TREATMENT WETLANDS - CHAPTER 7 docx

A major function performed by wetland ecosystems is the removal of suspended sediments from water moving through the wetland.. SOLIDS CHARACTERIZATION The suspended solids entering a tre

A major function performed by wetland ecosystems is the

removal of suspended sediments from water moving through

the wetland These removals are the end result of a

compli-cated set of internal processes, including the production of

transportable solids by wetland biota

Low water velocities, coupled with the presence of plant

litter (in FWS wetlands) or sand/gravel media (in HSSF and

VF wetlands), promote settling and interception of solid

materials This transfer of suspended solids from the water to

the wetland sediment bed has important consequences for the

quality of the water, as well as the properties and function of

the wetland ecosystem Many pollutants are associated with

the incoming suspended matter, such as metals and organic

chemicals, which partition strongly to suspended matter In

FWS wetlands used for municipal wastewater treatment, the

accretion of solids contributes to a gradual increase in the

bottom elevation of the wetland However, wetlands used to

treat urban or agricultural stormwater, or those exposed to

periodic ancillary flooding, may have rapid accretions in the

inlet zone

In HSSF and VF wetlands, incoming suspended matter is

removed primarily through the mechanisms of interception

and settling Although particle resuspension due to wind,

wave, or animal activity can play an important role in the

sediment cycle of FWS wetlands, these mechanisms are

min-imized in HSSF and VF wetland systems As a result,

par-ticulate matter tends to accumulate in HSSF and VF wetland

beds, with profound consequences on hydraulic conductivity

and system performance

It should be noted that the concept of using VF filter

beds to remove incoming total suspended solids (TSS) as the

initial stage of a treatment process dates back to the 1960s

This concept originated with Dr Kathe Seidel, and came to

be known as the Max Planck Institute Process (MPIP) or

Krefeld Process (Seidel, 1966; Liénard et al., 1990; Brix,

1994d; Börner et al., 1998) The MPIP system consisted of

batch-fed vertical flow wetland beds followed by HSSF

wet-land stages for further effluent polishing

7.1 SOLIDS MEASUREMENT

TSS are measured gravimetrically after filtration and

dry-ing (Method 2540D; APHA, 1998), and reported in mg/L

The organic content is characterized as volatile suspended

solids (VSS), determined from the weight loss on ignition at

550°C The TSS method has been subjected to considerable

criticism by Gray et al (2000) for use on “natural” waters,

and these authors recommend a suspended sediment

concen-tration (SSC) analysis as a replacement (Method D 3977.97; ASTM, 2000) One fundamental difficulty is the representa-tiveness of aliquots, especially if they contain sand particles

A second difficulty is the wide variability of the TSS method

in low concentration ranges Gray et al (2000) quote the

Standard Methods precision as a 33% coefficient of variation

at 15 mg/L TSS measurements are likely to be biased low compared to SSC measurements

Turbidity in water is caused primarily by suspended matter, although soluble colored organic compounds can contribute Therefore, turbidity is sometimes used as a sur-rogate for gravimetric measurement of suspended matter The measurement technique involves light scattering The instrument is the turbidimeter, consisting of a nephelom-eter, light source, and photodetector The standard unit is the nephelometric turbidity unit (NTU) The correlation between TSS and NTU is often good for a specific wetland system, but care must be taken in the extrapolation from one site to another (Table 7.1) From these results, it may be concluded that the NTU–TSS relationships for FWS wetland effluents differ substantially from those for activated sludge effluents, and vary somewhat between natural systems

POTENTIAL FOR SAMPLING ERRORS

It is sometimes virtually impossible to sample interior land waters for TSS because of the disturbance of sediments caused by sampling Errors of one to two orders of magni-tude can easily occur This is the case in shallow zones of vegetated FWS wetlands If the water is deeper than about

wet-20 cm, accurate sampling is possible but not easy sion of a sampler may cause disturbance of bed sediments,

Immer-or the currents caused by water rushing into a sample bottle may disturb those sediments Ideally, the sample should flow into the sample bottle at the local velocity of the water in the

wetland This is termed isokinetic sampling, and is necessary

to prevent extraneous resuspension It is often not possible to achieve undisturbed sampling for TSS, and therefore difficult

to obtain proper flow-weighted or volume-weighted values of TSS at interior points For this reason, much of the available TSS data from wetland treatment systems consists of input and output measurements in pipes and at structures

This difficulty carries over to those chemical ents which partition strongly to the solids, or form an integral part of them Any interior water sample will likely contain an unrepresentative proportion of the locally agitatible, or trans-portable, sediments and particulates Subsequent analysis for the total amount of a partitioned or contained substance will yield an inaccurately high value

constitu-Similar sampling problems exist for HSSF wetlands Most

of the solids present within a HSSF wetland bed are an

accu-mulation of microbial biofilms, intercepted particulate matter,

and plant-root networks This accumulated material,

collec-tively called a biomat, occurs either as material attached to the

bed media and plant roots or as colloidal material within the

media pores Because the actual flow velocity, v (see Chapter

2), in an HSSF bed is very low, sampling events can induce

localized flow velocities at the point of sample collection that

are much higher than ambient flow velocities This disturbs

the in situ biomat and leads to sampling errors.

Introduction of sampling probes within the HSSF bed

disturbs the bed matrix, shearing biomat off bed particles,

which interferes with sample accuracy As a result, samples

taken within the HSSF bed are typically done using sample

ports fabricated from perforated pipe (the same applies for

VF wetlands) These sample ports are installed during struction and are a permanent feature of the HSSF wetland bed Depending on the orientation of the perforated section of the pipe (horizontal or vertical), these sample ports will pro-duce a sample that is width-averaged or depth-averaged over

con-a loccon-alized portion of the HSSF wetlcon-and bed A typiccon-al HSSF sample port assembly is shown in Figure 7.1; installation of the ports within an HSSF wetland is shown in Figure 7.2 However, the use of such pre-installed internal sampling ports does not guarantee that samples will be representative, because solids may still be selectively aspirated into the port Difficulties in sampling lead to large variability for interior TSS samples For instance, the coefficient of variation for TSS samples from the HSSF bed at Minoa, New York, was

Turbidity Range

Secondary effluent 0.37–0.50 — — — — Crites and Tchobanoglous (1998)

Secondary effluent 0.42–0.43 — — — — Metcalf and Eddy (1991)

Everglades 0.25 0.80 1–18 0.4–3.4 126 South Florida Water Management District,

unpublished data River water 0.83 0.77 0–145 0–125 64 Des Plaines River Project, unpublished data River water 0.66 0.95 50–1,400 100–1,000 23 Harter and Mitsch (2003)

Agricultural runoff 0.75 0.52 — — 1,013 Everglades Nutrient Removal Project,

unpublished data Submerged vegetation 0.74 0.93 0–215 0–150 >100 James et al (2002)

Water hyacinths 1.39 0.54 4–18 6–21 12 Crites and Tchobanoglous (1998)

Oxidation pond 0.47 0.06 1–15 1–27 96 Gearheart et al (1983)

4 cm Ø PVC Conduit Spacer (typical)

Stainless steel band clamp (typical)

10 cm Ø Sch 40 PVC

Gravel layer Mulch/detritus layer

FIGURE 7.1 Example of a HSSF wetland sampling port This particular assembly is designed to allow sample collection at three

different bed depths and installation of a thermocouple at the base of the mulch layer.

72% (N = 534), with no apparent distance profiles Similarly,

the coefficient of variation was 145% (N = 215) in the Grand

Lake, Minnesota, HSSF system

As a consequence of these sampling difficulties, most of

the samples collected in HSSF and VF wetlands consist of

inlet and outlet samples, unless interior sampling ports were

installed in the wetland at the time of construction Because

of the low flow velocities encountered in these systems, inlet

and outlet works in contact with the water develop a biomat

coating Again, care must be taken not to disturb this

bio-mat coating If agitation of the water and sloughing of the

biomat occurs, the sample will be contaminated and is no

longer representative of the wastewater As a result,

high-energy devices such as dipping buckets and bailers should be

avoided The use of peristaltic pumps is one preferred

sam-pling method, as the rate of sample withdrawal can be

con-trolled, and the sampling tube can be carefully positioned to

collect a representative sample Small-diameter guide pipes

are sometimes installed to facilitate placement of the sampler

tubing away from side walls, tank bottoms, and other sources

of sample contamination

SOLIDS CHARACTERIZATION

The suspended solids entering a treatment wetland may display widely varying characteristics, according to the source water involved Domestic wastewaters at all pretreat-ment stages contain suspended materials that are primarily organic Runoff waters, both urban and agricultural, may contain high proportions of mineral matter Other source waters may involve highly specific characteristics, such as the colloidal materials that discharge from milking parlors The two principal ways of describing solids are: the soil type and the size distribution

Soil fractions are often also applied to suspended matter, especially for situations involving mostly mineral materials These fractions are: organic, clay, silt, and sand The VSS fraction of the solids is usually taken to be a measure of the organic fraction (Table 7.2), and the remaining nonvolatile sus-pended solids (NVSS) are assumed to be the mineral fraction

of the overall TSS For incoming waters derived from runoff from mineral soils, the fraction organic may be rather low

At the Des Plaines site, river water entering averaged 11–16%

FIGURE 7.2 Four-cell HSSF wetland at the University of Vermont White pipes extending from the wetland beds are sampling ports.

TABLE 7.2 Organic Content of Various Source Waters Entering Treatment Wetlands

TSS Inlet

Houghton Lake, Michigan Lagoon 25 56

Tarrant, Texas Sedimentation basin 37 20 Connell, Washington Potato processing 350 94

Note: NVSS = non-volatile suspended solids

organic, whereas water leaving the treatment wetlands

aver-aged 16–26% organic Harter and Mitsch (2003) reported 9%

organic for both entering and leaving waters from the

Olen-tangy River wetlands However, the Houghton Lake natural

peatland showed 77% organic, and after lagoon

wastewa-ter addition showed 56% organic (unpublished data) As an

extreme example, the fraction VSS in a potato wastewater

treatment wetland was 94% (unpublished data) Obviously,

no generalizations may be made across the spectrum of

treatment wetlands and source waters, but it should be noted

that organic materials may be subject to decomposition after

deposition

Mineral constituents may be defined by size ranges

(Lane, 1947; Brix, 1998; Braskerud, 2003):

Silt: 2 µm < size < 60 µm

Sannd: 60 µm < size < 2 mm

Gravel: 2 mm < size < 64 mm

These mineral particles have relatively high densities, Rsy

2–2.5 g/cm3, and the larger sizes settle readily In contrast to

organics, these materials accrete without decomposition

Neither the particles entering the wetland nor those

leav-ing are of a sleav-ingle size Frequency distributions of particle

sizes are always present (Figure 7.3) As a result, particle

pro-cessing also becomes distributed, with large particles

behav-ing differently from small

7.2 PARTICULATE PROCESSES

IN FWS WETLANDS

FWS wetlands process sediments and TSS in a number of

ways (Figure 7.4) After the suspended material reaches

the wetland, it joins large amounts of internally generated

suspendable materials, and both are transported across the wetland Sedimentation and trapping, and resuspension, occur en route, as does “generation” of suspended material

by activities both above and below the water surface For example, algal debris may form at one location and deposit downgradient in the wetland

P ARTICULATE S ETTLING Single Particles

The slow-moving waters in the FWS wetland environment often permit time for physical settling of TSS The settling velocity of the incoming particulates, combined with the depth of the wet-land, gives an estimate of the time and travel distance for those solids

Solids sink in water due to the density difference between the particle and water For single, isolated spherical particles, the terminal velocity is reached quickly:

D s

where

d C

particle diameter, mdrag coefficient,

acceleration of gravity, m/

terminal velocity, m/sdensity of wat

FIGURE 7.3 Particle size distributions for two FWS wetlands At Des Plaines (EW3), the outlet particles are larger than those entering At

Houghton Lake (HL), the discharge area particles are larger than those in wetland background areas (From unpublished data.)

where the particle Reynolds number is:

If all physical properties are known, Equations 7.1–7.3

com-bine to determine the settling velocity This calculation is

easily automated on a spreadsheet, with the results shown in

Figure 7.5

In the laminar flow region, Rep < 1.0, the drag cient is inversely proportional to the particle Reynolds num-ber, and the settling velocity of the particle is then calculable

coeffi-from Stokes law:

w18Mgd2Rs R (7.4)where

d g

particle diameter, macceleration of gravvity, m/sterminal velocity, m/sdensit

2

w

R yy of water, kg/mdensity of solids, kg/

Plankton &

invertebrate litterfall

Macrophyte litterfall

Litter Resuspension

FIGURE 7.4 Processes affecting particulate matter removal and generation in FWS wetlands (Adapted from Kadlec and Knight (1996)

Treatment Wetlands First Edition, CRC Press, Boca Raton, Florida.)

0.0001 0.001 0.01 0.1 1 10 100 1,000 10,000

Particle Diameter (µm)

density = 2.00 density = 1.30 density = 1.10 density = 1.03 density = 1.01

FIGURE 7.5 Settling velocity of spherical particles in water at 20°C, for different particle densities.

discs (Figure 7.6) Although it is possible to correct for

non-spherical shapes (Dietrich, 1982), there is not a convenient

method for determination of the particle density Further,

particles may agglomerate to larger size, or be subject to

interference from neighboring particles

Settling of Mixtures

Settling of particulate matter may be described by a first-order

model (Equation 7.4) for each size fraction In general,

set-tling velocities are proportional to the square of particle size,

with variation including shape factors and particle density

Particle mass may be estimated to be roughly proportional to

the cube of size The time of fall of a particle through a

verti-cal distance (h) is determined from its velocity:

If the water is moving through the wetland length (L) at

velocity (u), the time of travel is:

when

fall

L u

h w

These concepts have been applied to mixtures in shallow

overland flow in grass (Deletic, 1999), and in wetlands (Li

et al., 2007), with mean particle diameter used to determine the settling velocity (w) Values of Nfall were found to be above

10 for complete removal, reflecting the difficulty of settling

of the small end of the particle size distribution (Figure 7.7).These relations also allow the conversion of a size dis-tribution to a settling velocity distribution, and ultimately to the size distribution remaining after some fixed settling time Procedures for such calculations may be found in Crites and Tchobanoglous (1998); however, there is rarely sufficient information on particle properties available Braskerud (2003) found considerable discrepancies when applying these procedures to mineral particles trapped in wetlands

Column Studies

Settling rates may also be determined experimentally cally, a large diameter column of water is charged with a well-stirred suspension of particles, and the concentration measured

Typi-at a sequence of times Typi-at a series of depths below the wTypi-ater surface Vertical profiles of TSS exist in differing shapes, depending on flocculation and particle–particle interference

A number of analytical techniques may be applied to such data (Font, 1991) Only the mean water column concentration of

FIGURE 7.6 Photomicrograph of suspended particulate matter

in the effluent from Des Plaines wetland EW3 (From Kadlec and

Knight (1996) Treatment Wetlands First Edition, CRC Press, Boca

Raton, Florida.)

0.0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1.0

Particle Falling Number

Grass Wetlands

FIGURE 7.7 Removal of TSS in shallow overland flow in grass

The particle falling number is (Lw/uh), in which w is the terminal

velocity of the mean particle diameter Original data centered on a

mean diameter of about 50 µm (Data from Deletic (1999) Water

Science and Technology 39(9): 129–136; and Li et al (2007) nal of Hydrology 338: 285–296.)

Jour-TSS will be considered here That concentration decreases as

time progresses Settling column data, for example, wetland

waters and other sources, indicate an exponential decrease

in concentration with time, and a time scale of a few hours

for the majority of settling to occur (Figure 7.8) The settling

velocities shown in Figure 7.8 range from w = 0.076 to 26.3

m/d Interestingly, exponential decreases are found for the

several sediments in Figure 7.8

Caution must be used in those applications where

col-loidal materials may be present in the inflow, because these

materials are stable or very slow to settle Very fine clay

suspensions and some milk processing wastewaters fall into

this category The settling velocity for planktonic solids was

found to be on the order of w = 0.076 m/d for the Wind Lake,

Wisconsin, wetland, which was dominated by algae

Column settling data provide estimates of the removal

time for TSS in the absence of dense vegetation

Confirma-tion of field applicability was found for wetland EW3 at Des

Plaines in 1991 The inlet zone was essentially unvegetated,

and the water velocity was on the order of 30 m/d Settling

column data (Figure 7.8) suggested that solids should

essen-tially be gone in eight hours, or after a travel distance of about

ten meters Transect information confirmed this estimate

“F ILTRATION ” VERSUS I NTERCEPTION

Conventional wisdom has it that the presence of dense

wet-land vegetation causes settling to be augmented by filtration

This is often not true in the usual sense of the term

filtra-tion It is trapping of sediments in the litter layer that prevents

resuspension, and thus enhances the net apparent suspended

sediment removal Macrophytes and their litter form a homogeneous “fiber bed” in the wetland context The void frac-tion in the stems and litter is quite high; straining and sieving are thus not typically the dominant mechanisms Submerged biomass additionally traps sediment in sheltered microzones, thereby lessening the potential for resuspension Confirmation

non-of sedimentation as the principal mechanism was provided in

the laboratory studies of Schmid et al (2005).

However, there are wetland circumstances in which the dominant mechanism is particles striking immersed objects and sticking The three principal mechanisms of fiber-bed filtration are well known and documented in handbooks (see, e.g., Perry et al., 1982; Metcalf and Eddy, 1991):

1 Inertial deposition or impaction—particles ing fast enough that they crash head-on into plant stems rather than being swept around by the water currents

mov-2 Diffusional deposition—random processes at either microscale (Brownian motion) or mac-roscale (bioturbation) which move a particle to an immersed surface

3 Flow-line interception—particles moving with the water and avoiding head-on collisions, but passing close enough to graze the stem and its biofilm, and sticking

The efficiencies of collection for these mechanisms depend

on the water velocity, particle properties, and water ties, as well as the character of submerged surfaces A typical wetland “fiber” is a bulrush stem of about 1 cm diameter

proper-Houghton Lake (HL) Discharge

HL Control HL Discharge EW4 Out Wind Lake

Bar El Baqar

w = 0.076 m/d

R 2 = 0.86

FIGURE 7.8 Examples of settling characteristics of TSS derived from wetlands and other natural contributing sources The mean settling

velocities range from 0.076 m/d for the Wind Lake wetland TSS, to 26.3 m/d for the clay alum mix (Data for HL Control, HL Discharge,

EW3 In, EW3 Out, EW4 Out, EW5 In, EW5 Out, and Wind Lake: authors’ unpublished data; data for Clay/Alum: ASCE (1975)

Sedimenta-tion Engineering Vanoni (Ed.), American Society of Civil Engineers (ASCE): New York; data for Bar El Baqar: PLA (1993) 1993 Field Program for the Egyptian Engineered Wetland Report prepared for the United Nations Development Programme, New York, P Lane and

Associates, Ltd (PLA).) (Graph from Kadlec and Knight (1996) Treatment Wetlands First Edition, CRC Press, Boca Raton, Florida.)

A typical particle might be on the order of 1–100 µm A

typi-cal water velocity is on the order of 10–100 m/d Under these

conditions, the collection efficiencies of Mechanisms 1 and

2 are predicted to be vanishingly small There is evidence

that Mechanism 3 is operative and significant Lloyd (1997)

examined the submerged surfaces of bulrushes

(Schoeno-plectus (Scirpus) validus) and found particles as small as

0.5–2.5 µm sticking to biofilms (Breen and Lawrence, 1998)

Saiers et al (2003) studied the movement of very small (0.3

µm), unsettleable particles of TiO2 in the Florida Everglades

They concluded that 29% of the particle impacts on

periphy-ton-coated stems resulted in sticking in a plant (Eleocharis

spp.) density of 1,150 per m2 These stems were only 0.2 cm

in diameter, resulting in 99% porosity Saiers et al (2003)

defined a first-order rate constant for removal by sticking,

which on an areal basis is:

n d

2

2

14

Settled particles may not “stay put” for a number of reasons

Hydrodynamic shear forces may tear particles loose from the

sediment bed, which is a dominant mechanism in streams and

rivers However, wetlands provide an environment in which

other processes may occur as well Wind and wave action

are major drivers of resuspension in lakes, and may also be

operative in open water areas of FWS wetlands Additionally,

biological activity may result in the movement of particles

from the sediments to overlying water

Unvegetated Surfaces

Much is known about the resuspension of particulates from

flat surfaces (ASCE, 1975) Most interpretations are made

in terms of the force per unit area (shear stress) required to

tear a particle loose from the sediment surface The concepts

involve purely physical forces and apply most readily to

min-eral substrates and river systems Most theoretical results are

for planar sediment bed bottoms with no extraneous objects

Vegetated wetland bottoms do not fit these conditions

In the treatment wetland environment, physical

resus-pension (due to high flow velocities) is not a dominant

process Water velocities are usually too low to dislodge a

settled particle from either the bottom or a position on

sub-merged vegetation However, in design, it may be necessary

to avoid wetland aspect ratios that produce excessively high

linear velocities The potential for erosive velocities exists for highly loaded wetlands with high length-to-width ratios Estimation of the velocity required to foster resuspension may be based on the settling characteristics of the solids and the frictional characteristics of the wetland, combined with known correlations of the critical shear stress for particle dislodgment (ASCE, 1975) Modifications are needed for the case of laminar flow, which is the general case for wetlands (Mantz, 1977; Yalin and Karahan, 1979)

Velocities that cause erosion in open channels are high compared to wetlands For instance, French (1985) lists rec-ommended maximum (nonscouring) velocities for 14 canal

materials in the range 0.46 < u < 1.83 m/s Such

consider-ations resulted in a maximum canal velocity design constraint

of 0.76 m/s for Everglades protection wetlands conveyance canals (Burns and McDonnell, 1996) In anticipation of more erodable particulates inside the wetlands, wetland velocities were limited to no more than 0.03 m/s (2,600 m/d) These large wetlands had lengths up to 2,500 m, which therefore created a design detention minimum of one day The annual average design detention time was 30 days No erosion has been noted in this project or its companions of comparable size and detention

Effects of Vegetation

It is known that vegetation increases the retention of lates in both lake and stream environments For instance, Horp-pila and Nurminen (2003) found that beds of submerged plant

particu-species—butter cup: Ranunculus circinatus; coontail: tophyllum demersum; and pond weed: Potamogeton obtusifo- lius—in a lake environment effectively prevented resuspension,

Cera-which they attributed to a reduction in wind and wave action

Horvath (2004) studied the effect of macrophytes—rushes: cus spp.; bur-reed: Sparganium spp.; forget-me-not: Myosotis

Jun-spp.—on retention of particulate matter in a small stream, and found enhanced trapping in proportion to biomass

It is logical that these same effects are prevalent in ment wetlands Dieter (1990) found about a threefold reduc-tion in resuspension from open water to vegetated areas in a prairie pothole wetland Hosokawa and Horie (1992) demon-strated enhanced removal in both laboratory channels with

treat-dowels and in field flumes in a reed bed (Phragmites tralis) In fully vegetated wetlands, the litter and root mats

aus-provide excellent stabilization of the wetland soils and ments This limits, but does not eliminate, resuspension

sedi-The Floc Layer

Some treatment wetlands, such as those used for low-level nutrient removal, develop very flocculent sediment beds These sediments are positioned on top of the consolidated soils, and may be interwoven with plant detritus Bulk densi-ties of such floc layers may range downward to 0.03–0.05 g/cm3 of dry matter (James et al., 2001; Coveney et al., 2002)

Depths of these loose and unconsolidated materials have been found to exceed 30 cm in some situations (Table 7.3)

Despite low bulk density, the amount of floc dry matter is

substantial For instance, the Sacramento data in Table 7.3

convert to about 9,700 g/m2 of dry matter present as the floc

The origins of floc are not well understood, but it has

been found to occur in both macrophyte-dominated

(Sac-ramento) and SAV-dominated (ENRP Cell 4) wetlands It

likely contains a significant microbial detrital component, as

well as algal and macrophyte detritus Floc also occurs in

the ultra-low nutrient, unimpacted Everglades (Gaiser et al.,

2005), where it is presumably the result of an active

periphy-ton biological cycle

There is not an accepted common terminology for the

floc Nolte (1997) called it the “A layer,” and described it as

follows:

The A layer consists of a slurry of dark, decomposing, loosely

structured detrital material that pours out when the

sam-pler is tipped The material in the A layer has settled to the

bottom, but has not been integrated into the matrix of the

basin floor.

This material is not subject to transport under most ambient

conditions, but is very mobile if disturbed For example,

dis-turbance resuspension tests were conducted at the Houghton

Lake treatment wetland A bottomless sharp-edged cylinder

was twisted down into the soil, and the interior biomass (live,

dead, litter) was removed The remaining, isolated water was

gently agitated, and then sampled for solids content The

mobile material averaged 880 o 100 g/m2 (mean o SE)

Other Resuspension Mechanisms

The wetland environment provides an opportunity for three

other mechanisms of resuspension: wind-driven turbulence,

bioturbation, and gas lift In open water areas, wind-driven

currents cause surface flow in the wind direction and return

flows along the bottom in the opposite direction These

recir-culation velocities can far exceed the net velocity from inlet to

outlet For wetlands with large open water zones, waves add

to the overall process of resuspension Lake studies suggest both processes are wind-dependent For instance, Malmaeus and Hakanson (2003) suggest resuspension is proportional to the square of the wind speed Additionally, fetch and water depth are controlling factors

Animals of all types and sizes can cause resuspension to occur Feeding carp (Kadlec and Hey, 1994) and nesting shad (APAI, 1995) have been observed to cause problems The carp rooted in the sediments for food, and thus resuspended large amounts of sediments Control was by drawdown and freezing The shad fanned nests on the wetland bottom, and resuspended sediments Control was by drawdown and avian predation Beaver activity can cause stirring, often at the out-let of the wetland, in conjunction with attempts to dam the outlet Human sampling activities in the interior of treatment wetlands may also result in locally-elevated concentrations

of suspended solids For instance, the passage of a drifting boat can cause extreme resuspension (Figure 7.9)

Gas lift occurs when bubbles of gas become trapped in or attached to particulate matter Wetland sediments are often

of near neutral buoyancy; so a small amount of trapped gas can cause “sinkers” to become “floaters.” There are several gas-generating reactions in a wetland environment Most important are photosynthetic production of oxygen by algae and production of methane in anaerobic zones

CHEMICAL PRECIPITATES

Several chemical reactions can produce particulate matter within wetlands under the proper circumstances Some of the more important are the oxyhydroxides of iron, calcium carbonate under aerobic conditions, and divalent metal sul-fides under anaerobic conditions As conditions of chemical composition, pH, and redox change in the wetland, these and other compounds may undergo dissolution and be removed from the sediment bed

TABLE 7.3

Floc Thicknesses and Bulk Densities for the Everglades Nutrient Removal Project (ENRP),

Lake Apopka, Florida Project, and the Sacramento California Demonstration Wetlands Project

Thickness (cm) Bulk Density (g/mL)

Source: Data from Nolte and Associates (1997) Sacramento Regional Wastewater Treatment Plant Demonstration Wetlands Project.

1996 Annual Report to Sacramento Regional County Sanitation District, Nolte and Associates: Sacramento, California; Coveney et al.

(2002) Ecological Engineering 19(2): 141–159; and South Florida Water Management District, unpublished data.

Iron Flocs The iron oxyhydroxides are typically flocs,

with the possibility of coprecipitates They may form under

conditions of elevated dissolved ferric iron and oxygen-rich

water The processes may be represented as (Younger et al.,

2002)

2 2

14

12

Fe3 + 2H O2 l FeOOH(sus) 3H+ (7.10)

FeOOH(sus) l FeOOH(sed) (7.11)

These precipitates are characterized by an unmistakable

blood-red color (Figure 7.10) As indicated by the chemistry,

formation is inhibited by low pH and by low dissolved oxygen

Formation may be abiotic, or mediated by microorganisms

such as Thiobacillus ferrooxidans However, at pH > 9, the

rate of the abiotic reaction is so fast that formation is

con-trolled by the rate of oxygen supply (Younger et al., 2002)

In the pH range 6 < pH < 8 that generally typifies treatment wetlands, rates are slow enough to be a design consideration This set of reactions forms the basis for phosphorus removal

by addition of ferric chloride to wastewaters, and the panying co-precipitation of the phosphorus Consequently, the subsequent fate of these solids in polishing treatment wetlands is of considerable interest

accom-Aluminum Flocs The aluminum oxyhydroxides are

also typically flocs, with the possibility of co-precipitates They may form under circumneutral pH conditions, and do not require oxygen The processes may be represented as (Sobolewski, 1999):

FIGURE 7.9 Passage of a drifting boat can stir up a cloud of floc This site is in the interior of the A.R Marshall Loxahatchee National

Wildlife Refuge The water was about 45 cm deep, and the vegetation was sparse.

FIGURE 7.10 (A color version of this figure follows page 550) Venting groundwater at this Wellsville, New York, site contains iron,

which oxidizes upon contact with air.

These precipitates are characterized by their formation of a

“pin floc” material that does not readily settle in FWS

wet-lands (Bachand et al., 1999) This set of reactions also forms

the basis for phosphorus removal by addition of alum to

wastewaters, and the accompanying co-precipitation of the

phosphorus Consequently, the subsequent fate of these solids

in polishing treatment wetlands is of considerable interest

Calcium Carbonate Calcium carbonates may be formed

in wetlands, under conditions of elevated pH and dissolved

calcium The operative chemistry may be summarized as

This reaction may occur abiotically, but perhaps more

impor-tantly it may be mediated by algae Algal activity can drive

up pH, and create conditions that foster creation of

calcium-rich solids (Vymazal, 1995) Indeed, this process has

con-tributed to the formation of marl prairies as a form of natural

wetlands New sediments in Everglades protection treatment

wetlands contain a significant fraction of calcium compounds

(Dierberg et al., 2002).

Metal Sulfides Many metals form very insoluble

sul-fides, including mercury, lead, cadmium, and zinc, as further

discussed in Chapter 11 These precipitates are important in

the processes of metal removal in wetlands, and follow the

general chemistry (Sobolewski, 1999):

SO4 2CH O2 lHS H2HCO3 (7.14)

However, for many treatment wetland applications, metals

are present at only very low concentrations Consequently,

the formation of insoluble sulfides does not usually create

measurable additions to the sediments of the wetlands

B IOLOGICAL S EDIMENT G ENERATION

Wetlands produce sediments via processes of death, litter

fall, and litter attrition This occurs for biota at a number

of different size scales, ranging from macrophytes on down

to bacteria Algal productivity can be a major generator of

suspended solids A second set of processes adds pollen and

seeds to the water The TSS produced is organic in

charac-ter, resulting in a high carbon content and a high proportion

of VSS The chlorophyll and pheophytin (dead chlorophyll)

content is high if the algal pathway is dominant

Some TSS originates from leaf and stem litter For

instance, annual leaf litterfall in a natural sedge-shrub

peat-land was found to be 60–70 g/m2 (Chamie, 1976) Some part

of this material contributes to TSS, either via direct attrition,

or via microbial decomposition

The generation of sedimentary material is a very

impor-tant internal process in nutrient-rich treatment wetlands The

generous supply of nutrients assures a large production of a

wide variety of transportable organisms and associated dead

organic material Such wetlands are characterized by high

water chlorophyll content and high sediment accumulation

Bacterial and algal growth is promoted, and decomposition products form a new pool of suspendable material A host of

wetland invertebrates, such as Daphnia and waterboatman (Corixidae), also die and contribute to the sediments, and

they may be present in pumped lagoon water

These processes are virtually impossible to predict and quantify But it is important to recognize that they exist, because they contribute to a background level of TSS in a wetland

ACCRETION

Trapped TSS, plus material generated within the wetland, will accrete as either movable sediment or the consolidated immovable new soil produced from the sediments Not all

of the dead plant material undergoes decomposition Some small portions of both aboveground and belowground nec-romass resist decay, although these are typically shredded

by microbial and other invertebrate processes Underground processes form nonsuspendable accretions, some part of which is stable and does not fully decompose The origins of new sediments may be from remnant macrophyte stem and leaf debris, remnants of dead roots and rhizomes, and from indecomposable fractions of dead microflora and microfauna (algae, fungi, invertebrates, bacteria)

Measurement of Accretion

The processes above combine to determine the amount of sediment at various locations within the wetland as a func-tion of time and the TSS concentration in the wetland efflu-ent Cup collectors may be placed on the wetland bottom

(Jordan and Valiela, 1983; Fennessy et al., 1992; Braskerud,

2001a); these typically intercept the downward vertical flux of sediment but prevent shear-induced resuspension Plate collectors may be placed on the wetland bottom, fol-lowed by sediment harvest above that horizon at a later time (Kozerski and Leuschner, 1999; Braskerud, 2001a) Alterna-tively, neutral density particulate material may be laid down

in a layer, and retrieved by coring and sectioning (Harter and Mitsch, 2003) Another technique involves the elevation of a blunt-footed rod, which is lowered to the sediment surface

A reference rod, driven deep into stable soils, provides the local datum (Reeder, 1990) Other quantitative studies have relied upon atmospheric deposition markers such as radio-active cesium (137Cs) or radioactive lead (210Pb) (Kadlec and

Robbins, 1984; Craft and Richardson, 1993; Robbins et al.,

2004) These techniques require several years of continued deposition for maximum accuracy

Cup collectors typically yield much more sediment than

plate collectors For instance, Schulz et al (2003b) found

30 o 3 g/m2·d collected in cups in a riverine bed of Sagittaria sagittifolia, compared to 8 o 2 g/m2·d collected on plates This is presumably due to the prevention of resuspension in cups, whether it be due to fluid shear or to bioturbation For mineral sediments, the difference between cups and plates is less, probably because of the lesser importance of resuspen-sion of heavier particles (Braskerud, 2001a)

Amount and Distribution of Accretion

Accretions measured in various wetlands vary from a few

millimeters per year to over a centimeter per year (Table 7.4)

These accumulated solids represent the potential for filling

of a constructed wetland It is an easy calculation to allocate

the removed TSS to the buildup of new solids in the FWS

wetland For municipal wastewater polishing, typical

opera-tions lead to an accumulation of 1–2 mm/yr of new solids

(50 mg/L removed at q = 5.5 cm/d at a bulk density of

0.5 g/cm3 yields 2.0 mm/yr) But that material is augmented

by internally generated solids and decreased by

decomposi-tion of the organic pordecomposi-tion of sediments and soils The net

increase may total up to 10 mm/yr in a highly eutrophic marsh (Table 7.4) Even more accumulation can result from the trap-ping of mineral solids from urban or agricultural runoff.For high amounts of sediment trapping compared to gen-eration and resuspension, buildup typically occurs preferen-tially in the inlet section of the wetland Therefore, a “delta” of accreted sediments builds in the inlet region of the wetland For example, food processing wastewaters can contain very high TSS concentrations, which in turn can fill a treatment wetland with solids Van Oostrom (1995) reported that one third of

the volume of a floating Glyceria mat wetland was filled after

20 months of operation (Figure 7.11) The wastewater was

TABLE 7.4

Accretion Rates in FWS Wetlands

(mg/L)

Accretion (cm/yr)

Louisiana Salt marsh DeLaune et al (1978) 137 Cs Low 1.1–1.35 Louisiana Forested Conner and Day (1991) Feldspar Low 0.84 Louisiana Forested Rybczyk et al (2002) Feldspar 0.05 0.14 Xianghai, China Open marsh Wang et al (2004) 137 Cs + 210 Pb Low 0.35 Xianghai, China Isolated marsh Wang et al (2004) 137 Cs + 210 Pb Low 0.65 Michigan Marsh Kadlec and Robbins (1984) 210 Pb 0.1 0.2 Norway Farm Runoff Marsh CW Braskerud (2001b) Plate 0.16 2 Norway Farm Runoff Marsh CW Braskerud (2001b) Plate 0.37 4 Everglades WCA2A Marsh Reddy et al (1993) 137 Cs 0.3 0.5 Everglades WCA2A Marsh Craft and Richardson (1993b) 137 Cs 0.3 0.4 Everglades WCA3 Marsh Craft and Richardson (1993b) 137 Cs 0.1 0.3 Everglades Marsh Robbins et al (1999) 210 Pb 0.3 0.5 Everglades Marsh Chimney (unpublished data) Feldspar 0.1 0.85 Sacramento, California Marsh CW Nolte and Associates (1998b) Visual 16 1.5 Houghton Lake, Michigan Marsh NTW Kadlec (unpublished data) Resurvey 10 1.0 Chiricahueto Runoff, Mexico Marsh Soto-Jimenez et al (2003) 210 Pb 14 1.0 Louisiana Forested NTW Rybczyk et al (2002) Feldspar 15 1.14

Note: CW = constructed wetland; NTW = natural treatment wetland.

0 5 10 15 20 25 30 35 40

FIGURE 7.11 The sediment “delta” developed in a small treatment wetland mesocosm (Data from van Oostrom (1995) Water Science and

Technology 32(3): 137–148.) (Graph from Kadlec and Knight (1996) Treatment Wetlands First Edition, CRC Press, Boca Raton, Florida.)

a nitrified meat processing effluent, with incoming TSS of

269 mg/L, and the removal rate was 5,300 g/m2·yr Accreted

sediments totaled 40% of the removed solids, 2,100

g/m2·yr, and these were concentrated near the inlet end of the

wetland The density of the solids was very low, around

0.03 g/cm3

In contrast, lighter loadings and open water areas may

foster the redistribution of suspendable material For instance,

Brueske and Barrett (1994) found a “delta” in a highly loaded

wetland (around 3.6 g/m2·d TSS), but little or no “delta” for a

lower loading (around 0.8 g/m2·d TSS) Both Harter and Mitsch

(2003) and Brueske and Barrett (1994) found greater sediment

accretion in open water areas, which may have been attributable

to most of the flow traveling through such areas, or to

bioturba-tion (Figure 7.12) In contrast, Benoy and Kalff (1999) found

a linear relation between sediment accumulation and biomass

for submerged species Myriophyllum spicatum, Potamogeton

spp., Ceratophyllum demersum, and Elodea canadensis beds

in Lake Memphremagog between Québec and Vermont It is

apparent that the processes involved in sediment accumulation

in wetlands are too complicated to permit generalities

In the long run, solids accretion may raise the elevation

of the wetland bottom, and thus impact system hydraulics

and treatment U.S EPA (2000a) suggests that accretion in

municipal wastewater treatment wetlands results from both

external and internal sources, which is conceptually correct

However, the U.S EPA (2000a) estimate of accretion from

external solids, 2–4 cm/yr, is based upon lagoon

accumula-tion rates, and is excessively high For example, the removal

of 30 mg/L of TSS at a hydraulic loading rate of 10 cm/d

results in solids storage of 1,095 g/m2·yr At a density of

0.2 g/cm3, this gives 0.55 cm/yr if there is no

decomposi-tion However, municipal TSS is about half mineral, and

half-decomposable solids (VSS, see Table 7.2), and hence

long-term external accretion would be about 0.27 cm/yr

U.S EPA (2000a) estimates internal accretion as the annual

deposition of macrophyte detritus to be 2.4 cm/yr However, that material too is subject to decomposition, leaving an esti-mated residual long-term buildup of 20% of the input, or 0.48 cm/yr In sum, the accretion in this example would be 0.75 cm/yr This is consistent with the measured accretions

min-eral content and loadings of TSS increase, so do accretions Highly loaded wetlands treating mineral solids have been observed to accrete 2–8 cm/yr (Braskerud, 2001a)

Accretion is typically spatially nonuniform, due to dients in deposition and productivity This has been found to

gra-be true even in wetlands of very low nutrient status (Reddy

et al., 1993) Inlet zones may therefore be subject to solids

accumulations that are double the wetland average However, some wetlands appear to redistribute solids fairly evenly from inlet to outlet

To the authors’ knowledge, only one municipal water polishing FWS wetland has been serviced for solids removal, the Orlando, Florida Easterly Wetland inlet cells

waste-(White et al., 2004) The one removal of accumulations

restored good hydraulic patterns, and restored original water quality performance

It was suspected that uneven accumulations of new ments were affecting flow patterns, and reducing efficiency (Sees, 2005) The inlet 9% of the wetland was excavated 45

sedi-cm, after 15 years of operation This overexcavation restored more than the original freeboard, and resulted in a great improvement in hydraulic efficiency, from 34% to 74% (see

Michi-gan (32 years, constructed), and Houghton Lake, MichiMichi-gan (30 years, natural), have experienced accretions in the range

of Table 7.4, but this has not jeopardized containment or operability However, the Tucson, Arizona, Sweetwater wet-land inlet cells have required solids removal after just a few years, because of the high suspended solids inlet water (see

FIGURE 7.12 Spatial distribution of plate sediment collection rates along the flow direction of a constructed marsh treating river water

(Data from Harter and Mitsch (2003) Journal of Environmental Quality 32(4): 325–334.)

7.3 TSS REMOVAL IN FWS WETLANDS

As for most treatment wetland water quality parameters,

the utilization of input and output data to compute percent

removals is an inadequate representation of the processes

which lead to those removals This is particularly true for the

removal of TSS

INTERNAL CYCLING: MASS BALANCES

Models of sediment transport have been developed and

veri-fied for estuaries (Hayter and Mehta, 1986; Nakata, 1989, for

example) These are 2- and 3-D models that allow for

disper-sion, settling, and resuspension; and generation is not

usu-ally an important term These models may be adapted to the

wetland situation In the short term, there are significant

fluc-tuations in TSS storage within the water column in response

to the variations in settling, resuspension, and generation

Childers and Day (1990) state: “Our results affirm the

vari-ability of short-term sediment transport and depositional

processes.…” Over a long period, however, changes in water

column storage are negligible compared to other inputs and

outputs The water column TSS mass balance then assumes

the character of a steady state model There is an

accompany-ing sediment bed balance, in which the change in storage is

the dominant feature The long-term, time-average profiles

calculated from the vertically averaged mass balances for

TSS in a linear flow wetland are (see Figure 7.14):

2

h P

su

2

2

S t u

w h

¤

¦¥

³µ´

T(7.21)

whereconcentration, g/m mg/Lconcentr

i

C C

wetland length, mnomin

w Nh

(7.22)

wherenumber of TIS

N

FIGURE 7.13 Excessive TSS can fill the inlet deep zone to a

treat-ment wetland, as happened at the Tucson, Arizona, sweetwater

wetland Note the bird tracks that highlight the complete filling of

the deep zone with relatively high density solids Incoming waters

had high TSS from filter backwashes at the secondary treatment

plant that provided the source water.

Equation 7.21 contains a subtle message that bears on the

removal of nearly all pollutants in wetlands, not just TSS

The right-hand numerator contains the settling velocity times

the wetland length An increase in either will cause a faster

approach to C* The denominator contains the water

veloc-ity times the depth (uh) An increase in either of those will

cause a slower approach to C* The detention time does not

appear directly in this simplified mechanistic model, and the

reason is easy to understand If the water depth is doubled,

for the same incoming volumetric flow rate and wetland area,

the detention time will be doubled But the particles do not

fall any faster and now have twice as far to travel to the

bot-tom The extra detention time is used up by a greater vertical

travel time On the other hand, doubling the area of the

wet-land, all else being equal, will also double the detention time

The vertical settling distance is not increased, and the extra

time causes greater removal

A detailed gradient study to provide calibration of the

k-C* model (as discussed in Chapter 6) was done at the Hallam

Valley wetlands in Melbourne, Australia (Wong et al., 2006)

Exceedingly high water flows (nominal HRT < three hours)

were required to detail the rapid decrease of TSS Model fits

were excellent, with w-values in the range of 16–21 m/d, for

both vegetated and unvegetated channels However, the

C*-value for the unvegetated channel was about double that for

that for the vegetated channel (60 versus 33 mg/L) This is

consistent with resuspension being greater in the open channel

(Equation 7.19) The rates of TSS removal in other continuous

flow through wetlands are not quite exponential (Figure 7.15)

The rapid initial declines in concentration prevail for only a

brief time of travel, after which declines follow a slower pace

(The Hallam Valley study did not contain a long portion of

wetland that could display such a slow decline.)

Thus it is clear that the TSS leaving an FWS treatment wetland of moderate to long detention is more reflective of generation and resuspension than of unsettled incoming sol-

ids Therefore, for nearly all FWS data sets, the parameter w

cannot be determined accurately

INTERNAL CYCLING

The second feature of the mass balances is the ability to sure individual components of solids processing, and to com-bine them to infer other results Data from the Des Plaines may be used in this way Wetland EW3 was heavily loaded when the pump was operating and contained relatively sparse emergent vegetation Independent measurements were made

mea-in settlmea-ing columns, yieldmea-ing w = 9.7 m/d Measurements of

R were made utilizing sediment cups plus input and output data, which gave R = 46.0 g/m2·d Estimates of G = 1.6 g/m2·d (WRI, 1992) Accordingly, from Equation 7.19, the expected

value of C* = 4.9 g/m3 Thus both C* and w were estimated

independently from the transect data for TSS The predicted drop in TSS agreed quite well with the measurements.This same data gives allows an approximation for the resuspension rate, and the net accretion rate (gross accretion less decomposition; Figure 7.16) The generation rates in this balance were estimated from measurements of productiv-ity of the organisms in the water column and from biomass measurements The striking feature of the mass balance is the large amount of solid material that is cycled, compared

to inputs, outputs, or removals Other studies have produced similar results (Table 7.5)

It may be concluded that in most instances, the ent TSS from a FWS treatment wetland is determined by

S, Settling rate

FIGURE 7.14 Framework for mass balances on suspendable materials in the wetland environment (Adapted from Kadlec and Knight

(1996) Treatment Wetlands First Edition, CRC Press, Boca Raton, Florida.)

internal biological processes, and not by the removal

effi-ciency for incoming TSS As a corollary, the solids

leav-ing the wetland will very often not be related to the solids

entering, but rather to the detrital fragments originating

internal to the system

SEASONAL AND STOCHASTIC EFFECTS

Because wetland effluent TSS is strongly related to internal

ecosystem processes, random physical and biological events

have pronounced effects on effluent concentrations In tion, season and temperature are modifiers of the processes that generate and cycle solids These effects may be sepa-rated by detrending the data, which typically follow a mild annual cycle with superimposed variability The trend may

addi-be determined most accurately if there are data spanning many annual cycles, which may then be “folded” into one multiyear display and averaged

TSS data time series often display some degree of dal behavior through the course of a calendar year Therefore,

sinusoi-Gross sedimentation

0.3 g/m 2  d 6.5 g/m 2

26.6 g/m 2  d

Output Input

FIGURE 7.16 Components of the sediment mass balance for wetland EW3 at Des Plaines, Illinois The balance period is the 23-week pumping

period in 1991 (Data from WRI (1992) The Des Plaines River Wetlands Demonstration Project Report to U.S EPA, July 1992 Wetlands Research Inc (WRI), Chicago, Illinois.) (Adapted from Kadlec and Knight (1996) Treatment Wetlands First Edition, CRC Press, Boca Raton, Florida.)

FIGURE 7.15 Gradients in suspended solids along the flow direction in treatment wetlands (Data for Arcata, California: Gearheart et al.,

(1989) In Constructed Wetlands for Wastewater Treatment: Municipal, Industrial, and Agricultural Hammer (Ed.), Lewis Publishers, Chelsea, Michigan, pp 121–137; data for Listowel, Ontario: Herskowitz, (1986) Listowel Artificial Marsh Project Report Ontario Ministry

of the Environment, Water Resources Branch: Toronto, Ontario; data for Des Plaines, Illinois: unpublished data).