate method specifies a maximum loading rate per unit area for a given constituent. These methods are common in the design of oxidation ponds and land

treatment systems. Areal loading rates can be used to give

both planning level and final design sizing estimates for

FWS systems from projected pollutant mass loads. For

example, knowing the areal BOD loading rate, the expected

BOD effluent concentration can be estimated or compared

to the long term average performance data of other welldocumented, full-scale operating systems.

S v

( /n) (h )

1/2 

1 2/3 =
80

In section 4.1 each pollutant was discussed based on

areal loading vs. effluent concentration based on the

DMDB, the TADB, specific studies and mechanistic evaluations of other sources of information. Areal loading does

not always correlate to a reasonable design basis, especially with regard to nutrients and pathogen removal, and

other mechanistic explanations are necessary. However,

if typical municipal wastewaters are to be treated which

have total and filtered pollutant fractionation which are reasonably consistent from site to site, a rational design approach can be deduced for those parameters which can

be removed during the enhanced flocculation/sedimentation which occurs in the initial fully vegetated zone of a

FWS constructed wetland. Therefore, based on Figures

4-1 and 4-4 the following areal loadings can be employed

for this initial zone (zone 1) of the FWS :

Parameter Zone 1 Areal Loading Effluent Concentration

BOD 40 kg/ha-d 30 mg/L

TSS 30 kg/ha-d 30 mg/L

The relative areal loadings imply that unless the pretreatment process were to have a BOD concentration of

greater than 1.3 times the TSS concentration, the latter

would be the critical loading rate for the fully vegetated

zone if secondary standards are to be met by a fully vegetated FWS system.

If the FWS system were to have significant open areas

between fully vegetated zones, a better effluent quality

could be attained at areal loadings, based on the entire

FWS system area (Aw):

Parameter Areal Loading Effluent Concentration

BOD 45 kg/ha-d <20 mg/L

 60 kg/ha-d 30 mg/L

TSS 30 kg/ha-d < 20 mg/L

50 kg/ha-d 30 mg/L

These loadings are based on the entire system area,

not just zone 1. Therefore, with open-water zones which

provide aerobic transformations and removal opportunities, a better effluent quality is achievable than with a fully

vegetated FWS system. Although there are insufficient data

at this time to eliminate the need to provide effluent disinfection, more disinfection interferences are removed which

would facilitate that step. Conversely, the open water zones

would attract wildlife to a greater degree, and the impacts

created by their activities. Similarly, the need to and the

power required to reaerate the final effluent will at the least

be reduced. The advantages of this design concept have

been described by Gearheart and Finney (1996) to include

reduced “background” BOD concentrations in the effluent

owing to the aerobic biological removals in the open-water

zones. As with the fully vegetated systems, the TSS areal

loading is more critical. With more quality data these limiting loadings could be shown to be conservative, especially

the BOD loading for attaining secondary effluent standards

with open-water FWS systems.

4.5 Design

4.5.1 Design Sizing and Performance

Mechanisms

If a pretreatent system already exists, the type of influent characterization necessary has already been discussed, but at a minimum all pollutants which are of concern to the NPDES permitting authority should be measured as both total and filtered through a standardized glass

fiber filter prior to analysis. Ideally, a particle-size distribution analysis of the type described in Crites and

Tchobanoglous (1998) could be performed for all critical

pollutants to aid the designer in predicting what level of

removals of each pollutant are likely to be attained by an

FWS or other treatment processes. If the pretreatment

system does not exist, the designer will need to perform a

variety of investigations as described in several engineering texts (e.g., Crites and Tchobanoglous,1998; WEF,

1998).

A primary supposition of this manual is that a FWS constructed wetland is most likely to treat effluent from a stabilization or oxidation pond or from primary-treated (settled)

municipal wastewater. After the designer determines overall

size of the FWS system from these BOD and TSS areal

loading rates, he or she can return to evaluate the fate of

other constituents.

If the total and filtered analyses are available, it is a reasonable approximation to assume that the filtered analysis represents a rough approximation of effluent quality

attainable from treatment zone 1 (fully vegetated zone)

given that the filter pores are generally a bit larger than the

specific particle sizes indicated for the “colloidal/soluble”

fraction in Table 4-3. Internal loads in the soluble form

should also be added to this fraction in estimation of zone

1 effluent(see Figure 4-2 and 4-3). For more stringent effluent requirements than those cited above for BOD and

TSS, the designer should look at alternative polishing processes such as land treatment or slow sand filtration.

While a few physical and chemical processes occur uniformly over the entire wetland volume, many of the most

important treatment processes occur in a sequential manner and the wetland must be designed to accommodate

this characteristic. For example, TSS removal and removal

of associated BOD, Org N and P, metals, etc., occur in the

initial portion of the cell, while the subsequent zones can

impact certain soluble constituents. Given sufficient dissolved oxygen in open (unvegetated) areas, soluble BOD

removal and nitrification of ammonia can occur. If insufficient oxygen is present, soluble BOD is very slowly removed by anaerobic processes. Wetland design must also

consider the background level, or expected lower limit, of

water quality constituents in the FWS wetland effluent (see

Table 4-4). Particulate and soluble constituents are internally produced as a part of the normal decomposition and

treatment processes occurring in a constructed wetland.

Wildlife contribute fecal coliform and additional organic

compounds. During periods of intense activity, wildlife also
81

stir up settled solids contributing to an increase in turbidity, TSS, and BOD. Table 4-4 shows the typical background

levels for the constituents of interest recommended for

users of this document. Designs requiring effluent quality

close to the values in Table 4-4 must be aware of the natural fluctuations about the mean values, as shown in Figure

4-12. For more details on the numerical values in the table,

the reader is urged to refer to Reed, et al, (1995), Kadlec

and Knight, 1996, and Gearheart, 1992.

A similar approach to the one suggested here for designing FWS wetlands, referred to as the "sequential

model", has been developed by Gearheart and Finney

(1999). The overall approach of the model is to consider

the dominant physical and biological processes responsible for determining effluent quality from each distinctive

area or zone of the constructed wetland and allow the designer to specify areal requirements and wetland depth for

each of these specific functions. This methodology recognizes that while some of the constituent transformations

and removal mechanisms are to some degree occurring

simultaneously throughout the wetland, the majority of the

removal occurs in a sequential fashion, with one process

or mechanism providing the products for the next process

or mechanism. The total area required for treatment is then

a sum of each of the zones required to reach a specific

effluent objective. This approach allows the designer to

sequentially determine the range of effluent characteristics which are attainable in a given definable zone before

entering a subsequent reactor (zone) which has known

treatment capabilities.

Table 4-4. Background Concentrations of Water Quality Constituents

of Concern in FWS Constructed Wetlands

Range Typical

Parameter (mg/L) (mg/L) Factors governing

TSS 2 - 5 3 Plant types, plant coverage, climate,

wildlife activity

BOD1 2 - 8 5 Plant types, plant coverage, plant

density, climate, wildlife activity

BOD2 5 - 12 10 Plant types, plant coverage, plant

density, climate

TN 1 - 3 2 Plant types, plant coverage, climate,

oxic/anoxic conditions

NH4-N 0.2 - 1.5 1 Plant types, plant coverage, climate,

oxic/anoxic conditions

TP 0.1 - 0.5 0.3 Plant types, plant coverage, climate,

soil type

FC3 50 - 5000 200 Plant types, plant coverage, climate,

wildlife activity

1

FWS with open water and submergent and floating aquatic macrophytes.

2

Fully vegetated with emergent macrophytes and with a minimum of

open water.

2

Measured in cfu/100 ml

The sequential model approach recognizes that all the

treatment objectives beyond secondary require a minimum

of three general wetland "compartments" (see Figure 4-

13): (1) an initial compartment where the bulk of the flocculation and sedimentation will occur, (2) an aerobic compartment where soluble BOD reduction and nitrification can

occur, and (3) a vegetated polishing compartment where

further reductions in TSS and associated constituents and

nitrogen (via denitrification) can occur. Permanent phosphorus removal in wetlands is generally small and is largely

the result of phosphorus adsorption to solids and plant

detritus. Sedimentation and pathogen reduction are related

to detention time in zone 1, to retention time and temperature in zone 2, and to retention time in zone 3. As noted

earlier, the notion of "compartments" is artificial as the treatment processes overlap in time and space, and no specific physical compartment is necessarily implied. However,

separation of an FWS into a series of single-function zones

(cells) with individual outlet controls is not an unattractive

concept.

A rational overview of the FWS system is depicted in

Figure 4-14. It illustrates that the primary mechanisms in

zone 1, which is fully vegetated and anaerobic throughout

its depth during the growing season, are sedimentation

and flocculation, as determined by transect measurements

of dissolved oxygen and pollutant concentrations. Any extension of the HRT in zone 1 beyond approximately 2 days

at Qmax would be essentially wasteful since the anaerobic conditions will not result in any significant further removal of soluble constituents and flocculation sedimentation has been effectively completed. The TSS and associated constituents (particulate BOD, organic nitrogen and

phosphorous, metals and certain semivolatile organic compounds) have also reached this same status. Volatile organics are likely to be removed from the wastewater during the collection or oxidation pond treatment processes

(Hannah, et al, 1986), while most semivolatiles are removed

with the solids in the oxidation pond or in zone 1 of the

FWS system.

For many years it has been recognized that effluent flocculation is primarily a function of energy input from either

external sources or internal hydrodynamic forces, and that

reduced Reynolds’ Numbers (Re) induce optimal sedimentation of particles. Over the past several decades this phenomena has been applied in the development of hydrodynamic devices which accomplish excellent flocculation and/

or sedimentation without moving parts, such as pipe mixers and flocculators, tube and plate settlers, and pebble

bed and wedgewire outlet devices for clairifers. Flow

through the emerging vegetation is extremely tortuous and

is accompanied by a very small hydraulic radius. The

Reynolds Number (Re) is a direct function of the hydraulic

radius (diameter, if the path were round (as in a pipe). If

the Re falls in a range which corresponds to laminar flow,

sedimentation is maximized. Re is several thousand in large

basins, and even larger in non-vegetated ponds. No direct

measurements of Re or laminar flow have been made at

the time of this writing, but analogous results from studies
82

150

140

130

120

110

100

90

80

70

60

50

40

30

20

10

0

0 10 20 30 40 50 60 70

Mean

Median

Minimum

Maximum

Distance from Influent (m)

BOD (mg/L)

Figure 4-12. Mean, median, minimum and maximum transect BOD5 data for Arcata Pilot Cell 8

Figure 4-13. Elements of a free water surface (FWS) constructed wetland

Pretreated

(Lagoon)

Influent

Inlet Settling Zone

Floating and

Emergent Plants

Submerged

Growth Plants

Floating and Emergent

 Plants

Zone 3

Fully Vegetated

D.O. (–)

H ≤ 0.75 m

Zone 2

Open-Water Surface

D.O. (+)

H ≥ 1.2 m

Zone 1

Fully Vegetated

D.O. (–)

H ≤ 0.75 m

Outlet

Zone

Length (L)

Depth (H)

Width (W)

Variable

Level

Outlet
83

of tube settlers and particle-size removals support this

theory, given the large amount of wetted surface available.

(Sparham, 1970). This concept also supports the use of

fully vegetated areas immediately preceding outlet weirs.

In zone 2, which is primarily open-water, the natural

reaeration processes are supplemented by submerged

macrophytes during daylight periods to elevate dissolved

oxygen in order to oxidize carbonaceous compounds

(BOD) to sufficiently low levels to facilitate nitrification of

the NH4-N to NO3-N. These processes require large

amounts of oxygen and time in a passive system (no mechanical assistance). The maximum HRT in zone 2 is generally limited to about 2 to 3 days before unwanted algal

blooms occur. Therefore, more than one open zone may

be required to complete these reactions. If so, the result

would be a five (or more) zone design since each open

zone would be followed by a fully vegetated zone. The

reactions in zone 2 are essentially the same as in a facultative lagoon. Therefore, the equations which apply to those

systems might offer reasonable approximations to the rate

of transformations occurring in this open-water zone. Therefore, the first-order Marais and Shaw (1961) equation for

fecal coliform dieoff could be applied as an approximation,

along with its temperature dependancy:

 (4-7)

where: Co = influent FC concentration, cfu/100 ml

Ce = effluent FC concentration, cfu/100 ml

N = number of open-water zones in the

FWS

t = HRT (T)

Kp = fecal coliform removal rate constant

(T-1)

= 2.6 (1.19)T-20 (4-8)

where: T = temperature, °C

BOD removal in the open-water zone should also follow

existing equations such as (Crites and Tchobanoglous,

1998);

 (4-9)

Zone 1 Zone 2

+ ∆ = Waterfowl Input

– ∆ = Sunlight Killing

 (disinfection)

Zone 3

Phosphorus

Pathogens

Nitrogen

TSS, Metals, etc.

Inlet Outlet

100

0

% Removal

BOD

Resolubilization

(internal loading)

Aerobic Biological

Decomposition

+∆ = f (resuspension,

 periphyton generation)

Dominant Processes:

Sedimentation and Denitrification

Dominant Processes:

Flocculation and Sedimentation

Figure 4-14. Generic removal of pollutants in 3-zone FWS system

Ce

Co 1

(1 tK ) p

N = +

Ce

Co 1

(1 tK ) p

N = +
84

where:

C = BOD, mg/L

Kb = specific BODs removal rate constant

(T-1)

Kb = 0.15 (1.04)T-20 (4-10)

Therefore, in analyzing Figure 4-14 the downward stope

in FC and BOD in zone 2 can be approximated through

the above equations, without considering offsets from wildlife. As noted previously, the nitrifying bacteria can proliferate and convert ammonia-nitrogen to nitrate(NO3

-N) and

will be the primary nitrogen transformation role of zone 2.

However, the carbonaceous BOD must be low enough to

allow these reactions to occur. In rotating biological

contactors this concentration of BOD is about 15 mg/l

(USEPA, 1993). As noted by Gearheart (1992) increasing

the size of the open-water zone generally increases dissolved oxygen, pH, and NO3

-N, while decreasing soluble

BOD and ammonium.

U.K.’s Department of Environment has studied lagoon

systems treating similar quality influent to that of zone 2.

They have noted that algal growth generally starts to occur between days 2 and 3 (UK, 1973). Algal growth can

raise pH, interfere with FC kill and the growth of submerged

plants, increase NH3-N volatilization, and induce phosphorus precipitation. Also, the additional biomass and precipitates that must be removed in zone 3 will add to the internal loading on the FWS system. The primary goal of the

open-water zone is to provide dissolved oxygen to remove

BOD and convert NH4-N to NO3-N. Therefore, the optimium

sizing of this zone would be an HRT of 2 to 3 days. Assuming a Qmax/Qave of 2, the designer might choose an HRT of

2 days at Qmax or an HRT of 3 days at Qave. Climate would

likely be the final criterion, with the larger size favored in

northern areas and the smaller in southern ones.

The third zone is fully vegetated like zone 1 and has a

similar function. Zone 3, like zone 1 is also capable of denitrification if the influent flow contains NO3

-N. Where oxidation pond pretreatment of municipal wastewaters is employed, zone 1 of the FWS system is not generally required

to denitrify, but zone 3 will if zone 2 induces nitrification.

The primary energy source for successful denitrification is

the release of organic substrates from the detritus from

decaying plants. However, partially digested, previously

removed organics may also be available. Denitrifying bacteria perform only under anaerobic conditions and best

when attached to large surface areas, e.g., plants. Denitrification, like nitrification, is temperature-sensitive. Nitrification and denitrification are greatly impaired when water

temperatures are reduced below 10°C. Gearheart (1992)

showed total inorganic nitrogen in the Arcata Marsh to be

reduced from 25 to 5 mg/L. In 1995 he demonstrated pilotscale removal of NO3

-N from 130 mg/L to 6 mg/L using no

supplemental carbon sources in 80 hours at 15°C. The

primary limitation in a three-zone FWS system designed

to remove nitrogen is the rate of nitrification in the openwater zone. If the open-water zone succeeds in nitrifying

the NH4

-N, the system should be able to denitrify it. Reed,

et al, (1995) indicate that denitrification should require less

than one day hydraulic retention time (HRT) for denitrification from municipal wastewater concentrations to an effluent requirement of ≤ 10 mg/L. Kadlec and Knight (1996)

found that 1 to 2 days should suffice to reach 90% NO3

-N

removal. Therefore the previously- stated requirement for

zone 3 (HRT of 2 days) should meet this retention requirement and ensure significant denitrification. WEF Manual

of Practice FD-16 (1990) indicates that the denitrification

rate can be as high as 10 kg/ha-d. Loadings must be within

the limits of available labile carbon to proceed at the maximum rate.

As with zone 1 there is additional, temporary nutrient (N

and P) removal by plant uptake in zone 3, which may be

significant at certain times during the year, while release

of most of these nutrients occurs at other times. These

plant effects can mask the effects of other processes which

could be impacting the system performance at the same

times. Unfortunately, there are insufficient data to fully quantify the nutrient cycle for each zone of the FWS system.

4.5.2 Total Suspended Solids Removal

Design Considerations

Since prior discussion indicates that TSS removal (rather

than BOD removal) drives the sizing process, there is a

need to provide further discussion of the mechanisms involved and their implications on design. Treatment mechanisms which dominate in the vegetated inlet zone of a FWS

constructed wetland volume are flocculation, sedimentation and anaerobic decomposition. Discrete and flocculent

settling occurs as the wastewater flows through the initial

fully vegetated zone. Since the FWS was likely preceded

by an oxidation pond where most discrete settling has occurred already, the enhanced settling in zone1 is mostly

due to flocculation of large supracolloidal solids in passage through the emergent vegetation. The processes are

generally not temperature dependent and occur at relatively high hydraulic loading rates. TSS removal rates of

40 to 60% are common with a q of 0.06 m/day to 0.27 m/

day, but relative removals are more accurately determined

by influent characteristics and the hydrodynamics of the

initial vegetated zone.

The majority of incoming solids are removed in this initial settling volume. Hyacinth and duckweed systems are

similar to (but not as good as) zone 1 of an FWS in the

hydrodynamics which promote excellent flocculation and

sedimentation. The mechanisms of the fully vegetated zone

1 can be estimated from the use of particle size distribution analysis. Generally, wastewaters have been analyzed

in form size ranges:

Settleable (>100 μm)

Supracolloidal (1 to 100 μm)

Colloidal (0.001 to 1 μm)

Dissolved (< 0.001 μm)
85

Only one study has employed this approach (Gearheart

and Finney, 1996) with oxidation ponds followed by FWS

constructed wetland treatment. The results shown in Table

4-3, clearly demonstrate the essentially complete removal

of the settleable fraction (100% for BOD and 95% for TSS)

and progressively reduced removal of the supracolloidal

(91% of BOD and 92% of TSS) and colloidal (66% of BOD)

fractions. This progression runs counter to the frequently

noted biological reaction rate vs particle size relationship

(Levine, et. al., 1991). Therefore, the primary mechanism

for removal of TSS and associated pollutants (BOD, organic nutrients, metals and toxic organics) is not biological in nature. This would appear to be reinforced by the

lack of dissolved oxygen, high oxygen demand, and the

slow nature of anaerobic biological reactions which are

the predominant biological mechanisms.

The solids that are removed undergo incomplete anaerobic decomposition (acidification) resulting in a release of

nitrogen, phosphorus, and carbon in the form of volatile

fatty acids. The amount of accumulated internal load depends on the length of time the water temperature stays

below 5-10°C since this material does not undergo significant decomposition until the water temperature increases

above this threshold value. The longer the uninterrupted

period of less than 5-10°C, the greater the initial load and

its effect on dissolved BOD at temperatures above this

threshold. In most temperate North American climates, the

release of this accumulated organic material expresses

itself mostly in the late spring and in the early summer,

similar to oxidation pond "spring turnover". Some of these

impacts are noted by the comments within Figure 4-14,

which show how some of these phenomena might impact

removal patterns.

Non-degradeable material is removed, accumulates and

is compressed forming an organic layer of biologically recalcitrant material in the sediments of zone 1. The layer is

thicker near the influent end of the wetland and gets shallower in the direction of flow. This delta of accumulated

solid material can eventually reduce the HRT and the available solids storage volume of the wetland. These losses

are also acerbated by accumulated plant detritus. These

accumulated solids ("sludge" or "biosolids") will occasionally need to be removed and managed, e.g., directly land

applied and plowed under as a soil amendment or through

some other method as directed by the regulatory authorities.

The reduction in wetland volume due to settled solids,

living plants and plant detritus can be significant over the

long term. The rate of accumulation of settled suspended

solids is a function of the water temperature, mass of influent TSS , the effectiveness of TSS removal, the decay

rate of the volatile fraction of the TSS, and the settled TSS

mass which is non-volatile. The plant detritus buildup is a

function of the standing crop and the decay rate of the

plant detritus. Accumulation for emergent vegetated areas

of the Arcata enhancement wetlands was measured to be

approximately 12 mm/year of detritus on the bottom due

to plant breakdown and 12 to 25 mm/year of litter forming

a thatch on the surface (Kadlecik, 1996). The volume of

the living plants, specifically the volume of the emergent

plants, ranged from 0.005 m3

/m2 (low stem density, water

depth of 0.3 m) to 0.078 m3

/m2 (high stem density, water

depth of 0.75 m). This accumulation is more or less constant from year to year as the wetland matures. The total

volume reduction under the initial vegetated zone can be

estimated using a mass balance equation:

Vr

 = [(Vss)(t) + (Vd)(t)] Aw (4-11)

where:

Vr = volume reduction over period of analysis (m3

),

VSS = volume reduction due to non-volatile TSS and

non-degradable volatile TSS accumulation

(m3/ha-yr),

Vd = volume reduction due to non-volatile detrital

accumulation as A function of annual production

(m3/ha-yr),

Aw = fully vegetated wetland area (ha),

t = period of analysis usually (years).

The loss of volume per hectare over a ten year period

for a 1 hectare fully vegetated FWS wetland zone with a

depth of 0.75 m can be estimated by use of this equation.

Based on information in Middlebrooks, et al, (1982) and

Carre, et al, (1990) a reasonable default value for Vss when

treating raw wastewater in lagoons) would be 200 to 400

m3/ha-yr (2 to 4 cm/yr). Therefore, a conservative default

value of 150 m3/ha-yr can be used. One hundred percent

coverage of emergent vegetation was measured to contribute 120 m3

/ha- yr of bottom detritus, and 120 m3/ha-yr

of surface litter with a standing crop volume of 412 m3/ha.

Substituting into the equation for a 10-year analysis yields:

Vr =[150)(10) + (240)(10)] 1.0 = 3,900m3

Table 4-5 provides additional examples of wetland volume loss due to TSS and plants detritus. Based on the

actual Arcata experience, it is clear that use of equation 4-

11 is a conservative means of estimating volume reduction from TSS deposition and detrital accumulation.

Using the initial fully vegetated zone volume (V1) and

adding the standing crop (Vsc), the total loss of volume can

be estimated by addition to be 4,312 m3

. Since the original

volume is area (10,000 m2

) times depth(0.75 m) or 7,500

m3

, the total loss of volume would be 4,312/10,000 or 43%.

This corresponds to a new porosity (0) of 0.57. As noted

earlier dense, mature stands of emergent plants are assumed to have a porosity of about 0.65.

Table 4-5. Examples of Change in Wetland Volume Due to Deposition

of Non-Degradable TSS (Vss) and Plant Detritus (Vd) Based

on 100% Emergent Plant Coverage (Gearheart, et al, 1998)

Influent TSS.

Vss (m3/yr)ha) Vd(m3/ha/yr)

(mg/L) 50% removal 75% removal

40 75 113 240

60 80 112 150

168 225 240 240
86

The accumulation computed above indicate that the call

is nearly ready for residual solids removal, as its excess

storage capacity is essentially used up. However, the Arcata

facility for which the accumulation measurements were

made is still performing well after 12 years (USEPA, 1999).

The loss of volume and porosity computed by the previous method is obviously conservative, but illustrates how

one could conservatively estimate the loss of porosity in

the initial settling zone.

When designing the primary wetland cell treating oxidation pond effluent the designer should consider this cumulative problem by increasing the depth of the inlet zone

(up to 1.0m) to lengthen the period before solids removal

would be required. The designer should also provide for

easily accessible solids removal in this zone. There may

be a need to harvest vegetation and related detritus to

maintain fully vegetated and open-water areas in proper

proportion. Such controlled harvesting may be extended

into the fully vegetated zones to reduce the apparent loss

of effective treatment volume and delay the need to remove accumulated solids.

The fully vegetated, anaerobic zone 1 of the FWS wetland should be designed based upon the average maximum monthly flow rate (Qmax) to assure the potential for

effective removal of solids during periods of high flow. To

facilitate solids removal and handling, this initial compartment should be designed as at least two equally sized

wetlands with a 0.6 to 0.9 m operating depth, which can

be operated in parallel. This would allow taking one cell

out of operation for maintenance work such as for solids

removal, vegetation removal, or replanting.

4.5.3 Design Examples

Design Example 1 - BOD and TSS to meet secondary

effluent requirements

Design a FWS wetland to treat lagoon effluent to meet a

monthly average 30 mg/l BOD and TSS discharge objective. The community has a design population of 50,000

people with an average annual design flow of 18,920 m3/

day (5 MGD) (Qave). Use design loading factors from sections 4.4 and 4.5 to meet a 30 mg/l BOD and TSS effluent

standard. Since a single fully vegetated FWS system can

be employed with maximum areal loading rates for these

systems are 40 kg BOD/ha-d and 30 kg TSS/ha-d. Facultative lagoon effluent typically averages from 30 to 40 mg

BOD/L and 40 to 100 mg TSS/L, with the latter being much

more variable due to seasonal algal growth and spring and

fall overturn periods (WEF, 1998; Middlebrooks, et al,

1982). For this example the average FWS influent BOD is

50 mg/l at Qave (18,920 m3

/d), while the average TSS is 70

mg/L at this flow. At the maximum monthly flow (Qmax) of 2

x Qave the BOD is 40 mg/L and TSS is 30 mg/L.

Step 1 - Apply areal loading rates(ALR) to average (Qave)

and maximum monthly flow (Qmax) conditions to identify the critical conditions for sizing of the facility.

ALR = Q Co/Aw (4-12)

where:

ALR = areal loading rates: BOD = 40 kg /ha-d ;TSS

= 30 kg /ha-d

Q0 = incoming flow rate, in m3

/d

C0 = influent concentration, in mg/L

Aw = total area of FWS, in ha

which for BOD yields:

for Qave, Aw = (18,920 m3

/d)(1000 L/m3)(50 mg/L)/ (40

kg/ha-d)(106 mg/kg) = 24 ha

for Qmax, Aw = (37,840) (40)/ (40) (106) = 38 ha

Similarly, for TSS:

for Qave, Aw = (18,920) (70)/(30)(106) = 44 ha

for Qmax, Aw = (37,840)(30)/(30)(106) = 38 ha

Therefore, the limiting condition is the TSS loading at

average flow conditions, where 44 hectares are required

to meet secondary effluent standards with a fully vegetated,

single-zone FWS system. However, it has been previously

shown that open water zones permit higher areal loading

rates (from section 4.4), so the sizing can be recomputed

on that basis following the same procedure. From that

analysis, with BOD and TSS loading rates of 60 and 50

kg/ha-d, respectively, the critical condition is still the average flow condition and the TSS areal loading, but with a

requirement of 26 ha instead of 44 ha.

Step 2 - Determine the theoretical HRT(days) using equation 4-4, assuming h = 0.6 m and 0 = 0.75 in vegetated zones (1 and 3) and h = 1.2 m and 0 = 1.0 in

the open zone (2). The combined estimate is an average depth of 0.8 m and an average 0 = 0.8. Therefore,

the first estimate is for overall HRT, followed by individual cell estimates.

for Qave,

 = 8.9 days

for Qmax t = 4.5 days

This last calculation implies that at the maximum monthly

flow the overall HRT may not be adequate for the necessary treatment mechanisms to perform. If these relatively

equal-sized zones are employed as a first approximation,

there would be less than one day of theoretical HRT in

each at this maximum flow condition. For zone 1 the minimum HRT at Qmax should be about 2 days, making 4 days

at Qave.

For zone 2 there is an upper limit which depends on

climate and temperature. In this example, the concept is

t

Vw

Q

Aw h

Q

(26 ha)(10,000 m / ha)(0.8 m)(0.80)

18,920 m d

ave ave

2

3

== = ε ε
87

that the open-water area will have an HRT exceeding that

which is required for an algal bloom, creating a significant

additional loading on zone 3. The time required for this

condition varies with temperature, e.g., shorter in hotter

climates. For most U.S. conditions a maximum HRT of

about 2 to 3 days should avoid most blooms. On the other

hand, the longer the HRT in zone 2, the better the reduction in soluble organics, ammonia-nitrogen and fecal

coliforms. Therefore, the designer should consider isolation of this zone from the fully-vegetated zones that precede and follow it and provide some flexibility in HRT control independent of the other zones to optimize this zone’s

treatment performance. For this exercise a minimum HRT

of 2 days is chosen, making the average HRT 4 days at

average flow.

Zone 3 should be provided the same design considerations as zone 1, since it functions in the same manner.

Depending on the performance of zone 2, it may also provide denitrification in addition to flocculation and sedimentation. Therefore, it should have approximately the same

HRT as zone 1.

The minimum HRT at Qmax is therefore 2 + 2 + 2 = 6

days, and the average 12 days with the assumptions chosen in this example. This would then require (using equation 4-4)an overall wetland area of:

Aw = (t)(Q)/(h)(0)

= (12 d)(18,920 m3

/d) / (0.8 m)(0.80)(10,000 m2/ha)

= 35 ha

Applying the normal additional area for buffers and setbacks of 1.25 to 1.4, the total area required for the FWS

facility is 45 ha (135 acres).

Step 3. - Configuration

Given the high TSS of the influent stream and the potential for short circuiting, the system should be designed

with two parallel treatment trains of a minimum of three

cells in each. The first cells in each train should normally

each get 50% of the flow and may have a deeper (1.0 m)

inlet area directly adjacent to the inlet structure to handle

any discrete solids settling which might occur at this location. This design option would add approximately one day

to the overall HRT, and would only be chosen in situations

where pretreatment is likely to allow escape of readily settleable particulates. Multiple cells allow for redistribution of

the primary cell effluent in the subsequent cell which reduces short-circuiting. Flexible intercell piping will facilitate maintenance without a major reduction in the necessary HRT to produce satisfactory effluent quality. Aspect

ratios of the cells should be greater than 3:1 and adapted

to the site contours and restrictions. Additional treatment

will likely be required after the FWS system to meet fecal

coliform and dissolved oxygen permit requirements.

Design Example 2 - BOD and TSS ≤ 20mg/L Effluent

Requirements

To meet this effluent quality an open-water zone will be

required in the FWS. From Section 4.4.3 the critical loadings to meet these effluent concentrations are 45 kg BOD

/ha-d and 30 kg TSS /ha-d. Using the same influent conditions in the first example, the steps of preliminary sizing

are the same.

Step 1 - Apply areal loading to determine the FWS system’s

critical sizing conditions using equation 4-12.

BOD: at Qave, Aw = (18,920)(50)(1000)/(45)(106)

= 21 ha

at Qmax, Aw = (37,840)(40)(1000)/(45)(106)

= 34 ha

TSS: at Qave, Aw = (18,920)(70)(1000)/(30)(106)

= 44 ha

at Qmax, Aw = (37,840)(30)(1000)/(30)(106)

= 38 ha

The limiting condition is again the TSS loading at Qave,

where 44 hectares are required. This is a larger requirement than in the previous example, as would be expected

since more stringent effluent requirements are being met.

Step 2 - Determine the theoretical HRT (t) required for the

entire 3-zone FWS system and each specific zone

using equation 4-4, assuming an overall average depth

(h) of 0.8 m and an overall porosity (0) of 0.8.

for Qave,

 = 14.9 days

for Qmax t = 7.4 days

Returning to individual zones and assuming an equal

minimum HRT in each, equation 4-4 is used in dimensioning at maximum flow conditions:

 = 7.9 ha

Therefore, the area for zones 1 and 3 are:

A3

 =A1 = (44 - 7.5 )/2

 = 18 ha

The overall FWS system area, including buffers, would

be about 58 hectares (145 acres).

Step 3 - Configuration

Again the use of parallel trains is encouraged for all the

same reasons as noted in the previous example. Parallel

trains of 3 cells in each are recommended which allow any

single cell in a train to be removed from service with transfer of its influent to the same zone cell.

By using an aspect ratio in the range of 3 to 5:1 and

complete-cell-width inlets and outlets the intercellular transfers should be simplified.

t (44 ha)(10,000 m / ha)(0.8 m)(0.8)

19,920 m / d

2

3 =

A (2.5 d)(37,840 m / d)

(1.0)(1.2 m)(10,000 m / ha) 2

3

2
88

Design Example 3 - Estimating from examples 1 and 2.

For design example 1, if the influent to the FWS system

was facultative lagoon effluent, it would be reasonable to

expect the characterization in Table 4-6,with cognizence

that the impact of climate and season on the performance

of lagoons and the characteristics of municipal wastewaters vary by orders of magnitude. However, the numbers

in the table are within the normally expected ranges. To

test the areal loading information presented earlier in this

chapter, the loading factors for each chemical constituent

can be computed using equation 4-12. These yield a TKN

loading of 10.8 kg/ha-d for example 1 and 8.6 kg/ha-d for

example 2. Similarly, the TP loadings are 2.7 kg/ha-d and

2.1 kg/ha-d for examples 1 and 2, respectively.

Comparing these loadings to Figure 4-5, it is impossible

to accurately predict effluent quality for example 1, but it

appears that little removal could be expected had the original fully vegetated approach been taken. Mechanistically,

the primary mechanisms which are available for TKN removal in zone 1 are flocculation and sedimentation of organic N. Since there are only 4 mg/L of organic N the realistic maximum expectation would be a removal of all but 1

mg/L, while the NH4

-N may be mostly assimilated by the

plants during the active growing rate computed for this

example, but most would be returned to the water column

during senescence. Therefore, the effluent TKN could average anywhere between about 15 and 20 mg/L,depending

on the season, and an overall removal of about 2 to 3 mg/

L. With the final open-water design of this example, it is

not possible to predict removal without more data from

open-zoned systems. As in example 2, the three-zone FWS

is likely to produce an effluent TKN of greater than 4 mg/L

since all three such systems on the figure were loaded at

a lower rate. Actual removal would depend upon nitrification accomplished in zone 2, which would be a function of

temperature, HRT and dissolved oxygen in that zone. Since

the TKN loading rate is about 4 to 5 times the highest one

in the figure for open water systems, a conservative approach might be that one-fourth of the nitrogen might be

nitrified in the open zone and denitrified in zone 3. This

would yield an additional 4 mg/L to the 3 assumed for example 1. This would yield a removal of 7 mg/L and an effluent TN of about 13 mg/L which could vary from about 10

to 16 mg/L during the year depending on plant condition

and temperature. Conversely, the systems shown in the

figure may have had excess capacity in zone 2 to fully

nitrify all the ammonium-nitrogen and the same effluent

concentration for the lightly loaded open-water systems

could also be attained. By having an open-water zone

where nitrification can occur, the inherent denitrification caTable 4-6. Lagoon Influent and Effluent Quality Assumptions.

Parameter Raw Wastewater Lagoon Effluent

TKN(mg/L) 40 20

NH4-N(mg/L) 10 16

TP(mg/L) 7 5

FC(#/100ml) 106 104

pability of the subsequent fully-vegetated zone creates a

potent opportunity for nitrogen control. It is also feasible

in the open-water zone to enhance NH3 volatilization, but

this mechanism is less likely to be significant owing to the

limited size of zone 2 which may not permit increased pH

which would enhance volatilization. Such estimates are

extremely tenuous until more data is generated on higher

loadings to these open water FWS systems, and in the

interim the designer would be wise to perform pilot studies

where nitrogen limits are part of effluent permit requirements.

Areal loading data on total phosphorus (TP) in Figure 4-

6 are inconclusive. The TADB (USEPA, 1999) would suggest that these example loadings could produce an effluent of 3.0 to 4.5 mg/L. At the loadings indicated in these

examples, the data of Gearheart (1993) would allow for an

overall annual average removal of approximately one mg/

L. This would provide a similar effluent for both examples

of about 4 mg/L. The dominant removal mechanisms in

both examples are flocculation and sedimentation of organic phosphorus, but plant uptake and release will cause

the effluent to vary from background levels in the growthphase to levels at or above the influent concentration during the senescent-phase. This discussion does not include

the startup-phase where TP removal will occur for several

months until the soil’s phosphorus adsorption capacity is

reduced to an equilibrium level by satisfaction of the soil’s

calcium, aluminum and iron adsorption sites and completion of the initial growth phase of the plants.

Fecal coliform (FC) removal is limited by the natural background which is depicted in Table 4-4. Figure 4-10 shows

that FC removal is based on enhanced sedimentation and

flocculation in the fully-vegetated zone of an FWS. Therefore, approximately one log (90%) of removal can be safely

estimated in that zone. With an open-water zone, the FWS

can take advantage of the natural solar disinfection which

is described in the international lagoon literature (Mara,

1975). This additional kill of FC is limited by the HRT in the

open-water zone and is a temperature-dependent function with first-order kinetics. Time limitation and single-cell

hydraulics will likely limit additional kill to about one log.

Since in zone 3 FC removal would be by sedimentation,

less than one additional log of removal could be expected.

Based on the prior analyses the total removal for examples 1 and 2 would be 2+ logs of kill, with an expected

effluent FC count of ≤100/100ml. A fully vegetated system

which has no open zone would likely remove somewhere

between 1 and 2 logs to produce an effluent with several

hundred FC/100ml. Both would experience a natural variation about those means as discussed earlier in the chapter. One major reason for periodic increases will be wildlife

attraction to open water zones. However, with the requirement that the outlet be located at the terminus of the subsequent fully-vegetated zone 3, the impacts of wildlife

should be minimized. However, some spikes of fecal

coliform may still be evident.

As noted earlier, the impact of the example designs on

metals and toxic organics will vary also. Most metals will
89

likely be removed with the TSS by physical means, and

effluent metals will probably be similar in ratio to the two

TSS effluent concentrations. The average pond effluent

had a TSS of 70 mg/l and the two FWS designs should

yield 30mg/L and 20mg/L, respectively. Therefore, design

examples 1 and 2 should remove about 70% of the heavy

metals. Since each metal reacts differently this type of

analysis has little meaning. Typically, nickel, boron, selenium and arsenic are more resistent to removal by sedimentation than most of the other commonly measured

metals. A similar discusion can be provided with regard to

semi-volatile toxic organic compounds. Both classes of

pollutants may be associated with certain effluent particlesize fractions, which will cause them to follow the removal

patterns discussed earlier in concert with Table 4-3.

4.6 Design Issues

This subsection describes issues that are important in

the design and layout of a FWS constructed wetland.

These design issues are separate from wetland area determinations already described. However, it is important

for the engineer/designer to understand that design issues

and wetland area determinations are both important. The

design issues outlined here are intended to maximize the

treatment potential of FWS constructed wetlands and may

impact the wetland area determined from the wetland sizing steps, which should be considered the starting point

for a FWS constructed wetland design. Many of the design issues outlined below are based on experience with

FWS constructed wetland systems currently in operation.

4.6.1 Wetland Layout

4.6.1.1 Site Topography

In many cases, the topography of the site will dictate the

general shape and configuration of the FWS constructed

wetland. On sloping sites, for example, constructing the

long dimension of the wetland parallel to the existing ground

contours helps minimize grading requirements. With proper

design, sloped sites can reduce pumping costs by taking

advantage of the existing hydraulic gradients.

4.6.1.2 Aspect (length to width) Ratio

The aspect ratio (AR) or length to width ratio (L/W), of a

FWS wetland system is defined as the average length divided by the average width, and can be expressed as:

 (4-13)

where:

L = average length of wetland system, and

W = average width of wetland system.

FWS constructed wetlands have been designed with ARs

from less than 1:1 to over 90:1. Generally, FWS constructed

wetlands are designed and built with an AR greater than

1:1. It has been suggested that wetlands with higher ARs

help to minimize short circuiting, and force the wetland to

more closely conform to plug flow hydraulics (Gearheart,

1996: Dombeck, 1998). However, results of dye studies

on existing FWS constructed wetlands have shown that

many wetlands deviate from ideal plug flow hydraulics independent of the AR. For wetland systems with very high

length to width ratios, careful consideration needs to be

given to headloss and hydraulic gradient considerations

to avoid overflows of confining dikes near the influent end.

Use of equation 4-6 and the material in section 4.3.2.2 will

permit the designer forced to use high AR cells to evaluate

each assumption and make corrections as necessary.

When conducting a hydraulic grade line analysis to determine if the backwater is at an acceptable elevation near

the inlet, the outlet level is normally assumed to be at the

midpoint.

4.6.1.3 Wetland configuration

The shape of a FWS constructed wetland can be highly

variable depending on site topography, land configuration,

and surrounding land use activities. FWS constructed wetlands have been configured in a number of shapes, including rectangles, polygons, ovals, kidney shapes, and

crescent shapes. There is no data that supports one FWS

constructed wetland shape as being superior in terms of

constituent removal and effluent quality, over another

shape. However, any wetland shape needs to be designed

and configured following the general guidelines of this report. Design concerns such as hydraulic retention time,

short-circuiting, headloss, inlet/outlet structures, and internal and surface configurations can significantly impact

wetland effluent quality.

4.6.1.4 Multiple cells

It has been shown in both the design of oxidation ponds

and FWS constructed wetlands, that a number of cells in

series can consistently produce a higher quality effluent.

This is based upon the hydrodynamic characteristics that

constituent mass is gathered at the outlet end of one cell,

and redistributed to the inlet of the next cell. This process

also minimizes the short circuiting effect of any one unit,

and maximizes the contact area in the subsequent cell.

For treatment and water quality purposes, it is recommended that a FWS constructed wetland should consist

of a minimum of three cells in series. Open water zones

have also been used to redistribute flows, but their value

in this regard has been overshadowed by their other attributes.

Large wetland cells can have internal berms running

parallel to the flow direction, effectively creating smaller

parallel cells with better hydraulic properties. Multiple cells

with appropriate piping between them offer greater operational flexibility. In the event that a wetland cell needs to be

taken off line for maintenance reasons, the remaining cell

or cells can remain operational. This is made even more

important if cells are sized to coincide with zoning. Completely vegetated and completely open cells are easier to

maintain and are more flexible when sequencing or independent cell HRT adjustments or maintenance is required.

AR = L

W
90

4.6.2 Internal Wetland Components

4.6.2.1 Open water/Vegetation ratio

The location of emergent vegetation, the type and density of this vegetation, and the climate as it relates to plant

senescence are important factors in the design of a FWS

constructed wetland. Providing adequate open water areas is an important, but often overlooked, component in

the design and implementation of FWS constructed wetlands (Gearheart, 1986; Hammer, 1996; Hamilton, 1994,

Stefan et al., 1995). Open water is defined as a wetland

surface which is not populated by emergent vegetation

communities, but may contain submergent aquatic plants

as well as unconsolidated groupings of floating aquatic

plants. Historically, many FWS constructed wetlands were

designed and built as fully vegetated basins with no designated open water areas. Many of these systems proved

problematic with very low or no water column dissolved

oxygen, that resulted in odor production and vector problems.

Natural wetlands generally contain a mix of open water

and emergent vegetation areas. The open water areas

provide many functions such as oxygenation of the water

column from atmospheric reaeration, submerged macrophytes, and algal photosynthesis. They also permit predation of mosquito larvae by fish and other animals and provide habitat and feeding areas for waterfowl. Open water

areas in FWS constructed wetlands will not only provide

the same functions as for natural wetlands, but will also

provide the opportunity for increased soluble BOD reduction and nitrification of wastewater because of the increase

in oxygen levels. It is recommended that a FWS constructed

wetland not be fully vegetated, but should include some

open water areas. Open water areas in a FWS constructed

wetland will result in a more complex, dynamic, and selfsustaining wetland ecosystem, that better mimics a natural wetland. Open water wetlands have lower background

BOD than fully vegetated wetlands (see Table 4-4) , which

reflects their improved treatment potential.

The ratio of open water to emergent vegetation depends

on land availability, costs, and the function and goals of

the FWS constructed wetland system. Generally speaking zones 1 and 3 should be 100% vegetated with the zone

2 surface having > 50% to 100% open water. If denitrification is required, the 3rd zone which is 100% vegetated will

accomplish it.. The open-water zones with an HRT in excess of 3 days may invite algal blooms. As long as this

zone is followed by a fully vegetated zone with an HRT of

2 or more days, this should not represent a problem beyond increased biomass management requirements.

The most effective method used for creating open water

areas in a single cell is to excavate a zone that is deep

enough to prevent emergent vegetation colonization and

migration. Some have periodically raised water levels to a

depth that limits emergent vegetation growth, but this is

operationally demanding and may have negative treatment

impacts. The type of dominant macrophyte (i.e., emergent or submergent) can be controlled by controlling the

operating depth. Water column depths greater than approximately 1.25 to 1.5 meters planted with submergents

such as Potomogeton spp., will not rapidly be encroached

upon by emergent macrophytes like bulrushes reeds, and

cattails. If the water column depth is between 0.5 to 1.0

meters and planted with emergent vegetation, such as

bulrush and cattails, they will prevail over submergents

and most other emergents by filling in the surface area

through rhizome and tuber propagation. The seasonal

change in water levels (hydroperiod) is also a determinant

in establishing various aquatic macrophyte communities.

Due to the lack of shading, significant blooms of algae

can occur in large open water areas, which can have negative effects on effluent quality. To help minimize the potential for algal growth, open water areas should be designed

for less than 2 to 3 days hydraulic retention time. In general, the growth cycle of algae is approximately 7 days, so

providing open water areas with less than 2-3 days retention time will help minimize algal growth in the open-water

zone of the wetland. Sufficient standing crops of

submergent macrophytes may also limit algal regrowth in

these zones. Conversely, excessive algal growth may impair the performance of the submergent macrophytes by

limiting the solar energy which reaches them.

Guidelines for designing a FWS constructed wetland in

terms of vegetated covering are as follows: Begin with an

emergent vegetation zone covering the volume used in

the first 2 days of retention time at maximum monthly flow

(Qmax) to provide for influent solids flocculation and separation.. The emergent zone should be followed by an open

water zone covering days 3 and 4 in the retention time

sequence at Qmax. The open water zone should be designed

to facilitate production of dissolved oxygen to meet CBOD

and NBOD demands. The final 2 days of hydraulic retention volume at Qmax should be an emergent wetland to

reduce any solids (algae, bacteria, etc.) generated in the

open water and to supply carbon (decomposing plant material) and anoxic conditions for denitrification. It is also

recommended that this final stand of emergent vegetation

be as close as possible to the outlet of a FWS constructed

wetland. This provides a final level of protection just before the effluent leaves the wetland to minimize the impact

of wildlife on effluent quality. This recommendation may

heighten the maintenance requirements for the outlet device, but it will result in less variability in the effluent quality.

4.6.2.2 Inlet Settling Zone

Depending on the pretreatment process, a substantial

portion of the incoming settleable and suspended solids

may be removed by discrete settling in the inlet region of a

FWS constructed wetland. For example, if the FWS system is to follow an existing treatment facility which is prone

to produce high concentrations of settleable TSS, an inlet

settling zone should be used. If the pretreatment facility

does a good job of solids capture, but has a high concentration of soluble constituents, an inlet settling zone is of
91

no value. Given that FWS systems are generally preceded

by lagoon systems which seasonally produce large TSS

concentrations, primarily due to algal mass, the need for

an inlet settling zone would be marginal since algal solids

will require flocculation and sunlight restriction before becoming settleable.

If an inlet settling zone should be desired, it should be

constructed across the entire width of the wetland inlet. A

recommended guideline is to design a settling zone which

provides approximately 1 day hydraulic retention time at

the average wastewater flow rate, as most settleable and

suspended solids are removed within this time period. The

settling zone should be deep enough to provide adequate

accumulation and storage of settled solids, but shallow

enough to allow the growth of emergent vegetation, such

as bulrush and cattails. Recommended depth is approximately 1 meter.

Most accumulated organic solids will slowly decay and

reduce in volume. This decay is one of the two major

sources of internal loading and background constituents

in the effluent. However, at some time in the future the

remaining accumulated solids will need to be removed from

the settling zone.

4.6.2.3 Inlet/Outlet Structures

Placement and type of inlet and outlet control structures

are critical in FWS constructed wetlands to ensure treatment effectiveness and reliability. To effectively minimize

short-circuiting in a FWS constructed wetland, two goals

concerning cell inlet/outlet structures are critical: (1) uniform distribution of inflow across the entire width of the

wetland inlet, and (2) uniform collection of effluent across

the total wetland outlet width. Both of these should minimize localized velocities, thus reducing potential

resuspension of settled solids. It is important that any outlet structure be designed so that the wetland can be completely drained, if required. Some of the common types of

wetland inlet/outlet systems in use today, and general

guidelines regarding their design are further discussed in

Chapter 6.

Depending on the type of wastewater influent, the inlet

structure discharge point could be located below or above

the wetland water surface. Perforated pipe inlet/outlet structures can be difficult to operate and maintain when they

are fully submerged. All inlet distribution systems should

be accessible for cleaning and inspection by using

cleanouts.

Outlet structures represent an operational control feature that directly affect wetland effluent quality. It is important that outlet structures facilitate a wide range of operating depths. By adjusting the outlet structure, both the water depth and hydraulic retention time can be increased or

decreased. This and the need to accommodate cell drainage usually results in locating the outlet manifold at the

bottom of the outlet zone. The differences in water quality

between water depths can also be highly variable. An outlet

structure design which allows for maximum flexibility of

collection depths may be desirable, but may not always

be compatible with collection devices that minimize shortcircuiting. With this type of design, the outlet structure can

be adjusted to draw wetland effluent from the water depth

with the best water quality. This alternative design usually

involves multiple drop boxes with openings at different

depths. In most cases however the uniform collector set

at the bottom is favored owing to its inherent advantages

in terms of improved effluent quality and facilitation of cell

drainage.

Two types of inlet/outlet structures are commonly used

in FWS constructed wetlands. For small or narrow (high

AR) wetlands, perforated PVC pipe can be used for both

inlet and outlet structures. The length of pipe should be

approximately equal to the wetland width, with uniform

perforations (orifices) drilled along the pipe. The size of

the pipes, and size and spacing of the orifices will depend

on the wastewater flow rate and the hydraulics of the inlet/

outlet structures. It is important that the orifices be large

enough to minimize clogging with solids. Perforated pipes

can be connected to a manifold system by a flexible tee

joint, which allows the pipes to be adjusted up or down. In

some cases wetland designers with this type of inlet/outlet

structure will cover the perforated pipes with gravel to provide more uniform distribution or collection of flows. This

type of inlet/outlet structure requires periodic inspection,

some operation and maintenance to maintain equal flow

through the pipe, and access at the end to clean clogged

orifices.

For larger wetland systems, multiple weirs or drop boxes

are generally used for inlet and outlet structures. Weirs or

drop boxes are generally constructed of concrete, but

smaller PVC boxes are also available. These structures

should be located no further apart than every 15 m (center

to center) across the wetland inlet width, with a preferred

spacing of 5 to 10 m. The same spacing requirements apply

for the outlet weirs or drop boxes. Depending on the source

of the wastewater influent, the inlet weirs or drop boxes

can be connected by a common manifold pipe. Whatever

the configuration, it is important that the manifold pipes

and weirs be hydraulically analyzed to attain reasonably

uniform distribution. Simple weir or drop box type inlet structures are relatively easy to operate and maintain.

Weir overflow rates have not been considered in the

design of most wetlands. Weir loading rates of existing

wetlands are significantly higher than those required in

most biological solids removal processes (i.e., 120 to 190

m3

/m.d ) (WEF, 1998). Excess weir rates can cause high

water velocities near the outlet which could entrain solids

which would otherwise be removed from the effluent.

Therefore, weir loading rates should be designed to meet

the above range for best performance until more quantitative data are generated.

4.6.2.4 Baffles

Baffles are internal structures installed either perpendicular or parallel to the direction of flow. Baffles can be
92

effective in reducing short-circuiting, for mixing waters of

different depths, and for improved flocculation performance.

Properly designed and placed open water zones can also

act as baffles by allowing mixing and redistribution of wastewater before it enters into the next wetland vegetated zone.

The use of baffles depends on cell configurations, aspect

ratios, treatment goals, and permit compliance. In general,

except for special circumstances unforseen in typical municipal wastewater treatment application the use of such

structures is not recommended. However, their use in correcting problems which are due to hydraulic flow difficulties (short circuiting, dead zones, etc.) make them useful

to the operator-owner.

4.6.2.5 Recirculation

Recirculation is the process of introducing treated effluent back to the inlet or to some other internal location of

the wetland. Recycling effluent can decrease influent constituent concentrations and increase dissolved oxygen

concentrations near the inlet. The increased dissolved

oxygen concentrations can help reduce inlet odors, lower

BOD, and enhance nitrification potential in open-water

zones. If recirculation is to be considered, the effects of

recirculation on the wetland water balance and wetland

hydraulics need to be analyzed. In general, the ability to

recycle, like the ability to drain each cell, could be considered part of the need to have flexible piping , multiple cells,

and multiple trains of cells. The value of recirculation has

not been shown to date to be a major factor in improving

FWS performance.

4.6.2. 6 Flow Measuring Devices

Many existing wetland systems do not have accurate

flow measuring devices. Even if accurate estimates of inflows and/or outflows to the treatment plant are known,

internal flow distribution to individual wetland cells is not

known or measured. Without accurate flow measurements

to individual wetland cells, it is impossible to determine

internal flow rates, average velocities, and hydraulic retention times for each cell, thus making system performance adjustments difficult. It is recommended that some

type of flow measuring device be either installed in or available to be installed in each cell of a FWS constructed wetland. This includes separate flow measuring devices on

each inlet for multiple wetland cell configurations. Some

examples of flow measuring devices include simple 90o Vnotch or rectangular weirs, and more sophisticated Parshall

flumes for larger systems. Depending on the size and layout of the wetland, cell inlet/outlet structures should be

designed to be compatible with available flow measuring

devices.

4.6.3 Pretreatment Requirements

Examples of treatment that should precede FWS constructed wetlands include all types of stabilization ponds

and primary sedimentation systems. The use of wetlands

to polish secondary effluent to less than 10 mg/l BOD and

TSS has been documented, but is not covered in detail

here. The reader is directed to USEPA (1999) for guidance in these applications. The effluent entering a FWS

constructed wetland should be free from floatable and large

settleable solids, and excessive levels of oil and grease.

Also important to a FWS constructed wetland is the incoming metal concentrations. While a FWS constructed

wetland does remove and immobilize many heavy metals

along with the TSS, excessive influent concentrations could

result in residuals which are unacceptable for subsequent

land application. A source reduction program and/or an

industrial waste pretreatment ordinance are required if

excessive metals concentrations are present in the raw

wastewater.

4.7 Construction/Civil Engineering Issues

Specific construction/civil engineering design issues that

should be considered early in the planning and design

phase of a FWS constructed wetland project include site

topography and soils, berm construction, impermeable liner

materials, wetland vegetation substrate, and internal drainage. Many of these issues should be considered during

the site selection process, as they may become difficult or

costly to correct later in the actual design and construction

phases of the project. The construction/civil engineering

requirements for a FWS constructed wetland are similar

to other earthen water quality management systems such

as oxidation ponds, and are discussed in Chapter 6 and in

USEPA (1983) and Middlebrooks, et al (1982).

4.7.1 Site Topography and Soils

In general, level land with clay soils affords the optimal

physical setting for a FWS constructed wetland. Potential

wetland sites with other physical conditions can be used,

but may require more substantial engineering, earthwork,

construction requirements, and liners. In order to overcome

site limitations, the cost of a FWS constructed wetland will

also increase proportionally as the wetland site further

deviates from optimal site conditions.

FWS constructed wetlands can be built on sites with a

wide range of topographic relief. Construction costs are

lower for flat sites since sloped sites require more grading

and berm construction. Site topography will generally dictate the basic shape and configuration of the FWS constructed wetland.

The principal soil considerations in siting and implementing a FWS constructed wetland are the infiltration capacity

of the soils and their suitability as berm material and wetland vegetation substrate. In most cases FWS constructed

wetlands are required to meet stringent infiltration restrictions depending on the state regulations for groundwater

protection. An exception are wetland systems designed to

incorporate infiltration as part of the treatment and discharge process. In these cases, the underlying soil must

have infiltration rates compatible with the design discharge

rates. If native site soils are not suitable, separate infiltration trenches can be added to increase the infiltration surface area. In some cases, it will be necessary to import

berm and/or bottom materials or use synthetic liners (see

Chapter 6) to prevent infiltration.
93

Interior berms containing FWS wetland cells should be

built with up to 3:1 side slopes as the soil characteristics

allow. A minimum freeboard of 0.6 m above the peak flow

operating depth in the wetland is required. For wetlands

that will receive exceptionally high peak inflows, additional

freeboard may be required to ensure that berm overtopping does not occur. Additional freeboard may also be designed to accommodate long-term solids and peat buildup

during the operation of the wetland, and to allow appropriate water depths to be maintained as sludge builds up in

the initial cells over time.

All FWS-cell external berms should have a minimum top

width of 3 m, which provides an adequate road width for

most standard service vehicles. In some cases, internal

berms can have smaller top widths, as routine operation

and maintenance can be carried out by small motorized

vehicles, such as ATVs. Road surfaces should be an all

weather type, preferably gravel, which also minimizes direct runoff into the wetland.

Berm integrity is critical to the long term operational effectiveness of FWS constructed wetlands. Common berm

failure causes include burrowing by mammals, such as

beaver nutria and muskrat, and holes from root penetration by trees and other vegetation growing on or near the

berms. Several design features can eliminate and/or minimize these problems. A thin impermeable wall, or internal

layer of gravel, can be installed during construction, which

will minimize mammal burrowing and/or root penetration.

Planting the berm using vegetation with a shallow root system can also be effective. Unlike oxidation ponds, berm

erosion in fully vegetated zones and/or cells from wave

action is generally not a concern due to the dampening

effect of the wetland vegetation. However, in larger cells

with open zones it could be an issue, and stabilization pond

texts should be consulted for solutions (Middlebrooks, et

al, 1982)(USEPA, 1983).

In the design and site selection process, an important

consideration is the amount of additional area required for

berms. In general, the higher the aspect ratio for a FWS

constructed wetland, the more area that will be required

for the berms and for the entire wetland system. This increase in required total wetland area to accommodate

berms is more pronounced for smaller wetlands than for

larger wetlands. A factor of 1.2 to 1.4 times the cell area is

usually employed to determine the total site area for the

FWS system.

4.7.2 Impermeable Liner Materials

A concern with FWS constructed wetlands is the potential loss of water from infiltration and contamination of

groundwater below the wetland site. While there are some

wetland applications where infiltration is desirable, the

majority of the applications require some type of barrier to

prevent groundwater contamination. Under ideal conditions, the wetland site will consist of natural soils with low

permeability that restrict infiltration. However, many wetlands have been constructed on sites where soils have

high permeability. In these cases, some type of liner or

barrier will likely be required to minimize infiltration. Liner

requirements can also add significantly to the construction

cost of a FWS constructed wetland.

Existing natural soils with permeability less then approximately 10-6 cm/s are generally adequate as an infiltration

barrier. For site soils with higher permeabilities, some type

of liner material will likely be required. Some examples of

wetland liner materials include imported clay fill, bentonite

soil layers, chemical treatment of existing soils, asphalt,

and synthetic membrane liners such as PVC or HDPE. In

some instances, it will be possible to compact the existing

site soils to acceptable permeability. Due to their ability to

be placed in shaped wetland cells, clay liners are generally a more sustainable component of the wetland than

synthetic membrane liners. Whatever liner material is chosen, an important consideration is to provide adequate soil

cover and depth that protects the liner from incidental damage and root penetration from the wetland vegetation (see

Chapter 6).

4.7.3 Soil Substrates for Plants

Aquatic macrophytes generally reproduce asexually by

tuber runners. Soils with high humic and sand components

are easier for the tubers and runners to migrate through,

and plant colonization and growth is more rapid. The soil

substrate for wetland vegetation should be agronomic in

nature (e.g. loam), well loosened, and at least 150 mm (6

inches) deep. Depending on the liner material, deeper soil

substrates may be required to protect the liner. If this type

of soil layer exists at the site, it should be saved. After the

wetland basin, berms and other earthen structures are

constructed, and the liner is installed (if required), the original soil substrate can be placed back into the excavated

region. To meet soil specifications, it may be necessary to

amend the saved soils with other materials.

While soils such as loam and silt are good for plant

growth, they can allow large vegetation mats to float when

large water level fluctuations occur in the wetland. Floating vegetation mats can significantly alter the treatment

capabilities of FWS constructed wetlands by allowing

wastewater to flow between the floating mats and substrate,

not in contact with any vegetation treatment media. To circumvent this potential problem, denser soil substrates such

as a sandy loam, or a loam gravel mix can be used. This

will be more important in FWS constructed wetlands where

large water depth fluctuations will be part of the operation

and maintenance procedure.

4.7.4 Internal Drainage and Flexible

Piping

In the event a FWS constructed wetland needs to be

drained, the wetland bottom should have a slope of 1% or

less. Drainage may be required for maintenance reasons

such as liner repair, sludge removal, vegetation management, and berm repair. Deeper channels may be employed

to allow for drainage and/or continued use when serial cells

are taken out of service. Channels can also be used to
94

connect deeper open water areas where these are part of

a larger cell, rather than separate cells. In general the more

complete the intercellular piping, the greater the operational flexibility is for the entire system.

4.8 Summary of Design Recommendations

A summary of the design recommendations for FWS

wetland treatment systems is presented in Table 4-7. As

more quality-assured data become available allowable

pollutant areal loadings will likely be revised.

4.9 References

Balmer, P. and B.Vik.. 1978. “Domestic Wastewater Treatment with Oxidation Ponds in Combination with Chemical Precipitation,“ Prog. Water Tech, Vol 10, No. 5-6,

pp867-880.

Carre, J., M. P. Loigre, and M. Leages 1990. “Sludge Removal from Some Wastewater Stabilization Ponds.”

Water Science Technology, Vol 22, No 3-4, pp 247-

252.

Cole, S. 1988. "The Emergence of Treatment Wetlands".

ES&T. Vol 3, No. 5, pp. 218-223.

Crites, R.W., and G. Tchobanoglous. 1998. “Small and Decentralized Wastewater Management Systems, WCB

– McGraw-Hill, NY.

Dombeck, G. 1998. Sacramento Regional Wastewater

Treatment Plant Demonstration Wetland Project. 1997

Annual Report, Nolte and Associates, Sacramento, Ca.

Frankenbach, R.I and J.S Meyer. 1999. Nitrogen Removal

in A Surface Flow Wetland Wastewater Treatment Wetlands, 1999, Wetlands, Volume 1a, No. 2, June 1999

pp.403-412.

Gearheart, R.A., and B. Finney. 1999. “The Use of Free

Surface Constructed Wetlands as An Alternative Process Treatment Train to Meet Unrestricted Water Reclamation Standards”, Wat. Sci. Tech. Vol. 40, No. 4-5,

pp. 375-382.

Gearheart, R.A., B. A. Finney, M. Lang, and J. Anderson.

1998. “A Comparison of System Planning, Design and

Sizing Methodologies for Free Water Surface Constructed Wetlands”. 6th International Conference on

Wetland Systems for Water Pollution Control.

Gearheart, R.A. and B. A. Finney. 1996. Criteria for Design of Free Surface Constructed wetlands Based

Upon a Coupled Ecological and Water Quality Model.

Presented at the Fifth International Conference on

Wetland Systems for Water Pollution Control, Vienna,

Austria.

Table 4-7. Recommended Design Criteria for FWS Constructed

Wetlands

Parameter Design Criteria

Effluent Quality BOD ≤ 20 or 30 mg/L

TSS ≤ 20 or 30 mg/L

Pretreatment Oxidation Ponds (lagoons)

Design Flows Qmax (maximum monthly flow) and

Qave (average flow)

Maximum BOD Loading 20 mg/L: 45 kg/ha-d

(to entire system) to Meet: 30 mg/L: 60 kg/ha-d

Maximum TSS Loading 20 mg/L: 30 kg/ha-d

(to entire system) to Meet: 30 mg/L: 50 kg/ha-d

Water Depth 0.6 - 0.9 m Fully vegetated zones

1.2 - 1.5m Open-water zones

1.0m Inlet settling zone

(optional)

Minimum HRT (at Qmax) 2 days fully vegetated zone

in Zone 1 (and 3)

Maximum HRT (at Qave) 2 - 3 days open-water zone

in Zone 2 (climate dependent)

Minimum Number of Cells 3 in each train

Minimum Number of Trains 2 (unless very small)

Basin Geometry (Aspect Ratio) Optimum 3:1 to 5:1, but subject to

site limitations

AR > 10:1 may need to calculate

backwater curves

Inlet Settling Zone Use Where pretreatment fails to retain

settleable particulates

Inlet Uniform distribution across cell inlet

zone

Outlet Uniform collection across cell outlet

zone

Outlet Weir Loading ≤200 m3/m-d

Vegetation

Emergent - Typha or Scirpus (native species

preferred)

Submerged - Potamogeton, Elodea, etc (see

chapter 2).

 (continued)

Table 4-7. Continued

Parameter Design Criteria

Design Porosities 0.65 for dense emergents in fully

vegetated zones

0.75 for less dense stand of

emergents in same zones

1.0 for open-water zones

Cell Hydraulics Each cell should be completely

drainable

Flexible intercell piping to allow for

required maintenance

Independent, single-function cells

could maximize treatment
95

Gearheart, R. A. 1995. Watersheds - Wetlands - Wastewater Management. In Natural and Constructed Wetlands

for Wastewater Treatment. Ramadori, R., L. Cingolani,

and L. Cameroni, eds., Perugia, Italy, pp 19-37.

Gearheart, R. A. 1993. “Phosphorus Removal in Constructed

Wetlands”. Presented at the 66th WEF Conference and

Exposition, Anaheim, Ca.

Gearheart, R.A. 1992. Use of Constructed Wetlands to Treat

Domestic Wastewater, City of Arcata, California”, Wat.

Sci. Tech., Vol. 26, No. 7-8, pp. 1625-1637.

Gearheart, R.A., F. Klopp, and G. Allen. 1989. Constructed

Free Surface Wetlands to Treat and Receive Wastewater Pilot Project to Full Scale, In D.A. Hammer (ed.) Constructed Wetlands for Wastewater Treatment, pp. 121-

137, Lewis Publisher, Inc., Chelsea, MI

Gearheart, R.A., B. A. Finney, S. Wilbur, J. Williams, and D.

Hall. 1984. “The Use of Wetland Treatment Processes

in Water Reuse”, Future of Water Reuse, Volume 2, Proceedings of Symposium III Water Reuse, AWWA Research Foundation, pp. 617-638.

Gearheart, R.A., S. Wilbur, J. Williams, D. Hull, B. A. Finney,

and S. Sundberg. 1983. City of Arcata Marsh Pilot

Project: effluent quality results-system design and management. Final report. Project No. C-06-2270, State

Water Resources Control Board, Sacramento, CA.

Gregg, J., and A. Horne. 1993. "Short-term Distribution and

Fate of Trace Metals in a Constructed Wetland Receiving Treated Municipal Wastewater", Environmental Engineering and Health Sciences Laboratory Report No.

93-4. University of California, Berkeley, CA.

Hammer, D, 1992,"Creating Freshwater Wetlands", Lewis

Publishers, Chelsea, MI

Hannah, S.A, B.M. Austern, A.E. Eralp, and R.H. Wise, 1986.

Journal WPCF, Vol 55, No 1, pp 27-34.

Hovorka, R.B. 1961. An Asymmetric Residence-time Distribution Model for Flow Systems, Dissertation, Case Institute of Technology.

Kadlec. R.H. 2000. The Inadequecy of First-Order Treatment

Wetland Models. Ecological Engineering, vol. 15, pp 105-

109;

Kadlec, R.H. and R.L. Knight. 1996. Treatment Wetlands.

Boca Raton, FL: Lewis-CRC Press.

Kadlecik, L. 1996. Organic Content of Wetland Soils, Arcata

Enhancement Marsh, Special Project, ERE Department

Wetland Workshop.

Levine, A. D., G. Tchobanoglous and T. Asano, 1991. Size

Distributions of Particulate Contaminants in Wastewater and Their Impact on Treatability. Water Research,

Vol. 25, No 8, pp. 911-922.

Linsley, R.K. Jr., M.A. Kohler, and J.L.H. Paulhus, 1982.

Hydrology for Engineers, 3rd Ed., McGraw-Hill, NY.

Mara, D.D. 1975 Proposed Design for Oxidation Ponds in

Hot climates. Journal ASCE-EE , Vol.101, No 2, pp 296-

300.

Marais, C. V. R., and V. A. Shaw. 1961. A Rational Theory for

the Design of Sewage Stabilization Ponds in Central and

South Africa. Transactions South African Institute of Civil

Engineers, Vol. 3, pp. 205ff.

Middlebrooks, E. J., C. E. Middlebrooks, T. H. Reynolds, G.Z.

Watters, S.C. Reed, and D.B.George. 1982. Wastewater Stabilization Lagoon Design, Performance and Upgrading, MacMillen, New York, NY.

Mitsch, W.J. and J.G. Gosselink.1993. Wetlands. Van

Nostrand Reinhold, NY.

NADB (North American Treatment Wetland Database). 1993.

Electronic database created by R. Knight, R. Ruble, R.

Kadlec, and S. Reed for the U.S. Environmental Protection Agency. Cincinnati, OH.

Odegaard, H. 1987. Particle Separation in Wastewater Treatment. In Proceedings EWPCA 7th European Sewage

and Refuse Symposium, pp. 351- 400.

Reckhow, K., and S. S. Qian. 1994. Modeling Phosphorus

Trapping in Wetlands Using General Models. Water

Resources Research, Vol. 30, No. 11, pp. 3105-3114.

Reddy, K. R., and W.R. De Busk. 1987. Nutrient Storage

Capabilities of Aquatic and Wetland Plants for Water

Treatment and Resource Recovery. Magnolia Pub., Inc.,

Orlando, FL.

Reed, S. C., R. Crites, and E. J. Middlebrooks, 1995 ,"Natural Systems for Waste Management and Treatment,

McGraw-Hill, San Francisco, Ca.

Reed, S.C., R.W. Crites, and E.J. Middlebrooks. 1995. Natural Systems for Waste Management and Treatment. 2nd

Ed., McGraw-Hill, NY.

Sartoris, J., J. Thullen, L. Barber, and D. Salas. 1999. Investigation of Nitrogen Transformations in a Southern California Constructed Wastewater Treatment Wetland, Ecological Engineering, Vol 14, pp. 49-65.

SCRSD. 1998. Constructed Wetlands Demonstration Project.

1997 Annual Report, Sacramento, CA.

Sparham, V.R. 1970. “Improved Settling Tank Efficiency by

Upward Flow Clarification,” JWPCF, Vo. 42, No 5, pp

801-811.

Tchobanoglous, G., R. W. Crites, R.A. Gearheart, and S. C.

Reed. 2000. A Review of Treatment Kinetics for Constructed Wetlands. Presented to WEF Specialty Conference Disinfection, 2000. New Orleans, LA.
96

Tchobanoglous, G., F. Meitski, K. Thompson and T.H.

Chadwick. 1989. Evolution and Performance of

City of San Diego Pilot-Scale Aquatic Wastewater Treatment System Using Water Hyacinths.

Research Journal WPCF, Vol 61, No. 11-12, pp

1625-1655.

Tchobanoglous, G. and E. D. Schroeder. 1985. Water Quality: Characteristics, Modeling, Modification. Addison-Wesley, Reading, MA.

United Kingdom, Dept. of Environment. 1973. Treatment of Secondary Sewage Effluent in Lagoons.

Notes on Water Pollution #63. London, U.K.

United States Environmental Protection Agency.

1999. FWS Wetlands for Wastewater Treatment:

A Technology Assessment. EPA 832/R-99/002.

Office of Water, Washington, DC.

U.S. Environmental Protection Agency. 1988. Design

Manual. Constructed Wetlands and Aquatic Plant

Systems for Municipal Wastewater Treatment.

EPA/625/1-88/022. Cincinnati, OH.

U.S. Environmental Protection Agency, 1983. Design

Manual - Municipal Wastewater Stabilization

Ponds. EPA/625/1-83-015. Cincinnati, OH.

U.S. Environmental Protection Agency, 1993.

Manual: Nitrogen Control. USEPA Publication No.

EPA/625/R-93/010. Cincinnati, OH.

Water Environment Federation, 1998. Design of Municipal Wastewater Treatment Plants, 4th Ed,

MOP#8. Alexandria, VA.

Water Environment Federation, 1990. Natural Systems for Wastewater Treatment. MOP FD-16, Alexandria, VA.
97

5.1 Introduction

The pollutant removal performance of vegetated submerged bed (VSB) systems depends on many factors including influent wastewater quality, hydraulic and pollutant loading, climate, and the physical characteristics of the

system. The main advantage of a VSB system over a free

water surface (FWS) wetland system is the isolation of the

wastewater from vectors, animals and humans. Concerns

with mosquitoes and pathogen transmission are greatly

reduced with a VSB system. Properly designed and operated VSB systems may not need to be fenced off or otherwise isolated from people and animals. Comparing conventional VSB systems to FWS systems of the same size,

VSB systems typically cost more to construct, primarily

because of the cost of media (Reed et al., 1985). Because

of costs, it is likely that the use of the conventional VSB

systems covered in this manual will be limited to individual

homes, small communities, and small commercial operations where mosquito control is important and isolation fencing would not be practical or desirable.

A conventional VSB system is described in Chapter 2

and depicted in Figure 2-3. The typical components include (1) inlet piping, (2) a clay or synthetic membrane

lined basin, (3) loose media filling the basin, (4) wetland

vegetation planted in the media, and (5) outlet piping with

a water level control system. The vast majority of VSB systems have used continuous and saturated horizontal flow,

but several systems in Europe have used vertical flow.

Alternative VSB systems are defined here as VSBs that

have been modified to improve their treatment performance

(George et al., 2000, Young et al., 2000, Behrends et al.,

1996). Typical modifications involve some type of cyclic

filling and draining of the system to improve the oxygen

input into the media. The potential improvement in performance with alternative VSB systems is offset to some degree by a more complex and expensive operating system.

It is too early to predict whether alternative VSB designs

will prove to be more cost effective or practical than conventional VSB systems, although they appear to provide

significantly better removal of certain pollutants.

This chapter will discuss VSB systems that treat (1) septic

tank and primary sedimentation effluents, (2) pond effluents, and (3) secondary and non-algal pond effluents. The

Chapter 5

Vegetated Submerged Bed Systems

most common VSB systems in the U.S. treat septic tank

and pond effluents for BOD and TSS removal. In Europe,

VSB systems are most often used to treat septic tank effluents, although they have also been used extensively in

the U.K. for polishing activated sludge and RBC effluents,

and for treating combined sewer bypass flows (Cooper,

1990, Green and Upton, 1994).

This chapter provides a summary of the theoretical and

practical considerations in the design of conventional VSB

systems. VSB systems, like other natural treatment systems, are less understood than highly-engineered waste

treatment systems because they (1) have more variable,

complex, and less controllable flow patterns, (2) have reaction rates and sites within the system that vary with time

and location, and (3) are subject to the inconsistencies of

climate and growth patterns. This complexity makes the

development and use of design equations based on idealized reactor and reaction kinetic theory difficult, if not impractical and unrealistic. Furthermore, because pollutant

removal performance can be quite variable, designs must

be conservative if a guaranteed effluent quality is required.

5.2 Theoretical Considerations

5.2.1 Potential Value of Wetland Plants in

VSB Systems

In several recent studies that have compared the pollutant removal performance of planted and unplanted VSB

systems, it has been found that plants do not have a major

impact on performance (Young et al, 2000, George et al.,

2000, Liehr et al., 2000). There is however significant cost

and time associated with the establishment and maintenance of the wetland plants in a VSB system. Nevertheless, planted systems have a significant aesthetic advantage over unplanted systems and may be of value as wetland habitat in some cases. Unfortunately, the aesthetic

value of plants and the value of VSB wetland systems as

wetland habitat are difficult to quantify, and no mitigation

credit is given by the USEPA for the habitat value they

provide. In the following sections the potential value of

wetland plants in VSB systems is discussed in more detail.

5.2.1.1 Type of Wetland Plants

Several studies have attempted to determine if pollutant

removal performance differs with various types of wetland
98

plants (Gersberg et al., 1986, Young et al., 2000). Although

some researchers have claimed a relationship, these

claims have not been substantiated by others (Gersberg

et al., 1986).

It is not clear if it is desirable to maintain a single plant

species, or a prescribed collection of plant species, for any

treatment purpose. Single plant (monoculture) systems are

more susceptible to catastrophic plant death due to predation or disease (George et al., 2000). It is generally assumed that multiple plant and native plant systems are

less susceptible to catastrophic plant death, although no

studies have confirmed this assumption. Plant invasion and

plant dominance further complicate the issue; in several

cases researchers have found that, with time and without

operator intervention, one of the planted species or an invader species has become the dominant species in all or

part of the system (Young et al., 2000, Liehr et al., 2000).

This occurs less frequently and more slowly in VSB systems than in FWS systems.

The impact of wetland plants on pollutant removal performance appears to be minimal based on current knowledge, so the selection of plants species should be based

on aesthetics, impacts on operation, and long-term plant

health and viability in a given geographical area. Local

wetland plants experts should be consulted when making

the selection.

5.2.1.2 Plant Mediated Gas Transfer

Wetland plants can facilitate gas transfer both into and

out of the wastewater of a VSB system. The focus of most

studies has been oxygen transfer into the wastewater.

However, methane and other dissolved gases in the wastewater can be transferred out of the wastewater by wetland

plants. The mechanisms of plant-mediated gas transfer

are described in detail in Chapter 3. The potential amount

of oxygen transferred by plant roots into the wastewater

depends on many factors including dissolved oxygen concentration in the wastewater, root depth in the wastewater,

air and leaf temperatures, and plant growth status (rapid

growth vs. senescence). Most studies to determine the

rates of plant-mediated oxygen transfer have been performed in laboratory microcosms or mesocosms under

controlled conditions.(George et al., 2000, Liehr, et al.,

2000) It is not clear if these results are transferable to fullscale systems.

Based on a review of the literature, the likely rate of oxygen transfer is between zero and 3.0 g-O2

/m2-d (0 – 0.6

lbs/1000 ft2-d). While this maximum value is within the BOD

loading range of lightly loaded VSB systems, (3 g/m2

-d =

30 kgBOD/ha-d = 27 lb BOD/ac-d), there is very little evidence to support the assumption that plants add significant amounts of oxygen to VSB systems. Typical values

of dissolved oxygen in VSB systems are very low (<1.0mg/

L), but because of the difficulty in obtaining an accurate insitu oxygen reading, the actual values are probably even

lower. In VSB systems where oxidation-reduction potential (ORP) has been measured, values were typically quite

negative, indicating strong reducing conditions.

Unplanted systems have been found to perform as well

as planted systems in both BOD and ammonia nitrogen

removal (George et al., 2000, Liehr et al., 2000, Young, et

al., 2000). Furthermore, investigations of root depth and

flow pathways have found that the roots do not fully penetrate to the bottom of the media and there is substantially

more flow under the root zone than through it (Young et

al., 2000, George et al., 2000, Bavor et al. 1989, Fisher,

1990, DeShon et al., 1995, Sanford et al., 1995a & 1995b,

Sanford, 1999, Rash and Liehr, 1999, Breen and Chick,

1995, Bowmer, 1987). The oxygen supply from the roots

is also likely to be unreliable due to yearly plant senescence, plant die-off due to disease and pests, and variable plant coverage from year to year. Considering all of

these factors, it is recommended that designers assume

wetland plants provide no significant amounts of oxygen

to a VSB system.

Plants will also affect the other potential source of oxygen to VSBs — the direct oxygen transfer from the atmosphere to the wastewater. Researchers at TVA have estimated oxygen transfer from the atmosphere to be between

0.50 and 1.0 g-O2

/m2-d (0.1 – 0.2 lbs/1000 ft2-d) (Behrends

et al., 1993). Decomposing plant matter on top of the media would likely cause even lower rates of oxygen transport into the wastewater because the plant matter acts as

a diffusion barrier and, ultimately, an oxygen demand.

5.2.1.3 Nutrient and Metals Removal by

Wetland Plants

Wetland plants take up macro-nutrients (such as N and

P) and micro-nutrients (including metals) through their roots

during active plant growth. At the beginning of plant senescence most of the nutrients are translocated to the rhizomes and roots. A significant proportion of the nutrients

may also be exuded from the plant (Gearheart et al., 1999).

Estimates of net annual nitrogen and phosphorus uptake

by emergent wetland species vary from 12 to 120 gN/m2-y

and 1.8 to 18 gP/m2

-y respectively (Reddy and DeBusk,

1985). Reeds and bulrush are at the lower end of both

ranges while cattails are at the higher end. These estimates are based on annual growth rates and nutrient concentrations of the whole plant, but since in a VSB system

only the shoots can be harvested, the values should be

reduced by at least 50%. Plant uptake of metals can also

be estimated by this method. To maximize nutrient removal

by the plants in a VSB system, shoot harvesting must be

done before senescence. Harvesting of wetland plants is

not recommended during the growing season because the

warm temperatures may cause plant stress, substantial

stem death, and significant delay in re-growth in some

wetland plants (George, et al., 2000).

The expected maximum removal rates of nitrogen, phosphorus and metals by direct plant uptake and harvesting

are small compared to typical loadings in VSB systems.

Furthermore, nitrogen, phosphorus and metals removal by

plant uptake will vary with time. Most of the nutrient uptake occurs during rapid plant growth in the spring and

summer, and if the plant is not harvested before senes-
99

cence a significant portion of the plant-sequestered nutrients are released back into the water. Therefore, unless

the nutrient removal standards for a VSB system are also

variable and synchronous with plant uptake and release,

the presence of plants may be more harmful than helpful

in meeting nutrient removal standards. Finally, it is unlikely

that the nutrients or metals removal obtained by harvesting are worth the considerable time and labor required to

harvest and reuse or dispose of the biomass.

5.2.1.4 Plant-Supplied Carbon Sources for

Denitrification

Because of the inherent anaerobic conditions associated with VSB systems, they are good candidates for denitrification. The likely limiting factor for denitrification in VSB

systems is biodegradable organic carbon. The value of

plant-supplied organic carbon for denitrification in a VSB

system depends on the wastewater COD to nitrogen ratio

and the forms of nitrogen in the influent to the system.

Plant-supplied organic carbon is most important in VSB

systems treating nitrate-rich influents deficient in biodegradable organic carbon such as effluents from nitrifying

activated sludge plants. The minimum COD to nitrate-nitrogen ratio for denitrification is 2.3 g-COD/g-NO3

-N. Since

oxygen is used preferentially over nitrate as the electron

acceptor by the microbes that carry out denitrification, the

required COD/NO3-N ratio can be significantly higher if any

oxygen is present in the system.

Decomposing wetland plants and plant root exudates

are potential sources of biodegradable organic carbon for

denitrification but are also sources of organic nitrogen,

which is easily converted to ammonia. Plant root exudates

of organic carbon and nitrogen are the largest at the beginning of senescence. Because of the predominantly

anaerobic conditions in VSB systems, decomposition of

plant biomass within the media of a VSB system will likely

provide more organic carbon (and ammonia) to the wastewater than will decomposition of the plant biomass on top

of the media, which takes place in largely aerobic conditions. Some of the decomposition products of biomass on

top of the media (including nitrates) are transported into

the wastewater by precipitation infiltration.

In one study of a VSB system treating a nitrified secondary effluent, nitrate removal improved from 30% to 80%

when mulched biomass including straw, wetland plants,

and grass was applied to the top of the media (Gersberg

et al., 1983). Another study with a VSB system treating a

nitrified landfill leachate found that nitrate removal was limited by biodegradable organic carbon (Liehr et al., 2000).

5.2.1.5 Plant Role in Thermal Insulation

One potential advantage of a planted over an unplanted

VSB system is the role of plants in providing thermal insulation to the wastewater during cold weather. Dead plant

biomass on top of the media helps to limit both convective

heat losses from the wastewater and infiltration of melted

snow into the wastewater. Two researchers have developed methods to estimate the effect of plants in preventing heat loss from the wastewater of a VSB system (Reed

et al., 1995, Smith et al., 1997). However, it is not clear

how important this factor is in pollutant removal performance because 1) it has not been shown that planted VSB

systems perform better than unplanted systems, even in

winter, and 2) the dead plant material on top of the media

also acts as a barrier to oxygen transfer and a potential

source of biodegradable carbon and nutrients to the wastewater.

5.2.1.6 Plant Impact on Hydraulic Conductivity

(Clogging) and Detention Time

Several VSB systems have experienced conditions

called “surfacing” where a portion of the wastewater flows

on top of the media. Surfacing (1) creates conditions favorable for odors and mosquito breeding, (2) creates a

potential health hazard for persons and animals that may

come into contact with the wastewater, and (3) reduces

the hydraulic retention time (HRT) and performance of a

VSB system. Surfacing occurs whenever the hydraulic

conductivity of the media is not sufficient to transport the

desired flow within the usable headloss of the media. The

usable headloss is defined by the difference in the elevations of the outlet piping and the top of the media. Surfacing can result from a number of factors including (1) poor

design of the system inlet and outlet piping, (2) an inaccurate estimate of the clean hydraulic conductivity of the

media, (3) improper construction, and (4) an inaccurate

estimate of the reduction in hydraulic conductivity, or “clogging”, that will occur due to solids accumulation and/or

growth of plant roots. Several researchers have found that

clogging was the most severe within the first 1/4 to 1/3 of

the system (Young et al., 2000, George et al., 2000, Bavor

et al. 1989, Fisher, 1990, Sapkota and Bavor, 1994, Tanner and Sukias, 1995, Tanner et al., 1998). The hydraulic

conductivity was found to be less restricted and fairly uniform over the remaining length of the system.

Based on studies in Europe during the 1980s, some researchers proposed that plant roots significantly increased

the hydraulic conductivity in VSBs with soil media by opening up preferential pathways for the wastewater flow

(Kickuth, 1981). Later studies of these systems found that

a significant portion of flow occurred on top of the soil (Cooper et al., 1989). Based on recent studies, the presence of

plant roots in the gravel media of a VSB system will have a

negative effect on hydraulic conductivity (George et al.,

2000, Young et al., 2000, DeShon et al., 1995, Sanford et

al., 1995a and 1995b, Breen and Chick, 1995). Researchers at TTU compared the reduction in void volume due to

root and non-root solids. They estimated that the reduction in void volume due to root solids (2 - 8%) was much

larger than the reduction in void volume due to non-root

solids (0.1 - 0.4%). Even though the overall estimated reduction in void volume was small, there was a 98% reduction in hydraulic conductivity.

The primary functions of a plant’s roots are to supply

water and nutrients and to physically anchor or support

the above-ground portions on the plants. Water and nutri-
100

ents will be plentiful at all depths of a VSB, so the plant

roots will typically penetrate only 15 - 25 cm (6“ - 10") as

needed to anchor the plant. In most VSB systems the plant

roots do not fully penetrate the entire depth of the media

and the reduction in hydraulic conductivity in the root zone

results in the creation of “short-circuiting” under the root

zone, and more flow through the portion of the media without roots (Bavor et al., 1989, Fisher, 1990, DeShon et al.,

1995, Sanford et al., 1995a and 1995b, Sanford, 1999,

Rash and Liehr, 1999, Tanner and Sukias, 1995, Breen

and Chick, 1995). This situation may also lead to the creation of stagnant zones within the media ("dead volume")

which results in lower actual HRTs as the water preferentially flows through a smaller volume of the media. The

decrease in HRT will depend in part on the fraction of the

depth that is occupied by the roots; that is, deeper beds

will have more a greater proportion of the media that is not

impacted by roots.

From tracer studies, researchers have found significant

differences between actual and theoretical HRTs in their

VSB systems and attributed it to dead volume in the upper

zone of the media where the majority of the roots grow

(Liehr et al., 2000, Young et al., 2000, Bavor et al., 1989,

Fisher, 1990, DeShon et al., 1995, Sanford et al., 1995a

and 1995b, Breen and Chick, 1995). However, the researchers at TTU did not find a significant reduction in HRT

in three of the cells they studied. (George et al., 2000)

The researchers at North Carolina State University

(NCSU) attributed part of the dead volume in their systems to stratification of the water caused by less dense

rain water infiltration ponding within the media on top of

higher density leachate. This phenomenon has also been

reported by others (DeShon et al., 1995, Sanford et al.,

1995a and 1995b, Sanford,1999, Rash and Liehr, 1999).

NCSU found that the short-circuiting was greater in an

unplanted VSB system than in a planted system. While

rain water ponding may be a problem with some VSB systems, the effect at NCSU was magnified by the relatively

large catchment area (due to the shallow side slopes used

in the system) and the salinity of the leachate. The CU

researchers performed two sets of tracer studies in the

three cells of the Minoa system. The first set was performed

in the clean media of each cell before planting. The second set was performed after plant establishment on the

one cell that was half planted and half unplanted. From

the first study they concluded that there was short circuiting through the lower media and dead volume resulting in

the actual HRTs being only 75% of the theoretical values,

even in the clean media. They attributed these results to

media compaction during construction and intermixing of

the upper pea gravel with the lower larger media. From

the second tracer study they concluded that plants roots,

which penetrated only half of the media depth, resulted in

more short circuiting and dead volume than in the unplanted

media. Tanner & Sukias (1995) also reported more accumulation of solids in the root zone, which further contributed to preferential flow around the root zone.

5.2.2 Removal Mechanisms

5.2.2.1 BOD and TSS

VSB systems have been used for secondary treatment

(i.e. 30 mg/L of BOD and TSS) for a variety of wastewaters including: primary and septic tank effluents; pond effluent; and effluents from activated sludge, RBC, and trickling filter systems that don’t consistently meet secondary

standards. As discussed in Chapter 3, the primary mechanisms for BOD and TSS removal are flocculation, settling,

and filtration of suspended and large colloidal particles.

VSB systems are effective for TSS and BOD because of

relatively low flow velocities and a high amount of media

surface area. They typically do better at TSS removal, because TSS removal is a completely physical mechanism,

while BOD removal is more complex. Larger biodegradable particles that have been quickly removed by physical

mechanisms will be degraded over time and be converted

into particles in the soluble and small colloidal size range.

As such they become an internal “source” of BOD as they

degrade and reenter the water. Some material is also incorporated into microbial biomass.

Some material will accumulate in a VSB, but the amount

of long term solids accumulation is unknown. Tanner and

Sukias (1995) reported finding less solids accumulation

than would be expected based on the load in the influent

wastewater. Researchers at Richmond, Australia (Bavor

et al., 1989, Fisher, 1990) found that most solids were removed in the initial section of the VSB and that the “solids

accumulation front” stabilized after a year and did not advance. These findings support the idea that trapped material will degrade over time. VSB systems treating pond

wastewater are likely to accumulate more solids, and be

more susceptible to clogging, because TSS in pond wastewater is predominantly algae, which are slightly less biodegradable and degrade more slowly than typical primary

or secondary wastewater solids.

BOD and TSS in the effluent from a VSB are probably

not materials that have passed through the VSB, but rather

are converted or internally produced material. As such it is

likely to be quite different in size or composition from influent BOD and TSS. For example, the influent TSS in the

Las Amimas system were predominantly algal cells, but

there were almost no algal cells in the effluent even thought

the effluent TSS averaged 30 mg/L (Richard & Synder,

1994).

True BOD removal only occurs when the material causing the BOD is completely converted by anaerobic biological processes to gaseous end products. The two most likely

anaerobic pathways are methane fermentation and sulfate reduction. Because methane fermentation is severely

inhibited at temperatures below 10°C, sulfate reduction

probably predominates for soluble BOD removal during

colder months. However, seasonal performance does not

vary as much as would be expected based on the typical

temperature dependence of biological reactions. A likely

explanation, illustrated in Figure 5-1, is that biodegradable

particles that are physically removed during colder months
101

Influent TSS

Accumulated

Solids

Effluent TSS

Fall Winter Spring Summer Fall

TSS (mg/L)

Influent BOD

Internal BOD Load

Due to Conversion

of Accumulated Solids

Effluent BOD

Fall Winter Spring Summer Fall

BOD (mg/L)

Figure 5-1. Seasonal cycle in a VSB
102

are degraded more slowly and accumulate (Kadlec and

Knight, 1996). As the temperature warms up the rate of

degradation of trapped particles increases, leading to a

reduction of accumulated solids and a release of BOD.

This theory would explain why summer BOD removal rates,

based on influent BOD loading, do not appear to be significantly greater than winter removal rates. The need for

insulation of the surface of VSB systems in northern climates has been discussed, but the need has not been

quantified (Jenssen, etal, 1993).

Alternative VSB systems should achieve higher oxygen

transfer rates, so BOD removal should improve because

aerobic biological processes will become more prevalent.

However, microbial biomass production should also increase, which may lead to increased clogging problems.

The potential of alternative VSB systems for TSS and BOD

removal is unclear, but performance at Minoa, NY has been

very good (Reed and Giarrusso, 1999).

5.2.2.2. Nitrogen

Several conventional VSB systems have been designed,

built and operated to remove ammonia from various wastewaters. While partial ammonia removal has been achieved

in some systems, the removals have been less than predicted (George et al., 2000, Liehr et al., 2000, Young et al.,

2000). Ammonia can be removed by microbial reactions

or plant uptake. Because VSB systems are predominantly

anaerobic, microbial removal via nitrification is very limited. As discussed in Section 5.2.1.3, plant uptake is also

very limited. Very lightly loaded systems have achieved

partial ammonia removal (George et al., 2000, 1999; Young

et al., 2000), but if ammonia removal is required, a separate ammonia removal process should be used in conjunction with a VSB system.

The predominantly anaerobic condition of VSB systems

seems well suited for microbial removal of nitrate via denitrification, but there are relatively few studies to document

their use for this specific purpose (Gersberg, et al, 1983;

Stengel and Schultz-Hock, 1989). Systems treating well

oxidized secondary effluents or other carbon limited wastewaters may have inadequate carbon for denitrification to

proceed efficiently (Liehr et al., 2000). Systems treating

wastewaters with more carbon, and that have achieved

partial nitrification, typically achieve almost complete denitrification (George et al., 2000, Young et al., 2000). Crites

and Tchobanoglous (1998) suggest that significant denitrification of municipal wastewaters can occur in VSB systems at a detention time of 2 to 4 days, but Stengel and

Schultz-Hock (1989) demonstrated with methanol addition

that denitrification was carbon limited.

Alternative VSB systems should achieve higher oxygen

transfer rates, so they should be more efficient at ammonia removal via nitrification (George et al., 2000, Reed &

Giarusso, 1999, Behrends et al., 1996, May et al., 1990)

and less efficient for nitrate removal via denitrification than

conventional VSBs.

5.3 Hydrology

5.3.1 Evapotransporation and

Precipitation Impacts

The avoidance of surfacing is a major design criterion

and high amounts of precipitation or snowmelt can increase

the flow in a VSB system. In climates with extended periods of precipitation or heavy snowmelt, the runoff from the

total catchment area that drains into the VSB must be estimated and included in the design flow. Evapotransporation

(ET) decreases the hydraulic loading and will not contribute to surfacing.

Except in very wet climates, flows from precipitation

events will probably not adversely affect performance because VSB systems have a relatively small surface area

(compared to FWS wetlands) and effluent controls should

be sufficient to prevent surfacing. Precipitation dilutes pollutants in the system, temporarily raises the water level,

and decreases the HRT, while ET concentrates pollutants,

temporarily lowers the water level, and increases the HRT.

ET rates will vary depending on plant species and density,

but rates from 1.5 to 2 times the pan evaporation rate have

been reported in the literature (refs). Except in very wet or

dry climates, the two results are probably offsetting, reducing the overall impact on water level and effluent values. Unfortunately, the specific effects of ET and precipitation on VSB performance are not documented because

good estimates of ET and precipitation are hard to obtain,

and precise influent and effluent flow measurements are

seldom available, even in research systems.

5.3.2 Water Level Estimation

An important step in the design process is to estimate

the elevation of the water surface throughout the VSB to

ensure that surfacing of the wastewater does not occur.

As in all gravity flow systems the water level in a VSB system is controlled by the outlet elevation and the hydraulic

gradient, or slope, which is the drop in the water level

(headloss) over the length from the inlet to the outlet. The

relationship between flow through a porous media and the

hydraulic gradient is typically described by the general form

of Darcy’s Law (Eq. 5.1). This form assumes laminar flow

through media finer than coarse gravel, and many authors

have modified it for other applications including other media and turbulent flow. However, use of the general form

without modification is recommended as sufficient to estimate the water level within a VSB.

Q = (K)(Ac)(S) = (K)(W)(Dw)(dh/dL) (5-1)

or, for a defined length of the VSB,

dh = (Q)(L) / (K)(W)(Dw) (5-2)

where Q= flow rate, m3/d

K = hydraulic conductivity, m3/m2-d, or m/d

Ac= cross-sectional area normal to wastewater flow, m2

= (W)(Dw)

where W = width of VSB, m

Dw = water depth, m
103

L = length of VSB, m

dh = head loss (change in water level) due to flow re

sistance, m

S = dh/dL = hydraulic gradient, m/m

The water level at the inlet of a VSB will rise to the level

required to overcome the head loss in the entire VSB.

Therefore, the VSB must be designed to prevent surfacing. K for an operating VSB varies with time and location

within the media and will have a major impact on the head

loss. K is very difficult to determine because it is influenced

by factors that cannot be easily accounted for, including

flow patterns (affected by preferential flow and short circuiting), and clogging (affected by changes in root growth/

death and solids accumulation/degradation). Therefore, a

value must be assumed for design purposes. Typical values for various sizes of rock and gravel are shown in Table

5-1. Several of the references listed in Table 5-1 also noted

that K was much less in the initial 1/4 to 1/3 of the VBS

than in the remainder of the bed. Based on the studies

listed in Table 5-1 and many observed cases of surfacing

in VSB systems, the following conservative values are recommend for the long-term operating K values:

initial 30% of VSB Ki

 = 1% of clean K

final 70% of VSB Kf = 10% of clean K.

5.3.3 Hydraulic Retention Time and

Contaminant Dispersion

The theoretical HRT in any reactor is defined as the liquid volume of the reactor divided by the flow rate through

it. The liquid volume in a VSB system is difficult to accurately determine because of the loss of pore volume to

roots and other accumulated solids, such as recalcitrant

biomass and chemical precipitates. The lost pore volume

will vary with both location in the VSB and time, both seasonally and yearly, because of root growth and decay, and

solids accumulation and degradation. Preferential flow (see

section 5.2.1.6) as illustrated in Figure 5-2 will also have a

direct impact on HRT and has not been correlated with

changes in pore volume. For design purposes the volume

occupied by roots and other solids is assumed to be insignificant and the theoretical HRT is estimated using the

average flow (including precipitation and ET for very wet

or dry climates) through the system, the system dimensions, the operating water level, and the initial (clean) porosity of the media, which is either estimated or experimentally determined.

The actual HRT has been frequently reported to be 40-

80% less than the theoretical HRT (based on pore volume) either due to loss of pore volume, dead volume, or

preferential flow (Fisher, 1990, Sanford et al., 1995b,

Bhattarai and Griffin, 1998, Batchelor and Loots, 1997,

Rash and Liehr, 1999, Tanner and Sukias, 1995, Breen

and Chick, 1995, Tanner et al., 1998, Bowmer, 1987). A

rough approximation of the liquid volume can be determined by measuring the volume of water drained from an

operating bed, but water held in small pores or adhering to

biomass will remain in the system. Draining will also not

be able to account for preferential flow. Tracer studies are

recommended as a more realistic measure of the HRT in

a VSB system, using one of a variety of tracers (Young et

al., 2000, Young et al., 2000, George et al., 2000, Netter

Table 5-1. Hydraulic Conductivity Values Reported in the Literature.

Size and type1 “Clean”/”Dirty” Type of Wastewater Length of Notes & References

of Media K (m/d) (Typical TSS, mg/L)2 Operation

5-10 mm gravel 34,000/12,000 2° effluent (100) 2 years K = 12,000 is for downstream portion (last 80 m) of VSB

5-10 mm gravel 34,000/900 2° effluent (100) 2 years K = 900 is for inlet zone (first 20 m) of VSB

Bavor et al (1989), Fisher (1990), Bavor & Schulz

(1993)

17 mm creek rock 100,000/44,000 nutrient solution (neg) 4 months neg = negligible TSS

6 mm pea gravel 21,000/9000 nutrient solution (neg) 4 months Macmanus et al (1992), DeShon et al (1995)

30-40 mm coarse gravel NR/1000 2° effluent (30 w/a) 2 years w/a = with algae; pond effluent; gravel bed onlyno plants

5-14 mm fine gravel NR/12,000 2° effluent (30 w/a) 2 years coarse gravel is first 6m of bed; fine is last 9 m of

bed Sapkota & Bavor (1994)

20-40 mm coarse gravel NR/NR landfill leachate (neg) 26 months for coarse gravel, headloss was controlled by

outlet, not K

5 mm pea gravel 6200/600 landfill leachate (neg) 26 months Sanford et al (1995a & 1995b), Sanford (1999),

Surface et al (1993)

19 mm rock 120,000/3000 septic tank effluent (50) 7 months George et al (2000)

14 mm fine gravel 15,000/see note aerated pond (60 w/a) 2 years K of combined gravel (fine overlaid coarse) was

22 mm coarse gravel 64,000/see note aerated pond (60 w/a) 2 years 2000 at 50 m from inlet; 27,000 at 300 m from inlet

Kadlec & Watson (1993), Watson et al (1990)

1

Type as defined in the reference(s)

2

neg = negligible; w/a = with algae
104

and Bischofsberger, 1990, Fisher, 1990, Netter 1994,

Sanford et al., 1995b, Bhattarai and Griffin, 1998, Bowmer,

1987).

Some of the current design equations for VSB systems

assume plug flow conditions. However, tracer studies performed on VSB systems have found significant amounts

of dispersion as shown in Figure 5-3 (Sanford et al., 1995b,

Bhattarai and Griffin, 1998, Liehr et al., 2000, George et

al., 2000). Based on current data it appears that VSB systems can not be accurately modeled as either plug flow or

complete mix reactors. The simplest model that can provide a reasonable fit to the tracer curves is a series of

equal volume complete mix reactors. However, while this

model may mathematically fit the tracer data, it does not

realistically represent physical flow through porous media.

Intuitively it would seem that a plug flow reactor with dispersion would most closely represent the actual conditions

in a VSB. This model allows greater flexibility in determining a fit of the tracer data but typically results in a complex

mathematical model of pollutant removal. Estimates of the

dispersion number for VSB systems have ranged from

0.050 to 0.31 (George et al., 2000, Bhattarai and Griffin,

1998), with greater numbers for systems with small lengthwidth ratios. Dispersion numbers less than 0.025 are indicative of near-plug flow conditions while values above

0.20 indicate a high degree of dispersion. The modeling of

flow and dispersion is complicated by the non-uniformity

of flow and pore volume in space and time as previously

discussed, and by other factors including precipitation and

ET.

At this point in time there appears to be little justification

for using complex flow models, because of a lack of data

and the unpredictable and constantly varying conditions

within a VSB.

5.4 Basis of Design

5.4.1 Introduction

Attempting to fully describe pollutant removal in VSB

systems is at least as complex as trying to describe VSB

hydraulics. Many authors have examined several relationships as a model for pollutant removal, including zero and

first order reactions in both plug flow and complete mix

reactor models. None of the relationships were found to

reasonably fit of the all data that are available. Furthermore, data from VSB systems are typified by a wide variability, as would be expected of dynamic natural systems

that are influenced by many factors. This variability is evident in the plots of TSS, BOD, TKN and TP data in this

section. Data scatter is not reduced by comparing pollutant removal with a variety of factors (e.g. area, volume,

HRT, percent removals or loading rate), or by normalizing

the data (Ce

/C0). Expected trends, such as temperature

dependence for BOD removal or better removal with lower

Figure 5-2. Preferential Flow in a VSB

}

Flow

Outlet

Zone

0.6m

Outlet

Path

Flow

Major

Wetland Plants

Root Zone

(0.15 - 0.25m)

Not to Scale & Dimensions Are “Typical”

Intlet

Zone

Intlet

0.1 - 0.15m Small (Planting) Gravel
105

pollutant loading, are often not apparent due to the scatter

of the data. Therefore, the design approach recommended

here is to use the maximum pollutant loading rates that

have been shown to meet discharge standards. This approach yields a much more conservative design than other

common design approaches. As additional quality data becomes available in the future, it may be possible to extend

these conservative loading rates with confidence.

Two types of pollutant loading rates were considered,

an areal loading rate (ALR), g/m2-d, and a volumetric loading rate (VLR), g/m3

-d. Both ALRs and VLRs have been

used by researchers to describe VSB performance. ALR

is calculated by multiplying the influent flow rate (m3

/d) by

the influent pollutant concentration (mg/L = g/m3), and dividing by the surface area of the VSB system (m2

). Because sedimentation, plant growth and oxygen transfer are

theoretically dependent on the surface area, ALR may be

a characteristic parameter for some pollutants. VLR is calculated by multiplying the influent flow rate (m3

/d) by the

influent pollutant concentration (mg/L = g/m3), and dividing by the pore volume of the VSB system (m3

). Because

the removal of certain pollutants could be dependent on

the HRT, the VLR could be a characteristic parameter for

some pollutants. However, because the actual saturated

pore volume is seldom known and the HRT may not be

directly related to the pore volume due to preferential flow,

the utility of VLR for design purposes is limited. Also, a

comparison of Figures 5-4 through 5-7, which are typical

of scatter for all pollutants, shows that data scatter is not

reduced by the use of VLR. Therefore, the design recommendations in this chapter are based on ALRs.

Finally, because the type of pre-treatment has a major

impact on the characteristics of the wastewater being

treated, the following discussions are organized by the type

of wastewater being treated: septic tank and primary effluents, pond effluents, and secondary treatment effluents.

5.4.2 TSS and BOD Removal for Septic

Tank and Primary Effluents

Two recent studies, one conducted by Tennessee Technological University (TTU) and one conducted by Clarkson

University (CU) at the Village of Minoa, New York, have

provided the majority of data used to establish the design

recommendations for this section (George et al., 2000,

Young et al., 2000). These two studies were chosen because their research objectives were to provide design information, they utilized several VSBs with different measured loadings, and the data are of good quantity and quality. Influents in the TTU and CU studies were respectively

a low strength septic tank effluent and a fairly typical primary effluent. Each data point in the following figures represents a quarterly average of biweekly (every 2 weeks)

sampling for the TTU data, and a quarterly average of at

least two monthly samples for the CU data. The results

from one other VSB system treating septic tank effluent

studied by University of Nebraska - Lincoln (UNL) researchers are also included in these figures (Vanier & Dahab,

1997). After reviewing the literature, no other studies with

Figure 5-3. Lithium Chloride Tracer Studies in a VSB System (George et al., 2000)

Cell B

Cell N

Cell D

Cell F

Cell E

Cell M

Cell K

20

18

16

14

12

10

8

6

4

2

0

0 50 150100 200 250 300

Time (hrs)

Lithium Concentration (mg/L)
106

TTU1

TTU2

TTU3

UNL

CU

NADB

50

40

30

20

10

0

0

Effluent TSS (mg/L)

5 10 15 25 20

TSS Areal Loading Rate (g/m2-d)

Figure 5-4. Effluent TSS vs areal loading rate

Figure 5-5. Effluent TSS vs volumetric loading rate

TTU1

TTU2

TTU3

UNL

CU

600500400300200100

TSS Volumetric Loading Rate (g/m3-d)

0

0

10

20

30

40

50

Effluent TSS (mg/L)
107

Figure 5-6. Effluent BOD vs areal loading rate

Figure 5-7. Effluent BOD vs volumetric loading rate

TTU1

TTU2

TTU3

UNL

CU

NADB

70

60

50

40

30

20

10

0

0 2 1.6

Effluent BOD (mg/L)

4 6 8 10 12

BOD Areal Loading Rate (g/m2-d)

14

80

100

60

40

20

Effluent BOD (mg/L)

0

0 50 100 150 200

TTU1

TTU2

TTU3

UNL

CU

BOD Volumetric Loading Rate (g/m3-d)
108

septic tank or primary wastewater were found to have data

with similar quality and quantity as these three studies.

Data from the NADB for VSBs treating primary effluent,

some of which are of unknown quality, are also shown in

Figures 5-4 and 5-6.

TSS removal is quite good; effluent TSS was consistently less than 30 mg/L at TSS ALRs as high as 20 g/m2

-

d (Figure 5-4). The two data points in Figure 5-4 that are

above 30 mg/L are from systems at TTU that were intentionally overloaded to failure, and are not typical. Other

researchers have reported plugging of the surface of the

media (as opposed to clogging of the pore volume) when

excessively high TSS loadings were applied (Tanner &

Sukias, 1995, Tanner et al., 1998, van Oostrom & Cooper,

1990). Additional data may extend these limited ALRs when

it becomes available. However, it should be noted that the

typical sustained influent TSS concentrations for the data

plotted in these figures were less than 100 mg/L. It is recommended that TSS ALR be limited to 20 g/m2

-d, based

on the maximum monthly influent TSS. This would correspond to a loading of 2 cm/d for an influent concentration

of 100 mg/L of TSS, 4 cm/d for an influent concentration of

50 mg/L of TSS, and so on.

BOD removal is not as good as TSS removal, so the size

of a VSB designed to meet secondary treatment standards

will generally be controlled by the requirements for BOD removal. Effluent BOD values were found to periodically exceed 30 mg/L at BOD ALRs greater than 6 g/m2

-d (Figure 5-6).

It is recommended that BOD ALR be limited to 6 g/m2

-d,

based on the maximum monthly influent BOD, to produce

a maximum effluent BOD of 30 mg/L. Table 5-2 compares

the size of a VSB designed with this ALR compared to the

size of VSBs designed using several common approaches.

As expected the other design approaches result in VSB

systems significantly smaller than that using the conservative design approach presented here.

5.4.3 Nutrient Removal for Septic Tank

and Primary Effluents

Most of the organic nitrogen in septic tank and primary

effluents is associated with suspended solids that are easily

removed in VSB systems. It is generally assumed that the

organic nitrogen will be converted to ammonia in VSB systems, but spiked concentrations of urea (a soluble form of

organic nitrogen) were often not completely converted in

one study (George et al., 2000). Ammonia removal in VSB

systems is severely oxygen limited, and it is inversely related to the ultimate (carbonaceous and nitrogenous) BOD

loading. Also, the conversion of organic nitrogen into ammonia via ammonification or hydrolysis masks any attempt

to relate ammonia removal to other design factors. For this

reason Total Kjeldahl Nitrogen (TKN) data rather than

ammonia data are presented in Figure 5-8. The TKN removal performance is generally poor and highly variable.

Therefore, VSB systems should not be used alone to treat

pre-settled municipal wastewaters if significant amounts

of ammonia must be consistently removed.

Although the data are not presented here, if any nitrate

is produced in VSB systems treating septic tank and primary effluents, it is likely that the nitrate will be removed

by denitrification.

Table 5-2. Comparison of VSB Area Required for BOD Removal Using Common Design Approaches.

Design criteria

Flow (Q) = 400 m3/d (105,680 gpd)

Influent BOD5 (Ci) = 125 mg/L

Effluent BOD5 (Ce) = 30 mg/L

Design Rate Loading Other Factors Required

Approach Constant Constant Area

m2 (ac)

This Manual 6 g/m2-d 8,330 (2.0)

(54 lb/ac-d)

European KBOD = 0.1/d 5710 (1.4)

(Cooper, 1990)

Kadlec & 180 m/yr Background Concentration2 = 10 mg/L 1420 (0.4)

Knight (590 ft/yr)

(1996)

Reed, etal. Temperature Water Depth1 = 0.4 m (16")

(1995) Dependent2 Media Porosity1 = 0.38

K10 = 0.62/d at 10°C, 6090 (1.5)

K20 = 1.104/d at 20°C, 3400 (0.8)

TVA (1993) 5.3 g/m2

-d Derived from TVA design 9430 (2.3)

(48 lb/ac-d) Assumes septic tank effluent

1

Values chosen by user; these are not necessarily the values recommended by the design’s author. 2

Values calculated per instructions of design’s author.
109

Although phosphorus is partially removed in VSB systems treating septic tank and primary effluents, VSBs are

not very effective for long-term phosphorus removal (Figure 5-9). It should be noted that the phosphorus data shown

in Figure 5-9 are from VSB systems that are relatively new,

when it can be assumed that the phosphorus precipitation

and adsorption capacity of the media would be at its greatest. Because plant uptake of phosphorus is quite small

compared to typical loadings (Reed, et al, 1995; Crites &

Tchobanoglous, 1998), the phosphorus removal capacity

will decrease with time. Estimates of realistic long-term

phosphorus removal by plant harvesting is limited to about

0.055 g/m2

-d (0.5 lb/ac-d) (Crites & Tchobanoglous, 1998).

VSB systems should not be expected to remove phosphorus on a long-term basis.

5.4.4 TSS, BOD and Nutrient Removal for

Pond Effluents

There is much less quality data comparable to the TTU

and CU studies for VSB systems treating pond effluents.

Data from a study conducted at three experimental VSB

systems at Las Animas, Colorado by Colorado State University (Richard & Synder, 1994) were used to support

design recommendations for VSB systems treating pond

effluents (Table 5-3). The pollutant removal performance

for the Las Animas VSBs treating oxidized pond effluent

was not as good as the performance of the TTU and CU

systems. NABD data for VSBs treating pond effluent (Figure 5-10), which are not as reliable as the Las Animas

data, show similar performance. Several of the NABD systems have experienced surfacing caused by clogging of

the media surface (as opposed to clogging of the pore volume) by algae.

For Las Animas, the average TSS ALR was 6.2 g/m2

-d

(55 lb/ac-d) and produced an overall average effluent TSS

of 35 mg/L. The average BOD ALR was 2.0 g/m2

-d (18 lb/

ac-d) and produced an overall average effluent BOD of 25

mg/L. There was essentially no nitrogen or phosphorus

removal on average in the three VSBs. The poor overall

percent removal of BOD of 35% might be related to the

relatively high concentrations of algal cells in the influent

during several months of each year. The measured BOD

of pond effluent typically does not account for the true BOD

of algal cells because algal cells degrade more slowly than

other organic matter.

A VSB system in Mesquite, Nevada has been used since

1992 in parallel with overland flow and oxidation ditch systems to treat an aerated pond effluent. One year of monthly

data for the Mesquite system is summarized in Chapter 8.

Over the one-year sampling period the effluent BOD averaged 29 mg/L when loaded at an average BOD ALR of 2.5

g/m2

-d (22 lb/ac-d). Better BOD and TSS removals than at

Las Animas and Mesquite are reported in a 1993 EPA report for several VSB systems treating pond algal effluents.

However, the sparse amount of data of unproven quality

represented by the average values given in the report is

Figure 5-8. Effluent TKN vs areal loading rate

TTU1

TTU2

TTU3

UNL

CU

20

0

4 6 8 10 12 14

TKN Areal Loading Rate (g/m2-d)

Effluent TKN (mg/L)

10

20

30

40

50
110

Figure 5-9. Effluent TP vs areal loading rate

TTU1

TTU2

TTU3

CU

1.210.80.60.40.20

TP Areal Loading Rate (g/m2-d)

0

0.5

1

1.5

2.5

2

3

3.5

Effluent TP (mg/L)

inadequate to use with confidence. Sapkota and Bavor

(1994) report similar TSS removal, but do not report BOD

removal.

The upgrading of pond effluent with rock filters is similar

to the use of VSBs after ponds. However, because of variable results from rock filters, their use is generally cautioned due to a lack of reliable design information (Reed,

et al, 1995). Performance of rock filters is also plagued by

H2

S generation and high effluent ammonia. Illinois requires

effluent aeration and recommends disinfection before discharge for pond rock filter systems. Because of the limited

data and uncertainty about similar rock filter systems, VSB

systems are not recommended for treating pond effluents

if the system must consistently meet a 30/30 standard.

5.4.5 BOD, TSS and Nutrient Removal for

Secondary and Non-algal Pond

Effluents

There are very few quality-assured data available from

VSB systems in the U.S. treating secondary effluents. The

1993 U.S. EPA report included data collected over a three

month period from three systems. Additional data from one

of these systems, Mandeville, LA, is included in Chapter

8. The Mandeville VSB treats an aerated lagoon effluent

which has little or no algae. The average influent BOD and

TSS were 40 and 16 mg/L, respectively. The average effluent BOD and TSS were 5 and 3 mg/L, respectively at a

BOD ALR of 7.9 g/m2-d (70 lb/ac-d).

Representatives from Severn Trent Water, Ltd., have

reported on the performance of VSB ("reed bed") systems

treating activated sludge and RBC effluents in small treatment plants (less than 2000 people) in the U.K. (Green

and Upton, 1994). The goal for these VSB systems is to

provide additional treatment of secondary effluents so that

they consistently meet discharge limitations, which can vary

from 30/20 to 15/10 TSS/BOD. Essentially these systems

serve as aesthetic and sometimes economical substitutes

for tertiary filters for small treatment plants. In some cases

in the U.K. they have been used to treat storm water bypass flows at secondary treatment plants. While the Severn

Trent systems typically remove some nitrogen and phosphorus, they are not capable of meeting typical discharge

standards for nutrients in the U.S. The primary design basis used by Severn Trent is a hydraulic surface loading

rate of 0.20 m3

/m2-d (5 gpd/sq ft) for the average daily flow

(Green and Upton, 1994). This value is derived from the

design recommendation of the European task group on

VSB systems (Cooper, 1990). For systems with an average influent BOD < 40 mg/L, this results in average areal

BOD loading of less than 8.0 g/m2

-d (71 lb/ac-d). Typical

systems are 0.6 m (24 in) deep, 0.4 m wide per m3

/d (5 ft

per 1000 gpd) of flow, and 12.5 m (41 ft) long.

Based on the success of the Mandeville and Severn Trent

systems, it appears that VSB systems can be effectively

used to help small secondary systems consistently meet

secondary effluent standards. The recommended approach
111

Table 5-3. Data from Las Animas, CO VSB Treating Pond Effluent.

Time Inf. TSS Eff. TSS Inf. BOD Eff. BOD Inf. TKN Eff. TKN Inf. TP Eff. TP

Period1 mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L

Cell 1

Winter 91 89.7 43.0 26.7 37.5 ND ND 1.57 1.9

Srping 92 146.0 28.7 34.0 22.7 7.2 11.4 1 1.67

Summer 92 178.0 41.7 49.7 29.0 6.1 8.7 1.17 1.68

Fall 92 223.0 34.3 54.0 30.0 7.3 10.7 1.47 1.9

Winter 92 50.0 34.3 33.7 32.7 14.1 14.0 2.25 2.3

Spring 93 66.0 24.0 41.0 32.0 16.1 16.3 2.55 2.6

Summer 93 95.3 33.3 30.0 9.3 3.9 5.3 0.67 0.78

Fall 93 127.0 38.0 40.3 16.0 4.0 3.4 0.65 0.6

Average 121.9 34.7 38.7 26.2 8.4 10.0 1.42 1.68

Cell 2

Winter 91 89.7 51.0 26.7 37.0 ND ND 1.57 2.57

Srping 92 146.0 34.0 34.0 35.0 7.2 14.1 1 2.37

Summer 92 178.0 43.3 49.7 29.3 6.1 9.0 1.17 1.72

Fall 92 223.0 26.0 54.0 40.0 7.3 8.8 1.47 1.78

Winter 92 50.0 33.0 33.7 29.7 14.1 14.2 2.25 2.28

Spring 93 66.0 26.7 41.0 35.3 16.1 15.2 2.55 2.68

Summer 93 95.3 27.0 30.0 15.0 3.9 5.5 0.67 0.82

Fall 93 127.0 33.3 40.3 21.3 4.0 4.8 0.65 0.65

Average 121.9 34.3 38.7 30.3 8.4 10.2 1.42 1.86

Cell 3

Winter 91 89.7 46.0 26.7 11.3 ND ND 1.57 1.47

Srping 92 146.0 47.0 34.0 22.0 7.2 5.4 1 1.43

Summer 92 178.0 33.3 49.7 20.3 6.1 9.6 1.17 1.62

Fall 92 223.0 41.3 54.0 27.0 7.3 8.1 1.47 1.72

Winter 92 50.0 34.3 33.7 25.7 14.1 13.0 2.25 2.2

Spring 93 66.0 23.7 41.0 29.7 16.1 13.9 2.55 2.53

Summer 93 95.3 29.3 30.0 8.3 3.9 4.7 0.67 0.82

Fall 93 127.0 34.3 40.3 8.0 4.0 5.6 0.65 0.55

Average 121.9 36.2 38.7 19.0 8.4 8.6 1.42 1.54

1

Each of the values in the table is the average of three monthly samples.

Figure 5-10. NADB VSBs Treating Pond Effluent

BOD

TSS

2520151050

0

10

BOD or TSS Areal Loading Rate (g/m2-d)

NADB Systems Treating Pond Effluent

Effluent BOD or TSS (mg/L)

20

30

40

50

60
112

is to limit the BOD ALR to a maximum monthly value of 8

g/m2-d (71 lb/ac-d). However, VSB systems are not recommended as a remedy for inadequately operated activated sludge systems. Process upsets in poorly operated

activated sludge systems can quickly fill a VSB system

with mixed liquor solids, resulting in surface flow due to

clogging of the media.

5.4.6 Metals Removal for All Types of

Wastewater

Metals are removed in a VSB by two primary mechanisms. First, because many metals (e.g. Zn, Cr, Pb, Cd,

Fe, Al) are associated with particles (Heukelekian & Balmat,

1959; SWEP, 1985), the high efficiency of particulate separation in a VSB should remove these metals accordingly.

Second, sulfide precipitation occurs due to the reduction

of sulfates to sulfides in the absence of nitrate, and renders some metals insoluble, resulting in significant removals, as described in Reed, etal (1995) for Cu, Cd, and Zn.

As long as the system ORP remains low, which is likely

given the anaerobic nature of VSBs, it is unlikely that metals precipitated in the sulfide form will re-enter the water

column (Bounds, et al, 1998; Reed, et al, 1995). Some

metals such as Ni and Cd are more mobile and less likely

to be removed, but they are not normally present in toxic

quantities in municipal wastewater.

There is relatively little data on metals removal by VSB

systems and no known information from long-term studies. Gersberg et al. (1984) found significant removal of Cu,

Zn and Cd, and determined that plant uptake was responsible for only 1% of the Cu and Zn removal. In a study with

a VSB system treating a landfill leachate, researchers found

only a small increase in the Pb, Cd, and Cu levels on root

surfaces, and no increase in any of the metals measured

in any plant tissue compared to plants from a control system (Peverly et al., 1995). They concluded that the increased metal concentrations on the root surface was due

to metal precipitation and adsorption. Metal removal by

plant uptake should not be counted on in any VSB system

over the long term.

5.4.7 Pathogen Removal for All Types of

Wastewater

While pathogens will be partially removed in a VSB system, a disinfection step after the VSB will normally be required to meet discharge limits. Researchers in Nebraska

found a three log reduction in fecal coliforms from 106 to

103

/100mL in a VSB system treating a septic tank effluent

(Vanier and Dahab, 1997). Gersberg et al. (1989) found a

two log reduction in total coliforms in a VSB system treating primary effluent. The coliform removal in two VSB systems in England treating secondary effluents varied between 40% and 99%, but effluent values did not meet discharge requirements (Griffin et al., 1998). Fecal coliform

reductions were typically two logs (1 x 106

 to 1 x 104/100mL)

in several experimental VSB systems in Tennessee, except for two cells operated in a fill and drain mode. These

fill and drain cells achieved a three log reduction with the

same influent wastewater (George et al., 2000). For design purposes a two log reduction is a reasonable estimate of VSB performance.

5.5 Design Considerations

5.5.1 Media Size and Hardness

The media of a VSB system perform several functions;

they (1) are rooting material for vegetation, (2) help to

evenly distribute/collect flow at the inlet/outlet, (3) provide

surface area for microbial growth, and (4) filter and trap

particles. For successful plant establishment, the uppermost layer of media should be conducive to root growth. A

variety of media sizes and materials have been tried, but

there is no clear evidence that points to a single size or

type of media, except that the media should be large

enough that it will not settle into the void spaces of the

underlying layer. It is recommended that the planting media not exceed 20 mm (3/4 in) in diameter, and the minimum depth should be 100 mm (4 in).

The media in the inlet and outlet zones (see Figure 5-

11) should be between 40 and 80 mm (1.5 – 3 in) in diameter to minimize clogging and should extend from the top

to the bottom of the system. The inlet zone should be about

2 m long and the outlet zone should be about 1 m long.

These zones with larger media will help to even distribute

or collect the flow without clogging. The use of gabions

(wire rock baskets used for bank stabilization) to contain

the larger media simplifies construction. Gabions may also

make it easier to remove and clean the inlet zone media if

it becomes clogged.

Any portion of the media that is wetted is a surface on

which microbes grow and solids settle and/or accumulate.

Media in VSBs have ranged from soil to 100 mm (4 in.)

rock. Experience with soil and sand media shows that it is

very susceptible to clogging and surfacing of flows, even if

influent TSS concentrations are minimal, so soil or sand

media should be avoided. Gravel and rock media have

been used successfully, with smaller diameter media being more susceptible to clogging, and larger media more

difficult to handle during construction or maintenance.

Crushed limestone can be used, but is not recommended

for VSB systems because of the potential for media breakup

and dissolution under the strongly reducing environment

of a VSB, which can lead to clogging. Media with high iron

or aluminum will have more sites for phosphorus binding

and should enhance phosphorus removal, but only during

the first few months of operation. The limited removal capability is probably not worth an added expense if it is not

available locally at a reasonable cost. Alternative media

such as shredded tires, plastic trickling filter media, expanded clay aggregates and shale with potentially high

phosphorus absorptive capacity have been used, but there

is inadequate data to make a recommendation for or

against their use.

There does not appear to be a clear advantage in pollutant removal with different sized media in the 10 to 60 mm

(3/8 - 2 in.) range. Therefore, it is recommended that the
113

average diameter of the treatment zone media be between

20 and 30 mm (3/4 – 1 in.) in diameter as a compromise

between the potential for clogging and ease of handling.

To minimize settling of the media smooth, rounded media

with a Mohs hardness of 3 or higher is recommended if it

is available locally at a reasonable cost. Based on the data

in Table 5-1, the hydraulic conductivity of the 20 - 30 mm

diameter clean media is assumed to be 100,000 m/d.

5.5.2 Slopes

The top surface of the media should be level or nearly

level for easier planting and routine maintenance. Theoretically, the bottom slope should match the slope of the

water level to maintain a uniform water depth throughout

the VSB. However, because the hydraulic conductivity of

the media varies with time and location, it is not practical

to determine the bottom slope this way, and the bottom

slope should be designed only for draining the system,

and not to supplement the hydraulic conductivity of the

VSB. A practical approach is to uniformly slope the bottom

along the direction of flow from inlet to outlet to allow for

easy draining when maintenance in required. No research

has been done to determine an optimum slope, but a slope

of 1/2 to 1% is recommended for ease of construction and

proper draining (Chalk & Wheale, 1989). Care should be

taking when grading the bottom slope to eliminate low

spots, channels and side-to-side sloping which will promote dead volume or short-circuiting.

The slope of the berms containing a VSB should be as

steep as possible, consistent with the soils, construction

methods and materials. Shallow side slopes create larger

areas which capture and route precipitation into the VSB,

which may be detrimental to system performance. Also,

the site should be graded to keep off-site runoff out of the

VSB.

5.5.3 Inlet and Outlet Piping

The inlet piping must be designed to minimize the potential for short-circuiting and clogging in the media, and

maximize even flow distribution. For VSBs with length-width

ratios less than one, additional care must be taken to spread

the influent across the whole width of the VSB. Standard

hydraulic design principles and structures (e.g. adjustable

weirs and orifices) are used to split, balance evenly distribute flows (WEF, 1998). The recommended method to

evenly distribute flows is to use reducing tees or 90 degree elbows which can be rotated on the header (see Chapter 6). The main advantage of a rotating fitting is that it

allows the operator to easily adjust the distribution of the

influent, which may help in reducing media clogging. When

the potential for public access exists, a cover over the influent distribution system must be used. Possible covers

include half sections of pipe or cavity chambers, as used

in leach fields. If piping with orifices is used to distribute

flows instead of a pipe with rotating fittings, it is necessary

to minimize the headloss in the distribution piping so that

the headloss through the orifices controls the flow. This

requirement limits the number and size of orifices used,

and makes the distribution piping large enough so that the

velocity in it is low. The orifices should be evenly spaced

at a distance approximately equal to 10% of the cell width.

For example, a system 20 m (65 ft) wide should have orifices placed every 2 m (6.5 ft). If poor design causes wastewater to always discharge through only some of the orifices, clogging of the media or accumulation of a surface

Figure 5-11. Proposed Zones in a VSB

20 - 30mm gravel/rounded stone

Media Surface

Water Level

40 - 80

mm

Rock

40 - 80

mm

Rock

Inlet

Zone

2 m 30% of Length

Zone 1 Zone 2

70% of Length

Outlet

Zone

1 m

Outlet Treatment Zone

0.5m

0.6m

Side View
114

layer of solids near those orifices can become a problem,

especially for an influent with relatively high suspended

solids, such as pond effluent. Finally, the inlet piping should

be designed to allow for inspection and clean-out by the

operator.

The outlet piping must be designed to minimize the potential for short-circuiting, to maximize even flow collection, and to allow the operator to vary the operating water

level and drain the bed. For VSBs with length-width ratios

less than one, additional care must be taken to collect the

influent from the whole width of the VSB. A collection header

with orifices that is placed across the entire width of the

bottom of the VSB is recommended to promote even flow.

The collection header should be designed with the same

hydraulic principles used for inlet distribution piping. Slotted or perforated drainage pipe can be used if the collection header is not too long, but properly sized and spaced

orifices in a large diameter collection header allow a designer to use a longer collection header and still achieve

balanced flow collection. The recommended maximum distance between orifices in the collection header is 10% of

the cell width. The relative potential for clogging with slotted or perforated drainage pipe versus a longer collection

header with fewer orifices is unknown. Finally, the outlet

piping should be designed to allow for clean-out by the

operator.

A simple device to adjust the water level in a separate,

covered, outlet box is recommended to achieve variable

water level control (see Chapter 6). It is recommended that

there be only one collection header and adjustable-level

device per cell of a multiple cell VSB system. The adjustable device should allow the operator to flood the VSB to a

depth of 50 mm (2 in.) above the surface of the media (for

help in weed control), and to draw-down or drain the cell

for maintenance.

5.5.4 System Depth, Width and Length

The impact of water depth on pollutant removal is not

clear. One problem with almost all published information

on VSB systems is that even though the media depth may

be known, the actual operating water level is not known.

The TTU study found slightly better BOD removal with

greater media depth, when comparing 45 cm (18 in) with

30 cm (12 in) systems operated at same areal loading, but

it is unclear if this was due only to the increased HRT. No

other study has tested this result or determined the optimum depth for a VSB system (George et al., 2000). One

study suggested that total root penetration of the media

was critical to pollutant removal and recommended that

system depth be set equal to the maximum root depth of

the wetland species to be used in the VSB (Gersberg et al.

1983). However, as discussed previously, plants supplied

with abundant nutrients near the surface will not necessarily grow roots to their maximum depth. As a safety factor Kadlec and Knight (1996) recommend allowing room

for solids accumulation in the bottom of the VSB, but the

need for this has not been proven. Typical average media

depths in VSB systems have ranged from 0.3 to 0.7 m (12

to 28 in.), and various researchers have recommended

depths from 0.4 to 0.6 m (16 to 24 in.).

As discussed previously there is evidence for preferential flow below the root zone through media with a higher

conductivity. In order to minimize this flow, a shallower

depth would be required. On the other hand, a shallower

depth may require a greater area to achieve a desired HRT.

Until future studies provide better information on optimum

water depth, it is recommended to use a design maximum

water depth (at the inlet of the VSB) of 0.40 m (16 in.). The

depth of the media will be defined by the level of the wastewater at the inlet and should be about 0.1 m (4 in.) deeper

than the water.

The overall width of a treatment system using VSBs is

defined by Darcy’s Law, which is a function of the flow,

ALR, water depth and hydraulic conductivity. The width of

a individual VSB is set by the ability of the inlet and outlet

structures to uniformly distribute and collect the flow without inducing short-circuiting. The recommended maximum

width in a TVA design manual is 61 m (200 ft.). If the design produces a larger value, the user should divide the

VSB into several cells that do not exceed 61 m in width. As

discussed previously, several researchers have noted that

most BOD and TSS is removed in the first few meters of a

VSB, but some recommend minimum lengths ranging from

12 to 30 m (40 to 100 ft) to prevent short-circuiting. The

recommended minimum length for this manual is 15 m (50

ft).

Although much has been made of the aspect (lengthwidth) ratio in early constructed wetlands literature, the only

prerequisite for treatment is the area as defined by the

ALR. A study by Bounds, et.al. (1998) found that there was

no significant difference in TSS or CBOD removal in three

parallel VSB systems with aspect ratios of 4:1, 10:1, and

30:1. In all three systems the majority of TSS and CBOD

was removed in the first third on the VSB. Removals were

also unaffected by stressing the systems with large hydraulic spikes and intermittent loading. The TTU study also

found no significant difference in systems with 1:4 and 4:1

aspect ratios (George et al., 2000). Therefore, the aspect

ratio is not a factor in the overall design. However, the recommended values for maximum width and minimum length

discussed previously will tend to result in individual VSB

cells with an length-width ratio between 1:1 and 1:2.

5.6 Design Example for a VSB Treating

Septic Tank or Primary Effluent

The design has two basic assumptions. First, the total

VSB has four zones (see Figure 5-11). The inlet and outlet

zones were discussed in section 5.5.1. Based on the literature as discussed previously, the initial treatment zone

will (1) occupy about 30% of the total area, (2) perform

most of the treatment, and (3) have a big decrease in hydraulic conductivity (use K = 1% of clean K). The final treatment zone will occupy the remaining 70% of the area and

have little change in hydraulic conductivity (use K = 10%

of clean K). The second basic assumption is that Darcy’s
115

Law, while not exact, it is good enough for design purposes. The sizing of the initial and final treatment zones

follows these steps:

1) determine the surface area, using recommended

ALR

2) determine the width, using Darcy’s Law

3) determine the length and headloss of the initial treatment zone, using Darcy’s Law

4) determine the length and headloss of the final treatment zone, using Darcy’s Law

5) determine bottom elevations, using bottom slope

6) determine water elevations throughout the VSB,

using headloss

7) determine water depths, accounting for bottom slope

and headloss

8) determine required media depth

9) determine the number of VSB cells

For this example the following values are given:

• Maximum Monthly Flow (Q) = 200 m3

/d

• Maximum Monthly Influent (C0) BOD = 100 mg/L =

100 g/m3

• Maximum Monthly Influent (C0) TSS = 100 mg/L = 100

g/m3

• Required discharge limits = 30 mg/L BOD and TSS

Recommended values for VSBs (see Table 5-4) are:

• ALR for BOD = 6 g/m2

-d

• ALR for TSS = 20 g/m2

-d

• Use washed, rounded media 20-30 mm in diameter,

clean K = 100,000 m/d

• Hydraulic conductivity of initial treatment zone (Ki) =

1% of 100,000 = 1000 m/d

• Hydraulic conductivity of final treatment zone (Kf

) =

10% of 100,000 = 10,000 m/d

• Bottom slope (s) = _% = 0.005

• Design water depth at inlet (Dw0) = 0.4 m

• Design water depth at beginning of final treatment zone

(Dwf) = 0.4 m

• Design media depth (Dm) = 0.6 m

• Maximum allowable headloss through initial treatment

zone (dhi) = 10% of Dm = 0.06 m

5.6.1 Determine the Surface Area (As) for

Both Pollutants

As = (Q)(C0) / ALR

For BOD, As = (200 m3/d)(100 g/m3) / 6 g/m2-d = 3333 m2

For TSS, As = (200 m3/d)(100 g/m3) / 20 g/m2-d = 1000 m2

Use the larger area requirement, or 3333 m2.

The surface area for the initial treatment zone (Asi) =

(30%)(3333 m2) = 1000 m2

The surface area for the final treatment zone (Asf) =

(70%)(3333 m2) = 2333 m2

5.6.2 Determine the Width

Determine the minimum width (W) needed to keep the

flow below the surface, using Darcy’s Law (Eq. 5-1) and

recommended values for the initial treatment zone.

Q = (Ki

)(W)(Dw0)(dhi/Li)

where: Li

 = length of initial treatment zone = (Asi) / (W)

Substitute and rearrange equation to solve for W:

W2 = (Q)(Asi) / (Ki)(dhi)(Dw0) (5-3)

For this example:

W2 = (200 m3

/d)(1000 m2) / (1000 m/d)(0.06 m)(0.4 m)

= 8333 m2

W = 91.3 m

This is the width for which the headloss equals 0.06 m,

given all the parameters as defined. The designer must

use a width equal to or greater than this to ensure that the

headloss is less than or equal to the design value.

5.6.3 Determine the Length and Headloss

(Eq. 5-2) of the Initial Treatment

Zone (Li)

Li = (Asi) / (W) = (1000 m2) / (91.3 m) = 11.0 m

This is the length for which the headloss equals 0.06 m,

given all the parameters as defined. The designer must

use a length less than or equal to this to ensure that the

headloss is less than or equal to the recommended value.

dhi= (Q)(Li) / (Ki)(W)(Dw0) = (200 m3/d)(11.0 m) / (1000

m/d)(91.3 m)(0.4 m) = 0.06 m

5.6.4 Determine the Length and Headloss

of the Final Treatment Zone (Lf

)

Lf = (Asf) / (W) = (2333 m2) / (91.3 m) = 25.6 m

This is the length where the total area of the VSB will be

exactly equal to the value set by the ALR. The designer

must use a length equal to or greater than this to ensure
116

that the surface area is equal to or greater than the recommended value.

dhf

= (Q)(Lf) / (Kf)(W)(Dwf) = (200 m3/d)(25.6 m) /

(10,000 m/d)(91.3 m)(0.4 m) = 0.01 m

5.6.5 Determine Bottom Elevations

Ebe = elevation of bottom at outlet = 0 (reference point

for all elevations)

Ebf = elevation of bottom at beginning of final treatment

zone = (s)(Lf

) = (0.005)(25.6 m) = 0.13 m

Eb0 = elevation of bottom at inlet = (s)(Li

 + Lf) =

(0.005)(11.0 m + 25.6 m) = 0.18 m

5.6.6 Determine the Water Surface

Elevations

Ewf = elevation of water surface at beginning of final

treatment zone

= Ebf + Dwf = 0.13 m + 0.4 m = 0.53 m (Dwf = 0.4 m

was an initial recommended value)

Ewe = elevation of water surface at outlet = Ewf - dhf =

0.53 m - 0.01 m = 0.52 m

Ew0 = elevation of water surface at inlet = Ewf + dhi =

0.53 m + 0.06 m = 0.59 m

5.6.7 Determine Water Depths

Dw0 = depth of water at inlet = Ew0 - Eb0 = 0.59 m - 0.18

m = 0.41 m (about equal to design Dw0, so okay.)

Dwf = depth of water at beginning of final treatment

zone

= Ewf - Ebf = 0.53 m - 0.13 m = 0.40 m (equal to

design Dwf, so okay.)

Dwe = depth of water at outlet = Ewe - Ebe = 0.52 m - 0 =

0.52 m

5.6.8 Determine the Media Depth

The media depth will depend on whether the designer

wants a level media surface, or a minimum depth-to-water

(Dtw) throughout the VSB.

a) If a level surface is desired, the elevation must be

greater than the highest water elevation, which is at the

inlet, Ew0 = 0.59 m. A media elevation set at 0.65 m would

be reasonable, and the following media depths and Dtw’s

result:

Dm0 = depth of media at inlet = 0.65 m - Eb0 = 0.65 m -

0.18 m = 0.47 m

Dmf = depth of media at beginning of final treatment

zone = 0.65 m - Ebf = 0.65 m - 0.13 m = 0.52 m

Dme = depth of media at outlet = 0.65 m - 0 = 0.65 m

Dtw0 = depth-to-water at inlet = 0.65 m - Ew0 = 0.65 m -

0.59 m = 0.06 m

Dtwf = depth-to-water at beginning of final treatment

zone = 0.65 m - Ewf = 0.65 m - 0.53 m = 0.12 m

Dtwe = depth-to-water at outlet = 0.65 m - Ewe = 0.65 m

- 0.52 m = 0.13 m

The depth-to-water is small at the inlet (0.06 m) and the

designer may want to add an additional layer of media in

the first few meters of the initial treatment zone as an added

precaution against surfacing, even though the design ALR

and Ki

 values is very conservative. The resulting Dtw in the

final treatment zone would be 0.12 to 0.13 m, which should

not inhibit the growth of aquatic species.

b) If a constant depth-to-water throughout the VSB is

desired (e.g. 0.1 m), then the media depth would be calculated as follows:

Em0 = elevation of media surface at inlet = Ew0 + 0.1 m

= 0.59 m + 0.1 m = 0.69 m

Emf = elevation of media surface at beginning of final

treatment zone

= Ewf + 0.1 m = 0.53 m + 0.1 m = 0.63 m

Eme = elevation of media surface at outlet = Ewe + 0.1

m = 0.52 m + 0.1 m = 0.62 m

Dm0 = depth of media at inlet = Em0 - Eb0 = 0.69 m -

0.18 m = 0.51 m

Dmf = depth of media at beginning of final treatment

zone = Emf - Ebf = 0.63 m - 0.13 m = 0.56 m

Dme = depth of media at outlet = Eme - 0 = 0.52 m - 0 =

0.52 m

This approach would result in a drop in the media surface of (0.69 m - 0.51 m) of 0.18 m over the 11.0 m length

of the initial treatment zone (slope = 1.6%), which would

probably not impair operation and maintenance activities.

5.6.9 Determine Number of VSB Cells

It is recommended that at least two VSBs be used in

parallel in all but the smallest systems, so that one of the

VSBs can be taken out of service for maintenance or repairs without causing serious water quality violations. In

this example, the total size of the VSB system is 91.3 m

wide by 36.6 m long. Therefore, use two VSBs, each 46 m

wide and 37 m long could be used. Other combinations of

length and width that have the required surface area will

also work as long as the hydraulics conditions are meet.

Also remember that inlet and outlet zones will add to the

overall length of the VSB.
117

5.7 On-site Applications

A number of states, including Louisiana, Kentucky, Kansas, Arkansas, Texas and Indiana, have used VSBs for

on-site wastewater management. Kentucky alone lists over

4000 such installations (Thom et al., 1998). Most of these

states have adopted VSBs as a pretreatment step prior to

soil infiltration in an effort to protect groundwater by overTable 5-4. Summary of VSB Design Guidance.

Recommended for use after primary

sedimentation (e.g. septic tank, Imhoff

tank, primary clarifier) VSBs not

recommended for use after ponds

Pretreatment because of problems with algae

Surface Area

Based on desired effluent quality and areal loading rates as follows:

BOD 6 g/m2-d (53.5 lb/ac-d) to attain 30 mg/L

effluent

BOD 1.6 g/m2-d (14.3 lb/ac-d) to attain 20 mg/L

effluent

TSS 20 g/m2-d (178 lb/ac-d) to attain 30 mg/L

effluent

TKN Use another treatment process in

conjunction with VSB

TP VSBs not recommended for phosphorus

removal

Depth

Media (typical) 0.5 - 0.6 m (20 - 24 in.)

Water (typical) 0.4 - 0.5 m (16 - 20 in.)

Length As calculated (see design example);

minimum of 15 m (49 ft)

Width As calculated (see design example);

maximum of 61 m (200 ft)

Bottom slope 0.5 - 1%

Top slope level or nearly level

Hydraulic Conductivity

First 30% of length 1% of clean K

Last 70% of length 10% of clean K

Media

All media should be washed clean of fines and debris; more rounded

media will generally have more void spaces; media should be resistant

to crushing or breakage.

Inlet zone 40 - 80 mm (1.5 - 3.0 in)

[1st 2 m (6.5 ft)]

Treatment zone 20 - 30 mm (3/4 - 1 in) [use clean K =

100,000, if actual K not known]

Outlet zone 40 - 80 mm (1.5 - 3.0 in)

[last 1 m (3.2 ft)]

Planting media 5 - 20 mm (1/4 - 3/4 in)

[top 10 cm (4 in.)]

Miscellaneous Use at least 2 VSBs in parallel

Use adjustable inlet device with

capability to balance flows

Use adjustable outlet control device with

capability to flood and drain system

coming the inadequate purification capacity of certain soils.

They view VSBs as passive systems with low operation

and maintenance requirements.

A review of various on-site VSB design guidelines by

Mankin and Powell (1998) revealed that there was a large

variation among the designs. Recommended depth varied from 0.3 to 0.8 m (1 to 2.5 ft), with the great majority

between 0.3 and 0.5 m (1 and 1.5 ft). For a three bedroom

house typical VSB areas varied from < 10 m2

 to > 100 m2

(104 to 1088 ft2), HRTs varied from 1.3 to 6.5 d, and lengthwidth ratio varied from 71:1 to 1.8:1. Median values were

a depth of 0.45 m (1.5 ft), area of 30 m2

 (315 ft2), and HRT

of 4.7 d. Gravel size guidelines varied from 0.65 cm (0.25

in.) to 7.5 cm (3.0 in.)

These authors sampled three typical VSB systems in

Kansas and compared them with other reported data. Despite employing a larger than average area, the units failed

to meet a 30/30 BOD/TSS requirement at two of the three

sites.

Given the general lack of operation and maintenance

requirements and the potential aesthetic appearance of

VSB systems, their attractiveness to local and state regulators is quite predictable. As a passive system potentially

capable of meeting a 30/30 BOD/TSS requirement, they

have obvious advantages over mechanical systems which

require a significant management program and electrical

support to function satisfactorily. Also, their general reliability, when compared to mechanical systems, offers additional protection against clogging of the soil’s infiltrative

surface.

Based on the VSB design guidance presented previously,

a three bedroom home would require a VSB of 100 m2,

assuming six persons and a BOD loading of 100 g/cap-d.

As expected, due to the conservative nature of the design

approach presented in this chapter, this area is at the high

end of the areas found by Mankin and Powell (1998).

There should be minimal deviation from the recommendations of Table 5-4, except that simplified inlet and outlet

configurations, appropriate for small on-site systems, can

be used. As with larger VSBs, some means of post-aeration and disinfection will be required if surface discharge is

contemplated. Discharge to soil infiltration is more likely,

and soil absorption guidelines provided by the State will

apply.

5.8 Alternative VSB Systems

Alternative VSB systems are those that operate with

some schedule of filling and draining the media. Fill and

drain VSB systems are similar to sequencing batch reactors, intermittent sand filters, or overland flow systems in

that the flow into a single cell of a system is intermittent.

Draining a VSB system is a simple way to introduce more

oxygen into the media. Clearly plants are playing a lesser

role in these systems and they are inherently quite different from a natural wetland. Nevertheless they are discussed

here because they have been identified as constructed
118

wetlands and have evolved from conventional VSB systems. They are more complex to operate than a conventional VSB system.

One of the first and more unconventional alternative VSB

systems was developed in England and tested in England

and Egypt in the late 1980s (May et al., 1990). The system, called a gravel bed hydroponic system, is very similar to an overland flow process except that the wastewater

flows through 8-10 cm (3-4 in) of gravel. Loading is typically intermittent except when denitrification is desired.

Researchers at TTU (1999) experimented with alternating fill and drain VSB wetlands for one year in 1994 and

compared the results to conventional VSBs in side by side

testing. The alternating fill and drain cells followed a conventional VSB cell. Effluent from the fill and drain cells was

recycled back to the conventional cell. They found some

improvement in nitrogen removal but overall the results

were not as good as they had expected.

Researchers at TVA have developed (and patented) a

system in which the wastewater is quickly drained from

one wetland cell and pumped into a second parallel cell

(Behrends, et al., 1996). The draining and filling occurs

within 2 hours and then the process is reversed; the second cell is drained quickly and the first cell is refilled. The

reciprocating flow process is repeated continuously, with

a small amount of influent continually added to the first

cell and a fraction of the wastewater continually drawn from

the second cell as effluent. The reciprocating two cell system was compared with a conventional two cell system for

six months in side by side testing in late 1995 and early

1996 at Benton, Tennessee. Continued operation of both

two-cell pairs in the reciprocating mode has continued since

May of 1996. Comparing conventional operation to the

reciprocating mode, the reciprocating mode produced significantly lower effluent BOD and ammonia nitrogen.

One of the most studied full-scale VSB systems is located at the Village of Minoa in New York State (see Chapter 9). Two New York State agencies and the USEPA provided grant funds to the Village for incorporation of several special features in the VSB system and for a research

and technology transfer study of the system by researchers at Clarkson University, Potsdam, NY. The system was

originally designed and operated as a conventional VSB

system, but during 15 months treating a primary effluent,

the system performed very poorly compared to its design

expectations. Faced with numerous complaints from

nearby residents about hydrogen sulfide odors, the operators started operating the system with occasional drawdown periods to control odors. The drawdown significantly

reduced odors. In April 1997, when the experimental plan

for the system called for the three cells to operated in series, the Minoa operators decided to increase the flow by

100% and change the operation to a fill and drain mode.

The fill and drain operation included a resting period during the drained condition and continuous operation for some

time after filling. The fill and drain operation eliminated the

hydrogen sulfide odors and also resulted in a significant

improvement in the effluent quality. However, the operators were not satisfied with the improved performance and

experimented further. In 1998 they changed the operation

to a mode that continued to the writing of this manual. Two

of the wetland cells operate in a parallel fill and drain mode

very similar to sequencing batch reactors. The third cell is

operated in a conventional mode but in series with the first

two cells. This mode of operation has resulted in an additional significant improvement in effluent quality over the

previous fill and drain mode of operation. See section 9.8

for a more detailed description of the operation and the

results.

Within the last five years, several unsaturated vertical

flow systems have been constructed and tested in Europe.

Most have been used for tertiary treatment of secondary

effluents but they have also been used for treating septic

tank and sugar beet processing effluents. They appear to

perform significantly better than conventional VSB systems.

Recommended design loadings are approximately twice

that for conventional VSB systems.

5.9 References

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pilot scale subsurface flow wetland: 10 years after commissioning. Water Science & Technology, 35(5):337-343.

Bavor, H.J., D.J. Roser, P.J. Fisher, and I.C. Smalls. 1989.

Performance of solid-matrix wetland systems viewed as

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Wetlands for Wastewater Treatment. Chelsea, MI: Lewis

Publishers, pp. 646-656.

Bavor, H.J. and T.J. Schulz. 1993. Sustainable suspended

solids and nutrient removal in large-scale, solid matrix,

constructed wetland systems. In: G.A. Moshiri (ed.) Constructed wetlands for water quality improvement. Boca

Raton, FL:Lewis Publishers, pp. 219-225.

Behrends, L.L., Coonrod, H.S., Bailey E. and M.J. Bulls. 1993.

Oxygen Diffusion Rates in Reciprocating Rock Biofilters:

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of Texas at El Paso.

Behrends, L.L., F. J. Sikora, H.S. Coonrod, E. Bailey and C.

McDonald. 1996. Reciprocating Subsurface-Flow Constructed Wetlands for Removing Ammonia, Nitrate, and

Chemical Oxygen Demand: Potential for Treating Domestic, Industrial and Agricultural Wastewater. Vol 5, Pp

251-263. In: Proceedings of the Water Environment

Federation 69th Annual Conference. Dallas, TX.

Bhattarai, R.R. and D.M. Griffin, Jr. 1998. Results of tracer

tests in rock plant filters. Department of Civil Engineering, Louisiana Tech University, Ruston, LA.

Bounds, H.C., J. Collins, Z. Liu, Z. Qin, and T.A. Sasek. 1998.

Effects of length-width ratio and stress on rock-plant filter operation. Small Flow Journal, 4(1):4-14.
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Bowmer, K.H. 1987. Nutrient removal from effluents by

an artificial wetland: influence of rhizosphere aeration and preferential flow studied using bromide and

dye tracers. Water Research, 21(5):591-599.

Breen, P.F. and A.J. Chick. 1995. Rootzone dynamics

in constructed wetlands receiving wastewater: a

comparison of vertical and horizontal flow systems.

Water Science & Technology, 32(3):281-290.

Chalk, E. and G. Wheale. 1989. The root-zone process

at Holtby Sewage Treatment Works. Journal IWEM,

3:201-207.

Cooper, P.F. 1990. European Design and Operations

Guidelines for Reed Bed Treatment Systems, Rep.

UI17, Water Research Centre, Swindon, U.K.

Cooper, P.F., J.A. Hobson and S. Jones. 1989. Sewage

Treatment by Reed Bed Systems. Journal of the Institution of Water and Environmental Management.

3 (1) 60.

Crites, R. and G.Tchobanoglous. 1998. Small and decentralized wastewater management systems. San

Francisco, CA: McGraw-Hill.

Dahab, M.F. and R.Y. Surampalli. 1999. Predicting Subsurface Flow constructed Wetlands Performance: A

Comparison of Common Design Models. In: Proceedings of the Water Environment Federation 72th

Annual Conference. New Orleans, LA.

DeShon, G.C., A.L. Thompson, and D.M. Sievers. 1995.

Hydraulic properties and relationships for the design of subsurface flow wetlands. Presented at Versatility of Wetlands in the Agricultural Landscape

Conference, Tampa, FL, Sept. 17-20, 1995.

Fisher, P.J. 1990. Hydraulic characteristics of constructed wetlands at Richmond, NSW, Australia. In:

P.F. Cooper and B.C. Findlater (eds.) Constructed

Wetlands in Water Pollution Control. Oxford, UK:

Pergamon Press, pp. 21-31.

Gearheart, R.A. 1998. Use of FWS constructed wetlands

as an alternative process treatment train to meet

unrestricted water reclamation standards. Presented

at AWT-98, Advanced Wastewater Treatment, Recycling and Reuse, Milan, Italy, pp. 559-567.

Gearheart, R.A. et al. 1999. Free water surface wetlands for wastewater treatment: a technology assessment. USEPA, Office of Water Management, US

Bureau of Reclamation, City of Phoenix, AZ.

George, D.B. et al. 2000. Development of guidelines

and design equations for subsurface flow constructed wetlands treating municipal wastewater.

USEPA, Office of Research and Development, Cincinnati, OH.

Gersberg, R.M., B.V. Elkins and C.R. Goldman. 1983.

Nitrogen Removal in Artificial Wetlands. Water Research 17 (9) 1009.

Gersberg, R.M. et al. 1984. The Removal of Heavy Metals by Artificial Wetlands, In: Proc Water Reuse

Symp. III, Vol 2, AWWA Research Foundation, 639.

Gersberg, R.M., et al. 1986. Role of Aquatic Plants in

Wastewater Treatment by Artificial Wetlands. Water Research 20 (3) 363.

Gersberg, R.M., Gearheart, R.A., and M. Ives. 1989.

Pathogen Removal in Constructed Wetlands, Proc.

From First International Conference on Wetlands for

Wastewater Treatment, Chattanooga, TN, June

1988, Ann Arbor Press.

Green, M.B. and J. Upton. 1994. Constructed Reed

Beds: A Cost-Effective Way to Polish Wastewater

Effluents for Small Communities. Water Env. Res.

66 (3) 188.

Griffin, P., B. Green and A.Pritchard. 1998. Pathogen

Removal in Subsurface Flow Constructed Reed

Beds. In: Proceedings of the Water Environment

Federation 71st Annual Conference. Orlando, FL.

Heukelekian, H. and J.L. Balmat. 1959. Chemical composition of the particulate fractions of domestic sewage. Sewage & Industrial Wastes, 81:413-423.

Jenssen, P.T. M. Muehlan, and T. Kregstad. 1993. Potential use of constructed wetlands for wastewater

treatment in northern environments. In: Proceedings

of 2nd International Conference on Design and Operation of Small Wastewater Treatment Plants, pp.

193-200.

Kadlec, R.H. and R.L. Knight. 1996. Treatment Wetlands. Boca Raton, FL: Lewis-CRC Press.

Kadlec, R.H. and J.T. Watson. 1993. Hydraulics and solids accumulation in a gravel bed treatment wetland.

In: G.A. Moshiri (ed.) Constructed wetlands for water quality improvement. Boca Raton, FL:Lewis Publishers, pp. 227-235.

Kickuth, R. 1981. Abwasserreinigung in mosaikmatrizen

aus aeroben und anaerobenteilbezirken. In: F.

Moser (Ed), Grundlagen der Abwassereinigung, pp

639-665.

King, A.C., C.A. Mitchell, and T. Howes. 1997. Hydraulic tracer studies in a pilot scale subsurface flow constructed wetland. Water Science & Technology,

35(5):189-196.

Liehr, R.K. et al. 2000. Constructed wetlands treatment

of high nitrogen landfill leachate. Project Number

94-IRM-U, Water Environment Research Foundation, Alexandria, VA.
120

Macmanus, B.E., A.L. Thompson, and D.M. Sievers.

1992. Predicting water mounding in subsurface rock

bed wetlands. Presented at the Mid-Central Conference of the American Society of Agricultural Engineers, St. Joseph, MO, March 13-14, 1992.

Mankin, K.R. and G.M. Powell. 1998. Onsite rock-plant

filter monitoring and evaluation in Kansas. In: Proceedings of 8th National Symposium on Individual

and Small Community Sewage Systems, ASAE, St.

Joseph, MI.

May, E., J.E. Bulter, M.G. Ford, R.F. Ashworth, J.S. Williams and M.M.M. Baghat. 1990. Comparison of

Chemical and Microbiological Processes in Gravel

Bed Hydroponic (GBH) Systems for Sewerage

Treatment. In: Constructed Wetlands in Water Pollution Control. Cooper and Findlater (Eds)

Pergamon Press U.K.

Netter, R. and W. Bischofsberger. 1990. Hydraulic investigations on planted soil filters. In: P.F. Cooper

and B.C. Findlater (eds.) Constructed Wetlands in

Water Pollution Control. Oxford, UK: Pergamon

Press, pp. 11-20.

Netter, R. 1994. Flow characteristics of planted soil filters. Water Science & Technology, 29(4):37-44.

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treatment. In: Proceedings of 7th European Sewage and Refuse Symposium, EWPCA, pp. 351-400.

Peverly, J.H., J.M. Surface and T. Wang. 1995. Growth

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of constructed wetlands treating landfill leachate.

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Natural Systems for Waste Management and Treatment. 2nd Ed. NY: McGraw Hill.

Reed, S.C. and S. Giarrusso. 1999. Sequencing Operation Provides Aerobic Conditions in a Constructed Wetland. In: Proceedings of the Water Environment Federation 72th Annual Conference. New

Orleans, LA.

Richard, M. and J. Snyder. 1994. Results of the pilot

wetlands study at Las Amimas, CO. Report to the

City of Las Animas, Colorado State University, CO.

Sanford, W.E., T.S. Steenhuis, J-Y. Parlange, J.M. Surface, and J.H. Peverly. 1995a. Hydraulic conductivity of gravel and sand as substrates in rock-reed

filters. Ecological Engineering, 4:321-336.

Sanford, W.E., T.S. Steenhuis, J.M. Surface, and J.H.

Peverly. 1995b. Flow characteristics of rock-reed filters for treatment of landfill leachate. Ecological Engineering, 5:37-50.

Sanford, W.E. 1999. Substrate type, flow characteristics,

and detention times related to landfill leachate treatment efficiency in constructed wetlands. In: G.

Mulamootil, E.A. McBean, and F. Rovers (eds.) Constructed wetlands for the treatment of landfill leachate.

Boca Raton, FL:Lewis Publishers, pp. 47-56.

Sapkota, D.P. and H.J. Bavor. 1994. Gravel bed filtration

as a constructed wetland component for the reduction

of suspended solids from maturation pond effluent.

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Smith, I.D., G.N. Bis, E.R. Lemon and L.R, Rozema. 1997.

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artificial wetlands. In: D.A. Hammer (ed.) Constructed

Wetlands for Wastewater Treatment. Chelsea, MI:

Lewis Publishers, pp. 484-492.

Surface, J.M., J.H. Peverly, T.S. Steenhuis, and W.E.

Sanford. 1993. Effect of season, substrate composition, and plant growth on landfill leachate treatment in

a constructed wetland. In: G.A. Moshiri (ed.) Constructed wetlands for water quality improvement. Boca

Raton, FL:Lewis Publishers, pp. 461-472.

Tanner, C.C. and J.P. Sukias. 1995. Accumulation of organic solids in gravel-bed constructed wetlands. Water Science & Technology, 32(3):229-239.

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122

6.1 Introduction

Constructed wetland systems require infrequent operation and maintenance activities to achieve performance

goals if they are designed and constructed properly. This

chapter discusses construction details, start-up procedures,

and operation and maintenance activities for both free water

surface wetlands and vegetated submerged beds.

6.2 Construction

Construction of wetland systems primarily involves common earth moving, excavating, backfilling, and grading.

Most of the equipment and procedures are the same as

those employed for construction of lagoons, shallow ponds,

and similar containment basins. However, there are aspects that require special attention to ensure flow through

the wetland is uniform over the design treatment volume.

Also, establishment of vegetation is unique to the basin

construction and not always within the repertoire of construction contractors. It is the intent of this section to provide guidance on these special and unique aspects of

wetland construction.

6.2.1 Basin Construction

The basic containment structure of constructed wetlands

consists of berms and liners. The structural and watertight

integrity of the liner and surrounding berm are critical. Failure of either will result in loss of water, risk of ground water

pollution, and possible loss of plants due to the decline of

the water level in the wetland.

6.2.1.1 Basin Layout

The topography of the site will dictate the general shape

and configuration of the wetland. Constructing the wetland

on sloping sites with the long axis along the contour will

minimize the grading requirements. With proper layout, long

sloping sites can reduce pumping costs by taking advantage of the available fall.

6.2.1.2 Site Preparation

Clearing and grubbing, rough grading, and berm construction use the same procedures, techniques, and equipment used for lagoons and conventional water containment basins. If possible, it is desirable to balance the cut

and fill on the site to avoid the need for remote borrow pits

or soil disposal. If agronomic-quality topsoil exists on the

site, it should be stripped and stockpiled. In the case of

FWS wetlands, the topsoil can be utilized as the rooting

medium for the emergent vegetation and revegetation of

the berm surfaces. A soil-rooting medium is not required

for VSB systems.

To meet its performance expectations, it is critically important for the water to flow uniformly through the entire

wetland area. Severe short-circuiting of flow can result from

improper grading or nonuniform subgrade compaction. The

operating water depth may be 60 cm (2 ft) or less, so irregularities in the bottom surface can induce preferential

flow paths. Specified tolerances for grading will depend

on the size of the wetland. A very large FWS wetland of

several thousand acres cannot afford the effort to fine grade

to very close tolerances. Therefore, the wetland should be

subdivided into several smaller cells or the design should

incorporate a sizing safety factor to compensate for potential short-circuiting. For smaller wetlands of a few hundred hectares or less, it is usually cost effective to specify

closer grading tolerances. Bottom grades are an important consideration when converting existing lagoons to wetlands. Because of the design depths in lagoons, careful

grading of the bottom may not have been required. In many

cases in which conversions were made without careful

regrading, significant short-circuiting has occurred that reduced the wetland treatment performance.

Uniform compaction of the subgrade is also important to

protect the liner integrity from subsequent construction

activity (i.e., liner placement, soil placement for FWS wetlands, gravel placement for VSB systems) and from stress

when the wetland is filled. The loading on the liner is approximately 2,200 kg/m2

 (450 lbs/ft2) including the plant

mass. Short-circuiting of flow through a FWS wetland also

can result from ruts and low areas in the subgrades. The

subgrade should be uniformly compacted to the same levels used for native soils in road subgrades.

Fine grading and compaction of the native subgrade soils

also depends on the liner requirements. Most wetland cells

are graded level from side to side and either level or with a

slight slope in the direction of flow. Wetlands are often constructed with a bottom slope of 1% or less which is sufficient to drain the cell if and when maintenance is required.

Chapter 6

Construction, Start-up, Operation, and Maintenance
123

6.2.1.3 Berms

Berms in constructed wetlands contain water within specific flow paths. Exterior berms are designed to prevent

unregulated flow releases. Interior berms are used to augment flow distribution. External berms are typically built to

provide 0.6 to 1 m (2–3 ft) of freeboard with a width at

least 3 m (10 ft) at the top to permit service vehicle access. The amount of freeboard should be enough to contain a given storm rainfall amount. Side slopes should be

a maximum of 3:1; however, slopes of 2:1 have been used

for internal side slopes, particularly when liners or erosion

control blankets are used. Access ramps into each cell of

the system should be shallow enough for maintenance

equipment to enter. All berms should be constructed in

conformance with standard geotechnical considerations,

for they may be subject to local dam safety regulations.

Design considerations for internal berms, however, are less

critical since they are not designed for water containment.

See Figure 6-1 for typical design features of constructed

wetland berms.

Short-circuiting around the edges of cells has been experienced in some FWS wetlands where vegetation on the

berm slope is absent. This is a particular problem if synthetic liners are used. The liners do not provide a good

rooting medium and so may remain bare. The open water

gap between the berm and the vegetated area in the wetland proper provides a preferential flow path. A soil layer

can be placed on the berm side slope to establish vegetation, but the slope is very susceptible to erosion, particularly near the water line. The soil loss from erosion will

have the added impact of reducing the detention time in

the wetland. This has not been a problem in clay-lined

wetlands because the clay provides a good rooting medium.

6.2.1.4 Liners

Liners used for wetlands are the same as those typically

used for lagoons and ponds. The materials include:

Polyvinyl chloride (PVC)

Polyethylene (PE)

Polypropylene

Most systems typically use 30 mil polyvinyl chloride

(PVC) or high-density polyethylene (HDP). These may be

prefabricated for small, individual-residence wetlands, but

a) Exterior Berm

b) Interior Berm

Impervious Clay

Plug (optional)

Grassed

Slope

(2:1 to 3:1) Grassed Slope (≥ 3:1)

Grassed Berm

( ≥ 3:1 typical)

Gravel Road (≥ 3m wide)

Synthetic Liner

Synthetic Liner

0.6 - 0.9 m Freeboard

0.6 - 0.9 m 

Freeboard

0.6 - 0.9 m 

Freeboard

Figure 6-1. Examples of constructed wetland berm construction
124

they are usually constructed in place using conventional

procedures for assembly, joint bonding, and anchoring.

Liners also may include scrims, which are more costly. The

scrim is a woven nylon or polypropylene net embedded in

plastic or surrounding bentonite. Plastic liners with scrims

are marketed under trade names such as Hypalon or XR5. Several good resources are available for liner application and selection (EPA, 1993; EPA, 1994; Rumer and

Mitchell, 1996).

Liner punctures must be prevented during placement and

subsequent construction activity. If the subgrade contains

sharp stones, a geotextile fabric should be placed beneath

the liner. A geotextile fabric or a layer of sand approximately

5 cm (2 in) thick should be placed on top of the liner if

crushed rock is used in a VSB system. The engineer should

specify that the liner installer provides written approval of

the condition of the subgrade as a condition prior to liner

installation.

Many membrane liners currently used require protection from ultraviolet solar radiation. Conventional methods

can be used to achieve protection, but VSB systems should

not use a soil cover as UV protection since erosion may

wash soil into the bed and result in local media clogging.

Riprap material consisting of aggregate approximately 8–

15 cm (3–6 in) in size is recommended for this application.

This larger riprap will reduce the potential for weeds to

become established and spread into the wetland. It can

also withstand foot traffic for the life of the system.

Clay liners also have been used. Manufactured liners

using bentonite are common. The bentonite may be mixed

with the native soils and compacted, or it may be in the

form of pads or blankets consisting of bentonite between

two scrims of finely woven polypropylene or polyethylene.

Native soils may be used if they have sufficiently high

clay content to achieve the necessary permeability. Usually the state regulatory agency will specify the acceptable

permeability. Typically, the clay liner must be 0.3 m (1 ft) or

more in thickness to provide the necessary hydraulic barrier. In the case of a FWS, the surface of the clay layer

should be well compacted to discourage root penetration

by the emergent vegetation as the wetland matures.

6.2.1.5 Inlet and Outlet Structures

Inlet and outlet structures distribute the flow into the

wetland, control the flow path through the wetland, and

control the water depth. Multiple inlets and outlets spaced

across either end of the wetland are essential to ensure

uniform influent distribution into and flow through the wetland. These structures help to prevent “dead zones” where

exchange of water is poor, resulting in wastewater detention times that can be much less than the theoretical detention times.

In small- to medium-sized wetlands, perforated or slotted manifolds running the entire wetland width typically are

used for both the inlets and outlets. Sizes of the manifolds, orifice diameters, and spacing are a function of the

projected flow rate. For example, the first cell of the FWS

wetland in West Jackson County, MI, is designed for an

average flow of 2,270 m3/d (600,000 gpd). It uses a 300

mm (12 in)-diameter PVC manifold for the inlet that extends the full 76 m (250 ft) width of the cell. The manifold

is perforated with 50 mm (2 in)-diameter orifices on 3 m

(10 ft) centers. It rests on a concrete footing to ensure stability and discharges to a 150 mm (6 in)-thick layer of coarse

aggregate. A single inlet would not be suitable for a wide

wetland cell such as this because it would not be possible

to achieve uniform flow across the cell. Multiple weir boxes

could be used as an alternative. Splitter boxes using “V”

notched weirs or other methods can be used to divide the

influent flow equally between the individual weir boxes.

The weir boxes also can be used for measuring the influent flow. Examples of these types of structures can be found

in irrigation engineering textbooks.

Where possible, the inlet manifold should be installed in

an exposed position to allow access by the operator for

flow adjustment and maintenance. Several alternatives to

the simple drilled orifice can be used for flow distribution

control. See Figure 6-2 for examples of inlet manifolds.

In cold climates where extended periods of freezing

weather are possible or where public exposure is an issue, a submerged inlet is necessary. In these instances,

the simple perforated inlet manifold is used. Since it is not

possible to adjust the level of the submerged manifolds

after construction is completed, extra effort should be expended to compact and grade the inlet and outlet zones to

limit subsequent settling. It may be necessary to support

the manifold on concrete footings where the underlying

soils are potentially unstable. An accessible cleanout should

be provided at each end of the submerged manifold to

allow flushing if the manifold becomes clogged. Shut-off

devices should be provided on all inlets to permit maintenance or resting of the wetland.

In FWS wetlands, the encroachment of adjacent emergent vegetation may clog the manifold outlets with plant

litter and detritus. This problem may be eliminated by constructing a deep water zone approximately 1–1.3 m (3–4

ft) deeper than the bottom of the rest of the wetland. The

open area should be limited to 1 m (3 ft) in width. The

manifold also can be enclosed in a berm of coarse riprap

8–15 cm (3–6 in) in size. The coarse riprap inhibits plant

growth. The open water design, however, allows easier

access to the manifold for maintenance, but may encourage wildlife visitation and the potential effluent quality degradation that accompanies it.

Outlet structures help to control uniform flow through the

wetland as well as the operating depth. If submerged outlet manifolds are used, they must be connected to a level

control device that permits the operator to adjust the water

depth in the wetland. This device can be an adjustable

weir or gate, a series of stop logs, or a swiveling elbow

(Figure 6-3).

An alternative to submerged manifolds for inlet and outlet structures is multiple weir or drop boxes. These are
125

Figure 6-2. Examples of constructed wetland inlet designs

Control Valve

Gravel

Cleanout

a) Submerged Perforated Pipe

b) Gabion Feed

Wire Mesh Gabions

c) Swivel Tee

Cleanout (both ends)

Settled

Sewage

or Effluent

Reed Bed

Reed Bed

Level Surface

Liner

Liner Wire Mesh Gabion

with 60 - 100mm Stones 

Wire Mesh Gabion (optional)

with 60 - 100mm Stones 

0.5m

0.5m

Back-filled

with Stones

Soil Cover over Liner

Swivelling uPVC Pipe Tees

90° Tees with "O" Ring Seals Inlet
126

Figure 6-3. Outlet devices

a) Adjustable Weir

Outlet

Outlet

Debris

Screen

Debris

Screen

Adjustable Weir

Adjustable

Weir

Slotted Pipe

Collector

Liner

Interchangeable Section

of Pipe Fits "O" Ring Socket

Wire Mesh Gabion (optional)

with 60 - 100mm Stones 

Water

Level

b) Interchangeable Section

Slotted Pipe

Collector

Water

Level

"O" Ring Joint

c) 90° Elbow Arrangement

Wire Mesh Gabion (optional)

with 60 - 100mm Stones 
127

usually constructed of concrete, either cast in place or prefabricated. Several boxes must be installed across the width

of the wetland to ensure uniform flow through the wetland.

Preferred spacing varies from 5 to 10 m (15–30 ft) but may

be as much as 15 m (50 ft) on center depending on the

width of the wetland cell. Overflow rates should be limited

to <200 m3/m-d (16,000 gpd/ft2). Drop boxes do require a

deep water zone immediately around them to minimize

vegetation encroachment. In northern climates, the boxes

are more susceptible to freezing than are submerged manifolds.

Debris screens may be placed in front of FWS wetland

outlets. In Figure 6-3 there is an example of their placement in the outlet. The emergent vegetation in the wetland

will drop many leaves, and storm events can uproot entire

plants that float to the collection manifolds or outlet structures. The screens will prevent the debris from clogging

the downstream piping or treatment processes or impairing effluent quality.

6.2.1.6 Media

A soil or finer-rock medium is necessary as a matrix in

both FWS wetlands and VSB systems for supporting emergent vegetation. In FWS wetlands, a layer of soil at least

15 cm (6 in) deep is placed on the compacted bottom or

liner to create the rooting medium for the intended emergent vegetation. This soil can be the topsoil removed during initial site preparation and grading or can be imported.

Any loamy soil with acceptable agronomic properties is

suitable.

In VSB systems, the media provide the matrix for water

flow as well as the planting medium. Gravel is the most

commonly used media, but sand, crushed rock, and plastic media also have been used (Kadlec and Knight, 1996).

Large gravel media are typically recommended to prevent

clogging, whereas a smaller media layer can be used on

top of a bed of larger gravel to provide a better rooting

media (see Chapter 5).

Most local sand and gravel suppliers have concrete aggregate available that can be screened for the coarse fractions to provide an acceptable media for VSB systems.

The media should be washed to eliminate soil and other

fines that can contribute to media clogging. Rounded river

gravel is recommended over sharp-edged crushed rock

because of the looser packing that the rounded rock provides. Hard, durable stone (river gravel or crushed stone)

is recommended. Crushed limestone, which is soft and

easily disintegrates, should be avoided.

6.2.2 Vegetation Establishment

Vegetation and its litter is necessary for successful performance of FWS wetlands and contributes aesthetically

to the appearance of both FWS wetlands and VSB systems. Establishing this vegetation is probably the least familiar aspect of wetland construction for most contractors.

In recent years, a number of specialty firms have emerged

with the expertise for selecting and planting the vegetation

in these systems. Employing one of these firms is recommended for large projects if the construction contractor

does not have prior wetland experience.

6.2.2.1 Species Selection and Sources

For wastewater treatment, macrophytes selected for

planting should (1) be active vegetative colonizers with

spreading rhizome systems, (2) have considerable biomass or stem densities to achieve maximum velocity gradient and enhanced flocculation and sedimentation, and (3)

be a combination of species that will provide coverage over

the broadest range of water depths encountered (Allen et

al., 1990).

Wetland plants can be purchased from nurseries, collected in the wild, or grown for a specific project. No general recommendation can be made as to the best source

of plants for a particular project. However, wild collected

plants are usually the most desirable because they are

more closely adapted to local environmental conditions,

can be planted with limited storage, and offer a greater

diversity. For large projects, commercial seedlings may be

the most cost-effective alternative. The seedlings are supplied in suitable planting condition that allows use of mechanical agricultural equipment for planting.

6.2.2.2 Planting

Establishing vegetation in a constructed wetland involves

planting a suitable propagule at the appropriate time. Whole

plants or dormant rhizomes and tubers are typically planted.

Seeding has not been particularly successful because of

stratification requirements of wetland seed and loss of seed

from water action.

In temperate climates, the prime planting period begins

after dormancy has begun in the fall and ends after the

first third of the summer growing season has passed. The

planting period for herbaceous vegetation is broader than

for woody plants. Early spring growing-season plantings

have been the most successful (Allen et al., 1990).

Planting seedlings or clumps is the simplest method.

Some experience is necessary with rhizomes to identify

the node of the future shoot, which must be planted upward. Special anchoring may be necessary when the planting medium is soft, plants are buoyant, or erosion will disturb the system. The soil should be maintained in a moist

condition after planting. The water level can be increased

slowly as new shoots develop and grow. The water level

must never be higher than the tips of the green shoots or

the plants will die.

The macrophyte planting density can be as close as 0.3

m (1 ft) centers or as much as 1 m (3 ft). The higher the

density, the more rapid will be the development of a mature and completely functional wetland system. However,

high density plantings will increase construction costs significantly. If planted on 1 m (3 ft) centers, a wetland system will take at least two full growing seasons to approach

equilibrium and optimal plant-related performance objec-
128

tives. FWS wetlands should be planted more densely owing to the role of the plants in the treatment process. However, it may not be economical to plant very large FWS

wetlands on 1 m (3 ft) centers if the total wetland area is

intended to cover several thousand hectares. In such

cases, plantings should be done in separate bands extending the full width of the wetland cells to interrupt preferential flow of wastewater through the cells. VSB systems

can be planted less densely.

Water must be provided during the initial growth period.

This can be complicated in large FWS wetlands because

it may not be possible to plant the entire surface in a cell at

one time. If mechanical equipment is used for planting,

the unplanted areas should be kept dry until planting is

complete. Since the bottom is sloped toward the discharge

end, planting should start at the outlet and proceed toward

the inlet. Sprinklers and shallow flooding have been used

to keep the planted areas wet. If planting by hand, the whole

area can be flooded with a few centimeters of water. The

water depth can be increased gradually as the plant shoots

grow until the design level is reached. If the FWS wetland

is designed to treat a high-strength influent such as primary treated wastewater, a cleaner water source or diluting the wastewater with storm water or well water is recommended for the initial planting and growth period so the

plants are not overly stressed. If the intended influent is

close to secondary quality, it can be used immediately. If

an acceptable agronomic soil has not been used as the

rooting media, a preliminary application of commercial fertilizers may be necessary. Use of fertilizers should be carefully considered, however, because of the potential impacts

of the nutrients that might escape in the effluent to the

receiving water.

In VSB systems, it is typical practice to flood the wetland

cell to the surface of the media prior to planting and to

maintain that level until significant growth has occurred.

Later, the water level is lowered to the intended operating

level. If the wetland is designed to treat septic tank or primary effluent, clean water is recommended for the planting phase. The high oxygen demand of the wastewater

could inhibit initial plant growth. After a few weeks of plant

growth, wastewater can be introduced. A layer of straw or

hay mulch 15 to 20 cm (6–8 in) in thickness should be

placed on the gravel surface to protect the new plants from

the high summer surface temperatures that can occur on

bare gravel surfaces. The mulch also is useful for providing thermal insulation during the first winter of operation in

northern climates.

6.3 Start-up

Start-up periods for FWS wetlands are necessary to

establish the flora and fauna associated with the treatment

processes. The start-up period will vary in length depending on the type of design (FWS wetlands or VSB systems),

the characteristics of the influent wastewater, and the season of year. In FWS wetlands, the start-up period should

be sufficient for the vegetation to become well established

if the treatment objectives are to be met. The start-up period for VSB systems is less critical since its performance

is less dependent on vegetation.

6.3.1 Free Water Surface Wetlands

FWS wetlands will not attain optimum performance levels until the vegetation and litter are fully developed and at

equilibrium. The time required to reach this point is a function of the planting density and season of the year. A wetland with a high density planting that is started in the spring

is likely to be fully developed by the end of the second

growing season. A wetland with a low density planting

started in late fall in a northern climate may require three

years or more to achieve its intended treatment performance.

Under ideal conditions, start-up of a FWS wetland should

be delayed six weeks after planting to provide sufficient

time for the emergent plants to acclimatize and grow above

the working water level. When this is not possible, start-up

must control the water level at less than plant height. However, such rapid start-up will risk damage to the new plants

and may prolong the time required for the system to reach

optimum performance.

When start-up is initiated, the water level must be gradually raised to the design level by adjusting the flow control

device at the outlet of each cell. This is done to allow the

tops of the emerging vegetation to remain above water. If

the influent is high strength, such as primary or septic tank

effluent, it may be necessary to dilute the influent with clear

water or recycle treated effluent to slowly increase the

pollutant loadings to the wetland until the vegetation is

acclimatized.

During the start-up period, the operator should inspect

the wetland several times per week. Plant health and

growth should be observed, berms and dikes inspected

for structural problems, water levels adjusted, and mosquito emergence noted. In large areas of a FWS system

where growth of the vegetation has failed, the macrophytes

should be replanted to avoid the risk of short-circuiting of

flow. The experience developed during this period will be

helpful in determining the inspection frequency that will be

required during the mature phase of the wetland.

Treatment performance during the start-up period may

not be representative of long-term expectations. Poorly

established FWS wetlands with minimal vegetation will not

perform acceptably. Influent TSS and associated pollutants will not be properly removed. Large open areas will

permit algae blooms. The system will not perform differently from a maturation pond, in that only pathogen kill will

likely occur. Removal efficiency of TSS and its associated

pollutants can be expected to improve as the plant canopy

develops and increases in density. Removals of ammonia

and phosphorus may be greater during the start-up period

than after equilibrium is reached in new FWS wetlands,

which have a new soil surface and rapidly growing vegetation. Both conditions provide a rapid but short-term removal of these nutrients. Adsorption sites on soil particles
129

can take up both ammonia and phosphorus, and the nutrient uptake by the plants during the rapid growing phase

can be significant. Within one or two years of start-up, however, removal of phosphorus will decline. Removal of ammonia nitrogen will decline also unless the system is a

FWS with substantial open areas.

6.3.2 Vegetated Submerged Bed Systems

Treatment in VSB systems is primarily BOD and TSS

removal through the trapping of particulate material in the

media. Some BOD removal may be reintroduced from biochemical methanogenisis of captured organic solids in the

anaerobic environment. Biological denitrification also may

occur if nitrates are present in the influent. Plants may take

up nutrients in the wastewater, but this is usually not significant. Phosphorus removal during the start-up period will

occur, but typically becomes minor as chemical exchange

sites on the media are filled. Since the plants play an insignificant role in the treatment performance, equilibrium

should be reached in less than one year unless high-capacity media is employed.

During the start-up period, the operator is primarily responsible for adjusting the water level in the wetland. Typically, the VSB systems will be filled to the surface of the

media at the end of planting. As the plants begin to root,

the water level can be gradually lowered to the design

operating level. It may be necessary to add fertilizer during this period until sufficient nutrients are made available

by the addition of wastewater.

6.4 Operation and Maintenance

Constructed wetlands are “natural” systems. As a result,

operation is mostly passive and requires little operator intervention. Operation involves simple procedures similar

to the requirements for operation of a facultative lagoon.

The operator must be observant, take appropriate actions

when problems develop, and conduct required monitoring

and operational monitoring as necessary. The most critical items in which operator intervention is necessary

areAdjustment of water levels

• Maintenance of flow uniformity (inlet and outlet structures)

• Management of vegetation

• Odor control

• Control of nuisance pests and insects

• Maintenance of berms and dikes

6.4.1 Water Level and Flow Control

Water level and flow control are usually the only operational variables that have a significant impact on a welldesigned constructed wetland’s performance. Changes in

water levels affect the hydraulic residence time, atmospheric oxygen diffusion into the water phase, and plant

cover. Significant changes in water levels should be investigated immediately, as they may be due to leaks,

clogged outlets, breached berms, storm water drainage,

or other causes.

Seasonal water level adjustment helps to prevent freezing in the winter. In cold climates, the water levels should

be raised approximately 50 cm (18 in) in late fall until an

ice sheet develops. Once the water surface is completely

frozen, the water levels can be lowered to create an insulating air pocket under the ice and snow cover to maintain

higher water temperatures in the wetland (Kadlec and

Knight, 1996; Grits and Knight, 1990). This procedure is

used for both FWS wetlands and VSB systems.

6.4.2 Maintenance of Flow Uniformity

Maintaining uniform flow across the wetland through inlet and outlet adjustments is extremely important to achieve

the expected treatment performance. The inlet and outlet

manifolds should be inspected routinely and regularly adjusted and cleaned of debris that may clog the inlets and

outlets. Debris removal and removal of bacterial slimes

from weir and screen surfaces will be necessary. Submerged inlet and outlet manifolds should be flushed periodically. Additional cleaning with a high-pressure water

spray or by mechanical means also may become necessary.

Influent suspended solids will accumulate near the inlets to the wetland. These accumulations can decrease

hydraulic detention times. Over time, accumulation of these

solids will require removal. VSB systems cannot be

desludged easily without draining and removing the media. Therefore, VSB systems should not be considered for

treating wastewaters with high suspended-solids loads,

such as facultative pond effluents, which have high algal

concentrations.

6.4.3 Vegetation Management

Routine maintenance of the wetland vegetation is not

required for systems operating within their design parameters and with precise bottom-depth control of vegetation.

Wetland plant communities are self-maintaining and will

grow, die, and regrow each year. Plants will naturally spread

to unvegetated areas with suitable environments (e.g.,

depth within plant’s range) and be displaced from areas

that are environmentally stressful. Operators must control

spreading into open water areas that are intended by design to be aerobic zones through harvesting.

The primary objective in vegetation management is to

maintain the desired plant communities where they are

intended to be within the wetland. This is achieved through

consistent pretreatment process operation, small, infrequent changes in the water levels, and harvesting plants

when and where necessary. Where plant cover is deficient,

management activities to improve cover may include water level adjustment, reduced loadings, pesticide application, and replanting.

Harvesting and litter removal may be necessary depending on the design of the system. Plant removal from some
130

FWS wetlands may be required to meet the treatment

goals, but a well-designed and well-operated VSB system

should not require routine harvesting. Harvesting of

Phragmites at the height of the growing season and just

before the end of the growing season does help to remove

some nitrogen from the system, but phosphorus removal

is limited (Suzuki et al., 1985). Winter burning of vegetation can be used to control pests.

6.4.4 Odor Control

Odors are seldom a nuisance problem in properly loaded

wetlands. Odorous compounds emitted from open water

areas are typically associated with anaerobic conditions,

which can be created by excessive BOD and ammonia

loadings. Therefore, reducing the organic and nitrogen

loadings can control odors. Alternatively, aerobic open

water zones interspersed in areas between fully vegetated

zones introduce oxygen to the system. Turbulent flow structures such as cascading outfall structures and channels

with hydraulic jumps, which are employed to introduce

oxygen into the system effluent, can generate serious odor

problems through stripping of volatile compounds such as

hydrogen sulfide, if the constructed wetland has failed to

remove these constituents.

6.4.5 Control of Nuisance Pests and

Insects

Potential nuisances and vectors that may occur in FWS

wetlands include burrowing animals, dangerous reptiles,

mosquitoes, and odors. An infestation of burrowing animals such as muskrats and nutria can seriously damage

vegetation in a system. These animals use both cattails

and bulrushes as food and nesting materials. These animals can be controlled during the design phase by decreasing the slope on berms to 5:1 or using a coarse riprap.

Temporarily raising the operating water level may also discourage the animals. Live trapping and release may be

successful, but in most cases it has been necessary to

eliminate the animals. Fencing has had little success.

Dangerous reptiles are common in the southeastern

states. The most common are snakes, particularly the water

moccasin, and alligators. It is difficult to control these animals directly. Warning signs, fencing, raised boardwalks,

and mowed hiking trails can be used to minimize human

contact with the animals. Operators should be made aware

of the dangers and preventive actions that can be taken to

avoid dangerous situations.

Mosquito control is a critical issue in FWS wetlands. In

warm climates, wetlands have been seeded with mosquito

fish (Gambusia) and dragonfly larvae to control mosquitoes. Mosquito fish also can be used in northern climates,

but they need to be restocked each year. However, mosquito fish have difficulty reaching all parts of the wetland

when the accumulation of litter is too dense, particularly if

cattails are grown. Other natural control methods have included erecting bat and bird houses. Desirable birds include purple martins and swallows. Bacterial larvicides,

BTI (Bacillus thuringiensis israelensis), and BS Bacillus

sphaericus) have been used successfully in a number of

wetlands.

6.4.6 Maintenance of Berms and Dikes

Berms and dikes require mowing, erosion control, and

prevention of animal burrows and tree growth. When the

wetland is operated at a shallow depth, periodic removal

of tree seedlings from the wetland bed may be necessary.

If the trees are allowed to reach maturity, they may shade

out the emergent vegetation and with it the necessary conditions to enhance flocculation, sedimentation, and denitrification.

6.5 Monitoring

Routine monitoring is essential in managing a wetland

system. In addition to regulatory requirements, inflow and

outflow rates, water quality, water levels, and indicators of

biological conditions should be regularly monitored and

evaluated. Monitoring of biological conditions includes

measurement of microbial populations and monitoring

changes in water quality, percent cover of dominant plant

species, and benthic macroinvertebrate and fish populations at representative stations. Over time, these data help

the operator to predict potential problems and select appropriate corrective actions.

6.6 References

Allen, H.H., G.J. Pierce, and R. Van Wormer. 1990. Considerations and techniques for vegetation establishment in constructed wetlands. In: D.A. Hammer (ed.) Constructed wetlands for wastewater treatment, municipal, industrial and

agricultural. Chelsea, MI: Lewis Publishers, Inc.

Grits, M.A. and R.L. Knight. 1990. Operations optimization. In:

D.A. Hammer (ed.) Constructed wetlands for wastewater

treatment, municipal, industrial and agricultural. Chelsea,

MI: Lewis Publishers, Inc.

Kadlec, R.H. and R.L. Knight. 1996. Treatment wetlands. Boca

Raton, FL: CRC Press LLC.

Rumer and Mitchell, 1996. Assessment of barrier containment

technologies. NTIS report # PB 96-180583.

Suzuki, T., A.G. Wathugala, and Y. Kurihara. 1985. Preliminary

studies on making use of Phragmites australis for the removal of nitrogen, phosphorus, and COD from the waste

water. UNESCO’s Man and Biosphere Programme in Japan, Coordinating Committee on MAB Programme, pp. 95–

99.

U.S. Environmental Protection Agency (EPA). 1993. Report of

workshop on geosynthetic clay liners. EPA/600/R-93/171.

Office of Research and Development, Washington, DC.

U.S. Environmental Protection Agency (EPA). 1994. Seminar

Publication: Design, operation, and closure of municipal

solid waste landfills. EPA/625/R-94/008. Office of Research and Development, Cincinnati, OH.
131

7.1 Introduction

The major items included in capital costs of constructed

wetlands are

• Land costs

• Site investigation

• Clearing and grubbing

• Excavation and earthwork

• Liner

• Media

• Plants

• Inlet structures

• Outlet structures

• Fencing

• Miscellaneous piping, pumps, etc.

• Engineering, legal, and contingencies

• Contractor’s overhead and profit

Most of these costs are directly dependent on the design treatment area of the system, and the unit costs for

almost all are essentially the same for FWS and VSB systems. The major difference between the two concepts is

the media cost (Table 7-1). In the case of a VSB, a 60 cm

(2 ft) depth of gravel typically fills the bed, whereas the

medium for a FWS wetland consists of 15 cm (6 in) or

more of topsoil used as growth media for the wetland vegetation.

7.2 Construction Costs

7.2.1 Total Construction Costs

The cost data in this report were obtained from site visits to nine operational constructed wetland systems and

from related published sources. The nine systems were

Arcata, CA; Gustine, CA; Mesquite, NV; Ouray, CO; West

Jackson County, MS; Mandeville, LA; Sorrento, LA;

Carville, LA; and Ten Stones, VT. This group includes four

FWS wetlands and five VSBs with design flows ranging

from 0.3 to 175 L/s (6,700 gpd to 4 mgd). The start-up

dates for these subsystems range from 1986 (Arcata, CA)

to 1997 (Ten Stones, VT). In order to provide a common

base for comparison, all costs have been adjusted with

the appropriate Engineering News Record (ENR) Construction Cost Index (CCI) factor to August 1997 (ENR CCI =

5854).

Unfortunately, it is not possible to extract the individual

line-item construction costs listed in Table 7-1 for most

existing wetland systems because their construction contracts were let as lump sum bids for entire projects. In many

cases, the situation is further confounded since the lump

sum bids also may include preliminary treatment components, pumping stations, and community collection systems. In addition, local conditions and site characteristics

also significantly affect wetland system costs. VSB wetlands in southern Louisiana, for example, pay a high price

for imported gravel since none is available locally. Some

existing wetland systems were converted from existing lagoon cells. In these cases, the costs for clearing and grubbing and excavation and earthwork would be minimal. As

a result of these factors, it is not possible to derive general

nationally applicable cost-per-hectare unit cost. The best

that can be achieved is a range of costs that may be useful for an order-of-magnitude preliminary estimate.

Table 7-2 presents a summary of technical and cost data

for the nine constructed wetland systems included in the

EPA case study visitations. The costs listed in the table

are the estimated construction costs for the wetland component in each system at the time the system was constructed. It is difficult to draw general conclusions from the

data because of the many variables involved. The systems listed were designed with different design models and

procedures to achieve different water quality goals, so the

relationship between the treatment areas provided and the

design flow rates is not meaningful. The only nearly consistent factor is that land costs were zero in all cases except Ouray, CO, because the land was already in the possession of the system owner. The costs given are based

on actual construction costs and do not include a factor for

system design or site investigation. In most cases, the site

investigation costs for these nine systems were minimal

Chapter 7

Capital and Recurring Costs of Constructed Wetlands
132

since information pertaining to soil characteristics and ground

water conditions was already available.

There is some evidence of economy of scale in the tabulated

data; the 20.2-hectare (50-acre) FWS wetland expansion at West

Jackson County, MS, is the largest system listed and had the

lowest unit cost. Contributing factors are believed to be the lack

of a liner and minimal berm construction because of the small

number of relatively large cells selected for this project (seven

cells). Ouray, CO, had the highest unit costs listed for FWS wetlands. This system required the use of a membrane liner and

had significant construction costs for berms because of the number of small cells, in this case four cells on 0.89 hectares (2.2

acres). The system at Gustine, CA, shows a higher unit cost

than Arcata, CA, primarily due to the extra berm construction

involved. Other sources indicate that the cost per hectare for

large FWS systems is about one-third the cost of smaller FWS

wetlands, similar to the range presented in Table 7-2. The construction costs for FWS wetlands might therefore range from

about $34,600 per hectare to $237,200 per hectare ($14,000/

acre to $96,000/acre in 1997$), depending on the size of the

system, the number of cells and berms, and the need for a membrane liner. The size of the system will depend on the waterquality goals and local climatic conditions. Additional costs will

include any site investigation and engineering design, pre- and/

or post-treatment components, means for transferring wastewater to the treatment site and effluent from the site, and land

costs.

Economy of scale is apparent for the five VSB systems listed

in Table 7-2. The average unit cost for the two smallest systems is at least twice that of the three larger systems. AnTable 7-1. Cost Comparison of 4,645 m2 Free Water Surface Constructed Wetland and Vegetated Submerged Bed

Free Water Vegetated

Surface Wetland Submerged Bed

Item Units Unit Price Total Cost ($) % of Total Total Cost ($) % of Total

Excavation/ m3 $2.30 13,000 19.4 13,000 10.7

Compaction

Soil (45 cm) m3 $1.30 2,800 4.2 na —

Gravel (60 cm) m3 $20.95 Na — 51,900 42.6

Liner (30 mil PC) m2 $3.75 19,250 28.7 19,250 15.8

Plants Each $0.60 7,500 11.2 13,330 10.9

(60 cm o.c.) (45 cm o.c.)

Plumbing Lump sum 7,500 11.2 7,500 6.1

Control Structures Lump sum 7,000 10.4 7,000 5.7

Other Lump sum 10,000 14.9 10,000 8.2

67,050 100.0 121,980 100.0

* 4,645 m2 = 50,000 ft2

Table 7-2. Technical and Cost Data for Wetland Systems Included in 1997 Case Study Visitations

Startup Area Flow No. Berm Land Const. Adj. Cost3

Location Date (hectares) (m3/s) Cells Liner1 Const2 Cost Cost $/m2 $/m2

Free Water Surface Wetlands

Arcata, CA 1986 3.0 66.2 3 No No $0 $225,000 7.43 10.12

Gustine, CA 1987 9.8 22.8 24 No Yes $0 $882,000 9.04 12.27

Ouray, CO 1993 0.9 8.7 4 Yes Yes $55,000 $108,500 12.16 13.02

W.J.C. MS 1997 20.2 54.8 7 No Yes $0 $700,000 3.44 3.44

Vegetated Submerged Beds

Carville, LA 1986 0.3 3.4 1 No Yes $0 $100,000 38.64 52.64

Mandeville,LA 1990 2.3 34.2 3 No No $0 $590,000 26.05 32.18

Mesquite, NV 1991 1.9 9.1 12 No No $0 $515,000 27.13 32.83

Sorrento, LA 1991 0.07 1.14 1 No Yes $0 $75,000 103.44 125.18

Ten Stones, VT 1997 0.04 0.16 2 Yes Yes $0 $40,000 84.66 89.66

1

A ‘No’ response is compacted native soil or pre-existing pavement 2

A ‘No’ response had pre-existing lagoon berms 3

Adjusted to August, 1997 costs (ENR CCI = 5854)

 Hectares = 2.47 acres; m3/s = 22.83 mgd; m2 = 10.76 ft2
133

other major factor in these cost differences is due to the

significant differences in the local costs of the rock and

gravel media used. Sorrento, LA, for example, used

crushed limestone imported from Mexico because specified aggregate was not available in that part of Louisiana.

In comparison, Mesquite, NV, used available media from

a nearby gravel pit. Local gravel also was available for the

Ten Stones, VT, system, but the system is conservatively

designed for the cold winter climate and has a higher unit

cost than Mesquite, NV. If the Sorrento costs are omitted,

the construction costs for VSB wetlands may range from

$321,200 per hectare to $897,000 per hectare ($130,000/

acre to $363,000/acre in 1997$), depending on the size of

the system, the need for a liner, and the local costs of gravel.

The size of the system will depend primarily on the waterquality goals and to a lesser degree on local climatic conditions. Additional costs will include any site investigation

and engineering design, pre- and/or post-treatment components, means for transferring wastewater to and/or from

the treatment site, and land costs. These costs for VSB

wetlands are significantly higher than the highest costs cited

for FWS wetlands.

7.2.2 Geotechnical Investigations

Only four of the systems listed in Table 7-2 did not utilize

any preliminary geotechnical investigations. Three of these,

Arcata, CA; Gustine, CA; and Mesquite, NV, utilized existing lagoon cells, and the underlying soil conditions were

presumably already known. The fourth system, Ten Stones,

VT, was relatively small, and the state regulatory agency

required a membrane liner regardless of the underlying

soil characteristics, so a geotechnical investigation was

not considered necessary. The other five systems utilized

some shallow borings to verify expected soil conditions at

the wetland site. The only system that retained cost data

for this activity was Mandeville, LA, where approximately

$15,000 was allocated in 1989 for site surveys and soil

borings in the wetland area. The updated cost for surveying and soil borings at the Mandeville wetland site would

be about $2,720 per hectare ($1,100/acre) in 1997$.

7.2.3 Clearing and Grubbing

Three of the systems listed in Table 7-2 required clearing and grubbing as part of their site preparation. Technical details and related costs are listed in Table 7-3. The

cost data in Table 7-3 are compatible with experience in

the general construction industry. Costs for clearing and

grubbing on relatively level land can range from $4,940

per hectare ($2,000/acre) for brush and some small trees

to $12,355 per hectare ($5,000/acre) for a tree-covered

site. Campbell and Ogden (1999) indicate southwestern

costs at $2,965 per hectare ($1,200/acre).

7.2.4 Excavation and Earthwork

Excavation and earthwork typically includes grading the

wetland site to finished grade, constructing berms and

access ramps, and in the case of FWS wetlands, reserving and replacing topsoil in the bed to serve as the vegetation growth medium. Table 7-4 summarizes available cost

data from the 1997 survey.

An economy of scale is expected for earthwork costs

and is evident in the data for the small Ten Stones, VT,

project where excavation costs were three times greater

than the larger municipal-sized systems. All three sites

listed were on relatively level land, with soils ranging from

clay at West Jackson County, MS, to silty loam at Ten

Stones, VT. The average cost for the two municipal sysTable 7-3. Clearing and Grubbing Costs for EPA Survey Sites

Location Site Area Total Cost/ Adj. Cost1

(hectares) Cost Hectare ($/hectare)

West Jackson Co., MS2 20.2 $100,000 $4940 4940

Sorrento, LA3 1.6 $7,000 $4325 5235

Ouray, CO4 2.0 $22,700 $11,120 12,050

1Adjusted to August, 1997 $ (ENR CCI = 5854

2

Ground cover: brush and sparse tree seedlings

3Ground cover: brush

4Ground cover: trees

Hectares = 2.47 acres

Table 7-4. Excavation and Earthwork Costs for EPA Survey Sites

Site Area Quantity Total Adj. Cost1

Location (hectares) (m3) Cost $/Hectare $/m3 ($/m3)

W. Jackson Co., MS 20.2 52,000 $408,000 20,165 7.85 8.29

Gustine, CA 9.7 34,400 $200,000 20,600 5.81 8.03

Ten Stones, VT 0.04 355 $8,200 184,200 23.16 23.16

1Adjusted to August, 1997 $ (ENR CCI = 5854

 Hectares = 2.47 acres

 m3 = 1.31 yd3
134

tems shown is about $8.17 per cubic meter ($6.25/yd3).

About one-third of that cost could be assigned to excavation of the wetland bed to grade (on relatively level land),

with the remainder for berm and ramp construction and

reservation and replacement of topsoil for the FWS system. Campbell and Ogden (1999) suggest $1.96 to $3.27

per cubic meter ($1.50 to $2.50/yd3) as a default value for

preliminary estimates.

7.2.5 Liner Costs

A variety of materials, including the in situ native soils,

have been used as liner material depending on the requirements of the regulatory agencies. The majority of the

systems listed in Table 7-2 utilized the existing on-site soils

for their liner material. Two of the remaining systems listed

in Table 7-2 used plastic membrane liners. Table 7-5 summarizes these costs. These costs reflect the economy of

scale available for larger systems. The unit cost at Ouray

was $5.27 per m2

 ($0.49/ft2), while $10.01 per m2 ($0.93/

ft2) was found at Ten Stones (1997$). Ouray used a 30-mil

HDPE liner, and Ten Stones used a prefabricated 30-mil

PVC liner. Other liner materials are also available, and typical large system costs for some of these are presented in

Table 7-6. Where soils are rocky, a geotextile fabric or layer

of sand may be necessary beneath the synthetic liner to

protect it from punctures. The liner will add $0.54 to $0.86/

m2 ($0.05 to $0.08/ft2) to the costs presented in Table 7-6.

The costs of compaction and testing of clay liners can exceed $3.23/m2

 ($0.30/ft2).

7.2.6 Media Costs

The media in a FWS wetland are the soils placed on top

of the prepared bottom of the bed which serve as the growth

medium for the emergent vegetation in the system. A similar layer of topsoil is also usually applied to the berm slopes

to allow their revegetation. Placement of these soil layers

is usually included in the earthwork costs previously discussed.

The media used in a VSB are the gravel or rock placed

in the bed. They serve to support the growth of the vegetation and to provide physical filtration, flocculation, sedimentation, and surfaces for attached microbial growth and adsorption to occur. Several different sizes of rock and gravel

can be used in these systems. At the sites visited in the

EPA study, medium-sized gravel, 20–25 mm in diameter

(0.75–1 in), was used for treatment. Coarser rock, 40–50

mm in diameter (1.5–2 in), was used to surround the inlet

and outlet manifolds, and a layer of pea gravel, 5–10 mm

in diameter (1/4–3/8 in), was sometimes used to cap the

gravel in the treatment bed. Coarse stone, 10–15 cm in

diameter (4–6 in), is sometimes used to cover the exposed

liner on the side slopes and to reduce the risk of burrowing

animals. The unit cost of these materials depends on the

size of the material, the volume needed, and the distance

from the source to the wetland site. The media is usually

the most expensive part of the construction of a VSB, potentially representing 40 to 55% of total construction costs

(Table 7-1). Table 7-7 summarizes the costs for these materials as derived from the 1997 site visitations.

The local availability of gravel and the transport distance

to the VSB site are the key factors influencing media costs.

As a result, the cost data in Table 7-7 should only be used

for preliminary estimates. Local sources should be contacted for a detailed construction cost estimate. Based on

the data in Table 7-7, the media costs (for the main bed)

range from $74,130 to $133,440 per hectare ($30,000 to

$54,000/acre) depending on the local availability of suitable material.

7.2.7 Plants and Planting Costs

Plant materials sometimes can be obtained locally by

cleaning drainage ditches. It is also possible to develop an

on-site nursery at the wetland construction site if there is

sufficient advance time, or grow plant sprigs or seedlings

from seed and transplant these to the wetland cells. A large

and expanding number of commercial nurseries also exist

and can supply a large variety of plant species for these

wetlands systems. The majority of the systems listed in

Table 7-2 were planted with commercial nursery stock.

Small systems are typically planted by hand; large systems can use mechanical planters, and nursery-grown

sprigs or plant seedlings are advantageous for the purpose. Hydroseeding has been successful with Typha seeds

(Gearheart et al., 1998). Table 7-8 summarizes available

cost data for plants and planting from the 1997 survey sites.

Campbell and Ogden (1999) quote a range of $0.50 to

$1.00 per plant in place as a default value, while Gearheart

et al. (1998) estimate $12,355 per hectare ($5,000/acre)

for a total planting cost.

Table 7-5. Liner Costs for EPA Survey Sites

Treatment Area Total Cost/ Adj. Cost2

Location (hectares) Cost Hectare1 ($/hectare)

Ouray, CO 1.363 $64,000 $46,930 52,725

Ten Stones, VT 0.045 $4,500 $100,175 100,175

1

Represents cost per hectare of treatment area. Actual liner area is

larger to cover berms, etc.

2

Adjusted to August, 1997 $ (ENR CCI = 5854)

3

Lined area at Ouray, CO includes lagoons and wetland cells.

 Hectares = 2.47 acres = 10,000 m2

Table 7-6. Typical Installed Liner Costs for 9,300 Square Meter

(100,000 ft2) Minimum Area

Liner Material $/m2

Bentonite (9.8 kg/m2 and harrowed in place) 0.52-0.60

Clay impregnated geosynthetic 0.37-0.60

Asphalt concrete 0.60-0.75

Butyl rubber (50 mm thickness) 0.60

PVC (30 mil) 0.28-0.40

HDPE (40 mil) 0.35-0.40

Hypalon (30 mil 0.55

Hypalon (60 mil) 0.60-0.70

PPE (30 mil) 0.58-0.60

Reinforced PPE (30 mil) 0.65

XR-5 0.85-0.92

m2 = 10.76 ft2
135

Table 7-7. Media Costs for VSBs from EPA Survey Sites

Media Size Media Depth Media Quantity Cost Cost Adj. Cost1

Location (mm) (m) (m3/hectare) ($/m3) ($/hectare) ($/hectare)

Mesquite, NV Bed: 10-25 0.8 8,140 10.99 89,380 108,230

Carville, LA Top: 20 0.15 1,525 27.14 41,325 44,735

Bed: 40-75 0.60 6,100 20.21 123,160 133,320

Ten Stones, VT Top: 10 0.15 1,523 25.07 38,180 38,180

Bed: 20-25 0.60 6,095 12.01 73,180 73,180

Outlets: 50 0.60 725 7.85 5,680 5,680

RipRap: 100 0.12 445 18.31 8,130 8,130

1Adjusted to August, 1997 $ (ENR CCI = 5854)

 m3/hectare = 1.89 yd3/acre

Table 7-9. Costs for Inlet and Outlet Structures from EPA Survey Sites

Cost Adj. Cost1

Location Structure Type Weir Type ($/structure) ($/structure)

West Jackson Co., MS Concrete box Bolted plate 2,500 2,500

Carville, LA Concrete box Shear Gate 3,900 4,400

Gustine, CA Concrete box Stop log 1,125 1,500

Ten Stones,VT 100 mm PVC manifold None 1,500 1,500

1

Adjusted to August, 1997 $ (ENR CCI = 5854)

Table 7-8. Costs for Wetland Vegetation and Planting from EPA Survey Sites

Plant Adj. Planting

Plant Density Planting Plant Cost Planting Cost Cost1

Location Type (no./hectare) Method ($/plant) ($/hectare) ($/hectare)

Ten Stones, VT Cattails 35,830 Hand 0.23 Not available Not available

Bulrush 35,830 Hand 0.23 Not available Not available

Mandeville, LA Bulrush 15,000 Hand Not available 2,965 3,670

Carville, LA Pickerel Weed 34,595 Hand Local sources 2,470 3,370

Arrowhead 34,595 Hand Local sources 2,470 3,370

Sorrento, LA None None None None None None

Gustine, CA Bulrush 46,950 Mechanical Not available 3,460 4,595

Cattails 11,860 Mechanical Not available 1,975 2,695

Ouray, CO Cattails 13,465 Hand Local sources Not available Not available

Bulrush 13,465 Hand Local sources Not available Not available

West Jackson Co., MS Cattails 11,860 Mechanical 0.38 4,450 4,450

Mesquite, NV Bulrush Not available Hydroseeding Not available Not available Not available

Arcata, CA Bulrush 9,885 Hand Not available Not available Not available

1

Adjusted to August, 1997 costs (ENR CCI = 5854)

7.2.8 Cost of Inlet and Outlet Structures

The inlet and outlet structures for most small- to moderate-sized wetland systems are typically some variation of

a perforated manifold pipe. Large wetland systems typically use multiple drop or weir boxes for both inlets and

outlets. Adjustable water level outlet structures should be

used to control the water level in the wetland cell. If the

outlet is a pipe manifold, a water level–control structure

must be added, which should cost about the same as a

weir box. Table 7-9 summarizes the available cost data

from the 1997 survey.

7.2.9 Piping, Equipment, and Fencing

Costs

These items include the piping to transfer the wastewater to the wetland, the piping from the wetland to a dis-
136

charge point, and any pumps required for either of those

purposes. Fencing is typically installed around all municipal wastewater treatment systems, but has not usually been

required around the smaller VSB wetland beds due to the

low risk of public contact and exposure to the wastewater.

None of these features are unique to wetland systems,

and costs for these items were not available at the sites

included in the 1997 EPA survey.

7.2.10 Miscellaneous Costs

These costs include engineering design and legal fees,

construction contingencies, and profit and overhead for the

construction contractor. These costs are not unique to

wetland systems and are usually expressed as a percentage of the total construction costs when preparing an estimate. Mobilization and bonding are also typically included

in the construction costs. Typical values for miscellaneous

costs are as follow:

• Mobilization, 5% of direct costs

• Bonds, 3% of direct costs

• Engineering design services, 15% of capital costs

• Construction services and start-up, 10% of capital costs

• Contractor’s overhead and profit, 15% of capital costs

• Contingencies, 15% of capital costs

The following example illustrates the application of these

factors. Assume direct project construction costs (i.e., labor, materials, equipment, etc.) are $300,000, therefore:

• Direct construction costs $300,000

• Mobilization, 5% $15,000

• Bonds, 3% $9,000

• Capital cost of construction $324,000

• Engineering, 15% $48,600

• Start-up, 10% $32,400

• Overhead, profit and contingencies, 30% $97,200

• Total capital costs $502,200

7.2.11 Construction Cost Summary

The major cost factors for both VSB and FWS wetlands

are compared in Table 7-10. The tabulated data are drawn

from previous tables for an assumed 0.405 hectare (one

acre) wetland with a membrane liner. The cost data are in

terms of dollars per hectare, and the percentage data are

percent of total cost. The latter can be used to determine

which system components are likely to be the most expensive. The cost data shown do not include the costs of

the land, mobilization, fencing, landscaping, pre- or posttreatment units, or the transfer piping to and from the wetland site, and should only be used for preliminary, orderof-magnitude cost estimates.

The cost of gravel media for the VSB is the most expensive item in Table 7-10, followed by the membrane liner for

both types of wetlands. The cost of the gravel media in the

VSB controls the cost regardless of the type of liner used.

If site conditions allow for the compaction of native clay

soil to produce an acceptable ground water barrier in lieu

of a synthetic membrane liner, then the liner costs can be

eliminated. In this case, perforated pipe manifolds for inlet

and outlet structures are used instead of concrete weir

boxes.

7.3 Operation and Maintenance Costs

The operation and maintenance of constructed wetland

systems designed for wastewater treatment are relatively

simple and require minimal time. They are similar to, but

somewhat more than, the O&M requirements for a facultative pond. Most of the operator’s time at a wetland treatment system is spent servicing pumps, headworks, disinfection, and other conventional components in the process.

Animal (i.e., nutria, muskrats) control, vector (mosquitoes)

control, and NPDES monitoring are probably the most timeconsuming aspects of wetland operation and maintenance.

Crites and Ogden (1998) report the operating costs for FWS

constructed wetlands and VSBs range from $0.10 to $0.30

and $0.04 to $0.08, respectively, per 3,785 L (1,000 gal) of

treated water.

At the FWS wetland system at Ouray, CO, the O&M requirements for the wetland are as follows:

• Check berms for animal Once per week

damage and erosion

• Check and clean effluent Once per week

debris screens

• Observe and adjust water Once per month

levels and flow rates

• Remove sludge from inlet As required

zone

• Flush manifold pipes As required

• Mosquito control As required by local

health authorities

The 1997 monthly operating costs for the complete treatment system (aerated lagoon, FWS, chlorination/dechlorination) at Ouray, CO, were

• Power for lagoon aerators $1,400

• Lagoon sludge removal and disposal 800

• Miscellaneous supplies 125

• NPDES laboratory tests 300

• Wages 1,083

• Total $3,708
137

Table 7-10. Range of Capital Costs for a 0.4 Hectare (Membrane Lined VSB and FWS Wetland)

Item Vegetated Submerged Bed (VSB) Free Water Surface (FWS)

Low Cost Range High Cost Range Low Cost Range High Cost Range

Cost1 Percent Cost1 Percent Cost1 Percent Cost1 Percent

($/hectare) of Total ($/hectare) of Total ($/hectare) of Total ($/hectare) of Total

Survey/Geotechnic 2,718 1.2 5,436 1.3 2,718 1.9 5,436 2.1

Clear & Grub2 4,942 2.2 12,355 3.0 4,942 3.4 12,355 4.8

Earthwork3 18,039 8.1 29,900 7.3 18,039 12.4 29,900 11.7

Membrane Liner4

30 mil PVC 37,807 16.9 43,243 —- 37,807 26.1 43,243 —-

40 mil PE 43,243 —- 48,432 —- 43,243 —- 48,432 —-

40 mil PPE 53,869 —- 59,305 —- 53,869 —- 59,305 —-

45 mil Reinf. PPE 64,494 —- 69,931 —- 64,494 —- 69,931 —-

60 mil Hypalon 69,931 —- 80,803 —- 69,931 —- 80,803 —-

XR-5 102,301 —- 113,174 27.5 102,301 —- 113,174 44.4

Media 129,4835 57.8 199,4135 48.5 16,0626 11.1 20,0156 7.8

Plants & Planting7 8,649 3.9 17,297 4.2 8,649 6.0 17,297 6.8

Control Structures 4,9428 2.2 16,062 3.9 39,5379 27.3 39,600 15.5

Plumbing & Fencing 17,297 7.7 17,297 4.2 17,297 11.9 17,297 6.8

Totals: 233,87710 100 410,93511 100 145,05010 100 255,01211 100

1Adjusted to 1997 dollars (ENR CCI = 5854)

2Clearing and grubbing costs are higher for sites with large trees

3

Earthwork (excavation and compaction) costs are typically 2.00 to 3.25 per m3. A 0.9 m deep excavation was assumed. 4

For rocky soils, add 5,435 to 8,645 per hectare to the costs presented. For a site employing a minimum of 9,300 m2 of liner, deduct 5,435. Delete

the costs for the liner if native clay liner is used.

5

Reported cost range is for 60 cm of gravel media with 15 cm of pea gravel. Gravel costs typically range from 17 to 26 per m3 within 20 miles of the

project. Longer delivery distances will increase the cost.

6

Assumes 15 cm of topsoil over the wetland bottom.

7

Planting costs are typically 0.50 to 1.00 per plant. Values shown are for planting at 2.5 sq. ft centers. Adjust cost if different spacing is used.

8

122 m of 100 mm diameter PVC manifold plus one water level control box

9

Eight concrete weir box structures for inlets/outlets

10Uses 30 mil PVC liner

11Uses XR-5 liner

 0.405 ha = 1 acre

The tasks specifically related to the wetland components

are estimated to require about 16 hours per month or about

$3,000 per year. On an areal basis, this equates to $3,370

per hectare per year ($1,364/acre per year) for this 0.89-

hectare (2.2-acre) wetland system. These wetland costs represent about 7% of the total O&M costs for the entire treatment process. If more rigorous testing than the minimal

NPDES testing were required, then monitoring could become

the most expensive cost of all O&M categories shown.

The annual O&M expenses at the VSB system in Carville,

LA (0.15 mgd), are shown in Table 7-11. At Carville, the annual costs directly associated with the wetland components

are estimated to be $650 per year or $2,500 per hectare per

year ($1,015/acre per year). These costs are about 6% of

the total O&M costs for the entire treatment system.

The 1996 O&M costs for the Gustine, CA, sewer department were about $433,275, which included bond repayment,

engineering fees, and other contractual services. A single

major expense was $152,402 for electrical power for pumps

and the lagoon aerators. The direct O&M costs for the sewer

system, the lagoons, and the wetland were $280,873. Since

the wetland O&M tasks at Gustine are similar to those previously described for Ouray and Carville, it can be assumed

that the cost percentage determined at those systems is also

applicable at Gustine. Using a 7% factor, the annual wetland

O&M costs at Gustine would be $19,661 or approximately

$2,025 per hectare per year ($819/acre per year).

At Arcata, CA, it is estimated that the O&M tasks directly

related to the wetlands require 20 minutes of operator time

per day or 122 hours per year. The major tasks are weir

adjustments and berm inspections. At an assumed cost of

$30 per hour (including benefits, incidentals, and support

costs), the annual O&M costs for the Arcata wetlands would

be $1,205 per hectare per year ($488/acre per year).

For the small system at Ten Stones, VT (6,700 gpd), onehalf hour per month is estimated to inspect the system and

pumps and to adjust water levels if necessary. These efforts

will be voluntary on the part of the corporation members, so

there will be no actual cost for the service. However, if a

contractual service were retained at the previously assumed

rate of $30 per hour, the annual maintenance costs would

be $4,043 per hectare per year ($1,636/acre per year).

The remainder of the systems included in the 1997 EPA

Survey (West Jackson County, MS; Mandeville, LA; Mesquite, NV; and Sorrento, LA) did not have recorded data for

separate wetland O&M costs, and estimates were not provided. Table 7-12 summarizes the O&M cost data that was

available.

It is not possible to divide the total annual O&M costs shown

in Table 7-12 into ranked categories for the major O&M tasks.

The major O&M tasks are visual inspections of the berms

and of plant health, and adjustments in water levels and other

flow-control structures as required. Both nutria and musk-
138

rats can cause physical damage and leakage in the berms

and can destroy, as well as some insects, the plant cover in

the wetland cells. If routine visual observations indicate

damage, a more intense O&M effort will be necessary for

repair and animal or insect control.

Active mosquito control may be an issue in California

and in the southwestern and southeastern states, and the

FWS system O&M costs will increase accordingly. None

of the systems listed in Table 7-12 were making special

efforts for either animal or mosquito control. On a longterm basis, it will be necessary to remove accumulated

sediment from the wetland cells when it begins to interfere

with the hydraulic performance of the system. A ramp for

this purpose should be included in the design and construction of each wetland cell.

7.4 References

Campbell, C.S. and M.H. Ogden. 1999. Constructed wetlands in the sustainable landscape. New York, NY: John

Wiley and Sons.

Crites, R.W. and M.H. Ogden. 1998. Costs of constructed

wetlands systems. Presented to WEFTEC, WEF 71st

Annual Conference, Orlando, FL.

Gearheart., R.A., B.A. Finney, M. Lang, and J. Anderson.

1998. A comparison of system planning, design, and

sizing methodologies for free water surface constructed

wetlands. In: 6th International Conference on Wetland

Systems for Water Pollution Control. IAWQ.

Middlebrooks, E.J., C.H. Middlebrooks, J.H. Reynolds, G.Z.

Watters, S.C. Reed, and D.B. George. 1982. Wastewater stabilization lagoon design, performance, and

upgrading. New York, NY: Macmillan Publishing Co.

Table 7-11. Annual O&M Costs at Carville, LA (570 m3/d) Vegetated Submerged Bed

Annual Cost

Item (1997 $)

Electricity

Lagoon aerators 4,080

UV disinfection 316

Miscellaneous 350

Maintenance

Repair berms (30 man-hours) 478

Maintain UV system (60 man-hours) 956

Berm grass cutting (50 man-hours) 797

Parts & Supplies

Aerators 230

UV System 800

Flow meter 25

NPDES Monitoring

Labor (192 man-hours) 3,209

Materials 350

Total 11,591

Table 7-12. Annual O&M Costs for Constructed Wetlands

Location Design Flow Treatment Area Cost

(m3/d) (hectares) ($/ha-yr)1

Ouray, CO (FWS) 1375 0.89 3370

Gustine, CA (FWS) 3785 9.71 2025

Ten Stones, VT (VSB) 25 0.05 4045

Carville, LA (VSB) 565 0.26 2510

1

1997 $
139

In 1997, a series of site visits to constructed treatment

wetlands was performed for the U.S. EPA to compile background information and assess performance of free water

surface (FWS) and vegetated submerged bed (VSB) wetlands. Edited versions of these site visit reports are presented in this chapter to enrich the reader’s understanding with insight gained from actual construction and operation of treatment wetland systems. Most quantifying

terms are expressed in English units. Conversion factors

are as follows:

• 1 acre = 0.405 ha

• 1 mgd = 3,780 m3

/d

• 1 ft = 0.3 m

• 1 gal/day-ft2 = 4 cm/d = 0.004 m3/m2-d

• 1 in = 2.54 cm

• 1 lb = 0.454 kg

8.1 Free Water Surface (FWS) Constructed

Wetlands

8.1.1 Arcata, California

8.1.1.1 Background

Arcata is located on the northern coast of California about

240 miles north of San Francisco. The population of Arcata

is about 15,000. The major local industries are logging,

wood products, fishing, and Humbolt State University. The

FWS constructed wetland located in Arcata is one of the

most famous in the United States.

The community was originally served, starting in 1949,

with a primary treatment plant that discharged undisinfected

effluent to Arcata Bay. In 1957, oxidation ponds were constructed, and chlorine disinfection was added in 1966. In

1974, the State of California prohibited discharge to bays

and estuaries unless “enhancement” could be proven, and

the construction of a regional treatment plant was recommended. In response, the City of Arcata formed a Task

Force of interested participants, and this group began research on lower-cost alternative treatment processes using natural systems. From 1979 to 1982, research conducted at pilot-scale wetland units confirmed their capability to meet the proposed discharge limits. In 1983, the city

was authorized by the state to proceed with development,

design, and construction of a full-scale wetland system.

Construction was completed in 1986, and the system has

been in continuous service since that time.

The wetland system proposed by the city was unique in

that it included densely vegetated cells dedicated for treatment followed by “enhancement” marsh cells with a large

percentage of open water for final polishing and habitat

and recreational benefits. This combined system has been

successful since start-up and has become the model for

many wetland systems elsewhere.

Two NPDES permits are required for system operation:

one for discharge to the enhancement wetlands for protection of public access and one for discharge to the bay.

The NPDES limits for both discharges are BOD 30 mg/L

and TSS 30 mg/L, pH 6.5 to 9.5, and fecal coliforms of 200

CFU/100 mL. Since public access is allowed to the enhancement marshes, the state required disinfection prior

to transfer of the pond/treatment marsh effluent. The state

then required final disinfection/dechlorination prior to final

discharge to Arcata Bay. The effluent from the final enhancement marsh is pumped back to the treatment plant

for this final disinfection step.

The basic system design for the treatment and enhancement marshes was prepared by Dr. Robert Gearheart and

his colleagues at Humbolt State University. The design was

based on experience with the pilot wetland system that

was studied from 1979 through 1982.

The pilot wetland system included 12 parallel wetland

cells, each 20 ft wide and 200 ft long (L:W 10:1), with a

maximum possible depth of 4 ft. These were operated at

variable hydraulic loadings, variable water depths, and

variable initial plant types during the initial phase of the

study. Hardstem bulrush (Scirpus validus) was used as

the sole type of vegetation on all cells. The inlet structure

for each cell was a 60° V-notch weir, and the outlet used

an adjustable 90° V-notch weir, permitting control of the

water depth. Heavy clay soils were used for construction

of these cells, so a liner was not necessary and seepage

was minimal. The second phase of the pilot study focused

on the influence of open water zones, plant harvesting,

and kinetics optimization for BOD, TSS, and nutrient reChapter 8

Case Studies
140

moval. Some of the cells, for example, were subdivided

into smaller compartments with baffles and weirs along

the flow path. The results from these pilot studies not only

provided the basis for full-scale system design but have

contributed significantly to the state-of-the-art for design

of all wetland systems.

The full-scale treatment wetlands, with a design flow of

2.9 mgd, utilize three cells operated in parallel. Cells 1

and 2 have surface areas of about 2.75 acres each (L ≈

600 ft, W ≈ 200 ft), and cell 3 is about 2.0 acres (L ≈ 510 ft,

W ≈ 170 ft). The original design water depth was 2 ft, but

at the time of the 1997 site visit for this report they were

being operated with a 4-ft depth. Hardstem bulrush was

again used as the only plant species on these treatment

marshes. Clumps of plant shoots and rhizomes were hand

planted on about 1-m centers. Since nutrient removal is

not a requirement for the full-scale system, the treatment

marshes could be designed for a relatively short detention

time primarily for removal of BOD and TSS. The HDT in

these three cells is 1.9 d at design flow and a 2-ft water

depth. These treatment marshes were designed to produce an effluent meeting the NPDES limits for BOD and

TSS (30/30 mg/L) on an average basis. These wetland

cells utilized the bottom area of former lagoon cells. A schematic diagram of the operating system is shown in Figure

8-1.

The final “enhancement” marshes were intended to provide for further effluent polishing and to provide significant

habitat and recreational benefits for the community. These

three cells are operated in series at an average depth of

2.0 ft and have a total area of about 31 acres. Retention

time is about 9 d at average flow rates. The first cell (Allen

Marsh), completed in 1981, was constructed on former log

storage area and contains about 50% open water. The

second cell (Gearheart Marsh), completed in 1981, was

constructed on former pasture land and contains about

80% open water. The third cell (Hauser Marsh) was constructed in a former borrow pit and contains about 60%

open water. These 31 acres of constructed freshwater (effluent) marshes have been supplemented with an additional 70 acres of salt water marshes, freshwater wetlands,

brackish ponds, and estuaries to form the Arcata Marsh

and Wildlife Sanctuary, all of which has been developed

with trails, an interpretive center, and other recreational

features. The shallow water zones in these marshes contain a variety of emergent vegetation. The deeper zones

contain submerged plants (Sago pondweed) that provide

food sources for ducks and other birds and release oxygen to the water to further enhance treatment.

The construction costs for the entire system, including

modifications to the primary treatment plant, disinfection/

dechlorination, pumping stations, and so forth were

$5,300,000 (1985$). Construction costs for the treatment

wetlands are only estimated to be about $225,000, or

$30,000 per acre, or $78 per 1000 gpd of design capacity

(including removal of sludge from this site, which was previously a sedimentation pond for an aerated lagoon). This

does not include pumping costs to transfer final effluent

back to the chlorination contact basin, disinfection faciliPond

Treatment

Marshes

Enhancement

Marshes

2

1

3

Figure 8-1. Schematic diagram of wetland system at Arcata, CA
141

ties, or the pumping and piping costs to reach the enhancement marshes. Land costs also are not included since the

treatment wetlands were located on city-owned property.

8.1.1.2 Financial Arrangements

Construction costs for the Arcata system were funded

by a state/federal construction grant program with a grant

for 85% of the project costs. O&M costs for the system are

paid with a surcharge on the consumer’s water bills.

8.1.1.3 Construction and Start-up Procedures

Three of the final cells in the existing treatment pond

were selected for the treatment wetlands. This allowed the

use of city-owned land at no cost, a gravity flow connection from the ponds, and clay soils that eliminated the need

for liners and minimized the earthwork requirements. The

lagoon cells were drained and dried, local fill was placed

to the desired grade, and the wetland bottoms were graded

level. Shallow drainage channels were excavated to permit draining of the cell if desired. Construction of inlet and

outlet structures completed the construction activities. Each

cell has only one inlet and outlet structure. The wetland

effluent was collected from the bottom of the wetland in

each of these structures. The inlet structure has an adjustable weir so the flow to the three cells can be balanced.

The outlet weir is not adjustable and was originally designed to maintain a 2-ft water depth in the cell. Prior to

the 1997 site visit, timber sections had been bolted to the

top perimeter of the outlet box. This raised the water level

in the bed and converted this box to a four-sided overflow

weir. This allowed wetland effluent to be decanted from

the top of the wetland rather than off the bottom. The new

4-ft water depth was intended to suppress undesirable plant

species that had started to spread in the wetland.

The treatment wetland cells were hand planted with

hardstem bulrush with clumps of shoots and rhizome material planted on about 1-m centers. These plants were

obtained from the pilot wetland channels. The wetland cells

were flooded to a very shallow depth with tap water for

about three months to encourage new plant development

and growth. Wastewater was not applied at full depth until

the new plants were 3- to 4-ft tall and construction of the

rest of the system was complete (that is, effluent pump

station and other features).

The enhancement marshes were constructed on available waterfront land but at some distance from the basic

treatment system, so a pumping station and transmission

piping were required. These enhancement wetlands were

also sited on clay soils, so extensive soils and geotechnical

investigations were not necessary. The grading for these

wetlands was more complex than the treatment wetlands

because berms did not previously exist, and it was desired to produce a wetland with different water depths and

with several nesting islands in each cell. Each of these

cells also contains a single inlet and a single outlet structure. The final effluent comes through a highly vegetated

zone of emergent macrophytes with no open water. The

final cell is followed by a pumping station to return effluent

to the treatment plant for final disinfection and discharge.

The entire treatment system is operated and maintained

by three operators who work five days per week and are

on call on weekends. The only wetland-related O&M task

is adjustment of the inlet weirs to ensure that flow is properly balanced and to visually observe the status of the treatment wetlands; this might require 20 minutes per day. With

a total O&M effort of 87 hours per year at an assumed rate

of $30/hr for operator costs, the O&M costs for the treatment wetlands alone would be $2,600 per year. The O&M

costs for the pumping and the double chlorination would

add significantly to that, but these are unique to the Arcata

system and not generic to all wetland systems. Harvesting

or other special vegetation-management activities are not

practiced at this site.

8.1.1.4 Performance History

Performance data were collected for a two-year period

during the Phase 1 pilot testing program. This program

varied the flow rate and water depth in each of the two

cells to compare BOD removal performance at different

detention times and loading rates that would represent the

potential range for full-scale application at Arcata. These

data are summarized in Table 8-1. The BOD and TSS in

the pond effluent varied considerably during this period,

and not all of the cells were uniformly vegetated. Seasonal

variations in performance were observed, but Table 8-1

presents only the average effluent characteristics for each

of the cells over the entire study period. It is apparent from

the data that the wetlands were able to produce excellent

effluent quality over the full range of loadings and detention times used.

The long-term average performance of the Arcata system is summarized in Table 8-2. It is clear that both the

treatment and enhancement marshes provide significant

treatment for BOD and TSS. The long- term removals follow the pilot project results. Most of the nitrogen is removed

during the final stage in the enhancement marshes. This

is because of the long hydraulic detention time (HRT = 9

d), the availability of oxygen and nitrifying organisms in

Table 8-1. Summary of Results, Phase 1 Pilot Testing, Arcata, CA

(cells 1-4 had received two different hydraulic loading rates for one

year each-the higher loading occurred the first year)

Item HRT HLR BOD5 TSS FECAL COLI

(Actual) gal/ft2-d mg/L mg/L CFU/100 ml

Influent 26 37 3183

Effluent:

Cell 1 2.1/10.7 5.89/1.22 11 6.8 317

Cell 2 1.5/17 5.89/0.5 14.1 4.3 272

Cell 3 2.7/29 4.66/0.5 13.3 4.7 419

Cell 4 1.5/15 5.39/0.5 12.7 5.6 549

Cell 5 3.7 2.94 14.0 4.3 493

Cell 6 5.2 2.4 10.7 4.0 345

Cell 7 5.2 4.4 13.3 7.3 785

Cell 8 5.2 2.4 15.3 7.2 713

Cell 9 6.6 1.71 11.9 9.4 318

Cell 10 3.8 1.71 12.6 4.9 367

Cell 11 7.6 1.47 9.4 5.7 288

Cell 12 5.5 1.47 9.0 4.3 421
142

the open water zones, and anoxic conditions for denitrification in the areas with emergent vegetation.

8.1.1.5 Lessons Learned

The treatment wetlands (7.5 acres), with nominal HRTs

of three days, met weekly limits of 30 mg/L BOD and TSS

90% of the time. The enhancement wetlands (28 acres),

with a nominal HRT of 11 days, met weekly limits of less

than 5 mg/L BOD/TSS 90% of the time. Performance of

both wetlands results primarily from proper operation and

appropriate design that involves a combination of emergent vegetation and open water zones. TSS levels are

higher in cell effluents where outlets are located in open

water zones.

Wetland habitat values and opportunities for research

and environmental education provided by the enhancement marshes were optimized to gain state approval for a

near-shore discharge to Arcata Bay. Optimizing ancillary

benefits appears also to have complemented treatment

capabilities.

Nitrogen leaving the treatment wetlands in the ammonia

form is nitrified in the open water zones in the enhancement marshes, where deeper open water zones with submerged stands of Sago pond weed (Potemogeton

pectinatus) produce oxygen, and plant surfaces become

the substrate for attached- growth nitrifying organisms. This

plant also offers important habitat values since it is a major food source for many duck species. Long total detention times (~9 d) and alternating open and vegetated zones

resulted in excellent nitrogen-removal performance.

Duckweed that grows on the open water surfaces is prevented from becoming a permanent duckweed mat by sufficient wind action. In vegetated areas, duckweed does

mat, and it impedes reaeration.

Denitrification takes place in the fully vegetated anoxic

zones in the enhancement marshes.

Most of the fecal coliforms in the effluent are from birds

and other wildlife in the marshes and not from wastewater

sources.

Pilot-scale marshes produced better treatment than the

full-scale treatment wetlands, possibly due in part to the

different configurations of the two systems and the possibility for short-circuiting in the larger full-scale cells. A 3:1

aspect ratio for the full-scale cells is acceptable as long as

the influent is uniformly distributed over the full width of

the cell and the effluent collected in a comparable manner.

Short-circuiting probably occurs in the treatment cells,

but this could be corrected by replacing the single-point

inlet and outlet structures with perforated pipe manifolds

extending the full width of the cell.

8.1.2 West Jackson County, Mississippi

8.1.2.1 Background

The West Jackson County (WJC) wastewater treatment

system is owned and operated by the Mississippi Gulf

Coast Regional Wastewater Authority. It is one of several

treatment systems serving communities within the

Authority’s boundaries. The system is located near Ocean

Springs, MS, on the north side of I-10, about 20 miles east

of Biloxi, Mississippi.

The original WJC system included a 75-acre, multiplecell facultative lagoon for preliminary treatment followed

by 415 acres of slow-rate (SR) land treatment fields (growing hay). The land treatment site was underdrained, and a

portion of that recovered water was to be used to supplement dry-weather flow into marshes in the Mississippi Sandhill Crane National Wildlife Refuge. This system commenced operation in October 1987 with a design flow rate

of 2.6 mgd and a 56 d HRT. Problems developed soon

after start-up since the clay soils at the land treatment site

proved not to be as permeable as originally expected. After extensive investigations and discussions, the design

consultant agreed to design and construct a supplemental

free water surface wetland to treat the excess flow.

Three parallel wetland units were constructed with a total area of 56 acres to treat a design flow of 1.6 mgd. The

discharge from this new wetland was to Castapia Bayou,

with NPDES permit limits for BOD at 10 mg/L, TSS at 30

mg/L, NH4-N at 2 mg/L, DO at 6 mg/L, pH at 6.0 to 8.5, and

fecal coliforms at 2200/100 mL. Phase 1 of this new wetland was placed in operation in 1990 and Phase 2 in 1991.

Figure 8-2 is a schematic diagram of the 56-acre wetland.

Detailed costs are not available for the Phase 1 and 2

wetlands since they were funded privately by the design

consultant as part of the agreement with the Wastewater

Authority. Costs for Phase 3 (2.4 mgd) are available.

The Phase 1 wetland had two cells in series totaling 22

acres. The Phase 2 wetland (set 2) had three cells totaling

21.5 acres and another (set 3) which had two cells with a

total area of 12.5 acres. Local soils were all clays, so a

liner for these wetlands was not required. Bottoms of all

wetland cells were constructed with an average slope of

0.19% in the flow direction. At the end of each cell, multiple weir boxes were used as outlet structures with adjustable weir plates, allowing a maximum water depth in

the outlet zone of up to 2 ft. At the mean depth of 0.75 ft,

the design hydraulic residence times in the three wetlands

were 12.5 d in Phase 1, 10.1 d in Phase 2-2, and 10.7 d in

Table 8-2. Long-Term Average Performance, Arcata WWTP

Location BOD TSS TN

mg/L mg/L mg/L

Raw Influent 174 214 40

Primary Effluent 102 70 40

Pond Effluent 53 58 40

Treatment Wetlands 28 21 30

Enhancement Marshes 3.3 3 3
143

Phase 2-3. The inlets to all three sets of cells used a 12-

inch perforated PVC pipe manifold. The elevation of the

Phase 1 wetland was slightly higher than the lagoon, so it

was necessary to pump lagoon effluent to this wetland.

The Phase 2 wetlands were at a lower elevation, and gravity

was used as the motive force.

A unique feature of this wetland system was the incorporation of “deep zones” in each wetland cell. These consisted of trenches excavated perpendicular to the flow direction, with the bottom of the trench excavated about 5 ft

below the general wetland bottom surface. This provided

a water depth of about 6 ft in these “deep zones,” which

was sufficient to prevent colonization by the emergent

wetland plants. These trenches are about 20 ft wide at the

bottom and about 40 ft wide at the top. The purpose of

these “deep zones” was to redistribute the flow across the

width of the cell to minimize short-circuiting and to provide

an open water surface for atmospheric reaeration to supply the oxygen necessary for ammonia removal. The potential open water provided by these zones was only about

10% of the surface area in each wetland cell. The water

surface in these zones also quickly became colonized by

duckweed (Lemna spp.).

In larger open water bodies, duckweed is very susceptible to wind action; as a result, the water surface can remain available for atmospheric reaeration. At this location,

with the relatively narrow “deep zones” and the protection

provided by the adjacent emergent vegetation in the shallow portions of the marsh, wind was not sufficient to move

the duckweed mat, and oxygen transfer from the atmosphere did not develop.

Because there was insufficient oxygen in the system to

continuously support nitrification reactions, there were seasonal violations (particularly in late summer and early fall)

of the ammonia limits commencing in 1992. Corrective

action for this problem considered an external vertical-flow

filter bed for nitrification and submerged tubing aeration in

the “deep zones” to provide the necessary oxygen. The

latter was selected as the lower-cost alternative and was

installed in 1993. Problems again developed because nutria (an animal similar to a muskrat), which occupy the

wetland in large numbers, were attracted to the air bubbles

and destroyed the aeration tubing by gnawing on it. In subsequent discussions with the State of Mississippi, it was

decided that the ammonia limit would not be enforced until

the Castapia Bayou began to exhibit oxygen stress, so the

aeration tubing was not replaced.

The population is increasing rapidly in the communities

served by this system, and by 1996 the average flow into

these wetland units had reached 2.2 mgd. The WastewaPhase 2

Phase 1

Post Aeration

3A

2A

1B

2B

3B

2C

1A

Weir Boxes Deep Zones

Cell

1A

1B

 2A 

 2B 

 2C

 3A

 3B 

Area-Acres

12

10

10

 8

 4

 9

 3

(hectares)

(4.8)

(4.0)

(4.0)

(3.2)

(1.6)

(3.6)

(1.2)

Figure 8-2. Schematic diagram of Phase 1 and 2 wetland systems at West Jackson County, MS
144

ter Authority then authorized an upgrade and expansion of

the facility for a design flow of 5 mgd (1 mgd to land treatment, 4 mgd to wetlands). The design was completed in

1996 and construction began in August 1997. The expansion included modifications to the lagoon (providing aeration and baffles), 50 acres of additional wetland area, a

plastic-media trickling filter bed for nitrification, UV disinfection, and additional post-treatment aeration to ensure

adequate DO in the final effluent. The expanded wetland

system is shown in Figure 8-3. The trickling filter component has been designed but will not be constructed until

the State of Mississippi decides it will be necessary. Recent data, shown in Figure 8-3, indicate that the open water zones appear to be functioning well. A UV disinfection

system was added to the system because the NPDES

permit limits for fecal coliforms have been modified to 200/

100 mL in the summer and 2000/100 mL in the winter.

Additional features of the existing Phase 1 and Phase 2

wetlands include post-treatment aeration to satisfy the

NPDES discharge limit for dissolved oxygen (6 mg/L).

Multiple outlet structures with adjustable weirs are used

for cell-to-cell transfer and for final discharges from the

wetlands. A miniature “deep zone” was excavated around

each of these structures to prevent the growth of emergent vegetation in the immediate vicinity, as done for a

FWS system at Fort Deposit, AL. Published design models were not used in this case, and effluent quality (Figure

8-3) is excellent.

8.1.2.2 Financial Arrangements

The initial lagoon/land treatment system was funded

under a federal/state construction grant program in existence at the time. The Phase 1 and 2 wetlands were funded

directly by the design consultant. The Phase 3 expansion

was funded through a revolving loan fund as administered

by the State of Mississippi. Total costs for the entire Phase

3 project are estimated to be $2,758,000 (1997$). This

includes lagoon modifications, UV disinfection, the nitrification trickling filter, and post-aeration. The costs for just

the 50 acres of wetland expansion are estimated to be

about $700,000, or $14,000 per acre, or $250 per 1000

gallons of design flow capacity. Land costs for this project

are not included in this estimate since the land already

belonged to the Wastewater Authority. The O&M costs are

funded by a surcharge on each consumer’s water bill within

the Authority’s service area.

8.1.2.3 Construction and Start-up Procedures

The site for the original lagoon and land treatment site

was selected for its proximity to the Mississippi Sandhill

Crane National Wildlife Refuge, where the treated effluent

could be utilized in refuge marshes. The sites for the Phase

Phase 2

Phase 3

50 Acres

(20 ha)

Phase 1

Post Aeration

3A

2A

1B

2B

3B

2C

1A

Weir Boxes Deep Zones

Figure 8-3. Schematic diagram of Phase 3 wetland expansion at West Jackson County, MS
145

1 and 2 wetlands were selected to take advantage of available land already owned by the Wastewater Authority and

for proximity to Castapia Bayou for the system discharge

point. As shown in Figure 8-3, the Phase 3 wetlands were

then located to utilize the remaining land available on this

site.

The site investigation for the wetlands included a

geotechnical investigation which indicated that ground

water impacts or intrusion would be minimized by underlying clay subsoil. Borings and test pits verified the presence of these clay soils throughout the proposed wetland

area. Wetland sites originally were covered with scrub pine

and related ground cover, so clearing and grubbing of the

site was the first construction task. This was followed by

excavation to grade with typical highway construction

equipment and construction of both the external and internal berms with spoil material from the excavations. A liner

was not necessary due to the low permeability of clay soils,

but geotextiles were used adjacent to the inlet and outlet

structures to prevent erosion. The multiple concrete outlet

and transfer structures were cast in place.

The bulrush (Scirpus spp.) and cattails (Typha spp.) selected as the vegetation for this wetland system were obtained locally by cleaning drainage ditches within the

Authority’s jurisdictional area. The plants were brought to

the site and separated, and shoots were cut to about 1 ft in

length and planted by hand in individually augured holes.

The plants were placed on about 1-m centers. At that density, it would require about 227,000 plants to cover the original 56-acre wetland site. Observations at the time showed

that two laborers could prepare and plant about 1,200

plants per day. At that rate it would require about 54 manhours per acre to vegetate a wetland bed using this technique.

As soon as a zone was planted, it was flooded with a

very shallow depth of stream water to encourage plant

growth. Lagoon effluent at the design depth was not applied for at least 30 days after planting was completed. A

pattern was adopted for Phase 1, on which alternating

bands of bulrush and cattails were planted. This was abandoned for Phase 2, so whichever species was available

on a given day was planted; as a result, cattails were the

dominant species in Phase 2.

8.1.2.4 Performance History

At the original 1.6-mgd wetland design flow rate, the

flow was split between the three wetland units: 0.6 mgd to

Phase 1, 0.65 mgd to Phase 2-2, and 0.35 mgd to Phase

2-3. During the period 1992 to 1995, the average effluent

characteristics from the facultative lagoon were BOD 31

mg/L, TSS 33 mg/L, TKN 12.9 mg/L, and NH4-N 4.4 mg/L. During this same period, the combined final effluent from the wetland units met all NPDES limits on an annual average basis:

BOD 7.5 mg/L, TSS 4.6 mg/L, and NH4

-N 1.85 mg/L. On a

monthly basis, there were excursions; the BOD exceeded

permit limits eight times (18%) and ammonia exceeded limits 11 times (25%). The BOD violations were randomly distributed throughout the period and generally reflected

higher-than-normal loading. A more specific pattern was

shown by the ammonia, with the violations occurring in

late summer and early fall. By 1996 the flow rate to the

wetlands had increased to 2.35 mgd (47% higher than the

original 1.6 mgd design), and the excursions for both BOD

and NH4

-N were more frequent. Table 8-3 summarizes performance data for the period January 1996 through July

1997. It would appear from the data in this table that BOD

removal is slightly better in the warmer months, indicating

some dependence.

8.1.2.5 Lessons Learned

The inlet zone was submerged with a significant depth

of water, so the overland flow mode with very shallow sheet

flow did not develop. Ammonia removal provided by the

shallow sheet flow of water and the continuous availability

of oxygen intended for the system did not take place, even

with the weirs at their lowest setting.

“Deep zones” in each cell were intended to provide additional oxygen to support nitrification reactions, but this

benefit was not realized when duckweed mats formed over

the surface of the deep zones. A single large “deep water”

zone in each cell, instead of multiple narrow trenches,

should have allowed sufficient duckweed movement to

sustain atmospheric reaeration.

The Phase 3 design provides both a perforated manifold and an open ditch “deep zone” in the inlet area of the

cells to promote proper lateral distribution of the influent

and increased volume to capture incoming TSS.

Bulrush plants had been almost completely removed by

muskrats and nutria using these plants for food and nesting material, but damage to cattails was minor. Damaged

Table 8-3. Wetland Water Quality, West Jackson Co., MS

Date BOD5, mg/L TSS, mg/L NH4-N mg/L

In Out In Out In Out

1996

Jan 36 10 28 12 7 4

Feb 32 13 21 10 12 4

Mar 36 12 24 14 9 6

Apr 30 14 18 6 9 6

May 32 8 32 5 10 4

Jun 38 6 38 3 18 4

Jul 36 3 42 5 6 3

Aug 37 9 60 12 12 2

Sep 32 8 40 13 2 2

Oct 34 8 90 7 2 1

Nov 34 8 41 5 2 1

Dec 20 10 17 3 1 1

1997

Jan 33 13 18 5 1 1

Feb 23 9 17 6 2 2

Mar 43 10 24 5 16 8

Apr 41 9 23 4 9 6

May 28 4 24 2 9 7

Jun 47 6 23 2 1 2

Jul 42 4 43 3 3 2
146

sections in Phase 1 were planned to be temporarily drained

and replanted with cattails. In the new Phase 3 addition,

cattails were proposed as the only plant in the bottoms,

and a variety of attractive flowering wetland species were

planned for the inside perimeter of the new cells.

The system supports large numbers of birds and other

wildlife, even though special measures to enhance habitat

values were not taken in Phases 2 and 3.

Birds and other wildlife in the wetlands appear to have a

significant impact on effluent fecal coliforms from the system, with the final wetland effluent often higher than lagoon effluent entering the wetland.

8.1.3 Gustine, California

8.1.3.1 Background

Gustine is an agricultural community located in the Central Valley of California on the east side of I-5 and about 60

miles south of Stockton. There are several milk-processing industries in the community that impose high organic

loadings on the municipal wastewater treatment system.

The original treatment system consisted of an oxidation

pond with 14 cells operated in series (HRT ≈ 56 d, average pond depth ≈ 4 ft), with final discharge without disinfection to a small stream. Approximately one-third of the 1

mgd design flow originates from domestic and commercial sources; the remaining two-thirds come from dairy product industries. This combination produces a high-strength

wastewater with an average BOD of about 1200 mg/L and

TSS of 450 mg/L, and these characteristics resulted in frequent violations of the 30 BOD/30 TSS NPDES discharge

limits for the original lagoon system.

In 1981, a Facility Plan for the city was funded under the

Clean Water Act. This plan considered a number of alternatives for upgrading the existing treatment system. The

most cost-effective alternative was assumed to be a facultative lagoon followed by a constructed free water surface

(FWS) wetland for final polishing to consistently meet the

NPDES discharge limits. Since design criteria for FWS

wetlands were not well established in 1981, a pilot test to

develop final design criteria was recommended.

A pilot study was approved and was conducted from

December 1982 to October 1983. The pilot system modified an existing ditch that already contained a stand of cattails. The pilot cell was 39 ft wide and 900 ft long. Partially

treated water was taken from various intermediate pond

cells. The influent to the pilot wetland averaged 180 mg/L

BOD and 118 mg/L TSS. At the operational water depth of

6 in, the HRT in the wetland averaged 2.5 d. The BOD and

TSS in the wetland effluent stabilized at 30 mg/L after the

start-up period. Just prior to the pilot testing, one of the

dairy industries closed and was not expected to reopen.

This resulted in lower-strength wastewaters than had been

previously experienced, and these were expected to prevail in the future. The pilot results were used as the basis

for the design and sizing of the full-scale wetland component. Construction of the system commenced in March

1986 and was completed in October 1987.

Three of the 14 existing lagoon cells (plus some additional adjacent land) were selected as the site for the new

wetland component. This area was converted to 24 wetland cells operating in parallel. Each cell had a net area of

about 1 acre and was 38 ft wide and 1107 ft long (L:W =

29:1), similar to the size and the configuration of the pilot

wetland unit. Internal berms constructed to separate the

wetland cells were 10 ft wide and 2 ft deep. An exterior

levee 6.5 ft high was constructed around the entire wetland area to provide protection from the hundred-year flood,

as required by the State of California. Influent flow from

the lagoon passes through a distribution box where V-notch

weirs divide the flow into six equal parts. Each part of the

flow is then piped to a group of four cells. Gated aluminum

pipe is used to distribute flow across the width of each cell.

In order to provide flexibility for high-strength flows, a simple

step-feed arrangement was designed. Pipe manifolds were

located at the inlet to each cell and at the one-third point

along the flow path. Each manifold was valved so the operator could vary the amount of flow applied to each and

thereby avoid an overloaded inlet zone. An adjustable outlet weir at the end of each cell allows a water depth ranging from 4 to 18 in. These weirs discharge to a common

sewer, and the effluent is then pumped to the chlorine disinfection/dechlorination system. At the design flow of 1 mgd,

the design projected an average HRT at higher loading of

about seven days, which could be varied from 4 d in the

summer to 11 d in the winter, depending on the number of

cells in operation and on the water depth used. This operational flexibility allowed for each cell to be taken out of

service each summer for vegetation management or other

O&M, if required. The system is schematically depicted in

Figure 8-4.

At the time this system was designed, the capability to

effectively remove algae in FWS wetland systems was not

clearly established. This issue was a concern since very

high concentrations of algae were known to develop during the summer months in some of the lagoon cells at

Gustine. In order to provide the operator some control over

this situation, the new design incorporated separate outlet

structures in each of the last seven cells of the remaining

11-cell (in series) lagoon system. In this way, the operator

could visually observe which cell(s) had the least amount

of algae present and select those for discharge to the wetland.

Soon after the 1987 system start-up, the milk-processing industry in Gustine that had been closed for several

years was reopened and full-scale operations commenced.

This imposed a higher than expected organic load on the

treatment system; as a result, the lagoon/wetland system

could not consistently meet the 30/30 (BOD/TSS) NPDES

discharge limits, especially during the winter months. Following consent decree discussions with the U.S. EPA, the

City of Gustine evaluated the performance of the system

and recommended action that would bring the system into

compliance. The major system modification resulting from

this study was the addition of floating aerators to most of

the lagoon cells in order to reduce the organic loading.
147

The shallow 4-ft depth of the lagoon cells is not desirable

for efficient aeration but was too expensive to modify. Aspirator aeration equipment was selected for this project

because of the shallow water depth. This equipment was

installed in 1992, and the lagoon has performed acceptably since.

8.1.3.2 Financial Arrangements

Federal and state funding for the facility plan, the pilot

study, and the design and construction of the wetland system was provided under the Clean Water Act Construction

Grant Program administered in the early 1980s. The total

construction cost for this wetland system was $882,000

(August 1985$). This includes the cost of the multiple-pond

outlet structures and related piping and the 6.5 ft levee

around the perimeter of the wetland area. On an area basis (24 acres of treatment area), the cost would be $36,750

per acre. On a design flow basis (1 mgd system), the cost

would be $882 per 1000 gpd of treatment capacity. Land

costs are not included since the area was already owned

by the city. The gross area utilized was about 36 acres for

the wetlands, levees, and disinfection facilities.

For this facility, O&M costs are funded by a surcharge

on the consumer’s water bill. The City of Gustine budgeted

$433,275 in 1996 for operation and maintenance of the

wastewater treatment and sewerage systems in the community. Based on a 1-mgd design flow, the unit costs would

be $1.19 per 1000 gallons treated. The O&M costs for just

the wetland component are minimal and might represent

less than 10% of the total (e.g., $0.12/1000 gpd).

8.1.3.3 Construction and Start-up Procedures

There were no soils or geotechnical or ground water investigations at this site prior to or during construction. The

local soils were clays and the existing lagoons are unlined,

and it could therefore be assumed that the wetland cells

would not require lining. The site lay within the flood plain

of a small local stream, and the State of California did require a 6.5-ft high levee to protect the new wetland system

from the hundred-year flood.

Construction commenced with the draining and drying

of the three existing lagoon cells and clearing and grubbing of the adjacent land required to complete the system.

Figure 8-4. Schematic diagrams of the wetland system at Gustine, CA

1107 ft 38 ft

b) Typical Wetland Cell - Plan

a) Overall System - Plan

Wetland

Ponds

New Outlets
148

The entire wetland area was excavated to grade, with a

flat bottom, and then the interior berms were placed as

fill. Construction of the inlet and outlet structures and

the related piping completed the physical aspects of the

wetland system.

Cattails (Typha latifolia) and tristar bulrush (Scirpus

californicus) were selected for use on this wetland

project. The specifications required that 18 of the cells

would be planted with cattail rhizomes on 3-ft centers,

and the remaining six cells would be planted with bulrush rhizomes on 1.5-ft centers. During the first planting attempt in September 1986, rhizomes of both plant

species were obtained at local natural stands and spread

on the wetland surfaces and disked into the soil. Water

was not available for irrigation and very few plants

emerged the following spring. The second planting attempt occurred in June 1987 and consisted of mechanical planting of cattail seedlings obtained at a nursery.

The bed was then flooded with high-BOD pond effluent.

Almost all of the plants died in a short time. It is believed this was due to heat stress (the air temperatures

were 100%F) and the high oxygen demand from the

poorly treated water used to irrigate. The contractor also

seeded the wetland area by broadcasting mixed bulrush seed (hardstem and tristar). Some live cattail plants

were also transplanted to the wetland beds from local

drainage ditches. These more mature plants survived,

whereas the small seedlings did not. By the fall of 1987,

a few of the cells were almost completely covered with

bulrush plants, but the majority contained random stands

of bulrush and isolated patches of cattails.

As a result of these planting problems, the system

started up in 1987 with insufficient plant cover to provide the necessary substrate for treatment and an organic loading that was more than double the design load.

In the spring of 1988, about half of the cells had moderately dense growth over about 75% of the cell area. The

other half of the system contained only random patches

of bulrush and cattails. This situation improved slowly

during subsequent years; at the time of the 1989–1990

wetland evaluation, the wetland cells were still not completely covered with vegetation.

The hardstem and tristar bulrush gradually spread and

became the dominant species on the system. The wetland could be considered to be completely vegetated

since early 1993. However, during the 1997 site visit for

this report, patches of sparse vegetation and some open

water areas were still observed on some cells. As the

density of the vegetation increased, it began to create

hydraulic problems for operation of the system. Flow

through these FWS wetlands is thought to be governed

by Manning’s equation. The frictional resistance to flow

through a wetland bed is significantly higher than in a

normal grassed drainage channel since the vegetation

(and litter) exists throughout the full depth of the water

column. This resistance obviously increases and the

length of the flow path increases. A system with a high

aspect ratio (29:1 at Gustine) has the potential to develop a high enough resistance to force the water level

at the inlet to increase very significantly in order to provide the necessary hydraulic gradient. This occurred at

Gustine, and the water level at the inlets overtopped the

shallow berms. The operator had two options to solve this

problem: increase the height of the berms or reduce the

resistance to flow. He chose the latter course and got permission to burn the vegetation in late fall after senescence.

This immediately solved the hydraulic problem, and burning has become an annual occurrence at Gustine.

There were no special start-up procedures used at this

system, and pond effluents at the full rate were applied to

all of the wetland cells regardless of vegetation coverage

in the fall of 1987. However, until corrective action was

taken in 1992 to reduce the organic loading in the ponds,

the system frequently did not meet the NPDES discharge

limits.

8.1.3.4 Performance History

As described in the previous section, this system did not

consistently meet the NPDES limits at start-up and for several years thereafter. The problem was due to a higher

than expected organic load (particularly in the winter

months) and the lack of significant vegetation coverage

on most of the wetland cells. The vegetation and litter in

these FWS systems serve as the means for enhanced flocculation and sedimentation that actually perform the treatment. The importance of this vegetation and litter can be

seen by comparing the data in Table 8-4. These are performance results obtained during the 1989–1990 winter in

the special evaluation study, when several cells in the

Gustine wetlands were isolated and a careful performance

evaluation was conducted over a one-year period. One

cell (#6D in their set) was almost completely vegetated

with bulrush; the other cell (#2A in their set) listed in Table

8-4 was sparsely vegetated with some bulrush and cattails. The influent BOD was slightly different for the two

cells because different source ponds were in use, but during the period of concern the influent values were in the

same range. The HRT during this period was about 10 d

for both cells.

This FWS wetland system has had problems in meeting

its NPDES discharge limits since start-up. During the period 1987 to 1992, the problem was believed to be an organic overload on both the lagoon pretreatment and on

the wetland component in this system as evidenced by

the data in Table 8-4. The system design expected a BOD

concentration of 150 mg/L entering the wetland, but the

actual average wetland influent BOD during this initial period was close to 300 mg/L, with weekly excursions up to

630 mg/L during the winter months. This problem was compounded by the immature vegetative growth in the wetland cells that significantly reduced the flocculation/sedimentation treatment potential. The organic loading problem was corrected by the addition of aeration capacity to

the lagoon component, and the plant density has gradually increased on the wetland cells.
149

Table 8-5 presents current wetland effluent BOD and TSS

values for 1996–1997. Ammonia (NH3) and Kjedahl nitrogen (TKN) values also were measured during the first half

of 1996 and are shown in Table 8-5. Wetland influent characteristics were not measured during this period, but the

new aerators in the lagoon cells were operating continuously, so it can be assumed that the organic loading on

the wetland probably does not exceed the original design

expectations. During 1996 the average daily flow into the

treatment system was 1.02 mgd, which is essentially equal

to the design capacity, so the system is not overloaded

hydraulically. It is clear from these data that this wetland

system is still having difficulty meeting the NPDES discharge limits. The effluent BOD exceeded 30 mg/L twice

during 1996 and once during the first half of 1997; the effluent TSS met the 30 mg/L limit three times during 1996

(25% of the time) and three times during the first half of

1997 (50% of the time).

8.1.3.5 Lessons Learned

Poor performance evident in the 1996–1997 data may

not have been caused by an organic overload, since the

effluent BOD had been significantly below the discharge

limit most of the time, and the few BOD excursions were

relatively small and occurred mostly during the warm

months. Birds and other wildlife may be a contributing factor. Detritus and similar natural organic materials may also

be a source for the excess TSS.

Plant litter allowed to accumulate in the cells may improve water quality. With the litter burned each year, solids entrapment must depend on living plants. Also, the

tristar bulrush that dominates many of the wetland cells

has narrow stalks and no leaves, so the plants’ surface

area beneath the water surface is minimal, further reducing entrapment of solids. However, plant litter in the channels caused hydraulic failure when resistance to flow increased and the cells overflowed their banks at the entry

zone, owing to the excessive L:W ratio.

The hydraulic problem was corrected in 1995 when three

interior berms in each set of four cells were removed and

some of the surplus material was used to increase the

height of the remaining berms. This action reduced the

system to six larger cells, changed the aspect ratio from

26:1 to 5.5:1, and increased water depth near the entry

zone. Removing the three interior berms also opened up

an additional 30,000 ft2

 in each of the remaining cells to

serve as part of the wetland.

The initial selection of long, narrow wetland channels

was consistent with wetland design experience available

in 1983–1984, but experience has since shown that proper

treatment can be achieved in FWS wetlands with aspect

ratios as low as 2:1 to 3:1 as long as the system is properly constructed and the inlet and outlet structures allow

for uniform flow through the system.

Separate outlet structures at seven of the 11 lagoon cells

allow the operator to select the lagoon cell(s) with the least

algae for discharge to the wetland. This technique has been

very effective at algae removal as long as sufficient plants

and litter are present in the wetland, and as long as velocity of flow in large open water zones is sufficient to prevent

redevelopment and discharge of algae to the FWS system.

Toxicity discharge limits imposed by the State of California for un-ionized ammonia could not be met by the existing pond-wetland system in the present mode of operation. The existing point discharge is planned to be abandoned and the effluent from the wetland component to be

used for irrigation in a slow-rate land-treatment system.

8.1.4 Ouray, Colorado

8.1.4.1 Background

Ouray is located in southwestern Colorado, about 60

miles north of Durango, on State Route 50. Its population

is about 2,500 in summer and about 900 in winter. The

town is at an elevation of 7,580 ft in a mountain valley and

experiences severe winter conditions.

The free water surface (FWS) wetland at Ouray receives

influent from a two-cell aerated lagoon and provides secondary treatment prior to chlorine disinfection/dechlorination and final discharge to the Uncompahgre River. The

NPDES monthly average discharge limits are BOD5 30 mg/

L, TSS 30 mg/L, and fecal coliforms 6000 CFU/100 mL.

The wetland was designed for an expected winter water

temperature of 3°C and a summer water temperature of

20°C, with a 25% safety factor on sizing for BOD removal,

based on existing design equations.

Table 8-4. Performance Results in Mature Vegetated vs Immature Vegetated FWS Cells, Gustine, CA

Full Vegetation Partial Vegetation

BOD5, mg/L TSS, mg/L BOD5, mg/L TSS, mg/L

Month Temp In Out In Out In Out In Out

1989-90 °F*

Nov 50 54 8 72 13 49 20 80 24

Dec 43 154 17 103 6 150 66 104 42

Jan 44 525 35 116 24 515 153 113 71

Feb 45 483 29 79 19 478 123 86 49

Mar 56 215 16 100 17 185 21 94 20

 Avg - 286 21 94 16 275 77 95 41

* Average water temperature in the marsh cells
150

The design flow for the 2.2-acre wetland system is 0.250

mgd in winter and 0.363 mgd in summer. As shown in Figure 8-5, the wetland includes two parallel trains with three

cells each. The two trains operate in parallel, and one can

be taken out of service during the summer months for

maintenance if required. The curved configuration of these

wetland cells was selected in part because of the confined

site and in part for aesthetic reasons. The water depth in

the cells is adjustable via the outlet from a minimum of 8 in

to a maximum of 18 in. This water level is increased to the

maximum depth prior to the onset of winter to provide the

maximum possible detention time during the low temperature periods and to provide additional depth for ice formation on the water surface during the winter months.

A perforated manifold is used for both inlet and final outlet structures for the two sets of cells. Internal transfer from

cell to cell is accomplished with two parallel pipes through

each internal berm. The wetland cells are lined with 30-mil

HDPE membrane liners to prevent seepage since the local soils are sandy clay loams. The detention time in the

system depends on water depth and on the presence of

winter ice. At minimum water depth the HRT is 2.2 d; at

maximum water depth without ice, the HRT is 3.8 d. Locally obtained cattails (Typha spp.) and bulrush (Scirpus

spp.) were planted in the wetland cells. The vegetation is

continuous, and there are no intended open water zones.

The wetland was designed in 1992, constructed during

the spring and summer of 1993, planted in October 1993,

and placed in partial operation in November 1993. It has

been in continuous operation since that time.

The construction costs for the entire system, including

aerated lagoons, chlorine disinfection/dechlorination equipment, and miscellaneous features was $816,530 (1993$).

Construction costs for just the treatment wetlands is estimated to be about $108,500 or $49,300 per acre, or $300

per 1000 gpd of design capacity.

8.1.4.2 Financial Arrangements

The construction costs for the Ouray system were funded

by a state revolving-loan fund as administered by the State

of Colorado in 1993. O&M costs for the system are funded

with a surcharge on the consumer’s water bills. The total

monthly O&M cost for the entire system at Ouray is $2,625;

most of this is related to power costs, sludge removal from

the aerated lagoon, and laboratory testing for NPDES

monitoring. The average O&M costs for the wetland component is estimated to be about $200 per month for minor

maintenance tasks.

8.1.4.3 Construction and Start-up Procedures

A geotechnical investigation was undertaken at the site

for the new wetlands to determine underlying soil properties and ground water conditions. Soil borings to several

feet below the final wetland grade revealed the presence

of sandy clay loams with sand and gravel inclusions, with

an unconfined ground water aquifer at greater depth. These

soils were typical of the local flood plain and were considered too permeable, so a membrane liner was selected

for the wetland.

Clearing and grubbing was the first construction activity

at the new wetland site. This was followed by grading, berm

construction, and liner placement. Prior to liner placement,

the subgrade was leveled and compacted to 90% of Proctor density to preserve the intended grade during subsequent construction activities. After the liner was placed,

1.5 ft of local sandy clay loam was placed in the wetland

bottom to serve as the rooting medium for the wetland

vegetation. The curved configuration of the wetland cells

increased construction costs somewhat, but the site was

Table 8-5. Wetland Effluent Characteristics, Gustine, CA

Month Temp. °F BOD5, mg/L TSS, mg/L NH4-N mg/L TKN, mg/L

1996

JAN 48 21 20 13 15

FEB 54 20 22 16 18

MAR 55 16 24 11 16

APR 61 30 45 4 7

MAY 64 29 57 3 9

JUN 72 18 41 2 7

JUL 75 22 36 7 15

AUG 70 47 71 - -

SEP 63 32 39 - -

OCT 59 26 43 - -

NOV 52 28 42 - -

DEC 48 24 50 - -

1997

JAN 48 25 26 - -

FEB 50 22 27 - -

MAR 54 27 38 - -

APR 61 24 34 - -

MAY 66 22 26 - -

JUN 70 31 34 - -
151

too confined to permit construction of a typical rectangular

system with straight sides.

Treatment wetland cells were hand planted by correctional facility inmates with locally obtained bulrush and

cattail plants. The vegetation was planted on about 18-in

centers; at this density about 43,000 plants were required.

The bed was flooded with about 8-in of water and maintained in that condition until sufficient new plant growth

was observed. Some wastewater was applied during the

remainder of the 1993 winter, but full-scale operation did

not commence until the spring of 1994.

This system experiences subfreezing air temperatures

for extended periods each winter. An ice cover at least 6-

in thick persists for at least six months.

The inlet and outlet devices for each set of cells are 8-in,

perforated, Schedule 80 PVC pipe. These pipes were laid

in a 2-ft-wide, 18-in-deep trench extending the full width of

the cell. The trench bottom and sides are protected with 2-

to 4-in riprap. One end of each manifold has a 90% elbow

and a capped riser extending above the water surface to

serve as a cleanout if required. The effluent manifolds connect to a concrete outlet structure that contains adjustable

outlet riser pipes for controlling the water level in the cells.

There are no special O&M requirements for these wetland units, including harvesting or other plant management

procedures. Raising and lowering the wetland water levels on a seasonal basis and sampling for NPDES compliance are about the only O&M tasks required. There have

been no problems with muskrats or other animals damaging the plants as has occurred at several other wetland

systems. The wetland tasks listed in the O&M manual include weekly cleaning of effluent debris screens, weekly

checking of berms for erosion or muskrat damage, cleaning influent and effluent manifolds as required, and occasional muskrat control. Mosquitoes have not been a problem at this site.

8.1.4.4 Performance History

It is typical for most small systems, including the Ouray

system, to monitor only for NPDES limits, and for that reason to sample only the untreated (raw) wastewater and

the final effluent. As a result, the actual influent to the wetland component is not known. Data from the Ouray system for the 1995–1996 period is shown in Table 8-6.

Based on limited data, the aerated lagoon at Ouray is

estimated to remove about 54% of influent BOD5 and 65%

of influent TSS. On that basis, with the average wetland

influent in 1995 at 58 mg/L BOD5 and 63 mg/L TSS, the

wetland achieved an average removal of 83% BOD5 and

90% TSS. In 1996, the average wetland removal percentages were 88% for BOD5

 and 91% for TSS. Wetland average effluent fecal coliform concentrations during 1995 and

1996 were 570 CFU/100 mL and 1300 CFU/100 mL, respectively. All of the monthly values were well below the

NPDES limit of 6000 CFU/100 mL, so it was not necessary to operate the disinfection/dechlorination equipment

installed at the site.

8.1.4.5 Lessons Learned

• The Ouray system incorporated many improvements

learned from earlier FWS systems, including perforated

Figure 8-5. Schematic diagram of the wetland system at Ouray, CO

To Discharge

Manifolds

Chlorination/

Dechlorination

Facility

Transfer Pipes

From Lagoon

53' x 260'

60' x 256'

57' x 160'

58' x 104'

57' x 220'

54' x 234'

Manifolds
152

manifolds extending the full width of the wetland cells

for inlets and outlets, cleanouts on the ends of these

manifolds, and a simple adjustable outlet structure for

control of the water level in the wetland cells.

• The adjustable outlet structure for water-level control

was essential for the water level to be raised during

the winter months to accommodate expected ice formation.

• Large-sized riprap (4- to 6-in size) as a permanent sloping cover for both the influent and effluent manifolds

excludes clogging debris and prevents algae development. This technique precludes periodic cleaning

of a screen over the effluent manifold that would have

been installed to prevent accumulation of debris.

• Bats and dragonflies contribute to mosquito control during the warm summer months, so mosquitoes and similar insect vectors have not been a health problem at

this system.

• Odors occasionally noticed at the inlet end of the wetland cells are caused by accumulation of TSS and algae carried over from the final aerated lagoon cell

because the settling zone in this final lagoon cell is too

small to be completely effective.

• Ice cover and snow accumulation have provided acceptable thermal protection for the FWS system, and

the system has not needed alteration during the winter months. In response to State of Colorado concerns

that FWS wetlands would not sustain acceptable performance during low-temperature winter months, the

lagoon aeration system had been designed to allow

longer operational periods during winter months to

provide additional treatment so the wetland cells could

have been bypassed during winter months, if necessary.

• Chlorination/dechlorination equipment included in the

original design at the insistence of the State of Colorado has not been used, as the wetland effluent has

been consistently below permit limits.

8.2 Vegetated Submerged Bed (VSB)

Systems

8.2.1 Village of Minoa, New York

8.2.1.1 Background

The Village of Minoa is a small residential community of

approximately 3,700 in central New York state east of Syracuse. The average daily flow to the wastewater treatment

plant in 1993 was approximately 0.35 mgd, but peak flows

as high as 1.6 mgd had been recorded. Efforts between

1990 and 1993 to abate the high rates of infiltration and

inflow were unsuccessful, and the Village of Minoa was

forced into a consent order with the New York State Department of Environmental Conservation (NYSDEC) to

correct discharge violations.

In 1994 the village decided to use a VSB constructed

wetland system to treat primary effluent to secondary effluent standards, with an ultimate oxygen demand limit that

required at least partial nitrification. The VSB system also

would be used during wet weather conditions to treat

640,000 gpd of wet weather flow. The dry weather capacity of the VSB system was to be 160,000 gpd, but the actual constructed size of the system was smaller than the

original design, reducing the design capacity to approximately 130,000 gpd. The treatment goal also was changed

from a BOD5

 concentration of less than 30 mg/L and partial nitrification to BOD5

 alone.

Two New York state agencies and the U.S. EPA provided grant funds to the village for incorporation of several

special features in the VSB system and for a research and

technology transfer study of the system by researchers at

Clarkson University, Potsdam, NY.

The VSB system consists of three cells that can be operated in parallel, combined parallel and series, or series

modes. Cells 1 and 2 are approximately the same size

(0.17 ha or 0.42 acres). Cell 3 is significantly smaller and

is irregularly shaped (0.1 ha or 0.25 acres) (Figure 8-6). At

the inlet end, the media depth is 0.5 m and the bottom

surface has a slope of 1%, resulting in a bed depth of approximately 0.9 m at the outlet end and an average depth

of 0.76 m. The upper 7.6 cm of the beds consist of 0.6 mm

pea gravel, which allowed for the establishment of wetTable 8-6. BOD & TSS Removal for Ouray, CO

Date BOD In, BOD Out, TSS In, TSS Out,

mg/L* mg/L** mg/L* mg/L **

1995

Jan 84 5 124 11

Feb 78 8 122 8

Mar 84 7 216 10

Apr 132 9 182 5

May 66 8 152 11

Jun 174 15 196 4

Jul 180 14 170 5

Aug 216 10 316 5

Sep 204 18 296 4

Oct 132 16 144 6

Nov 96 6 146 5

Dec 78 3 98 4

Average 127 10 180 6

1996

Jan 90 4 176 4

Feb 92 6 154 6

Mar 95 2 184 2

Apr 60 5 178 2

May 78 13 68 11

Jun 96 11 109 6

Jul 162 7 334 9

Aug 168 11 226 5

Sep 120 5 102 4

Oct 126 8 127 5

Nov 108 3 121 3

Dec 78 2 160 4

Average 106 6 162 5

*Untreated wastewater

**Final system (wetland) effluent
153

land plants. The larger treatment media have an effective

size of approximately 1.9 cm and a measured porosity of

0.39. The cells are lined with a 60-mil HDPE liner.

Each cell is divided in half longitudinally by an extension

of the liner to the top of the media. Three of the half cells

were planted with Phragmites, two of the half cells were

planted with Scirpus, and the final half cell was left

unplanted. This planting scheme allowed for performance

comparisons of planted versus unplanted cells and Scirpus

versus Phragmites. The system is depicted in Figure 8-6.

In addition to the multicell design and multiple operational modes, the VSB system at Minoa incorporated several other special features, including trilevel observation

well clusters within each half cell, specially designed inlet

weirs, thermistors beneath one of the cell liners and at

various levels within the cell, a dual-level effluent withdrawal, and an adjustable water-level control.

The specific goals of the research/technology transfer

efforts were the following:

1. Establish optimum hydraulic, organic, and solids

application rates necessary to achieve Village of

Minoa NPDES permit limitations.

2. Conduct testing to determine the impact of wet-event

peak-day hydraulic impacts on treatment performance. One of the project objectives was to evaluFigure 8-6. Schematic of Minoa, NY, VSB system

a) Period One - Continuous Flow in Parallel

b) Period Two - Fill and Drain Flow in Series

c) Period Three - Combination Sequence (fill and draw [1 and 2] to continuous 3)

1 2 3

1 2 3

 1

area

0.17 ha

area

0.175 ha

area

0.10 ha

P = Phragmites

S = Scirpus

N = No Plants 2 3

P S

P

S

P

N

P

S

PN PS

PS PN SP
154

ate the performance of the system under the maximum hydraulic design condition of 640,000 gpd.

3. Conduct tracer dispersion testing to measure actual

bed HRT and “in-place” hydraulic conductivities, to

evaluate impacts due to clogging, and to determine

the extent of short-circuiting.

4. Correlate ambient and wastewater temperature data

with observed removal efficiency for BOD, UOD (Ultimate Oxygen Demand), and ammonia nitrogen.

5. Evaluate the effect of plants (vs. no plants) and specific plants (Scirpus vs. Phragmites) on treatment

performance.

6. Evaluate the effect of vegetative harvesting on nutrient removal efficiency.

7. Provide data for calibrating an existing VSB heatloss model that predicts the substrate temperature

at various locations in the system.

8. Evaluate effects of series- versus parallel-flow configurations on treatment performance.

9. Conduct a detailed energy audit to establish the

energy benefits of this system in comparison with a

conventional treatment approach.

Construction costs for the system are summarized in

Table 8-7. It should be noted that (1) the work at Minoa

was completed under adverse weather conditions and a

tight construction schedule because of the consent order

requirements, and (2) costs reflect all of the special features incorporated in the system for research.

 8.2.1.2 Financial Arrangements

The costs of the Minoa wetland system associated with

the research aspects of the project were funded by the

U.S. EPA and the two State of New York agencies. The

remaining capital costs of the project were funded with a

state revolving-fund loan under the innovative and alternative system program.

8.2.1.3 Construction and Start-up Procedures

As noted previously, the work at Minoa was completed

under adverse weather conditions and a tight construction

schedule because of the consent order requirements. During the establishment of the wetland plants throughout most

of 1995, the wetland cells received only secondary effluent from the existing trickling filter.

8.2.1.4 Performance History

The performance of the Minoa VSB system in treating

primary effluent can be divided into three periods. During

the first period of January 1996 to March 1997, the system

was operated as a conventional VSB system, with the three

cells in parallel. From April 1997 to March 1998, the three

cells were operated in series and in a sequential fill-anddrain mode. From March 1998 to the writing of this manual,

the system has been operated in a different fill-and-drain

mode. Two cells, cells 1 and 2, operate in parallel but in

alternating fill-and-drain mode, similar to sequencing batch

reactors. The third cell, cell 3, operates in series-flow, but

with a constant water level, following the other two cells.

Conventional Parallel Operation

The BOD5 removal performance of the Minoa VSB system in the conventional mode was very poor when compared with the original design expectations. The three cells

were operated in parallel flow, but with different HRTs. The

performance of the Minoa system in BOD5 removal during

the first 10 months of conventional operation is summarized in several of the figures (identified as CU) in Chapter

5 and can be compared with two other systems. The false

performance expectations for the system were based on a

design equation developed with limited data, mostly from

VSB systems treating lagoon and pond effluents. The equation assumed that BOD5

 removal performance is dependent on temperature. Pollutant removal was not found to

vary significantly with temperature at Minoa.

The performance of the Minoa VSB system in TSS, TKN,

and total phosphorus removal during this period was similar to the performance of other VSB systems treating septic tank effluents (see Chapter 5 figures). TSS and BOD

removal were reasonably good, whereas TKN and total

phosphorus removal was quite poor.

Tracer study results from Minoa were also very similar

to tracer study results from other VSB systems. After one

year of operation, a significant fraction of the wastewater

flowed under the shallow root zone of the system. Also

observed were substantial dead volumes and typical

amounts of dispersion within the media.

Comparing the treatment performance of planted and

unplanted half cells, the Clarkson researchers found that

the unplanted half-cell performance was equal to the

planted cells for all pollutants measured. They also found

that the Phragmites cells removed more COD, TKN, and

total phosphorus than the cells planted with Scirpus.

Series-Flow, Sequential Fill-and-Drain Operation

The three cells of the Minoa VSB system were operated

in series-flow, sequential fill-and-drain operation for approximately 12 months. The operation during this time made

use of the dual effluent piping to achieve the fill-and-drain

Table 8-7. Village of Minoa VSB Construction Costs (Fall, 1994)

Sitework $135,500

60 Mil HDPE Liner 82,500

Wetland Media 104,500

Wetland Plants 29,000

Piping & Distribution 179,000

Miscellaneous 25,500

Total $568,000
155

operation, even though the flow through all three cells was

continuous. The water surface in a cell was controlled by

opening and closing the bottom drain line valve. When the

drain valve was closed, effluent from a cell flowed through

the upper effluent piping.

A typical cycle started with wastewater flowing through

a filled cell 1. The drain valve in cell 1 was opened

while drain valves in the other cells 2 and 3 were

closed. Twenty-four hours later, the drain valve in

cell 1 was closed and the drain valve in cell 2 was

opened. After 24 hours in this configuration, the drain

valve in cell 2 was closed and the drain valve in cell

3 was opened. It should be noted that this mode of

operation was possible at Minoa because of the significant drop in elevation from cell 1 to Cell 3. At a

flow rate of 130,000 gpd, the draining of cells 1 and

2 from their upper levels would take four to five hours,

while cell 3 required only three hours. In filling, cells

1 and 2 would require 24 hours, while cell 3 required

12 hours.

The performance in BOD5

 removal during the sequential fill-and-drain operation was significantly better than

during the previous period of conventional operation. Effluent BOD5

 averaged less than 15 mg/L while the system

was treating a much higher flow, and performance improved

during the latter months of the period. TSS removal was

also good, but TKN and total phosphorus removal did not

improve significantly. One of the most important improvements in the operation of the Minoa system during this

period was the reduction in the hydrogen sulfide odors that

had plagued the system during the period of conventional

operation.

Alternating Parallel Fill-and-Drain/Series-Flow Operation

Operation since March 1997 has had cells 1 and 2 operating in an alternating fill-and-drain mode followed by cell

3 operating in a constant-saturated mode. The pollutant

removal performance for BOD5

 and TSS has remained

quite good, and there has been a significant increase in

nitrogen removal performance.

8.2.1.5 Lessons Learned

• Fill-and-drain operation can significantly increase the

BOD5 and nitrogen removal performance of conventional VSB systems.

• BOD5 removal is not temperature dependent in conventional VSB systems.

• Because of the potential for severe odor problems, conventional VSB systems must be designed to have lower

organic loading rates when sited near households.

8.2.2 Mesquite, Nevada

8.2.2.1 Background

Mesquite, Nevada, is located on I-15 near the NevadaArizona border, about 112 miles east of Las Vegas. The

original treatment system for the community included

coarse screening and aerated facultative ponds followed

by storage ponds and land application on 62 acres of alfalfa fields. The State of Nevada required an effluent with

BOD at 30 mg/L and TSS at 90 mg/L prior to land application. The effluent at the Mesquite facility often exceeded

these limits, so an upgrade was required.

A 1989 facility plan for the upgraded facility recommended an increase in total treatment capacity to 1.2 mgd,

additional aerated lagoons with lagoon effluent to either

overland flow terraces or a VSB, and either of these followed by rapid infiltration basins. The VSB concept was

selected for this system because a free water surface

(FWS) wetland would have required a larger land area,

might not have been as effective for algae removal, and

would have been more susceptible to mosquito problems.

The design flow to the VSB was 400,000 gpd, with the

remainder routed from the lagoon to the overland flow

slopes. The existing facultative pond contained multiple

cells, and three of these were selected for conversion to

VSBs. The total VSB area was 4.7 acres.

The modified aerated lagoons were expected to produce

an effluent with about 70 mg/L, and the VSB wetlands were

designed to produce an effluent with 30 mg/L BOD5

 in the

coldest month, which was January. The design model used

for BOD5 removal is temperature dependent, so the system was sized to produce the target effluent value during

the coldest month. There were no NPDES discharge limits for the VSBs since they were designed to discharge to

rapid infiltration basins and not to a receiving stream.

A schematic plan is shown in Figure 8-7 for one of the

three similarly configured VSB units. Each of the three

parallel units contained four parallel cells as shown on the

figure. The flow path in each of the four cells averaged 50

ft, and the cell width averaged 380 ft. This configuration

produces an average aspect ratio (L:W) of 0.13:1. This

very low aspect ratio was selected following observation

of surface flooding and related problems with VSB systems in Louisiana, Mississippi, and Oklahoma that had

aspect ratios of 10:1 or more and no provision for the necessary hydraulic gradient to overcome the frictional resistance of a very long flow path. In addition to the short flow

path distance provided at Mesquite, a bottom slope of 1%

was provided for the cell bottoms.

8.2.2.2 Financial Arrangements

Funding for construction of this new system was provided by a combination of municipal bonds and the State

of Nevada’s revolving-loan fund. The total construction

costs for the VSB component at Mesquite was $515,000

(1990$), or $109,600 per acre, or $1,287 per 1000 gallons

of treatment capacity. The area cost is less than the

$178,000/acre (1990$) at the comparable VSB system in

Mandeville, Louisiana (see Mandeville case study), and

the difference is probably due to the higher cost of rock

and gravel in Louisiana. Land and liner costs for the Mesquite project were zero because existing lagoon cells were
156

converted to VSB units. The O&M costs are funded directly by a sewer charge for each connected user; a singlefamily connection would pay approximately $8.63 per

month for this service.

8.2.2.3 Construction and Start-up Procedures

Construction of the new system components was completed in late 1990 and start-up occurred in April 1991.

The original lagoon cells were lined with asphaltic concrete. These were prepared for the new VSB units by draining and drying, and then placement and compaction of local clay soil backfill to a depth of about 2 ft. This backfill

was then graded to provide the desired 1% slope for the

bottom. This 2-ft of compacted soil also ensured the impermeability of the bottoms. The effluent manifolds were

placed and leveled on the bottom prior to gravel bed construction. Gravel for the bed was transported from the local pit, dumped in the wetland cell, and spread with a small

bulldozer. Trenches for the coarse inlet zone rock were

excavated and backfilled after placement of the entire 32-

in-deep gravel layer. The 2-in layer of fine gravel/coarse

sand was then placed on the surface of the bed, with the

exception of the inlet and outlet zones. Posts were then

driven into the gravel layer for support of the distribution

manifold pipes. Construction of external piping, outlet structures, and pumping stations then completed the work.

Flow distribution to the three VSB units utilized orifice

plates to split the flow, and 8-in perforated pipe manifolds

were used in each cell for both distribution and effluent

collection, as shown in Figure 8-7. The influent pipes were

elevated slightly above the bed surface, and the effluent

manifolds were at the bottom of the bed. The main VSB

bed consisted of a 32-in depth of washed river gravel ranging in size from 0.4 in to 1.0 in obtained at a local gravel

pit. An inlet zone underneath each inlet pipe contains 2-in

to 4-in rock to ensure rapid infiltration and distribution. This

zone is about 3 ft wide at the top and extends the full

depth of the bed. The gravel in the main bed was then

covered with about a 2-in layer of fine gravel/coarse sand

mixture to aid in the germination and growth of the vegetation.

There were no soils or geotechnical investigations at

this site since existing lined lagoon cells were to be used

for the new VSBs. The only geotechnical activity involved

with this project was to find a suitable source for the rock

and gravel required. The layer of fine gravel/coarse sand

was chosen because the intended method of planting was

hydroseeding. A layer of fine gravel/coarse sand mixture

was placed on top of the gravel in the VSB cells to serve

as a growth substrate for the intended hydroseeding. The

first bed was hydroseeded in July 1991 at a rate of 25 lb/

acre of seed mixed with 2500 lb/acre of mulch fiber. Sprinklers were then used to periodically flood the surface of

the bed to encourage germination and growth. By September 1991, only 20% germination could be observed.

Alkali bulrush (Scirpus robustus) was selected as the sole

vegetation type for all of the VSB cells. Again hydroseeding

was attempted but proved not to be successful. Planting

by hand with locally available plant materials (from ditches,

etc.) was successfully completed during the second year

of system operation. In 1997, the VSB cells were completely covered with healthy vegetation. There is no harFigure 8-7. Schematic diagram of typical VSB (one of three) at Mesquite, NV

1%

In Out

1%

1%

1%

50 ft

50 ft

50 ft

50 ft

200 ft

Outlet Manifolds Inlet Manifolds

380 ft
157

vesting or other vegetation- management procedures at

this site.

Water-level control in two of the three VSB units is provided by overflow weirs in the outlet structures. In the third

unit, water-level control depends on float-switch settings

for the discharge pump. In addition, the piping and distribution and collection system were designed to operate with

a continuous 0.4 mgd recycle flow (100% of forward flow).

As of August 1997, an additional plant expansion was

underway at the Mesquite system. The city is growing rapidly as a retirement/recreational community and a number

of golf courses are under construction or planned. To provide irrigation water for these golf courses, the wastewater plant expansion is including an oxidation ditch with nitrification/denitrification capability and UV disinfection to

meet the necessary bacterial limits for golf course irrigation. The VSB/overland flow/rapid-infiltration units at Mesquite will remain in stand-by use and will be operational

during high-flow winter months.

8.2.2.4 Performance History

The inlet orifice plates were provided to split the influent

flow proportionally to the surface area of each VSB unit

since there were slight variations in the size of the three

units. Table 8-8 presents average VSB performance data

during the period June 1992 through May 1993. Data are

not available on the performance of individual units or cells

within a unit.

Although the system met its effluent BOD target on an

annual average basis, there were monthly variations, as

shown in Table 8-9. However, these excursions had minimal impact on the rapid-infiltration system.

These 1992–1993 performance results were achieved

without recycling VSB effluent. However, at the time of the

1997 site visit, recycle at 400,000 gpd was practiced continuously and produced essentially the same performance

results shown in the tables. Recycle was only considered

to be essential in the very hot and dry summer months in

order to keep the plants on the beds alive and functional.

In the general case, algae forms in the lagoon and is

separated in the VSB, and the decomposition of the algae

releases additional ammonia and organics. As a result,

the effluent ammonia and organics are elevated due to

internal loading during the warmest periods. Removal during the warmer months of the year is believed to be offset

by plant uptake during the growing season. Subsequent

data would be very useful to help identify the annual cycle

over several years.

8.2.2.5 Lessons Learned

• The wetland configuration and cross section shown in

Figure 8-7 were designed to maximize the available

area in the former lagoon cell, while at the same time

minimizing the aspect ratio.

• Surface overflows are due to improper hydraulic design rather than clogging.

• Subdividing each VSB into four separate cells with the

right slope in each cell to ensure proper flows required

very careful grading of subgrade soils that significantly

increased the cost and complexity of construction.

• Subdividing each unit into two cells by applying influent along the centerline and collecting effluent along

the two sides would have produced an aspect ratio of

0.26:1.

• Converting each former lagoon cell to a single wetland bed, with application along one long side and effluent collection along the opposite side, would have

produced an aspect ratio of 0.5:1, with a level subgrade

and the water level and hydraulic gradient controlled

by an adjustable outlet.

• Continuously flooding the bed with a few inches of water after hydroseeding, rather than intermittently wetting it, may have improved germination, as would planting in a more moderate season in the desert climate.

• Hand planting of shoots or rhizomes in the gravel of a

VSB system is preferred. Potted shoots and rhizome

material for a wide variety of plant species are commercially available.

• An effluent recycle feature permitting 100% recycle is

not typical at most VSB systems and was not necessary for water quality purposes.

Table 8-8. Summary Performance, Mesquite, Nevada, VSB Component, June 1992-May 1993

Parameter Influent, mg/L Effluent, mg/L % Removal

BOD5 64 29 55

TSS 57 13 77

NH4-N 16 10 38

TKN 29 16 46

TP 7. 4 6.2 16

Table 8-9. Effluent Characteristics, Mesquite, NV, VSB Component,

June 1992-May 1993

Month Temp. BOD TSS NH4-N TKN TP

°C mg/L mg/L mg/L mg/L mg/L

1992

Jun 21.6 32 6 3.3 6.7 5.0

Jul 26.7 24 6 4.3 6.4 5.3

Aug 27.1 26 6 4.5 7.6 4.8

Sep 23.6 22 5 4.1 6.8 5.5

Oct 19.1 37 5 3.3 5.6 6.1

Nov 13.5 32 22 5.3 8.6 5.8

Dec 7.5 27 16 15.7 22.3 4.7

1993

Jan 8.1 24 14 19.8 29.7 6.1

Feb 12.7 24 18 21.9 29.9 8.0

Mar 13.9 23 16 22.1 29.9 9.2

Apr 16.2 49 17 12.4 23.6 7.1

May 20.3 27 21 6.0 9.5 7.0
158

• Routine maintenance requirements at this system are

minimal and consist of periodic pump inspections and

monthly cleaning of orifices in the influent distribution

manifolds.

8.2.3 Mandeville, Louisiana

8.2.3.1 Background

Mandeville, Louisiana, is located on the northern shore

of Lake Pontchartrain at the end of the causeway bridge

from New Orleans. The 1997 population of Mandeville was

about 10,000, and the suburban residential community was

expanding rapidly. A vegetative submerged bed (VSB) was

selected as a component in the new wastewater treatment

facilities at the recommendation of the State of Louisiana

and the U.S. EPA Region VI. The system was constructed

during 1989 and placed in operation in February 1990, with

a design flow of 1.5 mgd. The system discharges to Bayou

Chinchuba, which drains to Lake Pontchartrain. The

NPDES limits are BOD5 10 mg/L, TSS 15 mg/L, NH3/NH4

5 mg/L, fecal coliform 200/100 mL, and a maximum pH of

9.

The new system was constructed at the site of the

community’s original three-cell facultative lagoon, and one

of the original cells was retained for temporary treatment

and later abandoned at the completion of the new system.

A second original cell was deepened and converted to a

partial-mix aerated lagoon with three cells operated in series and submerged perforated tubing in the first two cells.

The hydraulic residence time (HRT) in this new lagoon was

about 15 days at design flow. The third original lagoon cell

was converted to a three-cell VSB gravel bed. The VSB

cells operate in parallel. Other new elements in the system included a headworks containing a bar screen and

grit chamber, final disinfection with UV, and an effluent

pumping station. All of these major system components

are shown in Figure 8-8.

At the time this system was designed, sizing criteria were

5 acres per mgd of design flow, a one- to two- day HRT in

the VSB, and an aspect ratio (L:W) of at least 10:1 to ensure plug flow conditions. These criteria assumed that the

VSB influent would contain about 30 mg/L of BOD5

 and

TSS following treatment in the aerated lagoon. All of these

criteria were applied at Mandeville except the 10:1 aspect

ratio, which could not be used due to the preexisting configuration of the facultative lagoon cell. The average aspect ratio of the three VSBs is about 2.5:1.

The three VSBs are separated by low internal earthen

berms that provide about 1.5 ft of freeboard above the

gravel surface in the bed. The external berms are the preexisting dikes of the former facultative lagoon. The bottom

surface area is 6 acres. The VSB bed is composed of a

1.5-ft depth of crushed limestone rock (2- to 4-in size) overlain by 6 in of granite gravel (0.5- to 1-in size). The surface

layer of gravel was considered necessary as a rooting

Cell 3: 188' X 384'

UV Disinfection

and Pump

Cell 2: 171' X 471'

Cell 1: 180' X 471'

Effluent Manifold

Floating Aerator

Submerged Aerator

Baffle Curtains

Lagoon Cells

140' X 540' X 10' Deep

VSB Cells

All 2' Deep

Figure 8-8. Schematic of VSB system at Mandeville, LA
159

medium for the vegetation. Softstem bulrush (Scirpus

validus) was selected as the sole vegetation type, and nursery-grown shoots were planted on about 4-ft centers. An

annual harvest of these plants was recommended by the

designers and was practiced for several years after startup.

A 20-in PVC pipe conveys lagoon effluent to the VSB.

This pipe connects to a PVC manifold extending the full

width of the three cells. At three equidistant points in each

cell, the manifold discharges to a 10-in outlet pipe that is

valved and extends 25 ft into the bed. These outlet pipes

are at the surface of the bed, and they each end in a 90°

“down” elbow that penetrates into the rock layer. These

nine gate valves were intended for flow control so that a

cell could be taken out of service and/or flow could be adjusted as required to produce a relatively uniform distribution of flow.

The effluent manifold for cells 1 and 3 is 21-in PVC and

24-in PVC for cell 2. These manifold pipes were buried

with the top of the pipe flush with the top of the coarse rock

layer. Four-in-diameter holes were drilled on 8-in centers

at the top center of these manifold pipes. These manifolds

connect to the UV disinfection chamber, which then discharges to the sump of the discharge pump. The top of the

gravel layer was graded level, as was the bottom of the

bed, and no adjustment was possible in the water level in

the bed, nor was it possible to drain the cells.

A special feature in all three wetland cells is the inclusion of buried 6-in perforated PVC pipes. Two of these

open-ended pipes are buried in each cell, about 6 in above

the bed bottom in the coarse rock media. Their apparent

purpose is redistribution of subsurface flow in case the entry

zone of the bed becomes clogged with solids. Each pipe

is 100 ft in length and is laid parallel to the flow direction;

the two pipes in each bed are located about 35 ft on each

side of the longitudinal bed centerline.

The construction costs for the entire system, including

the aerated lagoons, was about $3,000,000 (1990$). The

cost for the VSB cells was about $590,000 (1990$), with

about 70% of that for procurement and placement of the

rock media and gravel layer. The materials used at

Mandeville were barged from Arkansas, since rock and

gravel are not readily available in this part of Louisiana.

Other VSB projects in the vicinity have used rock and gravel

barged from Mexico. The VSBs are not lined since the

subsoils are clay and sandy clay. Since the exterior dikes

for the former lagoon were utilized, construction costs were

minimal (except for the cost of rock and gravel). Land costs

were zero since the preexisting lagoon was municipally

owned. The construction costs for this VSB were about

$590 (1990$) per 1000 gallons of design flow, or $105,400

per acre of treatment area for the 5.6-acre system.

8.2.3.2 Financial Arrangements

The construction costs for the Mandeville system were

funded privately through bonds issued by the City of

Mandeville. No grant or funding support was provided by

the State of Louisiana or the U.S. EPA. The apparent reason was the relatively low position of the city on the grant

priority list. The city, faced with the choice of curtailing community growth or funding a new system itself, chose the

latter option. The O&M costs for the system have been

obtained as a surcharge on the consumer’s water bill.

8.2.3.3 Construction and Start-up Procedures

Construction activities commenced with draining of the

existing facultative lagoon. The bottom was allowed to dry,

and then accumulated sludge was removed and disposed

of. The bottom was then leveled in preparation for backfilling with gravel. The concrete structures containing the UV

disinfection components and the effluent pump station were

also constructed at this time. The low interior earthen berms

were then constructed to divide the lagoon cell into three

parallel units. These interior berms permit foot traffic only.

The rock and gravel were hauled by truck from the barge

dock on Lake Pontchartrain to the site, dumped into the

bed, and spread with small bulldozers. The entire coarse

rock layer was placed and leveled before any gravel was

placed as the top layer. The inlet and outlet manifolds were

then installed and connected and rock backfilled around

them (the top gravel layer was not placed in these inlet

and outlet zones). The bed was then filled with water (with

effluent from the temporary lagoon) to the top of the coarse

rock. The bulrush shoots, obtained from a nursery in Mississippi, were planted by hand on 4-ft centers, with their

roots in contact with the water at the top of the coarse

gravel. About 15,000 plants were planted in the three cells.

Start-up of this system commenced immediately upon

completion of construction. In some systems of this type,

clean water is used to initially fill the bed, and the plant

shoots are allowed to grow for four to six weeks prior to

introduction of wastewater. In this case, lagoon effluent

was introduced during the planting stage, and daily flow

through the VSB commenced as soon as the aerated lagoons were operational. There were no special start-up

procedures used at this site. However, a unique maintenance procedure was adopted for several years, which

started with harvesting of weeds to encourage growth and

spread of the bulrush, which evolved into a complete annual harvest of all vegetation and the removal and disposal of the harvested material. That practice has now been

terminated.

8.2.3.4 Performance History

In 1991, the Mandeville system was selected by the U.S.

EPA for a detailed eight-week performance evaluation. This

effort included independent flow metering of system influent, VSB influent and effluent, tracer studies to verify HRT,

and weekly composite sampling and testing for BOD (total

and soluble), COD (total and soluble), TSS, VSS, TKN,

NH4

-N, NO3, TP, DO, pH, and temperature. The study period commenced in mid-June 1991 and was completed by

late September 1991. The average flow rate during this

period was 1.16 mgd, indicating that 77% of the system

design capacity was achieved in the second year of op-
160

eration. This is a reflection of the very rapid growth and

residential construction in the community. A summary of

the water quality performance data is given in Table 8-10.

The tracer study, conducted only in cells 2 and 3 because

the valves for cell 1 had been inadvertently closed, measured a flow rate of 1.352 mgd, which indicated an actual

HRT of 17.8 hours. This compared favorably with the theoretical HRT of 18 hours for the same flow rate, assuming a

porosity of 42% in the rock/gravel bed. At the time of the

tracer test, surface water was apparent on portions of the

wetland cells, but the majority of the flow was subsurface.

If cell 1 had been operational during the test, it is believed

that the actual HRT would have been close to the one-day

theoretical HRT for the full system.

Table 8-11 presents a summary of system performance

data collected in 1996 and 1997. The values shown are

the averages for the month shown.

As shown in Table 8-11, the current actual flow exceeds

the original design rate of 1.5 mgd, but the system continues to meet the discharge limits for BOD and TSS but exceeds the ammonia limit on a seasonal basis (i.e., noncompliance in the colder months). The routine compliance

with BOD5

 and TSS limits is in part due to the reliability of

these systems for removal of these parameters, but is also

in part due to significant modifications to the system made

in 1992. The present system configuration, with the surface aerators, the subsurface aerators, and the baffle curtains as shown in Figure 8-8, has been in place since 1992.

The lagoons as originally constructed had submerged,

partial-mix aeration tubing in the first and second aeration

cells, and there were no baffles in place. In effect, the first

baffle in the first lagoon cell converts the entry zone into a

complete-mix aeration component. The purpose of these

modifications was to obtain a more rapid removal of BOD5

and more effective settling of TSS in the lagoons, and to

subsequently permit more effective ammonia removal in

the lagoons and VSB component. This strategy has been

successful for BOD5

 and TSS, but not for ammonia. The

low ammonia values obtained in the EPA study during the

1991 summer are misleading. The records for the entire

year show a seasonal trend in effluent concentrations that

are similar to those shown in Table 8-11 for 1996–1997. In

1991, the effluent ammonia concentration averaged 3.2

mg/L during the warm months (March–November) and 7.8

mg/L during the colder months (December–February). The

system met the ammonia limit by a significant margin during that first year of operation.

There are no significant seasonal trends in the ammonia concentration in the untreated wastewater, but there

are in the lagoon effluent, which indicates that these higher

winter values are not treated effectively by the VSB. As a

result, the system effluent exceeds the discharge limit. This

condition suggests that the lagoon, as presently configured, does not provide effective nitrification during the

colder weather. That is a plausible hypothesis since the

nitrifier organisms are temperature sensitive and generally exist in relatively low numbers in these partial-mix aerated lagoons with no sludge return.

The city intends to increase the capacity of the system

to about 4 mgd to keep up with expected growth in the

community. Discussions are underway regarding the future system configuration to solve both the ammonia problem and permit capacity expansion with maximum utilization of the existing facilities.

8.2.3.5 Lessons Learned

• The internal hydraulics of the wetland cells force all of

the influent to enter the cell at three points, with a total

cross-sectional area of about 2 ft2, which is inadequate

to receive a design flow of about 350 gpm and results

in surface flow in the inlet zone. A perforated inlet

manifold that extended the full width of each cell should

have prevented surface flow. At the effluent end of the

cells, surface flow was caused by outlet ports installed

at the same elevation as the rock surface.

• A means of controlling water levels in the bed and allowing the bed to be drained for maintenance would

have improved the system.

• Modifications to the system, including additional orifices drilled in the effluent manifold in the side and lower

quadrant, additional surface gravel placed in the area

of the manifold, and a new pipe installed to permit drainage of the cells, resulted in a lowering of the water

level in the effluent zone of the wetland bed, so the

gravel surface in that area is generally dry.

• Surface flow was experienced almost immediately in

the inlet and outlet zones of this system and was not

caused by clogging, as confirmed by EPA investigations in 1991, but rather by lack of hydraulic gradient.

• Hydraulic gradient for a flat-bottomed system can be

provided with a water-level control device at the effluent end of the cell.

Table 8-10. Water Quality Performance, Mandeville LA Treatment

System, June/September 1991

 System Wetland Wetland

Parameter Influent Influent Effluent

BOD (Total) mg/L 154 41 10

BOD (Soluble) mg/L ND1 21 8

COD (Total) mg/L 349 79 43

COD (Soluble) mg/L ND 40 31

TSS mg/L 132 59 7

VSS mg/L ND 39 5

TKN mg/L ND 5 3

NH4-N mg/L ND 1.4 2.1

Organic N mg/L ND 3.1 1.1

NO3-N mg/L ND 4 0.2

TN mg/L ND 9 3

TP mg/L ND 3 4

Fecal Coliforms#/100ml ND TNTC TNTC2

DO mg/L ND 2.4 1.8

pH 6.9 6.9 7.0

Temperature °C ND 31.8 30.5

1

ND = No data available.

2

TNTC = Too numerous to count, sample taken prior to disinfection
161

• Bulrush planted in this system has attracted nutria and

muskrat, which favor bulrush for food and nesting

material. Nutria have eaten most of the bulrush plants

and bored through the interior berms, which causes

significant leakage between the cells. Small sacks filled

with a mixture of cement and sand have corrected the

leakage problem of nutria boring through interior

berms.

• The minimal availability of oxygen in VSB wetland beds

makes them ineffective for nitrification of ammonia, and

the Mandeville system can meet ammonia discharge

limits only when the aerated lagoon provides sufficient

ammonia removal.

8.2.4 Sorrento, Louisiana

8.2.4.1 Background

Sorrento, a small residential community in southeastern

Louisiana, is located about 50 miles southeast of Baton

Rouge. Prior to construction of the aerated lagoon wetland system, the community was served by on-site septic

tank systems. Many of these on-site systems were not functioning properly due to the difficult soil conditions in the

area. The new system was designed in 1990 and placed

in operation in late 1991. The lagoon component consists

of two 10-ft-deep aerated cells (first cell contains four 3-hp

floating aerators, second cell contains four 2-hp floating

aerators), followed by a 7-ft-deep settling pond. The two

aerated cells were designed for 10 d HRT at the potential

ultimate flow rate of 130,000 gpd. At the 1997 flow rate of

32,000 gpd, the HRT is about 40 d, and only a few of the

aerators were operated.

The VSB cell, with a bottom area of about 7800 ft2

, was

designed for a flow rate of 50,000 gpd with the intention of

adding a second parallel cell as the flow rate increases in

the future. The design HRT in this bed at 50,000 gpd would

be one day. The native soils are clays and silty clays, so

the bottoms of the lagoon cells and the VSB cells are not

lined. However, a geotextile liner is used on the inner slope

of all berms to prevent erosion and weed growth; the outer

slope of these berms is grassed. The system discharges

to Bayou Conway, and the NPDES discharge limits are

BOD5

 20 mg/L, TSS 20 mg/L, pH 6–9, and fecal coliforms

200–400/100 mL. There are no ammonia limits for this system.

As shown in Figure 8-9, the VSB cell is triangular in

shape, with the inlet zone about 60 ft wide and the flow

path to the outlet about 250 ft long. This shape was selected to minimize short-circuiting of flow. Previous designs

had large aspect ratios (10:1 or greater) but insufficient

hydraulic gradient to overcome the frictional resistance,

resulting in surface flow on top of the bed. At Sorrento, the

average aspect ratio is only 6:1, but all of the flow converges at the end of the triangular bed.

Table 8-11. Water Quality Performance, Mandeville, LA, Treatment System, 1996 -1997

Raw Wastewater Wetland Influent Wetland Effluent

 Avg. Flow BOD TSS NH4-N BOD TSS NH4-N BOD TSS NH4-N

Date mgd mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L

1996

Jan 1.57 133 115 14 15 8 14.3 5 2 11.9

Feb 1.60 156 120 15 14 8 14 6 2 12

Mar 1.75 126 116 13 16 11 16 2 2 11

Apr 1.26 145 148 14 11 8 12 4 1 8

May 1.11 137 133 18 45 25 10 3 2 9

Jun 1.72 138 132 18 27 18 1 2 2 0.5

Jul 1.33 131 115 16 24 24 1 2 1 0.7

Aug 1.71 64 70 20 30 48 7 10 7 5

Sep 1.32 61 155 31 59 20 12 9 7 5

Oct 1.69 86 137 21 43 18 14 7 4 5

Nov 1.51 119 126 44 62 8 8 8 5 4

Dec 1.57 116 122 37 43 11 28 6 2 19

Avg 1.51 118 124 22 32 17 11 5 3 8

1997

Jan 1.85 89 111 20 70 14 12 7 4 7

Feb 2.31 63 94 24 62 23 44 10 6 20

Mar 1.53 94 128 26 59 16 16 7 3 5

Apr 1.44 97 98 22 52 20 10 6 4 4

May 1.67 112 140 25 34 14 6 4 3 4

Jun 1.64 93 167 28 57 10 12 4 3 4

Jul 1.68 108 109 16 51 9 9 6 2 5

Avg 1.73 94 121 23 55 15 16 6 4 7

Weighted 1.59 109 123 22 40 16 13 5 3 8

Average

1996/97

1991 1.16 154 132 ND 41 59 1.4 10 7 2.1
162

The design engineer of the Sorrento system also incorporated several features to ensure that an adequate hydraulic gradient would always be available based on lessons learned from other systems. The bottom of the bed is

flat and level throughout its length, but the gravel depth is

3 ft at the inlet and 2.5 ft at the outlet, so the top surface of

the gravel slopes to provide for 0.5 ft of headloss. In addition, the single outlet structure contains an adjustable sluice

gate that allows a further increase in the available hydraulic gradient by adjusting the water level in the bed. The

inlet to the bed is an 8-in perforated pipe resting on the

bottom of the bed and extending the full width. The bed

effluent discharges to a concrete outlet box. There is also

an 8-in valved drain pipe at the outlet end of the bed to

drain the cell completely, if necessary. A chlorine contact

chamber is provided for disinfection prior to final discharge.

Two layers of aggregate are used in the Sorrento VSB.

The top layer is a 6-in depth of washed, 0.75-in gravel.

The main part of the bed is composed of crushed limestone imported from Mexico, ranging from 1.5 to 3 in in

size. Since ammonia removal was not required and because maintenance problems with vegetation were apparent at other systems, it was decided not to plant vegetation on the Sorrento VSB cell. At the time of the 1997 inspection for this report, weeds were growing around the

fringes of the wetland bed, but the general bed surface

was still free of vegetation.

This wetland system was selected for use at Sorrento

because the facility planning evaluation showed it to be

the most cost-effective process for meeting the NPDES

discharge requirements. The total construction cost for the

entire system was about $233,400 (1991$), with an estimated $75,000 for the VSB component. The unit construction cost for the VSB would then be about $1500 per 1000

gallons of design capacity (for the 50,000 gpd design flow).

On an area basis, the capital costs would be about

$419,000 per acre for the 0.18-acre VSB.

8.2.4.2 Financial Arrangements

The construction costs for the Sorrento system were

funded with federal and state money provided under the

U.S. EPA Construction Grant Program that existed at that

time. The O&M costs for the system are supported by sewer

fees from the connected users.

8.2.4.3 Construction and Start-up Procedures

A site for the new lagoon/wetland system was identified

on available land between the community and the final discharge point to Bayou Conway. Geotechnical investigations were undertaken to identify and characterize the in

situ soils. These proved to be clays and silty clays that

would provide adequate protection for ground water. There

also was no identified risk of ground water intrusion or surface water flooding at this site.

The site is relatively level, so the entire system was excavated, with excess material used to construct the berms.

A 3-ft freeboard was provided for the lagoon cells and the

VSB component. The rock and gravel were hauled by truck

from a barge dock on the Mississippi River, dumped into

the bed, and spread with a small bulldozer. The entire

coarse rock layer was placed and leveled before any gravel

was placed as the top layer. Inlet and outlet manifolds were

then installed and connected and rock backfilled around

Figure 8-9. Schematic of VSB system at Sorrento, LA

Aerators

Influent

Aerated Lagoons

Settling Pond

Vegetated Bed

60 X 250 ft

Discharge Chlorination
163

them. The bed was then filled with effluent from the lagoon

to the top of the coarse rock. Since vegetation was not

used on this system, start-up was immediate.

Routine maintenance procedures include servicing of the

lagoon aerators and the chlorine disinfection equipment;

there are no routine maintenance requirements for the

wetland bed. There have been problems with nutria burrowing in the banks of the lagoon cells; however, since

there is no vegetation and no exposed water in the wetland cell, these animals have not been a problem at this

site. Since maintenance has not been required for the

wetland cell, the O&M cost for this component is zero.

8.2.4.4 Performance History

Water quality data are not available for untreated sewage at Sorrento or for lagoon effluent entering the wetland

system. The 1997 flow is estimated to be in the range of

15,000 gpd. At that rate the HRT would be about 100 d in

the lagoons and 4 d in the VSB component. With such a

long HRT, lagoon effluent could be expected to have a

BOD5 of less than 20 mg/L and a TSS in the same range

(except for algal bloom periods). With inputs at this level,

the VSB with an HRT of 4 d could be expected to produce

background levels of BOD5 and TSS, as confirmed in Table

8-12.

8.2.4.5 Lessons Learned

• The triangular configuration of this VSB system causes

flow lines to converge at the end of the system in a

single outlet point, which is cost effective, but any such

design would need to evaluate weir loading rates per

unit length to avoid excessive velocity in the outlet zone

that could cause resuspension of TSS and its associated contaminants.

• The sloping surface of the gravel (0.2% grade) provides an additional 0.5 ft of potential head at the inlet

to help ensure that the hydraulic gradient is sufficient

to avoid surface flow on the bed.

• The adjustable outlet gate provides additional water

level adjustment; however, the outlet gate cannot be

lowered completely, so an additional drain pipe for dewatering the bed is necessary. A completely adjustable outlet may have eliminated both the additional

gravel necessary to produce the sloping surface and

the drain pipe for dewatering.

• The system is oversized for the current flow rate and

organic loading, so an additional VSB cell may not be

necessary for the system to handle flow rate increases

anticipated in the future.

• The lack of plants in this system does not appear to

affect removal of BOD and TSS, as was observed in a

1992 EPA performance evaluation of a vegetated system and a temporarily nonvegetated system.

• The lack of plants in effect equates the VSB concept

to a horizontal-flow, coarse-media, contact filter.

8.3 Lessons Learned

8.3.1 Design

Organic Loading

Organic loadings in the range of 10 to 25 lbs BOD/acre/

day to FWS systems have been shown to effectively meet

30 mg/L BOD and TSS monthly effluent standards, with

no need after 15 years of operation to remove the settled

material from a FWS system. For the majority of these

systems, this range of organic loading results in six to eight

days of theoretical hydraulic retention.

Data Gaps

The database for both VSB and FWS systems has a

continuing problem of not having enough quality-assured

data to evaluate removal rates, seasonal differences, water balances, and long-term treatment effectiveness. Insufficient data have been collected on contaminant loadings (both flow and concentration), incremental data

through the system (multiple data collection points), and

water column data (temperature, pH, and dissolved oxygen) at different locations.

Inlet/Outlet Works

Studies comparing the placement of inlet/outlet (I/O)

works as a function of cell geometry and outlet approach

conditions are lacking. As a result, there is not a complete

rational approach to the placement and design of inlet and

outlet works for these types of systems. For example, the

criterion of weir overflow rate typically has been used to

place outlet weirs and specify weir length in these low approach-velocity systems. Included in the outlet design are

bathymetric and vegetative conditions of the outlet zone

of FWS wetlands. Large collection areas immediately upstream from outlet works that have no emergent vegetaTable 8-12. VSB Effluent Water Quality, Sorrento, LA

Date BOD, mg/L TSS, mg/L Fecal Coli1 pH

2/23/94 <6 5 0 7.2

3/31/94 <6 4 0 7.2

5/18/94 <6 2 20 7.5

7/28/94 <6 4 100 7.4

8/10/94 <6 4 0 7.2

9/26/94 <6 4 6 7.8

12/30/94 <6 4 180 7.6

1/12/95 <6 4 TNTC 7.9

3/8/95 <6 4 4200 7.9

4/28/95 <6 4 666 7.1

6/16/95 <6 11 0 7.8

7/12/95 <6 4 0 7.5

8/10/95 <6 4 0 7.3

9/13/95 <6 4 0 7.5

10/11/95 <6 4 0 7.4

11/8/95 <6 4 0 7.5

12/13/95 <6 4 53 7.4

1/24/96 <6 4 350 7.4

2/14/96 <6 4 3 7.5

3/13/96 <6 4 6 7.0

1

Note: after chlorine disinfection, #/100 ml
164

tion have resulted in poor effluent quality. Relatively shallow collection zones with emergent vegetation have shown

less variability of effluent quality, but more O/M requirements.

Headloss

While headloss is a factor of concern in VSB systems, it

is not a major factor in FWS wetlands unless they have

extremely long and narrow flow reaches. Headloss is a

consideration in FWS wetlands only when L:W ratios are

great (10:1 or greater) and are combined with high hydraulic loading rates in heavily vegetated cells. Proper

placement of I/O works, sufficient berm height, and L:W

ratios less than 10:1 minimize headloss effects in FWS

wetlands. Clean-water headloss through a VSB wetland

system is quickly modified as pore spaces are filled with

separated solids and, to a lesser degree, rhizosphere development. Under conditions of plugging by these mechanisms, the liquid level will eventually surface and an undersized, fully vegetated FWS wetland condition will begin to develop on the surface.

Hydraulics

The internal hydraulics of a VSB system are critical to

treatment success. Systems constructed with high L:W

ratios, flat bottoms, and with effluent manifold ports located at the top of the gravel produced surface flow at the

effluent ends of the cell. These types of systems also did

not allow for water level control within the bed and cell

drainage. Multiple-ported influent manifolds extending the

width of the inlet zone, coupled with similar effluent collection manifolds with adjustable weirs or rotating elbows,

would allow for greater hydraulic contact in the basin and

more operational flexibility. Sloping the surface of the

gravel bed based on the design flow headloss may also

permit increases in VSB hydraulic loading and duration of

service prior to major inlet maintenance.

Aspect Ratio

High aspect ratio FWS wetland cells can produce significant operational problems at high hydraulic loading

rates and/or with dense stands of emergent vegetation.

Headloss effects are additive and are greatly aggravated

with high length-to-width ratios in totally vegetated FWS

wetlands. Both types of systems with L:W ratios as high

as 30:1 have produced significant flooding at the influent

berms while dropping the effluent water elevation below

weir recovery levels. Parallel cells with lower L:W with

multiple I/O works can control this effect. The only design

limitation is that the HRT must be above some minimum

to assure removal of TSS and associated pollutants of

concern.

Ice Formation

In colder climates where ice forms on standing water,

sufficient freeboard and outlet control is essential to allow

for ice formation to cap the normal operating depth of the

free surface water column. In most cases, this distance is

less than 1 ft. The ability to operate the wetland with the

water column directly in contact with the ice, with no lowering of the water level once the ice forms, is another important design feature. Lowering the water under these

conditions could allow for a secondary ice level, with a

liquid level constraint, to form under the primary level.

8.3.2 Mechanisms and Processes

Oxygen Transfer

Oxygen transfer through the rhizosphere evidently is not

a major contributor to contaminant oxidation in vegetated

submerged bed systems, based on both research and fullscale studies. Oxygen demand associated with storage

products in roots and tubers is much greater than excess

oxygen available at the root hairs and other plant parts. In

FWS wetlands, epiphytes can colonize on stems and

leaves preferentially depending on oxygen exuding from

the gas transport plant structures.

Nitrification-Denitrification in VSB Systems

Cost-effectively sized VSB systems have not been shown

to significantly nitrify treated influent. It follows that VSBs

have not been shown to be able to denitrify an influent,

which is predominantly ammonia. Early studies that suggested that significant amounts of nitrogen can be removed

in a VSB wetland system have not been duplicated in subsequent studies and full-scale evaluations.

Plant Coverage

Coverage by emergent plants in FWS wetlands should

not be 100% because too much coverage by emergent

plants is negatively correlated with effluent quality. Placing open water (submergent aquatic macrophytes) between

areas of closed water (emergent macrophytes) is correlated with better effluent quality than a wetland that has

100% emergence. Submergent plants release oxygen to

the water column, and these open water zones allow for

more surface reaeration. Emergent plants also contribute

more internal BOD loading upon decomposition.

Ground Water Recharge

Siting and designing FWS wetlands with intentional discharge to ground water is a legitimate application of this

treatment technology. With proper design, FWS wetlands

for ground water recharge can remove nitrate nitrogen

(through denitrification) and indicator organisms. Data collected at “leaky wetland sites,” such as in Jackson Bottom, Oregon, have demonstrated the effectiveness of these

processes in locations where soils are sufficiently porous.

Plant Litter

Plant litter is an essential component of a FWS constructed wetland. While this material contributes to flow

resistance, it more importantly seals surface areas in fully

vegetated zones to assure anoxic conditions.
165

8.3.3 Vegetation

Development of Treatment Effectiveness

The full treatment potential of a FWS system may not be

realized until there is both full coverage of plants (as designed) and a layer of litter beginning to accumulate on

the surface of the water. Depending on the type of plants,

planting density, and time of planting, a minimum of two

growing seasons, or two to three years, may be needed.

This may be important when negotiating discharge permit

requirements.

Ammonia Nitrogen

In VSB systems, plants are not critical to the removal of

BOD and TSS, as shown in several systems. Evidence

suggests, however, that they are effective in removing

portions of the nitrogen and phosphorus due to uptake by

the plants during the growing season. Most of these nutrients are returned to the water column during the senescent period.

Aeration

Attempts to aerate outlet zones of FWS systems with

submerged tubing have resulted in attracting animals such

as muskrats and nutria, which may damage the tubing.

Short-circuiting

Vegetation predation by nutria and muskrats in FWS

wetlands can produce serious hydraulic short-circuiting.

Varying plant resistance as the wastewater moves through

the wetland also can cause short-circuiting. Preferential

flow routes can develop in these systems, which have relatively low velocities.

Seeding and Germination

Hydroseeding both VSB and FWS wetlands has only

been successful for cattails in some instances. Additional

studies of this low-cost approach to planting are recommended. Seeding with 25 lb/acre mixed with a mulch at

2500 lb/acre have been used for VSB systems. Continuous shallow-water inundation has consistently produced

higher germination rates than have sprinklers.

Plant Toxicity

Emergent vegetation species, such as cattails and bulrushes, are sensitive to deep anaerobic sludge banks. In

situations with large volumes of carryover solids and malfunctioning activated sludge units, emergent vegetation in

the inlet zone can die from sulfide toxicity in the rhizosphere.

Wastewaters with high levels of suspended and settleable solids should be pretreated upstream from a wetland

system through use of a settling pond.

8.3.4 Treatment Effectiveness

Nutrient Uptake

Both FWS and VSB wetlands have been shown to be

unable to reduce levels of BOD, dissolved phosphorus,

and ammonia nitrogen below certain minimums. This is

due to internal processes in a wetland, such as the solubilization of influent settleable/suspended solids and the litter layer of aquatic macrophytes. Depending on the climate, pulses of dissolved carbon (both degradable and

non-degradable), soluble reactive phosphorus, and ammonia nitrogen are taken up by the plants, and they are

released during periods of active decomposition in the

wetland. Colder climatic conditions with early falls, long

cold winters, and warm springs will pulse these materials

into the water column during the spring warm-up period.

Nitrification

Without operating in a fill-and-draw batch mode, it is not

economically feasible to attain aerobic conditions in a VSB

system to convert ammonia to nitrate. A VSB is anaerobic

throughout most of its depth, with little opportunity for nitrifying bacteria populations to develop. Internally loaded

ammonia from the decomposition of algal cells has also

been shown to be a factor when attempting to use a VSB

to meet ammonia standards.

Sheet Flow for Nitrification

Attempts to operate a fully vegetated FWS wetland in a

shallow mode to simulate conditions of overland flow have

not proven to be effective.

Measurement of Treatment Effectiveness

Most of the FWS wetlands in the NADB are used to treat

high-quality influents producing low organic loading conditions. In most of these cases, the internal load is more

significant than the influent load. In only a few cases with

high organic loading rates were the upper limits of treatment effectiveness measured, such as Gustine, California. While many viewed Gustine as a failure, it provided

well-documented data on a wide range of BOD, TSS, nitrogen, and coliform fully vegetated zone loading conditions. These data showed that the upper instantaneous

BOD loadings of 150 to 200 lbs/acre/day could still result

in less than 40 mg/L BOD in the effluent. Such loading

rates were based on a fully vegetated FWS system and a

specific wastewater, so the utility of this information is limited. For example, if a wastewater had a soluble BOD loading of this magnitude, it would not be prudent (or successful) to use a single fully vegetated cell for treatment.