The Brighton treatment wetlands

Ecological Engineering 47 (2012) 56–70

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The Brighton treatment wetlands Robert H. Kadlec a,∗ , John Pries b , Keith Lee c a

Wetland Management Services, 6995 Westbourne Drive, Chelsea, MI 48118, USA CH2M HILL, 72 Victoria Street South, Suite 300, Kitchener, Ontario N2G 4Y9, Canada c Municipality of Brighton, 67 Sharp Road, Brighton, Ontario K0K 1H0, Canada b

a r t i c l e

i n f o

Article history: Received 27 March 2012 Received in revised form 12 June 2012 Accepted 23 June 2012 Available online 20 July 2012 Keywords: Treatment wetlands Nutrients Pathogens Ancillary benefits Cold climate

a b s t r a c t The Town of Brighton, Ontario implemented a 6.2 ha marsh in 2000, for the purpose of improving water quality before discharge to receiving waters. The wetlands have successfully operated in this moderately cold climate for over ten years. Phosphorus removal of 2.3 gP/(m2 yr) was achieved, with an annual areal rate coefficient of 9.2 m/yr. The removal is strongly seasonal, with the greatest reductions occurring in spring. The total nitrogen loading was dominated by ammonia (208 gN/(m2 yr)), with smaller amounts of organic and oxidized nitrogen. Ammonia was reduced to 173 gN/(m2 yr). Implied areal rate constants were high for mineralization of organic nitrogen (29 m/yr) and denitrification (101 m/yr), but low for nitrification (4 m/yr). CBOD5 was reduced from 5.4 to 3.2 mg/L, and TSS was reduced from 13.2 to 7.2 mg/L, both with slightly higher values during late winter. The wetland was not effective in reducing pathogens, with Escherichia coli at 167 cfu/100 ml entering, and 132 cfu/100 ml leaving. Vegetation was sparse, likely due to muskrats and deep water. Macro-invertebrate diversity was lower than for regional wetlands. Bird use was very high, and birding was a popular human activity. The wetland has been designated as provincially significant. © 2012 Elsevier B.V. All rights reserved.

1. Introduction Lagoon or pond treatment is a frequently used option for wastewater treatment for small communities, with over 7000 systems in use in the USA (Crites and Tchobanoglous, 1998). Most of these were designed for the reduction of biochemical oxygen demand (BOD5 ) to about secondary levels, 30 mg/L. Total suspended solids (TSS) are also reduced, compared to entering raw wastewater; but ponds function via algal activity, which creates solids. Those solids are difficult to remove, and additional processes are often needed to upgrade lagoon effluents (Walmsley and Shilton, 2005). One option is the addition of treatment wetlands downstream of the lagoon, which has been in use for several decades (Polprasert et al., 2005). Nutrient control is increasingly being required of wastewater treatment facilities. Facultative or aerated lagoons are not very effective for phosphorus removal, and by themselves are not capable of meeting phosphorus discharge standards in the Great Lakes region, which are often lower than 1.0 mg/L. Total phosphorus (TP) may be removed by chemical precipitation, or by biological activity in follow-on wetlands. Facultative ponds produce some degree

∗ Corresponding author. Tel.: +1 734 475 7256. E-mail address: [email protected] (R.H. Kadlec). 0925-8574/$ – see front matter © 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.ecoleng.2012.06.042

of nitrification, but are not good at reducing the resultant nitrate. Adding aeration to the lagoon facilitates some limited ammonia conversion to oxidized nitrogen, but that too leads to residual nitrates. Consequently, total nitrogen (TN) reduction is generally limited to algal uptake and sedimentation. The demand for improved treatment, together with the limitations of each individual treatment type, makes the selection of multiple unit systems inevitable in many situations. Several wetland categories: free water surface (FWS), horizontal subsurface flow (HSSF) and vertical flow (VF); are all being used as “standalone” wastewater treatment technologies (Kadlec and Wallace, 2009; Vymazal and Kröpfelová, 2008). They each have the potential to provide effective treatment of a variety of pollutants as long as loading rates are appropriate and the desired plant community is maintained (Browne and Jenssen, 2005). Although these technologies work alone, it is often beneficial to combine one or more of these options into integrated natural treatment systems. For example, constructed FWS treatment wetlands provide excellent reduction of TSS and denitrification of nitrate and nitrite nitrogen, but they are not extremely effective at nitrifying ammonium nitrogen. Wetlands possess the potential to be a valuable adjunct to treatment lagoons. They need not replace lagoons, because those aquatic systems provide reasonable treatment for many water quality parameters such as BOD5 and TSS. In North America, a combined

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system is in use, comprised of facultative or aerated lagoons or ponds for primary and secondary treatment, followed by a constructed free water surface wetland. Kadlec (2003) lists 21 such systems, and more have since been developed (see, for example, McKenna et al., 2002; Kadlec, 2008). However, this combination is also being adopted elsewhere in the world, such as New Zealand (Tanner and Sukias, 2003), and China (Wang et al., 2005). Pond wetland combinations may be further divided based upon seasonality of operation. In very cold climates, operation is only during the unfrozen season, but in more moderate climates, operation can be year-round. Winter storage may be in the treatment lagoon, or separate from it. TSS and BOD5 reductions in treatment wetlands are approximately independent of season, and for these the decision for winter storage involves only hydraulic concerns about freezing conditions. But, winter rate constants are about 1/3 to 1/10 the summer values for phosphorus and ammonia, respectively (Kadlec and Reddy, 2001). Thus, for nutrient removal, winter removal process slow-down adds to the hydraulic concerns. This paper describes the performance of the lagoon–wetland system at Brighton, Ontario, which operates year-round. Lagoon #1 employs aeration and chemical precipitation of phosphorus and lagoon #2 is a facultative pond. Prior to construction of the wetland, this wastewater system was meeting permit limits, but population growth predictions outstripped the lagoon design capacity. Wetlands were added to increase the rated capacity of the existing wastewater treatment system and to provide a performance enhancement to allow the system to continue to meet discharge targets to Lake Ontario.

2. Treatment system and goals 2.1. The Brighton wastewater treatment system Brighton’s Water Pollution Control Plant services a population of approximately 6297 consisting of 2800 residential and commercial accounts including Presqu’ile Park. The average daily wastewater flow to Brighton WWTP was approximately 0.49 m3 /capita in 2010, which includes domestic, industrial and commercial contributions as well as infiltration and inflow. This is based on a population of 6297 and an average daily flow of 3072 m3 /d. The population number includes an estimated 182 persons to account for the seasonal population of Presqu’ile Provincial Park. The sewage works in Brighton consists of a 0.68 ha aerated lagoon (lagoon #1) with two mechanical surface aerators. The effluent from the aerated lagoon passes through a chemical mixing chamber where alum is added before entering the 5.44 ha waste stabilization pond (lagoon #2). Effluent from lagoon #2 passes to the constructed wetlands. The wetlands have a total surface area of 6.2 ha, and consist of two parallel flow paths, each with alternating deep zones and vegetative terraces (Fig. 1). The effluent from the constructed wetland flow paths is combined, and discharged via a short surface stream to a natural wetland that borders Presqu’ile Bay located on the northeast edge of Lake Ontario. The aspect ratio of the flow paths is about 3.5 L:W. Each had deep zones at the inlet, outlet and two internal locations. These were excavated to a water depth of about 1 m to provide 1.3 m water depth, and a width of 6–8 m, to inhibit the growth of emergent macrophytes. The wetlands site was formerly agricultural land, owned by the Town, and typically planted in corn (Zea maize). The wetland cells were graded to provide material for berms and divider levees in early spring 1999. Planting of the wetland was by seeding in the fall of 1999, with aerial spreading at the east end of the north cell by local students and by hydroseeding using cattail (Typha

Fig. 1. Layout of the Brighton treatment wetlands and lagoons. Each wetland has two internal deep zones.

latifolia) seed heads collected from a local wetland. The surrounding wetland also provided a near continuous seed source supplementing the intentional seeding. During summer 2000, cattail and other plants and algae grew in water that was intentionally held at shallow depth. By August 2000, the wetland benches were completely vegetated with juvenile cattails, with a height of about 1 m. The deep zones remained free from emergent vegetation. The first litter formation resulted from the senescence of plants in the late autumn of 2000. 2.2. Regulatory considerations The treatment plant must meet Certificate of Approval (C of A) effluent concentration limits at the lagoon #2 discharge, set by the Ontario Ministry of the Environment (MOE), and the discharge following the constructed wetlands is expected to meet additional objectives (Table 1). Objectives rather than limits were assigned to the wetland discharge by the MOE to allow the Municipality to demonstrate the wetland technology for this level of contaminant reduction (NH3 -N and TP in particular). Limits and objectives are applied to monthly averages of weekly values for NH3 -N and TP. BOD5 and TSS must meet annual average values. Escherichia coli is sampled once a month from the waste stabilization pond for compliance purposes. Monthly E. coli measurements were also taken at the wetland outflow, although not an explicit objective. 2.3. Treatment wetland classification As a result of a regional wetland study by the Ontario Ministry of Natural Resources in 2006, the constructed wastewater treatment wetland was designated as a Class 1 wetland as part of the Provincially Significant Presqu’ile Marsh Wetland. That was a compliment

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Table 1 Certificate of approval limits for effluent parameters. Units are mg/L except E. coli cfu/100 ml.

CBOD5 Suspended solids Ammonia nitrogen (May 1 to October 30) Ammonia nitrogen (November 1 to April 30) Total phosphorus E. coli

Compliance frequency

Stabilization pond effluent limit

Wetland effluent objective

Annual Annual Monthly Monthly Monthly Monthly

30.0 40.0 14.0 17.0 1.00 200

15.0 15.0 10.0 15.0 0.80

to the ecological value of the system, but also carried potential consequences for the operation and maintenance of the treatment wetland. Certain activities, including modification and decommissioning of the wetland, would be considered “site alteration” under the Provincial Policy Statement (PPS). The Certificate of Approval for the wastewater treatment plant considers the wetland a component of the treatment works, and as such, subject to operations and maintenance requirements. Operations include varying the water level to optimize performance, reduce muskrat habitation, or raising the water level to maintain the hydraulic retention time during winter operations. Maintenance to constructed treatment wetlands can include vegetation mowing, dredging accumulated solids in deep zones, excavation to remove sediment buildup, and other. Under the PPS, a degree of control of the provincially significant wetland (PSW) now resides with the Ontario Ministry of Natural Resources (MNR). The PPS prohibits development and site alteration in a PSW, but changes in infrastructure are permitted within PSWs, provided it has been authorized under the Environmental Assessment Process. Infrastructure is defined as physical structures that form the foundation for development, and includes: sewage and water systems, septage treatment systems, and waste management systems. An MNR opinion stated that, provided approval under the Environmental Assessment Act was granted for the treatment plant, this approval would allow for ongoing operation and maintenance of this facility. However, the MNR also stated that it would be necessary to determine how these activities could be done with the least impact on the existing wetland values and functions. At the present time MNR opinion, as stated in correspondence with the Municipality, is that designation as a PSW does not restrict the O&M efforts at the Brighton wastewater treatment wetland site that are required by their C of A. 3. Methods

weather station, 17 km east of the wetlands (Environment Canada, 2010). Snowmelt was computed from snowfall and snow on the ground, with a water equivalent of one cm of water per 12 cm of snow. The catchment for precipitation was assumed to be 1.1 times the wetland wetted area. Evapotranspiration was estimated based on information for comparable climatic regions. Water quality sampling was conducted weekly at the wetland inflow and outflow. Routine water quality analyses included total phosphorus (TP), E. coli, carbonaceous biochemical oxygen demand (CBOD5 ), total suspended solids (TSS), total Kjeldahl nitrogen (TKN), total ammonia (NH3 N), nitrite (NO2 N), and nitrate (NO3 N), were measured by an accredited laboratory, Caduceon Environmental Laboratories, Kingston, Ontario. Oxidized nitrogen was calculated as the sum of nitrite plus nitrate nitrogen (NOX N = NO2 N + NO3 N). E. coli were sampled once a month at the effluents from the waste stabilization pond and the wetlands. Fieldmeasured parameters included pH and temperature, sampled at the inlet and outlet from the wetlands. Only spot checks were made for alkalinity and dissolved oxygen. Measurements of vegetation, invertebrates and water quality indices were made during summer 2007, for which methods and protocols may be found in McGauley (2008). 3.2. Seasonal cycles and water quality modeling Some wetland variables vary cyclically with season, and may therefore be represented by a sinusoidal trend model. For instance, temperature and dissolved oxygen follow such a pattern. The annual cycle of wetland water temperatures in mild to warm climates displays a summer maximum and a winter minimum. The onset of frozen conditions typically is accompanied by under-ice water temperatures of 1–2 ◦ C. The sinusoidal model, truncated for frozen conditions, is: For the unfrozen season (t1 < t < t2 ) :

3.1. Research and monitoring The wetland basins were completed by the end of 1999, and vegetation commenced in mid 2000. The wetland system was monitored from July 2000 onward. The second half of 2000 was deemed to be a startup period, in which plants and microbiological communities developed in the newly constructed wetland basins. The period of record (POR) for this analysis started in January 2001, and concluded with December 2010, to include ten full calendar years of operational data. Water flows were measured daily throughout the POR at the wetland inlet, with annually calibrated weirs. Water flows leaving the wetland were also measured during 2001, at the exit points for both the north and south cells. Wetland outflows were less than inflows by 6.7% (N = 132), which was deemed to be close to measurement accuracy, and within the data scatter (s.e. = ±7%). Accordingly, flow measurement was thereafter conducted only at the wetland inflow. Water depths were measured weekly in both wetland cells. Meteorological data (precipitation, snow depth, air temperature) was taken to be that for the Trenton A, Ontario

T = Tavg (1 + AT · cos[ω(t − tmax )])

For the frozen season (t2 < t < t1 ) :

(1)

T = To

(2)

where AT is the fractional amplitude of the sinusoid; t the time, Julian day; t1 the thaw time, Julian day; t2 the freeze up time, Julian day; tmax the time of annual maximum temperature, Julian day; T the water temperature, ◦ C; Tavg the annual average water temperature, ◦ C; To the under-ice water temperature, ◦ C; ω is the annual frequency, 0.01721 d−1 . The removal of nutrients was modeled as a first order areal process, using the PkC* model of Kadlec and Wallace (2009). Because of the high inflow hydraulic loading, compared to other losses and gains, the water flows were greatly dominated by the added lagoon water. For that situation, the PkC* model is: Co − C ∗ = Ci − C ∗



1+

k Ph

−P (3)

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Fig. 2. Water budgets for the Brighton wetlands. Values are 30-year means for precipitation and evapotranspiration, and the ten-year project means for wastewater inflows. Outflow is calculated. See text for further explanation.

where Ci is the inlet concentration, mg/L; Co the outlet concentration, mg/L; C* the background concentration, mg/L; h the water depth, m; k the apparent rate coefficient, m/d:m/(yr 365); P the apparent number of tanks in series;  is the nominal detention time, d. The background concentrations, the C* values in Eq. (1), are approximately zero for most contaminants, and were assumed to be 0.02 mg/L for both TP and NH3 N. The parameter “P” represents a combination of the effects of hydraulic efficiency and a distribution of removal rate coefficients for mixtures. The Brighton wetlands were assumed to have a P = 3 for both TP and NH3 N. The value P = 1 was selected for CBOD5 and for TSS, because these are both mixtures with broad distributions of removal rates. The rate coefficients for nitrogen compounds are expected to vary seasonally, due to temperature and plant growth. The temperature dependence of microbial process rate coefficients was modeled according to the modified Arrhenius relation: kT = k20  (T −20)

(4)

where kT is the rate coefficient at temperature (T) = T ◦ C; k20 the rate coefficient at 20 ◦ C;  the temperature coefficient. Growth processes rate coefficients were modeled with a seasonal dependence, keyed to the abundance of plants and other biota during the course of the year. These growth-related processes cause monthly variation, and a resulting sequence of rate coefficients: k = kj,

j = 1, 2, 3, . . . , 12

(5)

4. Wetland water quantities The wastewater flow to the wetlands was 3278 ± 78 m3 /d (monthly mean ± s.e.). High flows occurred in March and April, with

a ten-year maximum of 6270 m3 /d (Fig. 2). Low flows occurred in August and September, with a ten-year minimum of 1961 m3 /d. There was a slight increasing trend in the wastewater flows to the wetland, averaging 1.5% per year over the ten-year period of record. Wastewater dominated the water budget of the wetland in all seasons (Fig. 2). Precipitation and snowmelt averaged 5% of the total inputs to the wetland, and estimated evapotranspiration was 4% of the total outputs. These atmospheric inputs were, in total, fairly constant over the course of the year, at an annual average of 168 ± 6 m3 /d (monthly mean ± s.e.). Seepage was assumed to be nil, because the wetlands were sited in a tight clay soil. Water flows leaving the individual north and south wetland cells were measured on 132 dates during 2001. The flow leaving the north was 1056 ± 55 m3 /d, and leaving the south was 1076 ± 132 m3 /d, for a total of 2132 m3 /d. These are not significantly different (P = 0.05), and hence the flows were regarded as equally split. The inflow on those days averaged 2286 ± 132 m3 /d. That is not significantly different (˛ = 0.05) from the combined outflow, and hence the outflow measurements were discontinued. The water depths in the wetlands were variable, and different in the north and south cells. Water depth records from 2006 are shown in Fig. 3. Because of muskrat denudation in 2004 in the south cell, animals were removed, and water levels were lowered during the growing seasons, to promote vegetation regrowth. The north cell did not suffer comparable damage, and water levels were kept relatively constant year-round. The south cell averaged 36 ± 15 cm depth (monthly mean ± s.d.), with summer and winter depths of 24 and 48 cm respectively. The north cell was held at 50 ± 11 cm throughout the year. These depths are calculated from as-built topographic information, and are likely overestimates, because of soil swelling and sediment accretion in the wetlands.

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Fig. 3. Nominal water depths in the north and south wetland cells during 2006–2010. These have been corrected for accretion of 1.0 cm/yr.

Estimates of wetland water areas and volumes, based on design specifications, are given in Table 2. For the average water flow, the hydraulic retention time (HRT) in the wetlands was 9.4 days, and the hydraulic loading rate (HLR) was 5.3 cm/d. 5. Performance modifiers 5.1. Water temperatures The Brighton wetlands had long enough detention times so that water temperatures adjusted to the prevailing meteorological conditions. That adjustment typically takes less than 3–5 days, and the nominal hydraulic retention time (HRT) for the wetlands was more than nine days (Table 2). The value reached is determined by the balance of energy flows, and is termed the balance temperature (Kadlec and Wallace, 2009). Water temperature data display a strong annual trend, and thus require analysis via a model of those trends (Eqs. (1) and (2)). The trend for the wetland outlet water temperature is shown in Fig. 4. Table 3 summarizes the temperature trend parameters for the tenyear period of record (POR) (2001–2010), along with the historical averages for the ambient air. This model explains most of the variability of wetland data, with R2 values of approximately 0.90. These data fits are comparable to those reported by Kadlec and Wallace (2009) for other wetlands. The wetland outlet temperatures are significantly different from the wetland inlet water temperatures (˛ = 0.05). The lagoon water that enters the wetlands was about 2 ◦ C warmer than that leaving Table 2 Wetland dimensions and flow quantities. Bench areas are the shallow areas between deep zones (dz). HRT = hydraulic detention time, HLR = hydraulic loading rate.

Mean bench depth (m) Area (m2 ) Length (m) Width (m) L:W dz width (m) dz number dz area (m2 ) % dz area dz depth (m) dz volume (m3 ) Bench area (m2 ) Bench volume (m3 ) Total volume (m3 ) Mean flow (m3 /d) Nominal HRT (d) Nominal HLR (cm/d)

North cell

South cell

Total wetland

0.50 31,000 329 94 3.50 7.0 5 3294 0.106 1.30 3294 27,706 13,853 17,147 1639 10.5 5.3

0.36 31,000 329 104 3.50 7.0 5 3623 0.117 1.30 3623 27,377 9889 13,512 1639 8.2 5.3

0.43 62,000 329 198 1.67 14.0 6917 0.112 1.30 6917 55,083 23,742 30,659 3278 9.4 5.3

Fig. 4. Annual pattern of wetland outlet water temperature. The root mean square error of the weekly data about the model is 2.5 ◦ C.

the wetlands. However, the averages of the three-year side-byside data for the north and south wetland cells were not different (˛ = 0.05). As noted by Kadlec and Wallace (2009) for other wetlands, the Brighton wetland effluent water temperatures were linearly related to the mean air temperature during the unfrozen season: T = 3.07 + 0.923Tair

(6)

R2 = 0.995 where T is the water temperature, ◦ C; Tair is the air temperature, ◦ C. On average, the exiting wetland water is 2.2 ◦ C warmer than the mean air. The relative humidity at Brighton was approximately 74% (Environment Canada, 2010). For that humidity, the water temperature is expected to exceed the air temperature (Kadlec and Wallace, 2009). 5.2. pH The pH of the water entering the wetlands was 7.74 ± 0.01 s.u. (range 6.80–8.40) (mean ± s.e.), and that of the wetland effluent water was 7.63 ± 0.01 s.u. (range 7.00–8.50). Although these values are close, they are significantly different (˛ = 0.05). An analysis of the side-by-side data from the north and south cells shows that they were not different (˛ = 0.05). These pH values are higher than those for most other treatment wetlands, but not for constructed wetlands receiving nutrient-poor waters (Kadlec and Wallace, 2009). Speculatively, nutrient limitations can lead to sparse vegetation, and relatively abundant algae. Vegetation at Brighton was not dense, particularly in the south cell, and available light encouraged algae. Chlorophyll-a ranged from 17–64 ␮g/L in a synoptic study in 2007 (McGauley, 2008). This range is commensurate with that for a eutrophic water body (Wetzel, 1983). Such algal abundance is expected to elevate the pH of the water. 6. Water quality The period of record water quality entering and leaving the Brighton wetlands is summarized in Table 4. In general, modest reductions in the monitored pollutants were achieved, but a more thorough analysis is required to determine whether these reductions are within the bounds of the results from other wetlands.

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Table 3 Water and air temperatures at Brighton. RMSE = root mean square error about the model. The north and south cells are not significantly different, but the wetland inlet and outlet are significantly different (˛ = 0.05). Air Frequency Years Measured Mean Maximum Minimum Sinusoid model Mean Amplitude Maximum Minimum RMSE R2

7.0 38.9 −35.1 7.0 1.97 20.7 −6.8

North cell out

South cell out

Wetland inlet

Wetland outlet

Weekly 3

Weekly 3

Monthly 10

Monthly 10

11.30 26.6 −0.2

11.15 25.7 −0.3

11.73 28.3 −1.0

10.79 26.0 −1.0

10.7 1.12 22.8 1.5 2.50 0.893

10.6 1.10 22.3 1.5 2.53 0.898

11.30 1.01 23.7 1.5 2.39 0.917

9.08 1.24 21.9 1.5 2.46 0.908

6.1. Phosphorus Because the discharged water immediately enters the Provincially significant wetlands that border Presqu’ile Bay, phosphorus (P) is of great concern. For that reason, alum is added to lagoon 1 to achieve removal of the majority of the wastewater phosphorus. The wetland was expected to further lower the P content, and it did so. There is a strong seasonal pattern for total phosphorus (TP) entering the wetland and it is mirrored to an extent in the TP leaving the wetland at reduced levels (Fig. 5). The concentrations are highest in winter, and lowest in late summer (July–September). During that three-month late summer period, there was essentially no phosphorus removal in the wetlands. On average, entering TP was below the limit of 1.0 mg/L, and leaving TP was below the wetland objective of 0.8 mg/L. During the POR, there were three monthly exceedances of the 1.0 mg/L limit, directly attributable to clogging and breakage of the alum feed line to lagoon 1. Those malfunctions carried through to one exceedance of the wetland objective of 0.8 mg/L. The annual removal rate for TP averaged 2.26 ± 0.34 gP/(m2 yr) for the POR (mean ± interannual s.e.). This corresponds to an annual removal of 33%. Removal rate coefficients were calculated from Eq. (3), with C* = 0.02 mg/L and P = 3 apparent tanks in series. Therefore, the mean annual removal rate coefficient was 9.2 ± 1.5 m/yr R (mean ± interannual s.e.). Eq. (3) was also applied on a monthly basis, to determine the seasonal pattern in the P removal coefficient (Fig. 6). That pattern showed a maximum in early spring and

Fig. 5. Phosphorus concentrations by month during the ten-year POR for the Brighton wetlands. Error bars represent the standard error of the monthly means. The lagoon effluent limit is 1.0 mg/L and the wetland effluent objective is 0.8 mg/L on a monthly basis.

a minimum in late summer. That is contrary to the concept of P removal acting in concert with water temperature, and hence Eq. (4) does not apply. The negative rate coefficients in August and September indicate release of phosphorus during that period. The annual load removal was at the 33rd percentile of those for other wetlands (N = 282) as reported by Kadlec and Wallace (2009). The removal rate coefficient was at the 48th percentile. The startup period in the year 2000 has been excluded from the above analysis. The removal rate coefficient was 34.6 m/yr for that six-month period (July–December), or about 10 times higher than

Table 4 Water quality entering and leaving the Brighton treatment wetlands during the tenyear POR. These are means and standard errors (s.e.) of weekly data, except for flow and E. coli, which are monthly means and standard errors. Wetland inlet

3

Flow (m /d) Temperature (◦ C) pH (s.u.) CBOD5 (mg/L) TSS (mg/L) TP (mg/L) TKN (mg/L) Ammonia N (mg/L) Organic N (mg/L) Nitrite N (mg/L) Nitrate N (mg/L) Oxidized N (mg/L) E. coli mean (cfu/100 ml) E. coli geomean (cfu/100 ml)

Wetland outlet

Mean

s.e.

3278 11.73 7.74 5.4 13.2 0.378 13.60 11.31 2.24 0.31 0.53 0.88 2344 167

78 0.40 0.01 0.3 0.7 0.014 0.18 0.19 0.10 0.05 0.03 0.05 532 –

Mean – 10.79 7.62 3.2 7.2 0.255 11.20 9.42 1.85 0.12 0.34 0.49 967 132

s.e. – 0.41 0.01 0.1 0.3 0.009 0.20 0.20 0.08 0.02 0.02 0.03 290 –

Fig. 6. Seasonal pattern in the phosphorus removal rate coefficient. The water temperature is also shown for comparison.

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after startup. During the POR, those six months had k = 4.1 m/yr. As has been seen elsewhere, there was a startup premium for P removal, presumably due to sorption phenomena that later saturate. There is no evidence in the progress of annual P removals of an effect of the muskrat damage of 2003–2005. In fact the P removal rate coefficient for those three years was 10.3 m/yr, slightly higher than the overall average.

6.2. Nitrogen The lagoon system was not intended to greatly reduce nitrogen (N). The water leaving the lagoons contained principally ammonia nitrogen at 11.4 mg/L, and lesser amounts of organic and oxidized nitrogen (Table 4). Ammonia concentrations exhibited strong seasonal patterns, with both winter and summer peaks (Fig. 7). On average, the lagoons met the C of A limits; but there were 18 individual monthly exceedances for the 120 months of the POR (15%). On average, the wetlands met the C of A objectives, except for the month of October; but there were 27 individual monthly exceedances for the 120 months of the POR (23%). The POR total nitrogen loading to the wetlands was 266 ± 8 gN/(m2 yr), and 49 ± 3 gN/(m2 yr) were removed. There were only slight inter-annual trends. This load reduction places the wetlands in a category in which the vegetative uptake and return are important, as well as microbial processes. The wetlands bordered on the agronomic classification of Kadlec and Wallace (2009). A breakdown of the annual nitrogen conversions and flows is presented in Fig. 8. The fluxes are based on the ten-year mass balances, together with assumptions. These are: (1) ammonia is the sole source of N for plant growth; (2) plant uptake was 37.5 gN/(m2 yr), incorporated into 1500 g dw/m2 biomass production at 2.5% N dry weight; (3) 90% of the annual uptake is recycled as organic N, and

Fig. 7. Ammonia nitrogen concentrations by month during the ten-year POR for the Brighton wetlands. Error bars represent the standard error of the monthly means. The limits and objectives are monthly averages.

10% buried in new sediments; (4) ammonification converts organic N to ammonia; (5) nitrification converts ammonia to oxidized N; and (6) denitrification converts oxidized N to gaseous N products. It is seen that plant uptake and release are commensurate with the microbial processes (Fig. 8). Therefore, nitrogen removal was amplified in the spring growing season, as well as in the warmth of midsummer. Eq. (4) is expected to apply to microbial processes, and Eq. (5) to plant uptake. It was not feasible to attempt to directly measure the nitrogen content of the biomass and necromass, because of spatial variability and the large level of effort required (Reinhardt et al., 2006). However, the annual patterns of growth, senescence and burial can be estimated, which allows the month-by-month accounting for nitrogen in the biomass and necromass (Kadlec and Wallace, 2009).

Fig. 8. Annual fluxes of nitrogen in the Brighton wetlands in gN/(m2 yr).

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Fig. 9. Estimated nitrogen utilization and return from the biomass in the Brighton wetlands. Uptake is during the growing season, and return is during autumnal senescence.

The phytomass nitrogen mass balance for a fixed time period t is: (Ju − Jr − Jb ) · t = N

(7) gN/(m2

d); Jr where Ju is the uptake of nitrogen by phytomass, the release of nitrogen from phytomass, gN/(m2 d); Jb the burial of nitrogen from phytomass, gN/(m2 d); N the N storage in phytomass, gN/m2 ; t is the time interval, d. Uptake coincides with the growing season in north temperate environments. Data on biomass and nitrogen content have been reported for other treatment wetlands (e.g., Mueleman et al., 2002). At Brighton, that meant maximum uptake during April and May, and zero uptake in the frozen months. Return fluxes of nitrogen occur primarily in the late summer and autumn months, September through October (Kadlec, 2005). Less is known about the rate of conversion of necromass to refractory residuals (mineralization); there are no studies that have produced data. This mineralization, also called burial in some literature, has been here presumed constant year-round. These assumptions mean that monthly calculations are to some extent speculative (Kadlec and Wallace, 2009). The seasonal patterns for the 12 months were assumed to be: • Uptake: 0, 0, 0.04, 0.20, 047, 0.20, 0.07, 0.02, 0, 0, 0, 0. • Return: 0.03, 0.03, 0.03, 0.03, 0.04, 0.07, 0.10, 0.15, 0.23, 0.21, 0.05, 0.03. • Burial: uniform at 0.083 each month. These patterns were used to apportion the annual totals of 37.5 gN/(m2 yr) uptake, 90% returned and 10% buried. The resulting monthly fluxes of nitrogen are shown in Fig. 9. Uptake and storage occur in spring, and release occurs in autumn. During the spring growth period, more than the entire external nitrogen load removal was consumed to create the standing crop. The balance was assumed to be withdrawn from soil storage. The vegetative uptake and return, in conjunction with mass balances for the nitrogen species, allows computation of the monthly amounts of ammonification, nitrification and denitrification (Fig. 10). Because only small amounts of concentration reduction occurred, it is accurate to calculate rate coefficients as flux divided by mean concentration: J = k(C − C ∗ )avg

(8)

where (C − C*)avg is the mean concentration above background; J the nitrogen conversion rate, gN/(m2 d); k is the rate coefficient.

63

Fig. 10. Estimated microbial nitrogen processing in the Brighton wetlands.

Background concentrations were taken to be zero for ammonia and oxidized nitrogen, and 0.4 mg/L for organic and total nitrogen, the 10th percentile of the distribution of monthly values. Rate coefficients are tabulated in Table 5. The annual rate coefficient for ammonification, 28.5 m/yr, is at the 70th percentile of those for other wetlands as reported by Kadlec and Wallace (2009). But, the annual rate coefficient for nitrification, 3.8 m/yr, is close to the lowest observed in other wetlands. The maximum implied oxygen supply to support the low level of nitrification was 0.48 ± 0.06 gO/(m2 d) for the POR (mean ± interannual s.e.). This is about ten times lower than that for other similar wetlands. The carbonaceous oxygen demand removed was 0.12 ± 0.03 gO/(m2 d); thus the total oxygen required was 0.60 gO/(m2 d). This implied supply is toward the low end of experiences at other wetlands. Dissolved oxygen was not a routine monitoring parameter for the wetlands, but a field investigation was conducted in July 2007 that included DO measurements at several wetland locations (McGauley, 2008). These indicated little or no depletion along the flow paths, and an overall average level of 3.9 ± 0.4 mg/L in open water areas; and 5.3 ± 0.4 mg/L in vegetated areas. These values are not indicative of oxygen starvation of the nitrification process. Nitrification also requires alkalinity, which was not a routine monitoring parameter. Spot checks of alkalinity in the wetlands showed 220 ± 33 mg/L in the wetland influent; and 216 ± 18 mg/L in the wetland effluent, which is more than adequate to support nitrification. Denitrification is generally supposed to be the ultimate annual removal process for total nitrogen in excess of the net burial (Fig. 8). The amounts of inferred denitrification, computed from mass balances, are higher in the warm months than in the cold months (Fig. 10). Because there are only small amounts of oxidized nitrogen in the wetland water at all times (0.85 mg/L in; 0.46 mg/L out), it is likely that oxidized nitrogen is rapidly denitrified as soon as it is formed. The apparent removal of oxidized nitrogen, based on input and output, was 7.0 gN/(m2 yr), but mass balances indicate that the total denitrification was many times higher, at 45.2 gN/(m2 yr). The difference is due to production of oxidized nitrogen from nitrification. Denitrification occurs at low water concentrations of oxidized nitrogen, and consequently the rate coefficients are quite large, averaging 101 m/yr (Table 5). Total nitrogen reduction is the sum of burial and denitrification (Fig. 8), and the POR average was 49 ± 3 gN/(m2 yr) (mean ± interannual s.e.). The rate coefficient was restricted because of low nitrification, to 3.9 m/yr. This value is low compared to other wetlands, at the 10th percentile (Kadlec and Wallace, 2009).

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Table 5 Rate coefficients for pollutant removal in the Brighton wetlands. TP and TN removals are seasonal, and not governed by a temperature coefficient alone. CBOD5 and TSS removals are not seasonal (interannual mean ± s.e.). kAnnual (m/yr) TP removal TN removal CBOD5 removal TSS removal Ammonification Nitrification Denitrification

9.2 ± 1.5 3.9 ± 0.3 23 ± 6 41 ± 11 28.5 3.8 101

k20 ◦ C (m/yr)

40.8 5.6 170

 Not applicable Not applicable Not applicable Not applicable 1.046 1.052 1.088

R2

0.18 0.98 0.64

6.3. CBOD5

6.5. E. coli

Carbonaceous biochemical oxygen demand (CBOD5 ) was low entering the wetland, 5.4 ± 0.3 mg/L (monthly mean ± s.e.), and still lower leaving the wetland, 3.2 ± 0.1 mg/L. There was a seasonal variation in both the influent and effluent from the wetlands (Fig. 11), with higher concentrations and higher removals in winter and spring. The annual lagoon regulatory limit of 30 mg/L was not exceeded during the POR, nor was the wetland objective of 15 mg/L. The annual average rate coefficient was 23 ± 5 m/yr, for P = 1 and C* = 1.0 mg/L (Eq. (3)). This value is at the 40th percentile of those determined for 203 other wetlands (Kadlec and Wallace, 2009). Performance was better in colder temperatures, with a temperature coefficient of  = 0.960 (Eq. (4)).

E. coli was the bacterial indicator prescribed as a measure of fecal contamination of the lagoon and wetland waters. The wetland inlet E. coli was 152 ± 11 cfu/100 ml, and the wetland outlet was 129 ± 11 cfu/100 ml (ten-year geomean ± sd). Although the ten-year averages are below the regulatory limit and objective of 200 cfu/100 ml, exceedances were numerous for both wetland inlet and outlet. There were 41 exceedances for the inlet (out of N = 115) and 38 for the outlet (out of N = 113). The seasonal patterns showed high values in late winter and early spring (Fig. 13). The log10 reductions averaged 0.60 during cold months (December–May) and −0.47 during warmer months (June–November). Thus the wetland reduced E. coli for half the year, and increased E. coli for the other half of the year.

6.4. TSS Total suspended solids (TSS) was low entering the wetland, 13.24 ± 0.7 mg/L (monthly mean ± s.e.), and still lower leaving the wetland, 7.2 ± 0.3 mg/L. There was a seasonal variation in both the influent and effluent from the wetlands (Fig. 12), with higher concentrations and higher removals in winter and spring. The annual lagoon regulatory limit of 40 mg/L was not exceeded during the POR, and the annual wetland objective of 15 mg/L was also not exceeded. Occasional excursions typically originated with high TSS coming from the lagoon. However, the ten-year averages were well below limits and objectives (Fig. 12). The trapped solids contributed to sediment buildup in the wetlands. At an estimated density of 100 kg/m3 , these removed solids would have produced a total POR accumulation of 1.2 cm. That amount is likely much smaller than the solids generated within the wetland and trapped therein (Harter and Mitsch, 2003).

Fig. 11. Variation of CBOD5 during the course of the year. Data are from the folded 10-year POR, and the error bars indicate average interannual standard errors. The lagoon effluent limit is 30 mg/L and the wetland effluent objective is 15 mg/L on an annual average basis. The folded 10-year averages are 5.3 mg/L for the inlet and 3.3 mg/L for the outlet.

7. Ecology 7.1. Vegetation The shallow terraces of the wetland were initially seeded with cattails (T. latifolia), but over the course of time other species found their way into the wetland. As of 2009, 107 other flowering plant species were identified (Campeau et al., 2009). Some of these are generally viewed as nuisance species, such as thistles (Cirsium spp.; Sonchus spp.), purple loosestrife (Lythrum salicaria), and common reed (Phragmites australis); but most are common plants native to the region. The cattail stem counts reported for the Brighton wetland were very low compared to other wetlands, including those receiving

Fig. 12. Variation of TSS during the course of the year. Data are from the folded 10-year POR, and the error bars indicate average interannual standard errors. The lagoon effluent limit is 30 mg/L and the wetland effluent objective is 15 mg/L on an annual average basis. The folded 10-year averages are 13.2 mg/L for the inlet and 7.3 mg/L for the outlet.

R.H. Kadlec et al. / Ecological Engineering 47 (2012) 56–70

Fig. 13. Variation of wetland E. coli numbers during the course of the year. Data are geomeans of monthly samples from the folded 10-year POR. Error bars are the standard deviation of the log transformed data. The lagoon effluent limit is 200 cfu/100 ml and the wetland effluent objective is 15 mg/L on an annual average basis. The folded 10-year geomeans are 153 cfu/100 ml for the inlet and 131 cfu/100 ml for the outlet.

treated wastewater and those not (Table 6). The north cell was more densely vegetated than the south cell in 2007, but at only a small fraction of stem counts experienced elsewhere. Individual plants, gauged by aboveground live biomass, were of comparable size to those found in the comparison and Houghton Lake wetlands (Table 6). Individual plants were apparently of typical size, but plant densities across the Brighton wetland were much lower in 2007 than they ought to have been. Density of vegetation since this time appear to be increasing. This is especially true for the south cell where the water level was reduced to allow for seed germination and fresh growth has become rampant during the summers of 2010 and 2011. 7.2. Macroinvertebrates Aquatic invertebrates are commonly used to assess health of aquatic ecosystems, and wetlands in particular (Spieles and Mitsch, 2000; Nelson and Thullen, 2008). A study of macroinvertebrates in the Brighton treatment wetlands and nearby comparison wetlands was conducted in mid-summer 2007 (McGauley, 2008). The comparison wetlands were not constructed, but were impacted by land uses in their watersheds. These could be classified as “impacted natural” ecosystems. Kick and sweep sampling, and Hester Dendy samplers (Hester and Dendy, 1962), were used to enumerate the taxa and their relative abundance (Table 7).

Table 6 Vegetation density in the Brighton treatment wetland and in comparison wetlands. Data for the Brighton and comparison wetlands are from McGauley (2008), and are for 2007. The Houghton Lake wetland data span 16 years, for a wetland receiving lagoon effluent (Kadlec, 2009; Kadlec and Bevis, 2009). Stem count (#/m2 ) Brighton wetlands South cell North cell

9±4 18 ± 5

Live aboveground biomass (g dw/plant) 34 ± 6 37 ± 4

Comparison wetlands 48 ± 6 1 51 ± 5 2 62 ± 5 3

35 39 75

Houghton Lake wetlands Discharge 77 ± 8 35 ± 3 Control

35 ± 3 29 ± 5

65

Regardless of the method, the number of taxa in the comparison wetlands was about double that in the treatment wetland. As a result, diversity measured by either the Simpson (1949) or Shannon–Wiener (Magurran, 1988) indices was about triple in the comparison wetlands. The Hilsenhoff index accounts for the supposed pollution tolerance of the various invertebrates (Hilsenhoff, 1988), by weighting each taxon by its pollution tolerance. Because most of the dominant taxa in both kinds of wetlands were moderately tolerant, there was no difference between the Hilsenhoff indices. Both kinds of wetlands would be ranked as “fairly poor,” with “substantial pollution likely.” However, the Hilsenhoff weights are derived from “organic pollution” considerations, related to animals on the watershed or carboniferous wastewater plant discharges. The Brighton treatment wetlands received little CBOD5 , but considerable ammonia. Dissolved oxygen was higher in the comparison wetlands, at the time of sampling, 7.9 vs. 4.5 mg/L. Ammonia nitrogen was 50 times higher in the treatment wetlands, 10.9 vs. 0.22 mg/L. This created a nitrogenous oxygen demand in excess of the carbonaceous oxygen demand. If the details of the populations are examined, the differences are more apparent. The treatment wetland was dominated by Chironomidae (chironomids or non-biting midges), with secondary levels of Physidae (snails) and Corixidae (water boatmen). The comparison wetlands were dominated by Gammaridea (Amphipod crustaceans) and Asellidae (Isopod crustaceans). Secondarily in the comparison wetlands, there were sparse but more diverse mollusc assemblages, including the Sphaeriidae, Planorbidae and Bithynidae families. 7.3. Wildlife The Brighton treatment wetlands were not designed for biodiversity or wildlife habitat, but for polishing of the lagoon discharge. Nevertheless, it was known during the design process that additional benefits, and possibly problems, could ensue due to wildlife use. Some aspects of animal use of the treatment wetlands have been investigated (Campeau et al., 2009). An incomplete list of odonate species observed in the wetlands includes 23 dragonflies (Anisoptera) and 6 damselflies (Zygoptera). Twenty-eight species of butterflies (Lepidoptera) have been identified. Mammals that frequent the system include shrews, moles, weasels, fishers, mink, chipmunks, red squirrels, red foxes, coyotes, muskrats and white tail deer. It is the high level of bird use that has taken the naturalist community somewhat by surprise. The treatment wetland has proven, over the years, to be an important staging area for migratory species, which was anticipated. But it is also an important nesting area for Virginia rails (Rallus limicola), sora rails (Porzana carolina), and common moorhens (Gallinula chloropus). These three faired better in the Brighton wetland than in neighboring Presqu’ile Park (Campeau et al., 2009). In total, 110 species of birds have

Table 7 Macroinvertebrate abundance, richness and indices for Brighton and comparison wetland sites. Averages based on McGauley (2008). Treatment

Comparison

Kick&Sweep Kick&Sweep

Individuals Taxa

386 11

390 21

Hester Dendy Hester Dendy Hester Dendy Hester Dendy Hester Dendy

Individuals Taxa Simpson Index (1-D) Shannon–Weaver Index Hilsenhoff Index

322 7 0.28 0.69 6.47

136 18 0.86 2.35 6.49

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been seen in the wetlands. Some of these are rare in the region: the black-necked stilt (Himantopus mexicanus), the glossy ibis (Plegadis falcinellus), the red-necked phalarope (Phalaropus lobatus), the trumpeter swan (Cygnus buccinator) and the yellow-headed blackbird (Xanthocephalus xanthocephalus). A pair of ospreys (Pandion haliaetus) have successfully nested on a constructed platform in the treatment wetland each of the past two years (Morin, 2009). The Brighton treatment wetland is now a popular birding spot, with about a thousand visitors in 2009. This level of tourism stretched the ability of the Municipality to provide guided supervision, and a permit system was implemented beginning in 2010. Herbivorous muskrats (Ondatra zibethicus) caused a negative public perception. Muskrats prefer cattails for food and lodging (Kadlec et al., 2007), and they found the treatment wetlands very desirable. By the summer of 2003, many muskrat huts and feeding mounds were evident and vegetation density was suffering. An intensive trapping program in the autumn of 2003 caught more than 400 muskrats. By the summer of 2004, the vegetation at the west end of the south cell was completely gone. Citizens became concerned that water quality benefits would disappear with the cattails. However, there is no evidence in the Brighton water quality data time series to indicate that the intense muskrat activity of 2003–2004 had an influence on wetland performance. The Municipality hired a trapper during the ice-free seasons, resulting in fewer animals and less plant damage.

8. Discussion 8.1. Listowel vs. Brighton A very similar system operated at the Town of Listowel, Ontario, from 1980–1984. An aerated lagoon, with alum addition, discharged into a second, facultative lagoon. The Ontario Ministry of the Environment (MOE) built five pilot-scale constructed wetlands that received portions of the lagoon effluents, two receiving aerated lagoon effluent and three receiving facultative lagoon effluent. The Listowel wetlands were in an equivalent position in the treatment train as Brighton’s wetlands. The Listowel pilot wetlands were extensively studied over their four-year operational existence. The results were summarized in a detailed (but somewhat obscure) report (Herskowitz, 1986). The data from that report, for the three relevant pilot systems, were here analyzed in the same way as the Brighton data. Results in Table 8 show that the Listowel hydraulic loadings were less than half those at Brighton, although other conditions were similar. Concentrations of CBOD5 , TSS, and TP were somewhat higher, but ammonia was lower. The fluxes of nutrients were quite similar to those at Brighton; for instance, the amount of nitrification, calculated by mass balance, was nearly identical. Phosphorus removal at Listowel was about 80% of that at Brighton. Rate coefficients for TP were lower at Listowel, but the rate coefficients for nitrification were higher. Importantly, some parameters were measured at Listowel that were not studied at Brighton. In particular, dissolved oxygen was routinely measured at Listowel, but only a few synoptic measurements were taken at Brighton. These are substantively in agreement, with modest levels (3.5–5.5 mg/L) being present in surface waters. However, sediments were at lower oxygen (redox) status. Redox at Listowel was measured at the 1.5 cm depth in the sediments. Values ranged from +166 down to −194 mV, with a mean in ice free conditions of 58 mV for system 3. Lower-end values were found in mid-summer. These redox values, at the lower end of the range, are in a zone typically associated with sulfate reduction (Reddy and DeLaune, 2008).

At Listowel, sulfate entered 170–200 mg/L, a good portion of which, ca. 25%, originated from the alum addition. This sulfate was reduced in minor part to H2 S, with 1.1–2.4 mg/L of sulfide in the influent and 1.0–4.4 mg/L in the effluent (N = 15 monthly samples in three wetlands). These amounts of sulfate and sulfide may be troublesome for aquatic life, but are unlikely to impact the vegetation (Lamers et al., 1998). But importantly, sulfide can inhibit nitrification (Joye and Hollibaugh, 1995), and it may be that alum addition is detrimental to nitrification. 8.2. Water quality improvement Phosphorus removal in the Brighton wetlands was near the central tendency of other treatment marshes, with 2.26 gP/(m2 yr) removed. This was slightly higher than the Listowel pilot project, which removed 1.97 gP/(m2 yr); however, Listowel operated at a lower hydraulic loading and a higher incoming TP concentration (Table 8). The startup transient removed more TP, presumably to build the biomass in the wetland, and to load the sorption sites of the newly built system. The removal rate coefficient was maximum during the spring growing season (Fig. 5). Speculatively, the negative removal values in July and August may have been due to release at lower redox conditions, and to decomposition of necromass under warmer conditions. Ammonia reduction was low, which may have been due to a number of factors. Plant uptake was a key process in ammonia reduction, on a seasonal basis. The relatively sparse plant growth was responsible for less than optimal vegetative utilization. The other key process was nitrification, which also was at the low end of the anticipated spectrum. Speculatively, two conditions may have been responsible for low microbial ammonia conversion: (1) diffusional resistance due to deep water, and (2) sulfide toxicity to nitrifiers. In the first instance, ammonia must diffuse to the bottom sediments that are the location of the nitrifier populations. Additionally, oxygen must diffuse to those sediments from the overlying water. Deeper water impedes these movements, and has been shown to create lower ammonia removal rate coefficients. For example, data from Arcata, CA show a decrease of 50% in the rate coefficient for a depth increase from 30 to 50 cm (Kadlec and Wallace, 2009). Managing the water depth at the prescribed level is expected to improve the removal efficiency. Sulfide has been shown to be toxic to nitrifiers (Joye and Hollibaugh, 1995). At concentrations of less than 2.0 mg/L nitrification was reduced by 50%. Sulfide was not measured at Brighton, but at Listowel, such high levels of sulfide were found in surface waters (see above). Presumably, higher levels would be found in and near the sediments where nitrification was occurring. Nitrogen mass balances, as employed herein, require that most of the ammonia and organic nitrogen losses be ultimately removed via denitrification. That is because only minor amounts of the total nitrogen removal are to new sediment accumulation (see Fig. 8). This mass balance requirement can occur at very low levels of oxidized nitrogen, leading to very high apparent first order rate coefficients. Oxidized nitrogen buildups do not commonly occur in treatment marshes, with 28% of 129 such wetlands showing nitrate increases. The overall process has been termed “close coupling” of nitrification and denitrification (Blackburn et al., 1994; Kremen et al., 2005; Reinhardt et al., 2006). Mechanistically, close coupling implies denitrification is able to tolerate some oxygen, because diffusion distances need to be relatively small between oxic nitrification zones and anoxic denitrification zones. This situation is known to prevail in the root zone of wetlands (Reddy and DeLaune, 2008). The Brighton wetlands were ineffectual in reduction of the indicator E. coli, with reductions in cold months and additions in the

R.H. Kadlec et al. / Ecological Engineering 47 (2012) 56–70

67

Table 8 Comparison between the Listowel and Brighton performances. The numbers are averages of monthly data for the periods of record, four and ten years respectively. Numbers in italics are for July 2007 only. Listowel Inlet Concentrations (mg/L) CBOD5 TSS TP Organic N NH3-N NOXN Modifying parameters HLR Temperature mean Temperature max Depth pH DO Chl-a Annual fluxes (g/(m2 yr)) TP removal TN removal Ammonification Nitrification Denitrification Estimated burial Annual rate coefficients (m/yr) TP removal TN removal Ammonification Nitrification Denitrification CBOD5 removal

19.4 23.6 0.822 4.79 7.12 0.25

Brighton System 1 8.1 8.3 0.574 2.68 4.88 0.47 2.06

System 2 11.1 9.0 0.525 2.82 5.07 0.27 1.98

10.3 22.5 21 7.59 5.66 136

21

System 3 7.3 8.6 0.406 2.35 3.78 0.25 1.34 7.8 17.6 20 7.06 3.52 22

Inlet

Outlet

5.4 13.2 0.378 2.24 11.31 0.88

3.2 7.2 0.255 1.85 9.42 0.49

11.7 23.7 7.74 4.32 41

5.09 10.8 21.9 43 7.62 4.08 40

3.24

3.48

1.91 40.3 56.6 38.4 36.2 4.0

1.96 40.4 56.6 37.5 37.1 4.0

2.03 42.0 54.3 38.4 38.2 4.0

2.26 48.9 41.3 38.2 45.2 3.8

5.02 4.35 21.55 7.07 159 30

4.31 4.31 21.8 6.91 220 36

5.85 5.23 23.2 8.66 205 19

9.2 4.22 28.5 3.93 101 23

warm months. This may have been partially due to the low numbers leaving the lagoons, because other treatment wetlands with high incoming numbers of E. coli achieve considerable removal (Kadlec and Wallace, 2009; Boutilier et al., 2009). E. coli is problematic as an indicator in treatment wetlands for at least two reasons: (1) the test is prone to error, and (2) E. coli may both regrow and be generated within wetlands. The E. coli test has been found to respond to heterotrophic bacteria, which are numerous in wetlands (Olstadt et al., 2007; McLain and Williams, 2008). Regrowth of E. coli has been observed in the Tres Rios, AZ treatment wetlands (McLain et al., 2006). Those results suggested that bacterial growth was attributable to seasonal migratory patterns of avian populations and optimal regrowth conditions for the E. coli in the wetlands. Removal is consistently demonstrated in laboratory studies, and in small pilot wetlands without wildlife in similar climatic conditions (Boutilier et al., 2009). Both CBOD5 and TSS were reduced to wetland background levels in the wetland effluent (Kadlec and Wallace, 2009). Such background levels are the result of natural processes within a wetland, primarily decomposition of plant and algal residuals, and reflect complete removal of incoming solids and CBOD5 . 8.3. Winter operation Lagoon–wetland systems have been successfully operated in warm climates, where issues of frost are not encountered (see, for example, Gerke et al., 2001; Bays and Knight, 2002). The first subject that must be addressed is hydraulic operability. The Brighton system functioned in all seasons for ten years with essentially no difficulties associated with cold temperatures. In moderate climates, where ice formation allows hydraulic operation of the wetland, year-long operation may be employed. Winter operation is hydraulically possible for lagoons, even for fairly severe cold,

where ice thicknesses may range up to a meter (Heaven and Banks, 2005). Indeed, most lagoon systems receive water during the entire year. Winter operation is also possible for FWS wetlands, even in moderately cold conditions, because water can be managed to flow under-ice (Kadlec and Wallace, 2009). But at some point, as temperatures drop, complete freeze-up leads to over-ice water flows and ice buildup. Cold water temperatures lead to large reductions in microbial processes, and the vegetation cycle is dormant in frozen conditions. Oxygen transfer to under-ice water is blocked by ice and snow. These individual process effects combine in complex ways to produce the overall observed slow-down of treatment. It is worth noting that recommendations for other natural systems for wastewater treatment are not applicable to wetlands. For instance, overland flow treatment systems are recommended to have storage for all days expected to have below freezing air temperatures (WEF, 2001; Crites et al., 2006). That would mean at least three months storage for Brighton, which was not necessary for that lagoon–wetland system during the ten-year POR. Accordingly, it is useful to place the Brighton system in the wider perspective of other cold climate pond–wetland operations. The experiences at a number of pond–wetland facilities provide an operational strategy database (Table 9). These operate successfully with year-long discharges. All operate in freezing conditions, with a range of 81–171 days of mean daily temperatures below 0 ◦ C. Mean annual air temperatures range from 1.7 to 9.7 ◦ C, and the number of degree-days below zero ranges from 156 to 1836 ◦ C d. These three indicators are not independent, and all offer a means of discriminating between seasonal and successful year-long discharge strategies. In general, the hydraulic cold-limited condition is characterized by a mean annual air temperature of about 4 ◦ C, or 150 days of mean daily temperatures below freezing, or 1000 degree days below zero.

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Table 9 Climatological conditions for example lagoon–wetland systems. Sites are sorted by the number of degree-days below zero. Temperatures are mean daily air temperatures. State or Province

Town

Annual meanT (◦ C)

DaysTavg < 0 (d)

Degree days below zero (◦ C d)

Utah Ontario Ontario Colorado Michigan Michigan Colorado Nova Scotia Ontario Alberta Wisconsin Ontario

Logan Brighton Nanticoke Bennett LaPeer Onaway Ouray River Hebert Listowel High River Drummond Cobalt

9.7 7.0 8.8 7.2 8.2 7.2 6.2 6.6 5.1 3.8 5.1 1.7

81 94 95 102 106 113 121 121 137 143 144 171

156 697 259 285 364 446 467 514 765 985 986 1836

8.4. Vegetation Cattails began growing in the summer of 2000 and by autumn, the entire wetland was fully vegetated. There was little evidence of muskrat activity in the first two years. However, by the summer of 2003, many muskrat huts and feeding mounds were evident and vegetation was suffering. An intensive trapping program in the fall of 2003 caught more than 400 muskrats over a period of several weeks at a cost of $4450. By the summer of 2004, the vegetation at the west end of the south cell was completely gone. In spite of diligent trapping efforts, the entire south cell and a large portion of the north cell became devoid of vegetation by 2007. Two possible reasons are herbivory and excessively deep water. Aerial photography in 2007 showed 76% cover in the north cell and 23% in the south cell. As of the fall of 2010, the north cell is about 85% vegetated, and the south cell 40%. By the end of 2011, the south cell was 75% vegetated and the north cell remained at 85%. Cattails are sensitive to water depth. The upper depth tolerance of T. latifolia varies with the hydroperiod of the wetland. The Brighton wetland had a 100% hydroperiod, which would correspond to the minimum depth tolerance. For instance, Sharp (2002) found that water depths in excess of 40 cm were detrimental to T. latifolia. Sharp (2002) also found that a water depth of 20 cm was optimal, in contrast to 0 or 40 cm. If flooding was greater than 40 cm, cattail seeds had a harder time germinating; and seedlings in water depths greater than 40 cm showed a larger allocation to shoots and less to T. latifolia roots, which decreased survival rates. Grace and Wetzel (1981) found that T. latifolia did best in 26–29 cm of water, and was generally unable to survive in 50 cm. Rook (2002) suggests that optimal conditions are “boggy margins of ponds to shallow waters up to six inches deep.” These results suggest that deeper water (that greater than 40 cm) may not be suitable for maintenance of cattails. The north cell was operated at an annual mean depth of 44 ± 11 cm (N = 180) over the period 2006–2009. The south cell was operated at an annual mean depth of 36 ± 15 cm (N = 180), with summer depths of 25 cm (winter 47 cm), to foster regeneration of the plants. Increases in plant cover indicate this strategy was partially successful. But, the operational depths are at the upper end of those expected to foster growth and regeneration of Typha. Minor operational changes including managing the water depth at the prescribed level is expected to improve the plant density.

the logarithms of twelve water quality variables (Seilheimer et al., 2009). The WQI measures the degree of water quality degradation as a result of nutrient enrichment and road runoff. Values for the Great Lakes span a range from −3, qualitatively “highly degraded,” to +3, qualitatively “excellent.” It is supposed to be widely applicable to wetlands, and to produce measurements of wetland condition. McGauley (2008) calculated the WQI for the Brighton treatment wetlands, and for other wetlands in the vicinity. The average WQI for the treatment wetland was 1.19 ± 0.17 (mean ± s.e.) (range 0.82–1.76, N = 12 sampling places). The average for three comparison wetlands was 1.39 ±0.14 (range 1.12–1.55, N = 6 sampling places). These are all in the qualitative category labeled “very good” by Chow-Fraser (2006). The average value for the Presq’ile Park wetland was 0.77 ± 0.12 (range 0.6–1.0, N = 3 yr) (Chow-Fraser, 2006). That is in the qualitative category “good.” The highest score (interpreted as being the most pristine) for natural wetlands in the Lake Ontario/St. Lawrence watershed south of 45◦ North Latitude was WQI = 1.3 (Chow-Fraser, 2006). The Brighton treatment wetlands were at the 95th percentile in this group, which comprised 87 wetland-years. Therefore, the water discharge from the Brighton wetland is of a very high quality according to the WQI. Macroinvertebrates provide a separate means of assessing the ecological character of wetlands. As detailed in Section 7.2, McGauley (2008) studied invertebrates in the Brighton treatment wetlands and in three nearby comparison wetlands. The comparison wetlands were nutrient-impacted to some degree by agricultural runoff. There were more abundant individuals in the treatment wetlands, but the numbers of taxa were greater in the comparison wetlands. Both the Simpson and Shannon–Wiener indices were higher in the comparison wetlands, which had similar index values to the Olentangy, OH river water wetlands (Spieles and Mitsch, 2000). The Hilsenhoff Biotic Index (HBI) failed to differentiate between treatment and comparison wetland invertebrate communities (McGauley, 2008). This index assigns pollution values to various invertebrate taxa. All the Brighton wetlands had slightly lower HBI values (better) than the Olentangy wetlands. All the wetlands of the region are cattail-dominated systems, with patches of open water. Therefore, there is little differentiation to be expected from indices reflecting macrophyte vegetation numbers or taxa, and such indices were not computed.

8.5. Ecological indicators Several ecological indices have been developed to evaluate wetland quality in the Great Lakes region. These include water quality, wetland fish, wetland zooplankton, and wetland macrophytes (Seilheimer et al., 2009). The water quality index (WQI) was devised by Chow-Fraser (2006), and is calculated as a linear combination of

9. Conclusions The Brighton treatment wetland has provided and continues to provide water quality improvement, and thereby buffers the receiving waters from the lagoon effluent. The wetland also

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provides ancillary benefits for humans and wildlife. The principal advantageous features of the wetlands are: • Nutrient reductions were achieved, for phosphorus and nitrogen. The nutrient removal fluxes were very close to those observed at the precursor systems at Listowel. • COD5 and TSS were reduced to wetland background. • The wetland was designated as part of the Provincially Significant Presqu’ile Marsh Wetland. • The WQI for the Brighton treatment wetlands places them among the very best natural wetlands in the Lake Ontario/St. Lawrence watershed. • The ancillary benefits of the Brighton wetlands were very large for birds, and consequently for the human birder population. The Brighton treatment wetland is a very popular birding spot. Nutrient reductions were achieved, more effectively for phosphorus than for nitrogen. Wetland regulatory objectives were met on average, with some exceedances for ammonia. The relative ranking of the Brighton wetlands, in terms of water quality performance, was near the median for similar treatment wetlands. Phosphorus reduction was at the median, but ammonia reduction was lower. The wetlands did not serve a disinfection function, as determined by E. coli indicator numbers, which may have been due to extensive wildlife use and/or bacterial regrowth. Disinfection of the water entering the wetlands is not likely to help, based on studies at other wetland sites (McLain et al., 2006). The constructed marsh was successfully operated year-round in a moderately cold climate, from a hydraulic point of view. Some measure of treatment was achieved in the cold season, but the highest removals were observed in the spring growing season. There is a possibility that water depth may have contributed to moderately sparse vegetation, and to low nitrification. Both can likely be improved by managing water depth to lower values. It was also possible that excess sulfur, partly due to alum additions in the lagoons, may have contributed to microbial toxicity and hence to impairment of nutrient reductions. Closer monitoring of alum addition may reduce sulfur concentration and improve nitrification. These speculative ideas should be tested, either at Brighton or elsewhere, as resources permit.

References Bays, J.S., Knight, R.L., 2002. Fort deposit constructed treatment wetlands system: 10-year review. In: Pries, J. (Ed.), Treatment Wetlands for Water Quality Improvement, Quebec 2000 Proceedings. CH2M Hill, Waterloo, Ontario, pp. 35–44. Blackburn, T.H., Blackburn, N.D., Jensen, K., Risgaard-Petersen, N., 1994. Simulation model of the coupling between nitrification and denitrification in a freshwater sediment. Appl. Environ. Microbiol. 60 (9), 3089–3095. Boutilier, L., Jamieson, R., Gordon, R., Lake, C., Hart, W., 2009. Adsorption, sedimentation, and inactivation of E. coli within wastewater treatment wetlands. Water Res. 43, 4370–4380. Browne, W., Jenssen, P.D., 2005. Exceeding tertiary standards with a pond/reed bed system in Norway. Water Sci. Technol. 51 (9), 299–306. Campeau, M., Lee, R.V., Lee, K., 2009. Brighton Constructed Wetland. Ancillary Benefit Report to the Municipality of Brighton, October 2009. Chow-Fraser, P., 2006. Development of the water quality index (WQI) for assess effects of basin-wide land-use alteration on coastal marshes of the Laurentian Great Lakes. In: Simon, T.P., Stewart, P.M. (Eds.), Coastal Wetlands of the Laurentian Great Lakes: Health, Habitat and Indicators. Authorhouse, Bloomington, IN, pp. 137–166. Crites, R., Tchobanoglous, G., 1998. Small and Decentralized Wastewater Management Systems. McGraw-Hill, New York. Crites, R.W., Middlebrooks, E.J., Reed, S.C., 2006. Natural Wastewater Treatment Systems. CRC Taylor and Francis, Boca Raton, FL. Environment Canada, 2010. Canadian Climate Normals 1971–2000, http://climate.weatheroffice.gc.ca/climate normals (last modified September 28, 2010).

69

Gerke, S., Baker, L.A., Xu, Y., 2001. Nitrogen transformations in a wetland receiving lagoon effluent: sequential model and implications for water reuse. Water Res. 35 (16), 3857–3866. Grace, J.B., Wetzel, R.G., 1981. Habitat partitioning and competitive displacement in cattails (Typha): experimental field studies. Am. Nat. 118 (4), 463–474. Harter, S.K., Mitsch, W.J., 2003. Patterns of short-term sedimentation in a freshwater created marsh. J. Environ. Qual. 32 (4), 325–334. Heaven, S., Banks, C., 2005. Cold and continental climate ponds. In: Shilton, A. (Ed.), Pond Treatment Technology. IWA Publishing, London, pp. 381–407. Herskowitz, J., 1986. Listowel Artificial Marsh Project Report. Ontario Ministry of the Environment, Water Resources Branch, Toronto, Ontario. Hester, F.E., Dendy, J.S., 1962. A multiple plate sampler for aquatic macroinvertebrates. Trans. Am. Fish. Soc. 9, 420–421. Hilsenhoff, W.L., 1988. Rapid field assessment of organic pollution with a family-level biotic index. J. N. Am. Benthol. Soc. 7 (1), 65–68, http://www.rook.org/earl/bwca/nature/aquatics/typhalat.html. Joye, S.B., Hollibaugh, J.T., 1995. Influence of sulfide inhibition of nitrification on nitrogen regeneration in sediments. Science 270 (5236), 623–625. Kadlec, R.H., Pries, J.H., Mustard, H., 2007. Muskrats (Ondatra zibethicus) in treatment wetlands. Ecol. Eng. 29 (2), 143–153. Kadlec, R.H., 2003. Pond and wetland treatment. Water Sci. Technol. 48 (5), 1–8. Kadlec, R.H., 2005. Vegetation effects on ammonia reduction in treatment wetlands. In: Vymazal, J. (Ed.), Nutrient Cycling and Retention in Natural and Constructed Wetlands. Backhuys Publishers, Leiden, Netherlands, pp. 233–260. Kadlec, R.H., 2008. The lagoon–wetland treatment system at Onaway, Michigan. In: Proceedings of the 37th Annual Technical Symposium, Water Environment Association of Ontario, Milton, Ontario. Kadlec, R.H., 2009. Wastewater treatment at the Houghton Lake wetland: hydrology and water quality. Ecol. Eng. 35 (9), 1287–1311. Kadlec, R.H., Bevis, F.B., 2009. Wastewater treatment at the Houghton Lake wetland: vegetation response. Ecol. Eng. 35 (9), 1312–1332. Kadlec, R.H., Reddy, K.R., 2001. Temperature effects in treatment marshes. Water Environ. Res. 73 (5), 543–557. Kadlec, R.H., Wallace, S.D., 2009. Treatment Wetlands, 2nd ed. CRC Press, Boca Raton, FL. Kremen, A., Bear, J., Shavit, U., Shaviv, A., 2005. Model demonstrating the potential for coupled nitrification denitrification in soil aggregates. Environ. Sci. Technol. 39 (11), 4180–4188. Lamers, L.P.M., Tomassen, H.B.M., Roelofs, J.G.M., 1998. Sulfate-induced eutrophication and phytotoxicity in freshwater wetlands. Environ. Sci. Technol. 32 (2), 199–205. Magurran, A.E., 1988. Ecological Diversity and Its Measurement. Princeton University Press, Princeton, NJ. McGauley, E.K., 2008. Evaluating the ecological potential of constructed wetlands: a case study of the Brighton treatment wetland. MS Thesis. Trent University, Peterborough, Ontario. McKenna, G., Gregoriou, A., Pries, J., 2002. Treatment wetland for wastewater polishing at Ontario’ Power Generation’s Nanticoke generating station. In: Pries, J. (Ed.), Treatment Wetlands for Water Quality Improvement, Quebec 2000 Proceedings. CH2M Hill, Waterloo, Ontario, pp. 97–108. McLain, J.E., Williams, C.F., Rock, C.M. 2006. E. coli regrowth in a constructed wetland receiving treated sewage effluent: a threat to human health? Agronomy Abstract [CD-ROM P24484]. McLain, J.E.T., Williams, C.F., 2008. Seasonal variation in accurate identification of Escherichia coli within a constructed wetland receiving tertiary-treated municipal effluent. Water Res. 42, 4041–4048. Morin, B., 2009. Brighton constructed wetland. Ontario Birding News 14 (3), 1–2. Mueleman, A.F.M., Beekman, J.Ph., Verhoeven, J.T.A., 2002. Nutrient retention and nutrient-use efficiency in Phragmites australis stands after wastewater application. Wetlands 22 (4), 712–721. Nelson, S.M., Thullen, J.S., 2008. Aquatic macroinvertebrates associated with Schoenoplectus litter in a constructed wetland in California (USA). Ecol. Eng. 33 (1), 91–101. Olstadt, J., Schauer, J.J., Standridge, J., Kluender, S., 2007. A comparison of ten USEPA approved total coliform/E. coli tests. J. Water Health 5.2, 267–282. Polprasert, C., Koottatep, T., Tanner, C.C., 2005. Integrated pond/wetland systems. In: Shilton, A. (Ed.), Pond Treatment Technology. IWA Publishing, London, pp. 328–345. Reddy, K.R., DeLaune, R.D., 2008. Biogeochemistry of Wetlands, Science and Applications. CRC Press, Boca Raton, FL. Reinhardt, M., Müller, B., Gächter, R., Wehrli, B., 2006. Nitrogen removal in a small constructed wetland: an isotope mass balance approach. Environ. Sci. Technol. 40 (10), 3313–3319. Rook, E.J.S., 2002. Typha latifolia, common cattail. In: Flora, Fauna, Earth, and Sky..The Natural History of the Northwoods., http://www.rook.org/earl/bwca/nature/. Seilheimer, T.S., Mahoney, T.P., Chow-Fraser, P., 2009. Comparative study of ecological indices for assessing human-induced disturbance in coastal wetlands of the Laurentian Great Lakes. Ecol. Indic. 9, 81–91. Sharp, J.L., 2002. Managing cattail (Typha latifolia) growth in wetland systems. MS Thesis. University of North Texas, Denton, TX, August 2002. Simpson, E.H., 1949. Measurement of diversity. Nature 163, 688. Spieles, D.J., Mitsch, W.J., 2000. Invertebrate community structure in high- and lownutrient constructed wetlands. Wetlands 20 (4), 716–729.

70

R.H. Kadlec et al. / Ecological Engineering 47 (2012) 56–70

Tanner, C.C., Sukias, J.P.S., 2003. Linking pond and wetland treatment: performance of domestic and farm systems in New Zealand. Water Sci. Technol. 48 (2), 331–339. Vymazal, J., Kröpfelová, L., 2008. Wastewater Treatment in Constructed Wetlands with Horizontal Subsurface Flow. Springer, New York. Walmsley, N., Shilton, A., 2005. Solids and organics. In: Shilton, A. (Ed.), Pond Treatment Technology. IWA Publishing, London, pp. 328–345.

Wang, X., Bai, X., Qiu, J., Wang, B., 2005. Municipal wastewater treatment with pond–constructed wetland system: a case study. Water Sci. Technol. 51 (12), 325–329. WEF (Water Environment Federation), 2001. Natural Systems for Wastewater Treatment. WEF Manual of Practice FD-16, 2nd ed. Water Environment Federation, Alexandria, VA. Wetzel, R.G., 1983. Limnology, 2nd ed. Saunders College Publishing, Orlando, FL.