Use of created wetlands to improve water quality in the Midwest—Lake Bloomington case study

e c o l o g i c a l e n g i n e e r i n g 2 8 ( 2 0 0 6 ) 258–270

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Use of created wetlands to improve water quality in the Midwest—Lake Bloomington case study David A. Kovacic a,∗ , Richard M. Twait b , Michael P. Wallace a , Juliane M. Bowling a a

University of Illinois, Department of Landscape Architecture, 101 Temple Buell Hall, 611 E. Lorado Taft Drive, Champaign, IL 61820, USA b Bloomington Water Treatment Facility, Hudson, IL, USA

a r t i c l e

i n f o

a b s t r a c t

Article history:

Agricultural watersheds of the Midwest typically export nitrogen (N) and phosphorus (P) to

Received 4 August 2006

surface waters causing contamination of drinking water reservoirs and, ultimately, hypoxia

Accepted 8 August 2006

in the Gulf of Mexico. Two agricultural runoff wetlands, W1 (area 0.16 ha, volume 660 m3 ) and W2 (area 0.4 ha, volume 1780 m3 ), intercepting surface and tile drainage in the Lake Bloomington, Illinois, watershed were constructed in 1996 on forest soils (alfisols) between

Keywords:

upland cropland and Lake Bloomington. They were created to determine whether wetlands

Agricultural runoff wetlands

could reduce agricultural nonpoint source pollution before it entered the Lake Bloomington

Created wetlands

drinking water reservoir. Water (precipitation, tile inflow, surface inflow, outflow, seepage

Nitrate

and evaporation) and nutrient (N, P and carbon [C]) budgets were determined from 1 April

Phosphorus

1998 to 30 December 1999 for each wetland. Combined, the wetlands received 746 kg NO3 -N

Water quality

as tile loading, 104 kg as surface loading and exported 545 kg of NO3 -N as outflow and seepage. Mass NO3 -N retention was 36%. Following wetland treatment, overall volume-weighted NO3 − -N concentrations were reduced by 42% (W1) and 31% (W2). Combined P mass retention was 53%, and combined total organic carbon (TOC) mass retention was 9%. Wetlands were constructed in a sloping drainage (5%) where surface runoff was a major component of flow. Nutrient dynamics of P and C were affected by site slopes. Calculations made by extrapolating these results indicate that a wetland area of 450 ha would be required in the Lake Bloomington watershed to reduce N loading by 46%, at a construction cost ranging from 3 to 3.5 million dollars. Results support the growing evidence that agricultural runoff wetlands can effectively reduce NPS pollution loading in the Mississippi River Basin. © 2006 Elsevier B.V. All rights reserved.

1.

Introduction

The United States leads the world in food production, annually growing 40% of the world’s maize (Aquino et al., 2001) and 44% of the world’s soybeans (FAS, 1999). The Midwestern states of Iowa, Illinois, Minnesota, Indiana, Ohio, Missouri, and Wisconsin (in order of production) account for 58 and 70% of the maize and soybean production, respectively (USDA, 2002). This 40 million ha of land receives 6 million metric tonnes ∗

Corresponding author. Tel.: +1 217 244 5133; fax: +1 217 244 4568. E-mail address: [email protected] (D.A. Kovacic). 0925-8574/$ – see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2006.08.002

of fertilizer annually (Goolsby et al., 1999). If a conservative leaching loss estimate of 25 kg N ha−1 (David et al., 1997) is used, the Midwest would contribute at least 1 million metric tonnes of nonpoint source agricultural N to the Mississippi River annually (mainly in the form of nitrate, David et al., 1997). As nutrient-enriched waters flow to the sea, however, eutrophication of Gulf coastal waters has resulted. This problem, known as hypoxia, occurs regularly not only in the Gulf of Mexico (Goolsby et al., 1999), but throughout the world (Turner

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and Rabalais, 1991; Mitsch et al., 2001). Studies suggest that hypoxia may have detrimental effects on benthic organisms and other Gulf biota (Rabalais et al., 1992; Justic et al., 1996, 1997). Hypoxic conditions have caused collapses or reductions in fisheries in the Kattegat Sea, the Black Sea and the Baltic Sea (Diaz, 2001) and along the Oregon Coast (Grantham et al., 2004). State NPS assessment reports have clearly identified agriculture as the greatest NPS pollution problem in the U.S. (CAST, 1992; USEPA, 2002). Illinois water quality data reflect this same trend, with 96% of the annual nitrate load derived from agriculture (David et al., 1997). Similar results were found in Iowa (Keeney and Deluca, 1993). As a result of these high nitrate loads, many municipal drinking water reservoirs in the agricultural Midwest frequently exceed the EPA maximum contaminant level (MCL) for nitrate. Of primary concern is the prevention of methemoglobinemia (blue baby syndrome) which causes neural damage in infants 6 months old or less. The original natural ecosystems of the Midwest were stable late successional systems that retained nutrients and were dominated by perennial plants. In contrast, the agricultural systems that replaced them are highly disturbed immature (early successional) ecosystems of annual plants that characteristically “leak” nutrients (Loucks, 1979). As a result, streams and lakes in agricultural regions often contain excessive nutrient levels. Soil and above-ground carbon stores are low and levels of nutrients are high as a result of fertilizer application and ineffective uptake by monoculture cropping (Keeney, 1982; Simonis, 1988). These “leaky” systems are further exacerbated in the Midwest by extensive tile drainage. Tile drainage has decoupled wetland systems from their wetland/riverine interface on 37% of the agricultural lands in the Midwestern states discussed above, shunting contaminated upland drainage water directly to main river channels (Fausey et al., 1995). Bloomington, Illinois, is a prime example of a Midwestern city where drinking water from surface reservoirs is consistently plagued by high nitrate levels. The city relies on Lake Bloomington and Evergreen Lake for a large portion of its drinking water-sources which at times exceed the EPA MCL of 10 mg L−1 as a result of agricultural crop production on 86% of the watershed (USDA, 1991). From 1986 through 2003, the MCL for nitrate in Lake Bloomington was exceeded annually, and the concentration of the lake remained over the MCL from February through early July in 2001. Approximately 70,000 people receive their water from the Lake Bloomington municipal water treatment facility. The average demand is 20 million gallons per day (mgd) with a maximum of 24 mgd. When water in Lake Bloomington exceeds 10 ppm (the EPA MCL), the “solution is dilution” of the contaminated water with cleaner water from Evergreen Lake (the city’s backup water supply). Reducing nitrate levels in drinking water is important from the aspect of protecting human health, and a variety of methodologies exist to reduce nitrate levels. Often, expensive technical treatment makes sense for cities continually plagued with high nitrate levels in their drinking water and with the potential for being charged expensive fines. Failure to maintain drinking water quality standards resulted in a 1992 Illinois Environmental Protection Agency directive (authorized

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by the USEPA) to the city of Bloomington, Illinois, to reduce its maximum drinking water nitrate contaminant levels to below 10 mg L−1 . Although some progress was made, the city was not able to meet the directive’s deadline of June 1997. Two major issues in meeting the MCL were the cost of engineered systems designed to remove nitrates from drinking water and the potential effects these treatment costs would have on the economies of many agricultural communities. Ion exchange, the most cost-effective treatment available to remove nitrate is expensive and produces a salt (NaCl) brine that is expensive to treat (Letterman, 1999). The estimated cost in 2003 for construction of a 20 mgd plant was $8 million, with additional costs of up to $1 million for yearly maintenance and brine disposal (Jill Mayes, Bloomington, Illinois, Water Treatment Plant, personal communication). The plant would treat 40% or less of the water required to provide a blend concentration of 7 ppm NO3 − -N. In addition to being expensive, the construction of treatment facilities is also not in keeping with the EPA goal of “source water protection” (a watershed approach). Ion exchange solves the immediate problem of high nitrates, but it does not reduce overall nitrate loading to surface waters. Source water protection is the only way to reduce nitrogen loading to surface waters. Bloomington seeks less expensive, more ecologically sound watershed-based approaches to remove nitrate from water supply reservoirs. These include both fertilizer input management and wetlands (Mitsch et al., 2001, 2005). Constructed wetlands in tile-drained agricultural systems may prove to be a long-term practical, economical, and effective method to reduce surface water nitrate contamination. These “wetlands” are formed by berming an area adjacent to a stream and forming a small detention basin or holding pond that intercepts tile and surface drainage water before it enters the stream. The basin acts to reduce export of nitrate in drainage water through plant uptake and microbial transformation and degradation. Following wetland “treatment,” drainage water is slowly released to the stream through outlet flow. Research in Champaign County, Illinois, has shown that constructed wetlands established on former tall grass prairie soils (mollisols) can reduce total nitrogen loading from tile effluent by as much as 46% before it enters surface waters (37% of the N was denitrified (Xue et al., 1999) by the wetland pool, and an additional 9% was denitrified as it passed through the berm in seepage water) (Kovacic et al., 2000). Before wetlands systems are adopted as a universal approach to improve surface water quality in the United States, it is imperative to demonstrate their effectiveness and adaptability to key regions of the Midwest. The goal of this study was to construct, evaluate, and demonstrate the nutrient removal capacity of created wetlands on cropland soils (alfisols) that originally supported temperate deciduous forests. N, P and water input/output budgets were determined for two experimental wetlands constructed in conjunction with two experimental systematically tile-drained sub-basins in the Lake Bloomington watershed. The Bloomington constructed wetlands and their experimental watershed served as a valuable site to investigate nutrient runoff derived from both tile drainage and surface flow.

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2.

Methods and materials

2.1.

Description of the study area

The study site was located adjacent to Lake Bloomington in the northwest corner of McLean County, approximately 24 km north of Bloomington, Illinois. Lake Bloomington is characterized by numerous glacial moraines that form a rolling landscape. Slopes in the region normally range between 2 and 5% but can be as steep as 50%. The land area is 86% row crop agriculture in both maize and soybean, 4% pasture, 5% woodland, and 5% water and urban areas (U.S. Department of Agriculture, 1991). In this region, tile drains are installed in approximately 50% of the cropland area to allow drainage and to reduce high water tables, thus enhancing agricultural production (U.S. Department of Agriculture, 1991). Two experimental wetlands, wetland 1 (W1) and wetland 2 (W2), were created on Cambden silt loam forest soils. The watersheds were uniformly drained with subsurface drainage tiles in 1997. Drains were installed so that six 2-ha experimental plots (fields) were formed that could be monitored independently for flow and nutrient concentration. Drain spacing ranged between 22 and 27 m. Watershed 1, the smaller basin, contributed a total surface drainage of 3.76 ha, and a subsurface tile drainage area of 2.17 ha to W1 (Table 1, Fig. 1). W2 received drainage from watershed 2, a total surface area of 12.3 ha, and a total subsurface area of 12.1 ha. The W1 and W2 watersheds were in a maize/soybean rotation with maize planted in 1998. Watershed 1 received 197 kg N ha−1 as anhydrous ammonia (AA) for maize in 1998. Watershed 2 received 126 kg N ha−1 applied as AA for maize in 1998. In 1999, both

Table 1 – Wetland and watershed size and area description Wetland 1 Average depth (m) Volume (m3 ) Surface area to volume Tile drainage area (ha) Surface watershed area (ha) Wetland to tile drainage area Wetland to surface drainage area Wetland area (ha)

0.48 660 2.42 2.17 3.76 0.07 0.04 0.16

Wetland 2 0.52 1780 2.25 12.1 12.3 0.03 0.03 0.4

watersheds were planted in soybean and no fertilizer was applied.

2.2.

Wetland construction

Experimental wetlands were constructed in 1997. Wetland construction followed a modified conceptual design of Osborne and Kovacic (1993) (Fig. 2). Tile drainage lines (Fig. 1) were routed to the ground surface and a 3-m wide earthen berm was created to form wetlands that intercepted drainage waters. In addition, a surface water berm (water and sediment control basin, WASCOB) was built above the constructed wetland to collect surface runoff water (Fig. 2). Soil was excavated from the future wetland site to create the berms. Wetland bottoms consisted of excavated soils with a 6 in. layer of topsoil replaced. Deep zones (1.2–1.5 m deep) adjacent to the berms covered approximately 25% of the area in the wetlands. Wetland berms were constructed following USDA guidelines (Gray et al., 1992).

Fig. 1 – Conceptual plan view of wetlands showing basins and location of instrumentation.

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Fig. 2 – Conceptual design of experimental wetlands. (a) Modification of design from Osborne and Kovacic (1993) with water and sediment control basin (WASCOB) added.

Wetland bottoms were not planted immediately after construction; however, berms were seeded with oats following construction in the fall to provide a quick soil stabilizing cover. The following spring, berms and wetlands were seeded with barnyard grass (Echinochloa crusgalii L.). Wetland surface areas, volumes, and drainage area/wetland surface area ratios are given in Table 1.

2.3.

Instrumentation and wetland monitoring

Inlet and outlet structures were installed to monitor flow (Fig. 3). Two inlet stoplog structures (Inline Water Level Control StructuresTM Agri Drain Corporation, Adair, Iowa) were installed in each wetland. One was installed in the berm of the WASCOB and the other stoplog was installed at the tile dis-

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charge inlet. One outlet stoplog structure was installed in each wetland berm. All stoplog structures were fitted with a combination 30◦ v-notch/sharp crested weir. Each inlet and outlet monitoring station had a Campbell CR10 data logger, a Keller PSI pressure transducer, and an ISCO Model 2900 automatic water sampler. Data loggers continually recorded flow through the inlet structures on a 15 min interval, and automatic water samplers were used to collect flow proportional solution samples as frequently as every 15 min during runoff events. Outlet water samples were taken on a volumetric basis, with flow measured every minute in 1998 and every 2 min in 1999. Analyses indicated that tile and outlet flow nutrient concentrations were stable between precipitation events and often increased following major precipitation events. Tile discharge inlet, surface runoff, and outlet flows, along with estimates of seepage, evaporation, and precipitation, were used to determine water budgets for the wetlands. Precipitation was measured at the Lake Bloomington Water Treatment Plant, approximately 3.2 km from the site. Evaporation losses were determined using daily pan evaporation measurements provided by the Illinois State Water Survey in Champaign, Illinois. Evaporation from the wetland surface water was estimated using a pan coefficient (Ce ) of 0.70 (Brooks et al., 1991) and extrapolated over the surface of the wetlands to determine volume. Seepage was determined using the standard equation for saturated flow (K × A × i), where K is the hydraulic conductivity, A the area permitting seepage, and i is the hydraulic gradient. We used initial apparent K-values determined by Larson et al. (1999). An empirical water budget was used to predict the missing volume of water when inlet flow had ceased. The missing water volume was the difference

Fig. 3 – Design of berms, inlet, and outlet flow structures, and in-line flow control structure (stoplog). A 0.9 m deep by 1.2 m wide trench was excavated the length of the berm to form a key to retard seepage below the berm. Soil was applied to the berm in 15-cm layers and compacted with a sheep’s foot roller until the required height was reached. The berm in W1 was 2.4 m high, 3 m wide at the top and 19 m at the base, with a freeboard between the emergency spillway and the top of the berm of 0.6 m. The W2 berm was constructed 2.7 m high, 3 m wide at the top and 19 m at the base, with a freeboard between the top of the berm and the spillway of 0.4 m. The emergency spillways were 2.5 m wide. During construction of the berm, inlet and outlet pipes were installed as well as the in-line flow control structure. Berms to capture and allow the quantification of surface drainage were built above each wetland in a similar fashion to the main wetland berms described above using a water and sediment control basin (WASCOB) design (University of Illinois Extension Service, 2000). WASCOB berms were 1.8 m high, 3 m wide at the top and 12 m wide at the base and were similar to the berm depicted above. A tracked bulldozer with a pushing blade (Caterpillar Model D-8) was used to excavate and spread soil. The topsoil as well as the surrounding forest area provided a seed bank to revegetate the site.

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between initial standing volume and evaporation output over a 7 day time period. A final K-value was determined by adjusting the apparent K-value so the saturated flow equation results equaled the empirical evaporation estimate. The adjusted Kvalue was used along with the elevation estimates to estimate seepage. Fifteen-minute inlet and outlet flows were multiplied by inlet and outlet N, P and C concentrations to determine corresponding budgets.

2.4.

Sample analyses

Water samples were preserved for storage on the same day as collection following standard methods (APHA, 1995). Subsamples were filtered (Whatman GF/C glass fiber 1.2 ␮m pore size) and analyzed for NO3 − -N by ion chromatography; NH4 + -N by automated phenate; NO2 − -N by colorimetric techniques; dissolved P (orthophosphate phosphorus) by colorimetric techniques; and dissolved organic carbon (DOC) by DohrmannXertes DC-80 Analyzer (APHA, 1995). Dissolved P samples were filtered through 0.45 ␮m filters. Unfiltered sub-samples were analyzed for total N by the persulfate digestion method and total P by digestion with ammonium persulfate and H2 SO4 with detection of dissolved P. Organic N was calculated as total N minus the sum of nitrate, nitrite, and ammonium. Organic P was determined as the difference between total P and dissolved P.

3.

Results and discussion

3.1.

Water budget

Retention times averaged 10 and 14 d in W1; and 6.5–8 d for W2 in 1998 and 1999, respectively. During the 21-month study, approximately 69,000 m3 of combined flow entered the two wetlands with the maximum annual flow entering W2 in 1999 (32,892 m3 ) (Table 2). Expressed on a daily area basis, hydraulic loads were 0.03, 0.03, 0.05, and 0.04 m for W1 1998, W1 1999, W2 1998, and W2 1999, respectively. Flow events for W1 are depicted in Fig. 4 and represent the typical inlet and outlet flow patterns observed in the two wetlands during the study. Overall flow was greatest in June 1999. Peak flow was lowest in June 1998. Annual seasonal rainfall patterns were nearly identical for both years. Outlet flow contributed 64 and 68%, evaporation contributed 7 and 6% and seepage contributed 29 and 28% to the total output flow from wetlands W1 and W2, respectively. Surface area to watershed ratios were similar: 0.04 for W1 and 0.03 for W2. Estimated evapotranspiration losses were slightly larger than direct precipitation inputs to the wetlands in 1998, and slightly lower than precipitation inputs in 1999 (Table 2).

3.2.

Chemical budgets and concentrations

Annual wetland loading, export, and mass retention are presented for all constituents in Table 3. The 1998 estimates are based on 9 months of data collected from April through December. The 1999 estimates are based on a full year of data. Volume-weighted (flow-weighted) concentrations are presented in Table 4 for all constituents.

3.2.1. Nitrogen 3.2.1.1. Nitrate-N (NO3 − -N). Wetland NO3 − -N removal effi-

In 1998 and 1999, precipitation was 1038 and 957 mm, respectively. Flow monitoring for 1998 began 8 April, and ended 31 December 1998. During this period, precipitation totaled 699 mm. Although 63–67% of all precipitation occurred in winter and spring, the two seasons accounted for 92–100% of wetland hydraulic loading. During the course of the study, surface flow, tile flow, and precipitation were 47%, 43%, and 9% for W1; and 66%, 30% and 4% for W2, respectively. In 1998 and 1999, crop evapotranspiration increased through the late summer, when tile flow ceased and wetlands dried up.

ciencies ranged from 16% (W1 1999) to 43% (W2 1999). During the 21-month study, NO3 − -N tile loading, surface loading, and combined outlet plus seepage export were 746 kg, 104 kg and 545 kg, respectively. In all, NO3 − -N mass retention of the combined wetlands was 36%. Seepage accounted for only 2–3% of the total NO3 − -N export in W1 and W2, respectively. Low NO3 − -N seepage export was the result of excess hydraulic loads that were rapidly shunted to outlet flow during pulse flow events. While pulse flow is typical of Midwestern watersheds, the more sloping nature of the drainage probably accentuated runoff. In Champaign,

Table 2 – Water balances for Wetlands 1 and 2 from 16 April through 31 December 1998 and 1 January through 31 December 1999 Wetland and year

Tile discharge inlet

Surface

W1 1998 1999

2171 4449

2377 3649

W2 1998 1999

14351 21734

6112 10118

Outlet

Evaporationa

Seepage

559 765

−3612 −5197

−424 −440

−1062 −2868

9 357

0.18% 4.03%

1398 1040

−16381 −20762

−1061 −1100

−4212 −11703

206 −673

0.94% −2.05%

Precipitation

Values are in m3 . a b c

Estimated evapotranspiration. Balance refers to total inflow minus total outflow. Error resulting from flow monitoring and seepage determination.

Balanceb

Errorc

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Fig. 4 – Daily tile discharge inlet, surface inlet, and effluent flows for W1 and W2 from 8 April 1998 to 31 December 1999.

Table 3 – Constituent inlet load, surface load, outlet export, seepage export, and percent removal from two wetlands for sample period ranging from 8 April 1998 to 31 December 1999 Wetland

1998 Inlet load

W1 (kg year−1 ) NO3 -N NH4 -N Organic N

Surface load

Outlet export

1999 Seepage % Removal export

Inlet load

Surface load

Outlet export

Seepage export

% Removal

38.6 0.1 0.1

5.0 0.1 6.2

25.9 0.3 2.4

0.3 0.0 0.1

40 −50 60

60.9 0.2 0.5

5.8 0.3 5.6

54.5 0.4 0.9

1.6 0.0 0.0

16 20 85

38.8

11.2

28.6

0.4

42

61.6

11.6

55.1

1.6

23

PO4 -P Organic P

0.1 0.1

0.6 1.4

0.2 0.6

0.0 0.0

71 60

0.1 0.3

0.6 1.8

0.1 0.5

0.0 0.0

86 76

Total P

0.2

1.9

0.9

0.0

57

0.4

2.4

0.6

0.0

79

4.8

8.1

12.4

0.2

2

11.6

16.1

22.6

0.4

17

W2 (kg year ) 291.5 NO3 -N 0.3 NH4 -N Organic N 2.5

54.1 1.0 13.0

233.8 2.2 6.4

5.2 0.1 0.1

31 −77 58

355.1 0.4 2.4

39.1 0.7 9.2

214.8 1.1 6.5

8.8 0.0 0.6

43 0 39

294.3

68.1

241.9

5.5

32

358.5

49.0

222.3

7.2

44

1.1 2.0

1.0 2.5

1.0 2.3

0.0 0.0

52 49

0.3 0.6

0.6 2.7

0.6 1.9

0.0 0.0

33 42

Total N

TOC −1

Total N PO4 -P Organic P Total P TOC

3.1

3.6

3.4

0.1

48

0.8

3.4

2.5

0.0

40

43.4

28.0

62.9

1.3

10

43.8

41.2

78.6

0.0

8

Negative value indicates export from system greater than input.

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Table 4 – Flow-weighted average concentrations (mg L−1 ) for inlet and outlet constituents, April 1998 to 31 December 1999 Wetland

NO3 -N Inlet load

Surface load

NH4 -N

Total N

Organic N

Outlet export

Inlet load

Surface load

Outlet export

Inlet load

Surface load

Outlet export

Inlet load

Surface load

Outlet export

W1 98 99

17.8 13.1

2.1 1.5

7.2 10.5

0.07 0.04

0.05 0.08

0.09 0.04

17.9 13.3

4.7 3.1

7.9 10.6

0.04 0.14

2.59 1.49

0.65 0.11

W2 98 99

20.3 16.3

8.9 3.9

14.3 10.3

0.02 0.02

0.16 0.07

0.14 0.05

20.3 16.5

11.1 4.8

14.8 10.7

0.17 0.11

2.13 0.91

0.39 0.31

Wetland

PO4 -P Inlet load

Surface load

Total P

Organic P

TOC

Outlet export

Inlet load

Surface load

Outlet export

Inlet load

Surface load

Outlet export

Inlet load

Surface load

Outlet export

W1 98 99

0.04 0.03

0.24 0.13

0.06 0.03

0.10 0.10

0.82 0.56

0.26 0.13

0.07 0.06

0.58 0.42

0.18 0.10

2.2 2.6

3.4 4.4

3.4 3.3

W2 98 99

0.07 0.01

0.17 0.06

0.06 0.03

0.21 0.04

0.58 0.33

0.21 0.12

0.14 0.03

0.41 0.27

0.14 0.09

3.0 2.0

4.6 4.1

3.8 3.8

County, Illinois, the landscape is much flatter and seepage flow accounted for 15% of the NO3 − -N export (Kovacic et al., 2000). Average flow weighted concentrations of tile inflow ranged from 13.1 to 20.3 mg L−1 . Surface inflow ranged from 1.5 to 8.9 mg L−1 , and outflow ranged from 7.2 to 14.3 mg L−1 . Average annual outflow concentrations were consistently lower than tile inflow concentrations, but were always greater than surface inflow concentrations. For the study period wetland NO3 − -N concentration reduction was 42% (W1) and 31% (W2) (concentration reduction refers to the difference between wetland import and export concentrations). Outlet NO3 − -N concentration was reduced to below the EPA maximum contaminant level of 10 mg L−1 in W1 (1998) and close to the MCL in W1 (1999) and W2 (1999). These results are similar to those of Fink and Mitsch (2005), where NO3 − -N mass retention of an Ohio wetland draining forest soils was 41% and concentration reduction was 40%. Results are also similar to those of Champaign, County, Illinois (CCI) wetlands created on former tall grass prairie soils where overall wetland NO3 − -N retention was 38% and concentration reduction was 28% (Kovacic et al., 2000). Similar NO3 − -N mass retention ranging from 21 to 97% and concentration reduction ranging from 17 to 99% has also been reported in Illinois, Ohio, Louisiana, and North Carolina river diversion wetlands (Hey et al., 1994; Mitsch et al., 1998, 2005; Hunt et al., 1999). Wetland NO3 − -N inflow concentrations were similar to those reported by Kovacic et al. (2000). However, inlet NO3 − -N concentrations in most of the reported studies were generally lower than this study, ranging from 0.79 to 6.6 mg L−1 (Hey et al., 1994; Fink and Mitsch, 2005; Mitsch et al., 1998, 2005; Hunt et al., 1999). The high NO3 − -N concentrations reflect the large extent of tile drainage in central Illinois (David et al., 1997). Annual loading ranged from 27 g m−2 (W1 1998) to 99 g m−2 (W2 1999). Annual NO3 − -N retention ranged from 7 g m−2 (W1 1999) to 43 g m−2 (W2 1999). Loading was similar to that reported for created wetlands (Kovacic et al., 2000; Mitsch et al.,

2005; Fink and Mitsch, 2005). Daily NO3 − -N loading in W1 was 0.31 g m−2 d−1 and retention was 0.08 g m−2 d−1 . NO3 − -N loading in W2 was 0.58 g m−2 d−1 and retention was 0.22 g m−2 d−1 . In the CCI wetlands, Xue et al. (1999) identified denitrification as the major process removing NO3 − -N with a maximum denitrification rate of 0.28 g m−2 d−1 . Our results are within the potential denitrification range reported by Xue et al. (1999), and also within the range (0.001–0.48 g m−2 d−1 ) reported for river water treatment wetlands (Fleischer et al., 1994). Annual NO3 − -N export (the sum of surface runoff and subsurface/tile runoff loads) from watershed 1 was 19.1 and 29.6 kg ha−1 in 1998 and 1999, respectively. Export from watershed 2 was and 28.5 and 32.5 kg ha−1 in 1998 and 1999, respectively. Total nitrate export over the 2-year period was equal to 24 and 48% of the applied fertilizer N for watersheds 1 and 2. Other studies have reported similar watershed export of NO3 − -N in agricultural systems (Fausey et al., 1995; Drury et al., 1996; David et al., 1997).

3.2.1.2. Ammonium-N (NH4 + -N). In 1998 NH4 + -N export exceeded import load by 50% (W1) and 77% (W2). In 1999, W1 NH4 + -N mass retention was 20% while that of W2 was 0%. During the study period, tile loading, surface loading and export were 1.0 kg, 2.1 kg and 4.1 kg, respectively. Overall combined wetland NH4 + -N discharge was 32% greater than wetland NH4 + -N load. The lower levels of NH4 + -N in tile flow were expected. Ammonium is usually nitrified in field soils to form NO3 − -N. Only then does N derived from NH4 + -N subsequently leach through the soil and enter tile drains. Wetlands constructed for the treatment of tile drainage do not typically receive the large loads of NH4 + -N and that which does enter is derived from surface flow. Our results support this idea, with surface loads consistently greater than tile loads. In 1998, when more NH4 + -N exited the wetlands than entered them, NH4 + -N was probably derived from the suspension of unconsolidated sediments containing NH4 + -N and

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perhaps the decomposition of accumulated organic matter. In 1999, when wetland reduction of NH4 + -N occurred, it was most likely a result of sediment absorption, plant uptake and the nitrification of NH4 + -N to NO3 − -N (under aerobic conditions). Results of Hunt et al. (1999) and Kovacic et al. (2000) corroborate our findings reporting wetland export loads greater than import loads. However, Davis et al. (1981) found NH4 + -N removal efficiency in an Iowa marsh to be 78%. These results may be due to a lower hydraulic loading rate and increased residence time in the marsh. Ammonium-N export from the W1 watershed was 0.3 and 0.4 kg ha−1 in 1998 and 1999, respectively. Export from the W2 watershed was 2.3 and 1.1 kg ha−1 in 1998 and 1999, respectively. Except for W1 (1998), tile discharge inlet concentrations were lower than surface inflow concentrations. Overall flow weighted export concentrations were 46% (W1) and 68% (W2) greater than combined inflow concentrations.

3.2.1.3. Organic nitrogen. Wetland organic N removal efficiencies ranged from 39% (W2 1999) to 85% (W1 1999). Overall wetland organic N tile loading, surface loading and export were 5.5 kg, 34.0 kg and 17.0 kg, respectively. Overall the wetlands reduced organic N loading by 57%. Watershed export of organic N was 2.5 kg ha−1 (W1) and 6.5 kg ha−1 (W2) in 1998 and 0.9 kg ha−1 (W1) and 7.1 kg ha−1 (W2) in 1999. Tile inflow concentrations were always lower than surface inflow concentrations. Overall concentration reduction was 63% (W1) and 59% (W2). Organic N entered the wetlands primarily through surface runoff. The source of the organic N results from organic particulate matter and associated microbial biomass. Organic N in outlet water was probably derived from pre-existing plant and

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microbial pools as well as from the conversion of inorganic N to the organic N pool. Organic N accounted for only 4% of the total N entering the wetlands.

3.2.1.4. Total nitrogen. Total N tile discharge inlet loading, surface loading and outlet effluent export (Fig. 5) paralleled hydraulic loading. Low precipitation in 1998, resulted in low inputs of total N to the wetlands. The maximum daily total N inlet load for W1 (1998) was only 4 kg. In the dry spring of 1999 N movement in and out of the wetlands was minimal, however, pulse summer rain events resulted in both high N loading and export. Surface loading was higher in 1998, this probably resulted because spring rains saturated the soils resulting in greater runoff, while in 1999 the pulse events occurred in the summer, a time of dry soil and high evapotranspiration. N moved with subsurface flow and was carried by soil macropores into tile drains. Total N mass retention ranged from 23% (W1 1999) to 44% (W2 1999). Over the 21-month period the combined total N tile loading, surface loading and outlet plus seepage export were 753 kg, 140 kg, 548 kg and 15 kg, respectively. Overall total N mass retention was 37%. Watershed total N export in 1998 was 29 and 247 kg for W1 and W2, respectively. Total N export in 1999 was 57 and 230 kg for W1 and W2, respectively. Tile discharge inlet concentrations were always greater than both surface inflow concentrations and outlet flow concentrations (Table 4). Overall concentration reduction was 44% in W1 and 32% in W2. The observed fluxes closely paralleled hydraulic loading during the 2-year period. Flow weighted concentrations were similar to those of NO3 − -N (Table 4). Total N mass removal in our wetlands was nearly identical to that reported by Hunt et

Fig. 5 – Daily tile discharge inlet, surface inlet, and effluent N loading for W1 and W2 from 8 April 1998 to 31 December 1999.

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al. (1999) and Kovacic et al. (2000). However, it was lower than that reported in the Des Plaines wetlands (Hey et al., 1994) and the Eagle Lake marsh (Davis et al., 1981).

3.2.2. Phosphorus 3.2.2.1. Dissolved P. Retention of dissolved P ranged from 33% (W2 1999) to 86% (W1 1999) (Table 3). Overall dissolved P tile loading, surface loading, and export were 1.6 kg, 2.8 kg, and 1.9 kg, respectively. Wetland dissolved P mass retention was 57%. Watershed export of dissolved P in 1998 was 0.2 and 1.0 kg ha−1 for W1 and W2, respectively. In 1999 export was 0.1 and 0.6 kg ha−1 for W1 and W2, respectively. Tile inflow concentrations were always lower than surface inflow concentrations. Overall concentration was reduced by 72% (W1) and 53% (W2). During all 4 sample-years, dissolved P was effectively removed and volume-weighted concentrations of outflow were lower than inflow. Davis et al. (1981) reported a mass retention of 20% in a prairie pothole wetland receiving a load of 3.52 kg ha−1 year dissolved P. Mass retention for the CCI wetlands was 22%, while concentration reduction was negligible (Kovacic et al., 2000). The Bloomington wetlands, however, received a similar amount of dissolved P with greater removal efficiencies. Results indicate the Des Plaines wetlands and an Ohio agricultural wetland were similarly effective with dissolved P removal of 52–99% (Hey et al., 1994; Fink and Mitsch, 2005).

3.2.2.2. Organic phosphorus. Organic P removal ranged from 42% (W2 1999) to 76% (W1 1999). The 21-month combined wetland organic P tile load, surface load, and export were 3.0 kg, 8.4 kg, and 5.3 kg, respectively. Overall wetland organic P mass retention was 54%. Watershed organic P export in 1998 was 0.6 kg ha−1 for W1 and 2.3 kg ha−1 for W2. In 1999, the respective export was 0.5 and 1.9 kg ha−1 for W1 and W2. Tile discharge inlet concentrations were lower than surface inflow concentrations and lower than or equal to outflow concentrations. Overall concentration reduction was 61% in W1 and 50% in W2. No organic P was found in tile drainage to the CCI wetlands, and surface drainage was not measured because it was negligible on the site (Kovacic et al., 2000). In the present study, tile drainage did contribute organic P to the wetlands; however, surface drainage was the major source because of the greater slope. This was probably a result of residual organic matter carried in surface flow. 3.2.2.3. Total phosphorus. Total P mass retention ranged from 40% (W2 1999) to 79% (W1 1999). Overall combined wetland total P tile load, surface load, and export were 4.5 kg, 11.3 kg, and 7.5 kg, respectively. Overall wetland total P mass retention was 53%. Overall watershed export for the study period was 7.5 kg total P ha−1 . Tile inflow concentrations were always lower than surface inflow concentrations and lower than or equal to outflow concentrations. Overall concentration reduction was 58% in W1 and 50% in W2. The majority of the total-P entering the wetlands entered as surface runoff in the form of inorganic P adsorbed onto clay and sediment particles. Overall, 88% of the total P entered W1 as surface flow while 64% entered W2 as surface flow.

Results indicate that these wetlands were a significant sink for P. This is in contrast to the CCI wetlands which were neither a major source nor a sink for P (Kovacic et al., 2000). This contrast is probably a result of the difference in watershed topography and sediment derived through erosion. In the Champaign study, the wetland watersheds were extremely flat and surface runoff and therefore sediment export to the wetlands was low, while in the present study, the watersheds were steeper and sediment export to the wetlands was greater. The effective removal mechanism was probably sediment trapping (Higgins et al., 1993). Much of the P input was probably attached to sediment particles that were deposited in the wetlands. Plant uptake was negligible, because plant biomass production was initially low following the construction of the Bloomington wetlands. Davis et al. (1981) found that Eagle Lake marsh removed 11% of the overall 4.1 kg total P ha−1 entering as agricultural runoff. While the loading rates in the Bloomington wetlands were similar, the total removal was four to six times greater than in Eagle Lake marsh. The difference may be a result of wetland P saturation in the marsh. As wetlands age, the P budget tends to reach equilibrium, as the P leaving (from release of decomposing plant material) is equal to the P entering as sediment. Total P removal in the Des Plaines wetlands ranged from 52 to 99% (Hey et al., 1994). Loads were similar to those in this study, and much of the total P probably settled out and was removed by sedimentation. In a 0.6 ha wetland designed to remove total P from potato field surface runoff, 88% of a 48 kg total P load was removed during a 2-year period (Higgins et al., 1993). This system was highly effective because the P was associated with surface runoff and sediment and because it was designed with a grassed buffer for sediment reduction. Where high inputs of total P are of concern, an additional grassed buffer could help to improve P removal before entering the wetland. While our wetlands had a WASCOB basin associated with surface runoff collection, the vegetative cover may have been too sparse to be effective in P uptake.

3.2.3. Carbon 3.2.3.1. Total organic carbon (TOC). TOC mass retention ranged from 2% (W1 1998) to 17% (W1 1999). During the 21month study TOC tile load, surface load and export were 103.6 kg, 93.4 kg, and 178.4 kg, respectively. Overall wetland TOC mass retention was 9%. Watershed export of TOC in 1998 was 12.6 and 64.2 kg ha−1 for W1 and W2, respectively. In 1999, export was 23.0 and 78.6 kg ha−1 for W1 and W2, respectively. Tile discharge inlet concentrations were lower than both surface inflow and outlet export concentrations. Overall concentration reduction was 6 and 12% in W1 and W2, respectively. Our findings are in contrast to the CCI wetland study, where DOC retention was only 2% (Kovacic et al., 2000). Sixty percent of the TOC entered W1 in surface flow, while 44% entered W2 as surface flow during the study. It is assumed that allochthonus total organic C from tile flow entered the wetlands primarily as DOC. The bioavailability of this C derived from the upper 1.5 m of soil is probably lower than that of surface runoff derived C from the previous or present standing crop (McDowell and Likens, 1988; Baldock and Nelson, 2000; Royer and David, 2005). Autochthonus wetland C produced by

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algae and wetland plants, as well as surface runoff derived C, were probably most available to denitrifiers. Our concentrations are very similar to TOC concentrations reported for the CCI wetlands, with overall inlet concentrations that ranged from 2.1 to 3.6 mg L−1 and outlet concentrations that ranged from 3.4 to 4.6 mg L−1 . In an Iowa study, Davis et al. (1981) reported a wetland inflow DOC concentration of 18.0 mg L−1 from a maize/soybean watershed that also included both surface and subsurface drainage. Our results contrast with those of Davis et al. (1981), who found a prairie pothole wetland to be a major source of C, receiving 302 kg ha−1 of soluble organic C but exporting more than 516 kg ha−1 . The major source of C export was attributed to high biomass production in the pothole wetland as well as pulse flows following a period of drought.

3.3. Constructed treatment wetlands and water quality This study adds approximately 4 more water years to the relatively sparse agricultural cropland runoff wetland database. Trends in N dynamics between our wetlands constructed on temperate deciduous forest soils were nearly identical to those constructed on tall grass prairie soils (Kovacic et al., 2000), and were also similar to the removal rates of an Ohio agricultural runoff wetland also established on forest soils (Fink and Mitsch, 2005). In addition, N mass and concentration reduction were also in the range found in Ohio and Illinois diversion wetlands (Mitsch et al., 2005). Results suggest a commonality in Midwest wetland NO3 − -N removal dynamics. The partitioning of runoff into two components (tile inflow and surface runoff) has provided us with more insight into the dynamics of watershed topography and nutrient export. Besides NO3 − -N, all other nutrients entered the Bloomington wetlands primarily through surface flow. Apparently, the wetland/watershed nitrate dynamics between flat and sloping wetlands were similar, however, the dynamics of the remaining constituents appear to differ as a result of topography. The above information may help to develop more efficient wetland designs. For instance, the development of sediment and pre-filter basins may considerably reduce the input of NH4, organic N, organic P, and dissolved P into surface waters.

3.4. Contribution of wetlands to Midwest water quality Two important reasons for the creation of wetlands in the Midwest are water quality improvement, and the potential elimination of hypoxia in the Gulf of Mexico. There is a critical need to reduce nitrate loading in municipalities where drinking water reservoirs typically are in violation of drinking water standards during spring and summer. In many such areas, water discharged from agricultural runoff wetlands may be well below the MCL for drinking water. Annually, tile discharge inlet NO3 − -N concentrations in the Lake Bloomington watershed exceeded the EPA MCL of 10 mg L−1 . In 3 of 4 wetland years the flow weighted outlet NO3 − -N levels were reduced to below or very near to 10 mg L−1 . Absolute wetland loads were reduced, however, in all 4 wetland years.

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David et al. (1997) found that tile drainage accounted for nearly all the N entering the upper Embarras River watershed above Camargo, Illinois. Kovacic et al. (2000) estimated that the CCI wetlands with a wetland to watershed ratio of approximately 20:1 constructed on 50% of the tile systems would have a capacity to remove 18–23% of the inflow N. These wetlands would have had the potential to lower the annual nitrate concentration of tile drainage water entering the Embarras River from 7.8 to between 6.0 and 6.5 mg NO3 − -N L−1 in 1995, and from 10.1 to between 7.8 and 8.3 mg NO3 − -N L−1 in 1996. This is a significant reduction, considering these systems are relatively inexpensive, low-maintenance, gravity-fed systems that are long-lived due to low sediment loading. Our Bloomington study suggests that the effectiveness of constructed wetlands to reduce NO3 − -N loading to surface waters in steeper watersheds (where surface runoff contributes a larger component of flow and nutrients to export) is similar to those of the flatter Embarras River watershed CCI wetlands (Kovacic et al., 2000). Agricultural runoff wetlands that intercept tile drainage are a practical and simple tool to improve surface water quality and, in the long run, they should prove to be the most cost-effective method to reduce the nitrate problem in drinking water and N loading to the Gulf of Mexico. Doering et al. (1999) argue that if N fertilizer application restriction is used as a major method to reduce N export in the Mississippi Basin, then crop production would be transferred to another location. The use of agricultural runoff wetlands would allow farmers to maintain their current production practices and levels of fertilizer application while reducing nonpoint pollution export to surface waters. The costs of wetland installation can be considerable, but the cost share aspects of the Conservation Reserve Program (CRP) and the Wetlands Reserve Program (WRP) greatly increase the feasibility of wetland construction throughout the Midwest. The major expense in wetland development is the excavation and movement of soil. A theoretical 2acre (0.8 ha) 46 m wide wetland would require the movement of 1478 m3 of soil. Estimated excavation costs by traditional methods is $3.30–3.90 m−3 , for a total cost of between $4877 and 5764 per 0.8 ha wetland. Inlet and outlet flow control structures should also be incorporated at an estimated cost of $500/wetland (Inline Water Level Control StructuresTM Agri Drain Corporation, Adair, Iowa). The planting of wetland species has not been strongly supported by the WRP because results have shown that the seed bank and naturally dispersed seed frequently overwhelm the costly plantings. Therefore, the WRP relies on natural colonization for wetland development. Assuming construction costs will include excavation and control structures, and assuming the wetlands will be left to revegetate naturally, the cost of a 0.8-ha wetland would range from approximately $5377–6264. The area of the Lake Bloomington watershed is 180 km2 . If 5% of the tile-drained watershed (450 ha) was converted to 0.8-ha wetlands, including control structures, it would cost between $3,024,000 and 3,523,000 to construct treatment wetlands throughout the Lake Bloomington watershed. This could potentially reduce nitrogen loading by approximately 46% (Kovacic et al., 2000). Although this is a considerable cost, it is still only 38–44% of the cost to construct an ion exchange plant. Once constructed and stabilized by vegetation, wetland

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annual maintenance costs would be low. Of course, costs could be reduced by reducing the area converted to wetlands and the N mass retention goal. Constructed wetlands also have additional watershed benefits that technical treatment methods cannot offer. They can detain runoff, reduce flood peaks, and increase habitat for fish and wildlife and rare flora (Feierabend, 1989; Knight, 1992; Baker, 1992). In states such as Illinois, where 99.5% of the natural wetlands and wetland prairies have disappeared because of drainage and agricultural conversion (Mitsch and Gosselink, 1993), there is a critical need for the reestablishment of wetland habitat. In addition to improving water quality, constructed wetlands also provide areas for aesthetics and both nonconsumptive and consumptive (hunting and fishing) recreation, offering alternative agricultural business opportunities. As an example, the CCI wetlands site was also operated as a profitable private hunting preserve (Kovacic et al., 2000).

3.5.

Wetlands versus fertilizer management

Fertilizer management is one of the methods suggested to reduce loading in the Gulf of Mexico. While this can be an important method for nitrate removal, disagreement exists on the amount that fertilizer application rates must be reduced to affect a change in tile loading or watershed export. Using a retrospective model of the Mississippi River Basin, McIsaac et al. (2002) suggested that simply reducing N fertilizer application by 12% would reduce NO3 − -N export to the Gulf of Mexico by 33%. One factor attributed to this reduction was variable instream denitrification. In contrast to the McIsaac model, the U.S. Mathematical Programming (USMP) model (Doering et al., 1999) predicted that reducing N fertilizer application by 45% would reduce NO3 − -N export to the Gulf of Mexico by 20%. This model, which is based on edge-of-field studies, assumes a constant in-stream denitrification loss. Edge-of-field studies in tiledrained systems show that required reductions in N fertilizer use always exceed reductions in NO3 − -N export. Following a 67% reduction in N fertilizer use, Jaynes et al. (2001) reported a 40% decrease in NO3 − -N losses, while Clover (2005) reported a 49% decrease. Even when no N fertilizer was used, baseline tile drainage NO3 − -N losses were 9–22 kg ha−1 year−1 under corn/soybean agriculture (Baker et al., 1975; Baker and Johnson, 1981). Disproportional N loading, typical of much of the Midwest, is attributed to the area in agriculture and to the extensive areas of tile drainage (David and Gentry, 2000). For example, Illinois covers only 3% of the Mississippi River watershed but it contributes 15% to its annual total N load (David and Gentry, 2000). Edge-of-field studies (cited above) indicate that adoption of a 12% fertilizer use reduction program (in lieu of multiple mitigation strategies) would not substantially change NO3 − -N loading in low-order stream basins. Adopting such a program would ignore NO3 − -N loading in streams that feed many Midwestern municipal drinking water reservoirs. Dinnes et al. (2002) concluded that while fine-tuning fertilizer application for corn production has yielded reductions in NO3 − -N concentration, none will be sufficient to reduce

drainage water concentrations below 10 mg-N L−1 , the EPA maximum contaminant level for drinking water. Wetlands have proven effective in reducing tile drainage water NO3 − -N concentrations to acceptable drinking water levels (Kovacic et al., 2000). In this study, W1 reduced the volume weighted NO3 − -N concentration to below or near the MCL in 1998 and 1999, while W2 reduced the volume weighted NO3 − -N concentration to near the MCL in 1999. We believe that the creation of tile-drainage wetlands is one of the few solutions to water quality improvement in tile-drained regions of the Midwest. Results of Royer et al. (2004) support this idea. They found that previous studies of N transport to the Mississippi River basin (Alexander et al., 2000: David and Gentry, 2000) may have overestimated the loss of N through denitrification in headwater streams and its effect on N loading to the Mississippi. They suggest that habitats such as reservoirs and wetlands may support greater N removal than do headwater streams particularly in tile drained watersheds.

3.6. Wetlands for restoring the Mississippi, Ohio, and Missouri River Basins Mitsch et al. (2005) have determined that the removal of 40% of the N load to the Gulf of Mexico would require the creation/restoration of 22,000 km2 of wetlands in the Mississippi River Basin (approximately 1% of the basin land area). We believe this proposal is realistic and could provide a reasonable solution to Gulf hypoxia. However, studies conducted in central Illinois (Kovacic et al., 2000) indicate that a 46% mass retention of N would require an area equivalent to 5% of the corresponding drainage basin. The assumption for N mass retention differs somewhat from that of Mitsch et al. (2005), but the estimated area requirements are the same. We assume that the Midwest agricultural states are the major source of N to the Gulf of Mexico and created wetlands are the solution to source N reduction. If the Midwest covers 40,000,000 ha of land, using a 5% wetland to watershed ratio, 46% mass retention of Midwestern fertilizer N export would require 2,000,000 ha of wetlands. With increased information about the Midwest, this estimate can be modified and refined and the land area requirement may be reduced. If we can assume that an overwhelming majority of the N enters the Mississippi Basin through tile drainage systems, then we would focus our wetland creation in areas of tile drainage. If we know that 50% of the Midwest is tile-drained, then we would require only 1,000,000 ha of wetlands for a 46% N retention. This would leave 1,000,000 ha for wetlands that could be dispersed in other strategic areas throughout the Mississippi Basin. The cost of wetland construction in the Midwest would range from $6.7 to 7.8 billion. This investment would provide for long-term “environmental” homeland security.

3.7.

Research needs

We agree with Mitsch and Day (2006) that a program of adaptive management research should be undertaken. A series of studies is needed throughout the Mississippi River Basin, focusing on optimal locations for wetland creation and nitrogen removal. Before a large-scale created wetland project is funded and implemented, a number of critical questions need

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to be explored regarding location, design and function so the most likely wetland strategies can be developed, thus insuring the success of a long-term agricultural runoff and diversion wetlands program.

Acknowledgements The authors thank Jim Rutherford, Robert Hoffman, and the City of Bloomington for their cooperation and assistance in establishing the wetland study site. Thanks also go to Peggy Kovacic for editing this manuscript. This work was supported through grants from the Illinois Water Resources Center (Project Number: WRC-97-1) and EPA region 5 (Regional Geographic Initiative Program, grant # X985519-01).

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