TREATMENT WETLANDS - CHAPTER 24 doc

Type A wetlands are typically constructed free water surface FWS marshes and relatively low-tech constant-flow horizontal subsurface flow HSSF wetlands.. Specific examples of Type B tr

Treatment wetlands can be designed to be “au natural” or

complex Examples of designs along a full spectrum from

almost zero fossil-fuel input to highly engineered and

man-aged systems can be found around the world In this book

we term these design and operation strategies as Type A and

Type B wetlands Type A systems are at the most natural,

low-energy end of the scale Type B wetlands are

higher-energy and more highly engineered systems There are good

reasons to slide on this scale of design and management

complexity, depending upon project goals In this chapter we

describe some of the rationale in selecting the appropriate

level of management intensity

Type A treatment wetlands clearly cost less to operate

than Type B systems Type A systems may also cost less to

construct, even though they typically occupy a larger

foot-print area than Type B systems They are the preferred

solu-tion when a number of factors are not limiting:

Relatively inexpensive land area is abundant

Land is relatively level

There is insufficient treatment potential with a

Type A approach.

There is insignificant potential for unauthorized

human interaction with the wastewater in the

wetland

Habitat and aesthetic benefits outweigh the potential for

nui-sance conditions related to mosquitoes or other wildlife

Type A wetlands are typically constructed free water

surface (FWS) marshes and relatively low-tech constant-flow

horizontal subsurface flow (HSSF) wetlands

The Type B treatment wetland-management strategy

is preferred when one or more of the factors in the

preced-ing list are limitpreced-ing the design Specific examples of Type B

treatment wetlands include:

HSSF wetlands augmented with reactive media

intended for phosphorus or metals removal

Addition of aerators for treatment improvement

Operation of vertical flow (VF) wetlands in a

fill-and-drain mode for improvement of nitrification

and denitrification

Combinations of VF, HSSF, or FWS wetlands to

achieve specific process goals

Recycling of treated effluent to improve

denitrifi-cation and/or address influent toxicity issues

Step-feeding of influent to reduce the impact of

localized loading zones

Wetlands and aquatic plant treatment systems with

There are situations where these wetland modifications make sense from a process design and economic standpoint, as dis-cussed in this chapter However, the system will invariably require more operator attention, and there is often a reduc-tion in system reliability For instance, if a treatment wet-land requires an air blower to achieve regulatory compliance, system reliability is essentially governed by the mechanical blower, not the wetland ecosystem The same is true for dif-ferent operating regimes; if a treatment wetland requires alternating periods of loading and resting to avoid bed clog-ging, failure to follow the prescribed operating sequence will lead to system failure

Constructed wetlands have limits of performance, which have been discussed in previous chapters It is human nature

to attempt to relieve some of these constraints by modifying the basic wetland construct or its operation It is the purpose

of this chapter to explore some of the innovative ideas that have been implemented to enhance wetland performance These fall roughly into four categories:

Ecological or environmental modificationsChemical additions

Operational strategiesIntegrated sequences of natural systems, including one or more wetlands

These will typically move the wetland system from a passive

mode (Type A) to a mode of active management (Type B)

They will also usually involve more capital and operating cost, so the potential advantages should be weighed against other options for improving performance, such as alternative technologies or increased size

24.1 ECOLOGICAL OR ENVIRONMENTAL MODIFICATIONS

Natural wetlands have been regarded as “self-designing” systems They have also been ascribed certain functions that provide many kinds of benefits: habitat, flood control, biodi-versity, and water quality improvement Treatment wetlands target the water quality improvement functions It is there-fore logical—to humans—to try to nudge the wetland to a condition that fosters water quality For instance, it has been

suggested that magnetic fields can improve wetland

perfor-mance (Rasit, 2006) Ecological and environmental

modifi-cations that have been implemented to enhance the treatment

performance of wetlands are briefly discussed in the

follow-ing text

M ICROBIAL E NHANCEMENT

Microbial processing plays a very important role in the

reduction of many pollutants in treatment wetlands In

virtu-ally every case, among the many thousands of treatment

wet-lands, the presence of microbial populations that will deal

with the target pollutants is assumed in the design For the

major categories of contaminants, that presumption appears

justified: if there is ammonia, nitrifiers are there; if there is

nitrate, denitrifiers are there; if there is sulfate, sulfur

bac-teria are there However, these populations cannot establish

instantaneously, and a period of some weeks of grow-in may

be necessary In general, it does not seem necessary to

inocu-late a wetland with the desired microorganisms

It has been found that specialty bacteria—those that

focus on particular or unusual compounds—will also require

a period of time to develop This leads to the observation

of an induction period, during which little or no pollutant

reduction occurs For instance, Kadlec and Srinivasan (1995)

found no degradation of naphthoic acid for about 100 hours

in batch cattail mesocosms, but thereafter a steady decline

(Figure 24.1)

Runes et al (2001) found that bioaugmentation of

sedi-ment with small quantities of an atrazine spill-site soil (1:100

w/w) resulted in the mineralization of 25–30% of atrazine

under both unsaturated and water-saturated conditions

Atra-zine and its daughter products were almost undetectable

after 30 days Unbioaugmented sediment supplemented with

organic amendments (cellulose or cattail leaves) mineralized

only 2–3% of atrazine The population density of

atrazine-degrading microorganisms in unbioaugmented sediment

2 3 4 5 6 7 8 9 10 11 12

Time (hours)

FIGURE 24.1 Disappearance of naphthoic acid in cattail mesocosms A zero-order fit is appropriate after an induction period (R2 =

0.99) (Data from Kadlec and Srinivasan (1995) Wetland treatment of oil and gas well wastewaters Final Report DOE/MT/92010–10 (DE95000176), U.S Department of Energy: Bartlesville, Oklahoma; graph from Kadlec and Knight (1996) Treatment Wetlands First

Edition, CRC Press, Boca Raton, Florida.)

was increased from 102 to 104/g by bioaugmentation, and increased to 6 r 105/g) after incubation A high population of atrazine degraders (approximately 106 g−1) and enhanced rates

of atrazine mineralization also developed in bioaugmented sediment after incubation in flooded mesocosms planted with

cattails (Typha latifolia) and with atrazine at 3.2 mg/L Runes

et al (2001) concluded that bioaugmentation could be a way

to enhance constructed wetlands However, the persistence of degraders was not determined, and could be a constraint

W ILLOW W ETLANDS WITH Z ERO D ISCHARGE

The growth habit of willows allows their use to provide zero-discharge, zero-percolation treatment for small house-holds There are approximately 1,400 willow systems in use

in Denmark, and these have been described in Chapter 3

They rely upon an excess of evapotranspiration (ET) over

precipitation, which would not provide acceptable water

dis-posal for regional values of ET However, willow wetlands can be configured to enhance ET losses by using linear lay-

outs in a crosswind aspect (see Figure 3.27) Willow systems are small, serving 30 PE or less, and are essentially “wind

rows” that multiply regional ET by a factor of 2.5 due to the

clothesline effect described in Chapter 4 The gravel beds that support the willows serve as a reservoir to accumulate wastewater during the wet seasons and periods During dry periods, the interstitial water is drawn down Therefore, siz-ing is determined by the water balance, which in turn is spe-cific to a particular climatic zone (Gregersen and Brix, 2001; Brix and Gregersen, 2002; Brix and Arias, 2005) Approxi-mately 120–300 m2 are required for a single household under Danish climatic conditions

Raw sewage is pretreated in a sedimentation tank, and the supernatant is pumped onto or just under the surface of the plastic-lined bed Treated effluent may be collected in a system of drainage pipes, and water may be recirculated back

to the pumping well or the sedimentation tank Zero-discharge

willow systems are constructed with a width of eight meters

and a depth of 1.5 meters (Gregersen et al., 2003) A

drain-age pipe is placed at the bottom of the bed, which is used to

purge salt water from the bed after salinity rises too much

One third or one half of the willows are harvested every year

to keep them in a young and healthy state with high

transpi-ration rates (Brix and Arias, 2005) Harvesting at such small

scale can readily be accomplished by hand labor

The willow systems will meet all regulatory

require-ments because there is no outflow to any receiving water

bodies Willow systems may also be used in conjunction with

soil infiltration The willows will evaporate all wastewater

during the growing season, but during winter there is partial

discharge that is infiltrated into the soil

E NGINEERED P LANTS

In previous chapters it has been shown that the particular

species of plants used in treatment wetlands is not critical,

provided they create and sustain the desired biogeochemical

cycles However, there are differences among plants for some

pollutants, most notably for trace metals and trace chemicals

Thus, in addition to the possibility of selection of

hyperaccu-mulators, there is the longer-term possibility of genetic

engi-neering for the desired traits

In recent years, much attention has been focused on the

improvement of plants for the purpose of enhancing

phytore-mediation through genetic engineering One approach is to

influence the enzymatic pathways for assimilation and

vola-tilization by overexpressing genes of rate-limiting enzymes

in plants (Berken et al., 2002) Another approach is to

intro-duce additional metabolic pathways for hyperaccumulators

In this way the capacity of plants to take up, accumulate, and

volatilize compounds can be increased beyond that of any

naturally occurring plant species

An ideal plant for environmental cleanup can be

envi-sioned as one with high biomass production, combined with

superior capacity for pollutant tolerance, accumulation, and

degradation With genetic engineering, it may be feasible to

manipulate a plant’s capacity to tolerate, accumulate, and

metabolize pollutants and thus create an improved plant for

environmental cleanup Pilon-Smits and Pilon (2002) found

that mercury volatilization and tolerance was increased by

genetic engineering approaches by two- to threefold more

metal per plant Nandakumar et al (2005) found that Typha

latifolia could be genetically transformed by

Agrobacte-rium tumefaciens to enhance phytoremediation enzyme

production

Heaton et al (1998) proposed the use of transgenic

aquatic, salt marsh, and upland plants to remove available

inorganic mercury and methylmercury from contaminated

soils and sediments They noted that plants engineered with

a modified bacterial mercuric reductase gene were capable

of converting Hg(II) taken up by roots to the much less toxic

Hg(0), which is volatilized from the plant Plants engineered

to express the bacterial organomercurial lyase gene were

capable of converting methylmercury taken up by plant roots into sulfhydryl-bound Hg(II)

Results from lab and greenhouse studies are beginning

to show promise for transgenics, but field studies are needed

to establish their treatment potential, competitiveness in the wetland environment, and risks associated with their use

A RTIFICIAL E NCLOSURES

One extreme example of environmental modification is to place the treatment wetland in an artificial environment In fact, this is commonly done for bench- and pilot-scale treat-ment units because the convenience of working indoors (especially in cold climates) outweighs added operational costs such as artificial lighting (Figure 24.2)

In larger-scale applications, greenhouses have been used

as artificial enclosures to maximize the use of available sunlight (Figure 24.3) Due to the high cost of the building envelope; greenhouse systems are designed as mechanical activated-sludge processes (U.S EPA, 1997), although float-ing plant racks are used on top of the aeration tanks Plants used in the system may contribute to more stable operation due to retention of biosolids, and lower waste-activated sludge production has been reported (Austin, 2001)

FIGURE 24.2 Indoor bench-scale unit treating landfill leachate;

White Bear Lake, Minnesota Two 1,000-Watt lights are required

to provide the necessary light intensity to support plant growth

in these four mesocosms Care must be taken to ensure that the lights do not cause the plants to dry out or burn This setup was later modified by constructing a mini-greenhouse over the reactors using a PVC frame and a clear plastic sheet.

The energy input associated with the activated-sludge

process is quite high compared to treatment wetlands (see

Table 1.1), and if the system is located in a cold climate,

heat-ing demands can significantly increase energy costs (Brix,

1999)

Due to the high capital and operating costs associated

with indoor systems, they have been used in applications

where ancillary benefits such as aesthetics, visitor access,

and education are at a premium

24.2 CHEMICAL ADDITIONS

Wetlands do not necessarily supply sufficient reactants that

may be desirable or required to foster a particular removal

process Those additional reactants may be supplied via the

media for either FWS or SSF wetlands, or they may be

sup-plied as additional chemical streams entering the wetland

R EACTANTS VIA M EDIA

The selection of media to support specific chemical reactions

is a design issue In the case of FWS units, it is common to

ensure that adsorption capacity is present by the

incorpora-tion of organic material in the rooting medium Peat,

com-post, or organic soils components also condition the soil and

promote easier plant establishment and maintenance In the

case of HSSF systems, a much wider choice can be

consid-ered, but typically materials are in three categories:

Sorbents for phosphorus

Reactive media for metals, including

alkalinity/sul-fide generators to generate precipitation reactions

Materials that are designed to decompose and

gen-erate a source of organic carbon for denitrification

or sulfate reduction

Light expanded clay aggregates, zeolites, ferrous or calcitic

substrates, or organic materials can all be considered Their

phosphorus binding potential has been discussed in Chapter 10

•

•

•

FIGURE 24.3 Greenhouse-based treatment system near Land O’

Lakes, Wisconsin In the winter, considerable heating costs are

incurred to minimize snow loads on the roof.

Metals typically will bind to organics, and therefore mine drainage applications incorporate some form of compost as a layer in their substrate This zone of wetland bed also func-tions anaerobically to reduce sulfate to sulfide, which then precipitates metals (see Chapter 11) Mulch applied to the top

of an SSF wetland bed can be used to supply organic carbon

in excess of the plant biogeochemical cycle to fuel reactions such as denitrification if this is a process objective

All of these techniques of bed solids used as a cal addition have the implied need for periodic maintenance Sorbents become saturated, and surface mulches become depleted in soluble materials unless the plant geochemical cycle can produce sufficient organic carbon after start-up Therefore, these solid reactants must be replaced on some schedule, determined by the stoichiometry and loadings of the materials being treated Because of the cost and difficulty

chemi-of restructuring the bed, and reestablishing the wetland, the projected lifetime of a bed solid reactant typically needs to

be at least five to ten years Alternatively, reactant chemicals may be added as separate flows to the wetland

R EACTANTS VIA A DDED S TREAMS

Mechanical treatment plants often utilize chemical additions

to sustain particular portions of their treatment processes Aeration or oxygen addition fosters both biochemical oxy-gen demand (BOD) reduction and nitrification In plants that include denitrification, some means of resupplying the car-bon requirement is needed for conventional denitrification, such as utilization of an internal carbon-rich stream or the addition of an external carbon source (Metcalf and Eddy Inc., 1991) Removal of phosphorus in mechanical plants can involve the use of iron or aluminum salts to precipitate phos-phorus These process steps all have analogs in constructed wetland technology

The processing of nitrogen in a wetland requires oxygen for nitrification and carbon for conventional denitrification (see Chapter 9) Supplies of these necessary reactants may not be sufficient to support the full requirements of required removals Thus, HSSF wetlands are usually limited in their nitrification capabilities, and VF wetlands are usually limited

in their denitrification capabilities The capacity of a wetland

to foster nitrogen reduction may be increased by inclusion of stream additions of carbon or oxygen, often at some consid-erable operational expense The cost of such an enhancement should be weighed against other options, including the capi-tal cost of increasing the size of the system if necessary to be

closer to the Type A end of the spectrum.

Aeration of FWS Wetlands

Although the entire water surface of a FWS wetland is exposed to air, reaeration is still subject to mass transfer resistances as discussed in Chapter 5 Oxygen supply may therefore be a constraint on BOD reduction and nitrifica-tion It has been supposed that open-water areas may con-tribute to enhanced oxygen supply (U.S EPA, 2000a), but

that presumption is in question based on operating data

(Kadlec, 2003d; 2005e) Therefore, forced-air diffusers have

been considered for augmenting FWS performance, although

such attempts have been few At West Jackson County,

Mis-sissippi, air injection was attempted in the deep-zone cross

trenches No conclusive data was obtained because nutria ate

the plastic air diffusers

Ouray, Colorado

Situated at 2,365 meters above sea level, this municipal

con-structed wetland is surrounded by the San Juan Range, the

youngest and steepest range of the Rocky Mountains

(Camp-bell and Ogden, 1999; HDR/ERO, 2001) There is an average

annual flow of about 1,500 m3/d, higher in summer during the

tourist season The wetlands are arranged in two parallel paths

of three cells each Each path is 21 r 143 m, for a total system

area of 6,100 m2 Vegetation in all six cells is 80 to 90%

cat-tail (Typha spp.), with the exception of Cell 3, which is 80%

duckweed (Lemna minor) and 20% cattail (Typha latifolia).

Other species present include common reed (Phragmites

aus-tralis), curly dock (Rumex crispus), and lady’s thumb

(Polygo-num persicaria) Because of concerns of sulfate reduction in

the wetland inlet, the inlet deep zone of the wetlands is

aer-ated with submerged perforaer-ated air distributors

The water quality from the aerated lagoon pretreatment is

approximately secondary The wetlands reduced the

concen-trations of BOD to 4 mg/L, and total suspended solids (TSS)

to 6 mg/L, over the period 1996–2000 These are clearly

near background, and thus it is not possible to assign any

advantage or disadvantage to aeration in these reductions

Ammonia in the wetland influent was not monitored, but the

wetland effluent ammonia was 9 mg/L over the same

four-year period This is a high concentration, which exceeds that

expected for an aerated lagoon alone, and therefore no large

advantage could have been achieved in the wetland aeration

for ammonia reduction

Carbon Additions to FWS Wetlands

Under many circumstances, denitrification in FWS wetlands

is not carbon limited The decomposing dead biomass is a

moderately large carbon supply, although only a fraction of

that material is useable (see Chapters 9 and 17) However,

some waters contain large enough concentrations of nitrate

that removal is constrained by the carbon supply For instance,

Stengel et al (1987) worked with tap water at 30 mg/L nitrate

nitrogen and augmented that to 110 mg/L to emulate

ground-waters in northern Germany It was concluded that, under

some circumstances, insufficient carbon would be present to

support full denitrification

As noted in Chapter 9, a C:N ratio of 5 or 10 to 1 has been

deemed to be required for denitrification This ratio

deter-mines the potential need for an external carbon supply For

instance, for a HLR = 10 cm/d and a nitrate concentration

of 50 mg/L, the NOx-N load is 5 g/m2·d Decomposition of

1,000 g/m2·yr (dry weight) of biomass would yield about 500

gC/m2·yr, or 1.4 gC/m2·d Thus, the C:N ratio of 1.4:5 is far less than needed to support full denitrification

Lamb–Weston, Connell, Washington

The Lamb–Weston company built an integrated natural tem for treating and reusing wastewater from a potato process-

sys-ing facility near Connell, Washsys-ington (Kadlec et al., 1997; Burgoon et al., 1999) Two 1-ha free water surface wetlands

were constructed in series for denitrification of nitrified ent from VF beds (intermittent sand filters) (see Figure 24.18) The wetlands were designed to remove NO3-N from waste-water prior to land application During a five-month period of intensive study, the flow was 4,950 m3/d, producing a direct hydraulic loading of 25 cm/d When coupled with an inlet nitrate nitrogen of 45 mg/L, the denitrification wetlands received a very high nitrate load of 110 kg/ha·d The chemical oxygen demand (COD) in the nitrified water was insufficient

efflu-to support denitrification, even with the carbon generated

by the wetland plants Therefore, a small flow of untreated potato wastewater (COD ≈ 3,000 mg/L) was directed to the wetlands as a supplemental carbon supply This flow raised the hydraulic loading rate (HLR) to 28 cm/d Addition of this primary effluent resulted in a COD:NO3-N mass load ratio that ranged from 10 to 25 The NO3-N mass removal rate ranged from 50 to 99% of the influent load During the five months January to May 1998, average monthly effluent

NO3-N was 1.6 mg/L, and an average of 85% of the nitrate load was removed (Burgoon, 2001) Thus the feed-forward

of untreated water provided the necessary carbon to allow nearly complete denitrification at high loading rates

Sources of carbon are not always conveniently at hand Some wetlands have utilized methanol or molasses, but these are expensive Multiple wastewater sources can potentially provide synergism For example, the remediation of a high-nitrate source water could be enhanced in a wetland that also received high BOD municipal or food processing wastewa-ters At the time of this writing, such multisource systems have been proposed but not built

Aluminum and Iron Additions to FWS Wetlands

The phosphorus-binding capabilities of iron and aluminum, and their role in the wetland biogeochemistry of phospho-rus, are very well known (Reddy and DeLaune, 2007) Very early in the history of FWS technology development, it was observed that naturally occurring, blending streams of water containing iron were very effective in phosphorus removal

At Waitangi, New Zealand, phosphorus deposition occurred predominantly by reaction between treatment wetland waters (high in phosphorus, alkaline pH) with natural wetland waters (high in Fe and Al, acid pH) immediately below the confluence of their flow streams (Cooper, 1992; 1994).This study, coupled with the knowledge of mechanical plant phosphorus precipitation technology, led to investiga-tion of the direct addition of these metal salts to FWS wetlands for the purpose of enhancing phosphorus removal (Bachand

et al., 1999; Bachand and Richardson, 1999) Preliminary jar

test results suggested that treatment with ferric chloride and

alum could effectively remove phosphorus to between 8 and

30 µg/L (Metcalf and Eddy Inc., 1997) It was hoped that

chemical amendments directly within the treatment

wet-lands would create a synergistic effect in which phosphorus

is more effectively and efficiently removed than by either

of these two methods, chemical amendments and wetlands,

alone It was also hoped that addition of a polymeric

coagu-lant might not be necessary because the wetland would

effec-tively serve as a settling basin with very long HRT However,

small flocs formed, which did not quickly settle Total

phos-phorus concentrations in the water column remained elevated

This approach was therefore abandoned as a means of

enhanc-ing wetland operational effectiveness

Aeration of SSF Wetlands

Here for the first time we encounter a significant influence

of patents on SSF treatment wetland technology Aeration of

HSSF systems has developed both in the open literature and in

the proprietary sector Aerated subsurface flow wetlands are

available as a patented technology in North America (Wallace,

1998; Dufay, 2000; Wallace, 2001b; Wallace, 2002a; Flowers,

2003; Wallace and Lambrecht, 2003) We cannot access all

the proprietary data, and hence cannot independently

evalu-ate it For most pevalu-atented systems, only anecdotal, qualitative

information has been presented However, we can summarize

the results from the open literature and provide interpretation

of it Nothing here should be taken as supporting or refuting

the claims in existing patents

The growing realization in North America during the

mid-1990s that HSSF wetlands were inherently oxygen-

transfer-limited systems led to considerable interest in

artifi-cial aeration through the application of a mechanical blower

and diffuser tubing as a means to boost treatment

perfor-mance (especially for ammonia) in these systems One of the

most basic applications of the technology is shown

schemati-cally in Figure 24.4:

Coarse media

Impermeable liner Main bed media

Water level

Mulch layer Air header Air line

Air

FIGURE 24.4 Schematic of an aerated HSSF wetland.

In the most simplistic sense, this can be thought of as ing bubbles” inside a saturated-flow HSSF wetland However, implementation of the technology requires an understanding of the frictional loss of compressible fluids, hydrodynamics, and oxygen transfer theory Nevertheless, a variety of successful application areas have been developed, including petroleum hydrocarbons (Wallace and Kadlec, 2005), ammonia (Wallace

“blow-et al., 2006a), domestic wastewater (Wallace “blow-et al., 2006b), and airport de-icing runoff (Wallace et al., 2007a).

The decision to aerate a SSF wetland comes at a heavy operational cost, as operation and maintenance (O&M) of the facility is greatly increased From an economic viewpoint, aeration is only justified when the lifecycle cost of aeration

is sufficiently offset by the reduction in capital cost, as the net savings of reduced wetland size less the cost of aeration equipment

Aerated wetlands typically occupy a much smaller print than a nonaerated wetland, which translates into lower construction cost Or, sufficient land may not be available for

foot-a more pfoot-assive Type A wetlfoot-and, or not foot-avfoot-ailfoot-able foot-at foot-an

eco-nomically feasible price

At the time of this writing, performance information on aerated HSSF wetlands is a mixture of forensic analysis of full-scale systems, as well as analysis of performance data from pilot-scale systems or mesocosms

In the first generation of aerated wetland design, tions from other treatment technologies, such as lagoons, were often “borrowed” to support blower sizing and diffuser-tube placement in HSSF wetland systems As a result, these systems were not optimally designed In recent years, there has been an increased focus on pilot-scale systems, especially for industrial effluents A well-designed and implemented pilot study can produce data on rate coefficients (essential for first-order modeling), but questions about the hydrodynamic aspects of the pilot reactor versus the proposed full-scale sys-

assump-tem often remain (Noorvee et al., 2005b).

Apart from reaction rates, there is the issue of oxygen transfer rates These depend upon bed depth, interstitial

oxygen concentrations, diffuser geometry, and other factors

However, relatively gentle aeration can deliver something

on the order of 50 to 100 gO/m2·d to the HSSF bed This

rate of oxygen delivery is significantly greater than rates

that can be inferred from Type A nonaerated HSSF wetlands

(Tables 9.26, 9.27, Chapter 9), which are usually in the range

of 1 to 6 gO/m2·d

The energy requirement for such gentle aeration can

nev-ertheless be an important factor in overall energy usage (see

Table 1.1, Chapter 1), especially compared to passive Type

A wetlands, but is generally much lower than conventional

mechanical treatment processes

Domestic Wastewater Treatment

In a recent study, Wallace et al (2006b) compared the

treat-ment efficiency of 17 full-scale aerated HSSF wetlands and

22 full-scale nonaerated wetlands, using the loading chart

and statistical methods presented in Wallace and Knight,

2006 The conclusion of this study was that aeration resulted

in significant improvements in the reduction of both BOD5

and TSS With aeration, the equivalent wetland size needed

for BOD5 removal could be reduced by 67%; for TSS, the size

reduction was approximately 36%

Ammonia Oxidation

Aeration of SSF wetlands for ammonia removal has been

studied in recent pilot-scale systems For an effluent containing

a mixture of cyanide/iron/ammonia from a gold mine in South

America, aeration was found to increase the ammonia

oxida-tion rate over tenfold, from a volumetric rate (2 TIS) of 0.52

d−1 to 5.7 d−1 (Wallace et al., 2006a) A study of ammonia

oxidation at different temperatures for domestic wastewater

at a pilot-scale wetland in Canada (Wallace et al., 2006a)

yielded a temperature correction factor (Q) of 1.02,

indicat-ing that temperature sensitivity is of some importance in

aerated wetlands (compare with Table 9.17, Chapter 9) This

would be consistent with a view that heavily loaded aerated

HSSF wetlands are microbially dominated ecosystems

De-icing Runoff

Treatment of de-icing compounds (discussed in Chapter 13)

has been studied under aerated and nonaerated conditions in

HSSF mesocosms For removals of CBOD5 associated with

de-icing runoff, the volumetric rate coefficient was measured

at 0.6 d−1 (2 TIS); aeration at 0.85 m3 air per hour (per m3 of

wetland bed) improved the rate to 5.5 d−1 Additional aeration

at a rate of 2.0 m3 air per hour (per m3 of wetland bed) further

increased the degradation rate coefficient to 15.9 d−1 Based

on the range of temperatures studied (22 to 4°C), the

tem-perature correction factor was 1.03 (Wallace et al., 2007a).

Petroleum Hydrocarbons

Aeration has been demonstrated to improve removals of

petroleum hydrocarbons through aerobic degradation and

volatilization Early work was conducted at the Gulf

Stra-chan Gas Plant in Alberta, where an aeration system initially

designed for freeze resistance during winter operations resulted in improved removals of total petroleum hydrocar-bons (TPH) and benzene, toluene, ethylbenzene, and xylene

(BTEX) compounds (Moore et al., 2000a) A full-scale

aer-ated wetland system was constructed by Williams Pipeline for a terminal facility in Watertown, South Dakota, and has been operating successfully (Wallace, 2002b)

Pilot-aerated SSF wetland studies were conducted by

Ferro et al (2002) for the purpose of reducing petroleum

hydrocarbons in extracted groundwater Both aerated and nonaerated degradation rates were measured, as summarized

inTable 13.2, Chapter 13 (Wallace and Kadlec, 2005) These findings are consistent with the growing body of knowledge

of biodegradation and volatilization rates for hydrocarbons under aerobic and anaerobic conditions (Suarez and Rifai, 1999)

Aquaculture Waste

A study at the Botanical Garden of Montreal investigated the effects of aeration in mesocosms for the treatment of aqua-

culture waste (Ouellet-Plamondon et al., 2006) Aeration

improved the removal of TSS, COD, and total Kjeldahl gen (TKN) For instance, the first-order COD rate coefficient (plug flow) improved from 0.15 to 0.25 m/d for unplanted

nitro-mesocosms, and from 0.19 to 0.24 m/d for Typha-planted

mesocosms Aeration was only used in the inlet region of the HSSF bed The researchers concluded that aeration was a promising approach for improving the performance of HSSF wetlands

Oxygen Transfer

A recent study was conducted to measure the standard oxygen transfer efficiency of aerated SSF beds using off-gas measure-

ments (Wallace et al., 2007b) Oxygen transfer is a function

of the contact time between the bubble and the water column, which depends on the water depth It also depends on constit-uents within the wastewater that affect oxygen mass transfer

at the bubble/water interface (Metcalf and Eddy Inc., 1998) The average oxygen transfer efficiency in a HSSF wetland mesocosm using a standard sodium sulfite solution was 4.7% per meter of water depth As a frame of reference, coarse bubble diffusers have a transfer efficiency of approximately 2.6% per meter, and fine bubble diffusers have efficiencies

of approximately 6.6% per meter (U.S EPA, 1989) Because most HSSF wetlands have a bed depth of 0.3 to 0.6 m, this would put the oxygen transfer efficiency in the range of 1.4 to 2.8%, depending on the bed depth

The linear diffuser tubing used in this study was a fine bubble aeration device In an open-water tank, the expected oxygen transfer would be approximately 6.6% per meter The lower-than-expected oxygen transfer rate could speculatively

be from the influence of the granular media, which may serve

to coalesce bubbles and increase their size, reducing the oxygen transfer efficiency

Air-induced mixing of the wetland was greatly reduced

by the presence of the granular bed media; based on visual

observations, the influence of aeration was generally

restricted to a radius of approximately 15 cm from the point

of bubble introduction (Wallace et al., 2007b) Aerated

wet-lands are hydraulically inefficient, as shown by tests yielding

number of tanks in series (NTIS) of 1 to 3 It is apparent that

a very uniform aeration grid is required to maintain

oxygen-ated conditions in the wetland bed Figure 24.5 shows the test

of such an aeration system

Aeration of SSF wetlands is not a maintenance-free

activ-ity Without aeration, the system will fall back into the

oper-ating envelope of nonaerated wetlands However, because

regulatory compliance is dependent on the aeration system,

scheduled, proactive maintenance of blowers and other

aera-tion components need to be carried out by the system

opera-tor Wetland designers should also consider the fouling of air

diffusers within SSF wetlands, and provisions to replace or

chemically clean diffuser assemblies should be addressed

Carbon Additions to HSSF Wetlands

Early work on denitrification in HSSF wetlands was

con-ducted by Gersberg et al., 1984 HSSF wetlands were fed

with secondary effluent low in BOD (3.3 mg/L) and high

in nitrate (18.8 mg/L) In initial trials, the removal of total

nitrogen was limited by the availability of organic carbon,

and ranged between 9 and 19% at hydraulic loading rates

of 8.4 to 20 cm/d Utilization of methanol as a supplemental

organic carbon supply dramatically increased denitrification

efficiency, up to 97% at a hydraulic loading rate of 16.8 cm/d

(Gersberg et al., 1984) As a lower-cost carbon supply, plant

biomass was mulched and placed on top of the beds At a

hydraulic loading of 8.4–12.5 cm/d, the mean removal rate

was 89% for total nitrogen, although the carbon loading (0.09

kg/m3) was higher than that required for methanol (0.03 kg/

m3) This finding is consistent with more recent studies on

the bioavailability of plant detritus to fuel denitrification (see,

for instance, Ingersoll and Baker, 1998; Baker, 1998; Hume

et al., 2002a).

FIGURE 24.5 Testing of HSSF wetland aeration system for BTEX degradation; Casper, Wyoming Rows of bubbles represent the location

of linear air diffuser tubes, which were placed in radial arcs within this circular wetland The wetland basin was artificially flooded to allow visual observation of the bubble pattern.

Use of organic materials as a thermal insulator has the drawback of producing leachates with BOD that adds to the loading for that contaminant in domestic and municipal sys-

tems (Wallace et al., 2001) However, in systems treating

low BOD sources, such as many runoff and groundwater remediation projects, such an additional source of organics

is desirable and may be essential to the proper functioning of the wetland Surface mulching of HSSF wetlands is also an option for supplying carbon for denitrification of nitrate-rich

influents, as demonstrated by Gersberg et al., 1984

Mass-balance considerations to fuel denitrification processes are discussed in Chapter 9

VF wetlands are effective in oxidizing ammonia to nitrate, but typically are not efficient in denitrification pro-cesses (Cooper, 2001) One of the limitations of denitrifica-tion is availability or organic carbon; and VF wetlands are often combined with HSSF wetland stages for this purpose (Seidel, 1973; Brix, 1994d; Cooper, 2001)

24.3 OPERATIONAL STRATEGIES

The flow of water need not be continuous, nor need it be directional from inlet to outlet of a wetland or integrated nat-ural system There are other methods of operation, involving spatial or temporal sequencing, and the potential of recycle flows Some of these concepts are explored in the following text

uni-S TEP F EED

Step feeding refers to the introduction of water to a system

at several points along the flow path In activated-sludge mechanical plants, it refers to the introduction of settled wastewater at several points along the flow in a plug-flow aer-ation tank (Metcalf and Eddy Inc., 1991; Crites and Tchob-anoglous, 1998) In water hyacinth systems, as many as eight points of introduction have been used (Water Environment Federation, 2001) The Gustine, California, FWS system was

built with provisions for step feeding (U.S EPA, 1988b) The

reason for step feeding is, in all cases, stated to be the

avoid-ance of “overloading” of the inlet zone of the system Because

the concept of loading to an interior zone of the wetland is

fuzzy, a hypothetical situation is considered

The Gustine example is selected as the framework, for

which a step feed at the one-third point of the length is

pos-sible (Figure 24.6) The original design called for either one

third or two thirds of the flow to be introduced at this interior

point (U.S EPA, 1988b) It is known that the decline of BOD

through this system is exponential (see Figure 8.7, Chapter

8) Because of the high length-to-width ratio (337 m:11.6 m =

29:1), the plug-flow version of the k-C* model will be used

Calibration to the Gustine data of Walker and Walker (1990)

yields k = 48 m/yr and C* = 4 mg/L The average BOD

enter-ing the system was about 200 mg/L, and the hydraulic

load-ing was 3.3 cm/d Figure 24.7 shows the effect of the step

feed of one third of the influent to the one-third point The

front end of the wetland receives a lesser flow and does a

more effective job of reduction of concentration Upon

blend-ing with the step feed, the concentration goes up above that which would occur without step feed and stays above for the rest of the wetland travel distance The outlet concentration

is 29% higher for the step-feed option It may be shown that for any step-feed operation and first-order kinetics, step feed gives lesser performance

The overall BOD load to the wetland is not changed by step feeding because the area is fixed, and the inlet flow and concentration are fixed The question then becomes: what has been accomplished? If all that is considered is BOD, the answer is, nothing However, if it is presumed that part of BOD removal has to do with the sedimentation of solids, then TSS accumulation might factor into our thinking The Gus-tine system reduced TSS from 75 mg/L to about 30 mg/L Suppose that the entire accretion is to an inlet zone of 37 m length (1/9 of the length), at a bulk density of 0.5 Then the buildup would be about 1 cm per year By step feeding, one-third of that accretion would be transferred to the step-feed zone, and the inlet buildup would be only two thirds of a cen-timeter per year Therefore, the evaluation of step feeding as

FIGURE 24.6 The step feed distribution pipe in the Gustine, California, system.

FIGURE 24.7 Effect of step feed on BOD concentrations These are model-based calculations, based on calibrations to the Gustine,

California, system.

1 10 100 1,000

Distance (m)

No Step Step

an alternative depends upon the necessity for spreading out

the solids and the corresponding penalty in performance

R ECYCLE

Recycle can serve two important purposes in treatment

wet-lands: (1) it brings products of reactions in the wetlands from

the exit back to the inlet, and (2) it dilutes reactants entering

the system Both of these functions can be used to advantage

in some wetland situations

Dilution of Influents

Wetland vegetation can be sensitive to high concentrations

of some pollutants in the water Therefore, strong influents

can damage or destroy vegetation, and the problem will be

most severe in the inlet of the wetland before any treatment

has occurred For example, landfill leachates can be

inimi-cal to wetland vegetation (Figure 24.8) Recycling treated

water from the wetland system outlet can serve to dilute the

incoming water, but at some price in terms of cost and level

of treatment

FIGURE 24.8 A leachate seepage outbreak at the Saginaw, Michigan, closed landfill This water contained ammonia at concentrations

sometimes exceeding 1,000 mg/L, and was severely toxic to all vegetation.

0 20 40 60 80 100 120

FIGURE 24.9 The effect of recycle on a hypothetical wetland The fresh-feed Damköhler number is k/q = 2.0, P = 4, and C* = 0.

To illustrate the concept, a hypothetical wetland is sidered, with a recycle of water back to the inlet A fresh-

con-feed Damköhler number k/q = 2.0, P = 4 TIS, and C* = 0

are chosen, to be representative of a moderate degree of removal of the contaminant (80%) Figure 24.9 shows the calculated concentration profiles through the system for dif-fering amounts of recycle up to 200% of the incoming raw water Note that as the recycle ratio goes to very high levels, the effect is to create a single well-mixed unit because of the intense recirculation The limiting concentration reduction for this extreme is 67% Typical recycle rates are 200% or less, for which the reduction is 72% for this example Thus, a price in performance has been incurred by recycle However, the concentration in the inlet region is cut in half (52%) by 200% recycle

The economic consequences of a recycle pump and ing may not be large for very small systems, but for large FWS wetlands they may be a significant contribution to cost Recycle is being used on the large Everglades-con-structed wetlands to process seepage collected around their perimeters

pip-Recycled Nitrate

One of the dilemmas of once-through treatment wetlands is

that carbon (BOD or COD) is utilized in the front end of the

wetland, prior to the nitrification of a significant fraction of

any ammonia that may be present This is due to the more

rapid rate of elimination of BOD, compared to ammonia

Thus, although both processes occur more or less

simultane-ously in wetlands, the carbon is gone before a large share of

the nitrate is formed Therefore, denitrification is hampered

by a lack of carbon in the exit section of the wetland This

phenomenon is present in all types of treatment wetlands but

is especially problematic in VF systems One “cure” is to add

another wetland component with the capability of providing

the carbon needed, which would usually be a FWS system

with its litter decomposition as the carbon supply Another

possibility is to recycle the nitrate back to the inlet of the

wetland, where there is a supply of carbon associated with

the incoming fresh feed to the system

For example, Brix et al (2002a) performed

compari-son studies on a two-stage VF system treating municipal

wastewater, with and without recycle (Figure 24.10; see

also Figure 16.11, Chapter 16) Table 24.1 shows data from one pair of two-month campaigns, out of four that were conducted Several conclusions can be drawn from these results First, the wetlands are very effective in reducing BOD5 and TSS, whether or not there is recycle The sedi-mentation tank effectively reduces organic nitrogen, and the wetlands reduce it further to levels less than 1 mg/L, whether or not there is recycle The wetlands are very effec-tive in reducing ammonia, whether or not there is recycle The major effect of recycle is in the reduction of oxidized nitrogen and consequently total nitrogen leaving the system The effluent oxidized nitrogen was 38 mg/L without recycle, and 18 mg/L with recycle

Because of the success of this strategy, the Danish lines for small on-site treatment wetlands require recycle at a recycle ratio of 50% (Brix and Arias, 2005)

guide-FWS T IMED O PERATIONAL S EQUENCES

The flows to FWS wetlands need not be steady and ous Event-driven systems constitute one group of systems that receive nonsteady flows, and these have been discussed

continu-Vertical flow wetland #2

Vertical flow wetland #1

Sedimentation tank

FIGURE 24.10 Basic components of the experimental recycle VF system in Denmark (Adapted from Brix et al (2002a) BOD and Nitrogen

Removal from Municipal Wastewater in an Experimental Two-Stage Vertical Flow Constructed Wetland System with Recycling Mbwette

(Ed.) Proceedings of the 8th International Conference on Wetland Systems for Water Pollution Control, 16–19 September 2002; Comprint International Limited: University of Dar Es Salaam, Tanzania.)

TABLE 24.1

Comparison of Recycled and Once-Through Operation of a Two-Stage Experimental System

in Trige, Denmark

Source: Data from Brix et al (2002a) BOD and Nitrogen Removal from Municipal Wastewater in an Experimental Two-Stage

Vertical Flow Constructed Wetland System with Recycling Mbwette (Ed.) Proceedings of the 8th International Conference on

Wetland Systems for Water Pollution Control, September 16–19, 2002; Comprint International Limited: University of Dar Es

Salaam, Tanzania.