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Published online 8 August 2008
Published in J Environ Qual 37:1974-1985 (2008)
DOI: 10.2134/jeq2007.0637
© 2008 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA
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TECHNICAL REPORTS

Waste Management

Landfill Cover Soil, Soil Solution, and Vegetation Responses to Municipal Landfill Leachate Applications

Neil W. MacDonalda,*, Richard R. Rediskeb, Brian T. Scullc and David Wierzbickid

a Biology Dep., Grand Valley State Univ., Allendale, MI 49401
b Annis Water Resources Inst., Grand Valley State Univ., Muskegon, MI 49441
c Annis Water Resources Inst., Grand Valley State Univ., Muskegon, MI 49441
d Michigan Dep. of Environmental Quality, Remediation and Redevelopment Div., Grand Rapids, MI 49503

* Corresponding author (macdonan{at}gvsu.edu).

Received for publication December 7, 2007.

   ABSTRACT
TOP
 NOTES
 ABSTRACT
INTRODUCTION
 Methods
 Results and Discussion
 Summary and Conclusions
 REFERENCES
 
Municipal solid waste landfill leachate must be removed and treated to maintain landfill cover integrity and to prevent contamination of surface and ground waters. From 2003 to 2007, we studied an onsite disposal system in Ottawa County, Michigan, where leachate was spray irrigated on the vegetated landfill cover. We established six 20-m-diameter circular experimental plots on the landfill; three were spray irrigated as part of the operational system, and three remained as untreated control plots. We quantified the effects of leachate application on soil properties, soil solution chemistry, vegetative growth, and estimated solute leaching. The leachate had high mean levels of electrical conductivity (0.6–0.7 S m–1), Cl (760–900 mg L–1), and NH4–N (290–390 mg L–1) but was low in metals and volatile organic compounds. High rates of leachate application in 2003 (32 cm) increased soil electrical conductivity and NO3–N leaching, so a sequential rotation of spray areas was implemented to limit total leachate application to <9.6 cm yr–1 per spray area. Concentrations of NO3–N and leaching losses remained higher on irrigated plots in subsequent years but were substantially reduced by spray area rotation. Leachate irrigation increased plant biomass but did not significantly affect soil metal concentrations, and plant metal concentrations remained within normal ranges. Rotating spray areas and timing irrigation to conform to seasonal capacities for evapotranspiration reduced the localized impacts of leachate application observed in 2003. Careful monitoring of undiluted leachate applications is required to avoid adverse impacts to vegetation or soils and elevated solute leaching losses.

Abbreviations: AWRI, Annis Water Resources Institute • EC, electrical conductivity • MDEQ, Michigan Department of Environmental Quality • SRP, soluble reactive phosphorus • TKN, total Kjeldahl nitrogen • TOC, total organic carbon • VOC, volatile organic compound


   INTRODUCTION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
Methods
 Results and Discussion
 Summary and Conclusions
 REFERENCES
 
MUNICIPAL solid waste landfill leachate must be collected and treated to prevent the contamination of local surface and ground waters, often for many years after the closure and capping of a landfill. Landfill leachate disposal needs to be carefully managed because leachate contains varying concentrations of dissolved organic matter, plant nutrients, heavy metals, and potentially toxic organic compounds (Lema et al., 1988; Kjeldsen et al., 2002). Disposal of municipal solid waste landfill leachate is carefully regulated in Europe (Duggan, 2005) and in the United States must comply with the Resource Conservation and Recovery Act and Clean Water Act administered through the U.S. Environmental Protection Agency and authorized state agencies. After collection, landfill leachate typically is transported to sewage treatment plants, treated in on-site facilities, or recirculated through the landfill (e.g., Lema et al., 1988; Robertson et al., 1995; Townsend et al., 1996; Wiszniowski et al., 2006). Although these treatment methods are effective, they are expensive and may be unnecessary if landfill leachate can be spray irrigated on the vegetated landfill cover (Shrive et al., 1994; Adarve et al., 1998; Jones et al., 2006).

Spray irrigation of leachate offers advantages in that the required treatment is accomplished onsite through the combined actions of evapotranspiration, infiltration, microbial degradation, retention of pollutants by the soil, and plant uptake of nutrients, metals, and organics (Jones et al., 2006). In many cases, spray irrigation of leachate may improve the growth of landfill cover vegetation, increasing its effectiveness as a protection against soil erosion (Cureton et al., 1991; Shrive et al., 1994; Revel et al., 1999). Concern exists, however, that applications of highly concentrated leachates (Maurice et al., 1999; Stephens et al., 2000; Bowman et al., 2002) or leachates containing potentially toxic metals and organic chemicals (Adarve et al., 1998; Smith et al., 1999a, 1999b) may adversely affect soil properties, landfill vegetation, or ground water. Excessive application of leachate may be implicated in increased emission of greenhouse gases such as CH4 and N2O (Watzinger et al., 2005). The composition of municipal solid waste landfill leachate is highly variable from landfill to landfill (Lema et al., 1988; Kjeldsen et al., 2002) and over time within a landfill (Statom et al., 2004), necessitating careful monitoring of leachate chemistry before and during land application (Jones et al., 2006).

The Michigan Department of Environmental Quality (MDEQ) assumed the responsibility for completing the closure and capping of a small (9-ha) municipal solid waste landfill (SE 1/4 Section 12, T6N R13W, Ottawa County, Michigan; 42.9167° N, 85.7872° W) after the original owner and operator declared bankruptcy and abandoned the property. The landfill was capped with an impermeable membrane and covered with successive layers of sand, sandy loam subsoil, and sandy loam topsoil, and the landfill cover was seeded with a diverse mixture of native and agronomic varieties of grasses and forbs in the fall of 2002. After capping, leachate was pumped from four wells fed by a series of underdrains and stored onsite in a large tank. To dispose of this leachate, the MDEQ developed a spray irrigation system to apply it to the vegetated landfill cover. Environmental concerns about leachate applications were especially relevant to this site because of its location in the Grand River floodplain and close proximity to adjacent wetlands, parks, and residential areas southwest of Grand Rapids, Michigan. The MDEQ conducted preliminary onsite leachate applications during the fall of 2001 and summer of 2002 without any apparent adverse effects. The MDEQ expanded the onsite disposal to an operational system in 2003 and contracted with the Grand Valley State University Annis Water Resources Institute (AWRI) to study the effects of leachate application, with the primary concerns being the potential long-term impacts on soils and vegetation. Unlike many previous literature reports on the effects of landfill leachate irrigation, we studied an operational leachate disposal system as opposed to an experimental or trial system. The objective of our study was to determine the effects of leachate irrigation under field conditions on landfill cover soil and soil solution chemistry, plant responses, and solute leaching and to use these observations to help manage leachate applications at the study site. The results of this study are relevant to other situations where leachate disposal on the landfill cover is desirable or necessary.


   Methods
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Methods
Results and Discussion
 Summary and Conclusions
 REFERENCES
 
Experimental Design
In May 2003, we initiated a replicated study on the landfill cover to quantify the effects of the operational leachate applications. We studied three blocks of 20-m-diameter circular plots; one plot from each block was randomly selected to be spray irrigated as part of the operational disposal system, and one plot from each block remained as an untreated control plot. We did not irrigate control plots with additional water or nutrients because the relative effects of such treatments are well known (Shrive et al., 1994; Godley et al., 2004a; 2004b; 2005; Zalesny et al., 2007), and supplemental water and nutrients would not be applied separately from the landfill leachate in an operational disposal system. Untreated control plots thus provided a realistic comparison to evaluate the integrated effects of leachate application on the landfill cover soil–plant system under field conditions. The centers of paired plots were 30 m apart, and all plots were located on the landfill cover between elevations of 200 and 204 m above mean sea level. Because all plots were located on the upper slopes of the landfill cover, any subsurface or overland flow would move downhill and not toward an adjacent plot. Each block was similar with respect to aspect (block 1 {approx} 1°, block 2 {approx} 90°, block 3 {approx} 210°), and all plots had similar soil characteristics and slopes (13.2 ± 1.8%).

Leachate Application
The irrigated plots had stand-mounted spray heads that applied leachate to circular areas with effective spray radii of 10 m. In 2003, leachate was applied to the study plots on each irrigation date, approximately once per week between July and November at a mean rate of 2.14 cm per application. In 2004, 2005, and 2006, the study plots were included in a rotating irrigation schedule, and the plots were irrigated approximately once every 3 wk between June and September at mean rates ranging from 1.01 to 1.44 cm per application. To monitor leachate chemistry and volumes applied, leachate was collected in three systematically located bulk collectors per plot during each leachate application. Bulk collectors consisted of 4-L polyethylene bottles fitted with 17-cm diameter funnels placed at 3, 6, and 9 m from plot centers along one radius of each plot. On each irrigation date, leachate volumes were recorded for each collector, and the leachate from all plots then was combined to produce one composite sample per date for analysis. Bulk collectors were rinsed with deionized water after each use and stored in plastic bags between irrigation dates to prevent contamination. On each irrigation date, leachate samples were collected immediately after cessation of irrigation and transported in coolers to analytical laboratories.

Soil Sampling and Analyses
We sampled the surface (0- to 25-cm) and subsurface (25- to 50-cm) soils on all plots with bucket augers before spray irrigation commenced in 2003. Initial soil samples were taken in a random cardinal direction (N, E, S, or W) at a 1-m distance from three random points in each plot third and composited by plot thirds to form three subsamples for each depth per plot. We analyzed these samples for texture (hydrometer), pH (1:1 soil/water), organic carbon (H2SO4–K2Cr2O7 oxidation), electrical conductivity (EC) (saturated paste extract), and available water-holding capacity (to –1.5 MPa) following standard soil analytical methods (Page et al., 1982; Klute, 1986). Repeated measurement errors, based on 10% sample replication, were 2.0% for sand, 7.5% for silt, 4.7% for clay, 0.9% for pH, 5.1% for organic C, 5.8% for EC, and 14.0% for available water capacity. Initial soil samples from each plot were tested for total trace metals, and two pre-irrigation soil cores from the surface 10 cm of each plot were tested for the presence of volatile organic compounds (VOCs). Soil samples for metal analyses were stored in acid-washed, wide-mouth, 250-mL glass jars before digestion and analysis. Soil samples to be analyzed for VOCs were placed in 40-mL glass vials with Teflon septa and preserved in the field (10 g soil:10 mL methanol) before extraction and analysis using gas chromatography mass spectrometry.

We sampled surface (0- to 25-cm) soils again in October of 2003, 2004, and 2005 at the end of each irrigation season. The October soil samples were taken each year in sequential cardinal directions at 1-m distances from the original random points in each plot using bucket augers and composited by plot thirds to form three subsamples per plot. We analyzed the annual post-irrigation samples for pH and EC as before, obtaining repeated measurement errors of 0.6% for pH and 13.7% for EC. All post-irrigation composite surface soil samples were analyzed for metals. We tested one or more randomly located cores taken from the surface 10 cm of soil on each plot in October of each year for the presence of VOCs. Soils were not sampled in 2006 because of funding limitations.

Soil Solution Sampling and Solution Analyses
We installed three systematically located 60-cm ceramic suction cup soil water samplers (Soilmoisture, Inc., Santa Barbara, CA) on each plot in May 2003 to collect soil solution at a depth of approximately 50 cm, representing solutions that had passed through the primary rooting zone of the vegetated landfill cover. We collected samples from the soil water samplers on a weekly to bi-weekly basis from mid-May to mid-December in 2003 and from late March to mid-December in 2004, 2005, and 2006. Leachate and soil solution samples to be analyzed for VOCs were placed in 40-mL glass vials with Teflon septa, preserved with HCl, and refrigerated at 4°C before analysis. Samples for other analyses were stored in acid-washed polyethylene bottles, preserved with HNO3 (metals) or H2SO4 (total Kjeldahl nitrogen [TKN] and NH4–N) as needed, and refrigerated at 4°C or frozen (soluble reactive phosphorus [SRP]) before analysis. Soil solution and leachate samples were analyzed for pH, EC, Cl, SO4–S, NO3–N, NH4–N, Ca, Mg, SRP, total P, and total organic carbon (TOC) at AWRI following standard, quality-assured protocols (APHA, 1992). Chlorine, SO4–S, and NO3–N were determined with ion chromatography; Ca and Mg by flame atomic absorption spectrometry; NH4–N, SRP, and total P by automated colorimetry; and TOC using a total organic carbon analyzer. Matrix spikes and matrix spike duplicates for all analytes were analyzed at a frequency of 10% with precision limits of ±15% relative standard deviation and accuracy control limits of 90 to 110% recovery. The MDEQ Environmental Laboratory (Lansing, MI) analyzed leachate samples for TKN and soil solution and leachate samples for total trace metals (Ag, As, Ba, Cd, Cr, Cu, Hg, Pb, Se, Zn) and 72 VOCs following USEPA-approved analytical methods and QA/QC protocols (USEPA, 1996). After sample digestion, Hg was analyzed using cold vapor atomic absorption spectrometry, other metals were determined using inductively coupled plasma mass spectrometry, and TKN was determined by automated colorimetry. Volatile organic compounds were determined by gas chromatography mass spectrometry.

Plant Biomass Sampling and Tissue Analyses
In 2003, 2004, and 2005, we quantified aboveground plant biomass on all plots in September at the peak of biomass production, taking vegetation samples from the same locations scheduled for soil sampling in October of each year. All vegetation was clipped on nine 0.1-m2 quadrats per plot, dried at 70°C for 48 h, and weighed to estimate plot mean biomass. Biomass data were converted to a g m–2 basis to facilitate comparisons using a standard areal unit, and these data were statistically analyzed in that form. To monitor metal accumulation in plants, red fescue (Festuca rubra L.), timothy (Phleum pratense L.), red clover (Trifolium pratense L.), and white clover (Trifolium repens L.) were separately sampled on each plot as indicators in 2003 and 2004. These four species established on all plots with an average coverage of about 5 to 10% for each species within a matrix of other grasses and forbs. In 2005, as a result of extremely dry weather, not all of these species persisted on all plots, and we determined metal concentrations on four representative subsamples per plot taken from the plant biomass samples after weighing. Oven-dried (70°C) plant tissue samples were cut into short lengths with Teflon scissors, ground in a ceramic mortar and pestle, and stored in 125-mL acid-washed glass jars before digestion and analysis for total trace metals and TKN. Plant elemental concentrations were determined and are expressed on a dry weight basis. Vegetation was not sampled in 2006 because of funding limitations.

Leachate Constituent Deposition and Leaching Calculations
Leachate constituent deposition for each plot was calculated from mean constituent concentrations for each irrigation date and the irrigation depths recorded on each plot. We estimated total leachate constituent deposition on each plot for an irrigation season by summing the values calculated for each irrigation date. We did not statistically analyze leachate deposition data because the annual estimates largely reflect the annual variation in amounts of leachate applied and differences among years in leachate constituent concentrations. We estimated potential evapotranspiration using Thornthwaite's equation and calculated monthly leaching losses from the upper 50 cm of the landfill cover soils using the water balance method (Mather, 1978). Annual solute fluxes from the landfill cover soils were estimated from volume-weighted plot mean seasonal (July–December, January–June) soil solution concentrations and plot estimates of seasonal leaching losses (MacDonald et al., 1992). Climatic data for these calculations were obtained from an onsite rain gage (Universal Recording Rain Gage, series 5–780; Belfort Instrument Co., Baltimore, MD) and from the National Weather Service station in Grand Rapids, MI, which is approximately 15 km southeast of the landfill.

Statistical Analyses
Parametric one-way ANOVA was used to compare leachate inorganic chemistry among years. Kruskal-Wallis nonparametric one-way ANOVA was used for leachate VOC concentrations where data could not be analyzed using parametric statistics because of below-detection analytical results on one or more dates for these data. We used parametric one-way ANOVA to compare initial soil physical properties between treatments and parametric repeated-measures ANOVA to compare effects of treatments on soil, soil solution, plant biomass, plant metal, and solute leaching variables through the years sampled for each variable. These analyses incorporated a randomized complete block design, treating each of the three plot pairs as a block. Data were tested for homogeneity of variance with Bartlett's test. Most variables met this assumption, except for the mean annual volume-weighted soil solution concentration and seasonal solute leaching data, which met this assumption after being ln-transformed (ln X + 1 if data included zeros). We used Tukey's multiple comparison test to judge differences among means, with significance for all analyses accepted at P < 0.05. Statistical methods followed Steel and Torrie (1980), and statistical analyses were performed using SYSTAT (version 4; Wilkinson, 1989).


   Results and Discussion
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Methods
 Results and Discussion
Summary and Conclusions
 REFERENCES
 
Leachate Chemistry
Leachate chemistry varied significantly through time for most variables, with leachate TOC, Cr, Cu, Pb, and Zn concentrations tending to decrease during the study period (Table 1 ). Statom et al. (2004) also observed decreases in many inorganic ions and TOC through time in a Florida landfill cell while pH levels remained stable. In our study, annual mean leachate pH levels varied slightly around an overall mean of 8.03 (Table 1), indicating that the landfill was well into the methanogenic phase of refuse decomposition (Kjeldsen et al., 2002). The leachate had elevated concentrations of Cl, NH4–N, and TOC, whereas concentrations of NO3–N, SRP, and total P were much lower (Table 1), similar to other municipal landfill leachates (Fatta et al., 1999; Kjeldsen et al., 2002). Although concentrations of NH4–N varied slightly through time in our study, they typically remain high in leachate for years after landfill closure (Kjeldsen et al., 2002). As in other landfill leachates, most or all TKN was accounted for by NH4–N (Table 1) (Lema et al., 1988). Metals were present in the low concentrations (µg L–1) that are typical of many methanogenic-phase municipal solid waste landfill leachates (Table 1) (Fatta et al., 1999; Kjeldsen et al., 2002; Statom et al., 2004).


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  		Table 1. Chemical characteristics of leachate applied at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, 2003–2006.

 
Mean leachate EC increased slightly between 2003 and 2005 (Table 1), probably as a result of the landfill cover limiting infiltration of precipitation. High EC is problematic because it may adversely affect vegetation or soil properties if leachate is applied at high rates or over long periods (Jones et al., 2006). For example, Bowman et al. (2002) observed that concentrated landfill leachates (1.4–1.7 S m–1) needed to be diluted to <0.36 S m–1 to avoid damage to vegetation and soils during sustained leachate applications near Sydney, Australia. If leachate conductivity continues to increase over time in the landfill we studied, leachate may require dilution to avoid adverse effects. Alternatively, maintaining soil solution concentrations at safe levels by balancing leachate applications with precipitation inputs (Godley et al., 2005) may continue to be feasible if leachate conductivity stabilizes around 0.7 S m–1.

Most VOC concentrations in the leachate collected on the plots after spray irrigation were at the low end of the range reported for other landfill leachates (Kjeldsen et al., 2002), reflecting biodegradation and volatilization losses during pre-irrigation storage and additional volatilization during application (McBride et al., 1989a; Thorneby et al., 2006). Most VOCs in leachate collected after spray irrigation were only sporadically present above detection limits, and not all compounds were consistently observed in all years (Table 2 ). Compounds like 2-butanone, 2-propanone, tetrahydrofuran, and tertiary butyl alcohol are common organic solvents that are water soluble, volatile, and degradable (Nyer, 1992; Verschueren, 2001). Although diethyl ether is volatile, it is the only VOC we detected in leachate that is considered resistant to biodegradation (White et al., 1996; Hardison et al., 1997). Fuel-related compounds (e.g., 1,2,4-trimethylbenzene; 1,4-dichlorobenzene; chlorobenzene; 2-methylnaphthalene; ethylbenzene; m-, o-, and p-xylene; toluene; and benzene, which were observed sporadically) are among those most frequently found in landfill leachate (Kjeldsen et al., 2002). Bromodichloromethane, bromoform, and dibromochloromethane, observed only in 2004, are typically formed as byproducts of water chlorination and are not commonly found as environmental contaminants (Verschueren, 2001). If not volatilized or biodegraded, most of the VOCs detected can leach to ground water. None of the VOCs identified in the spray-irrigated leachate to date are considered to present major environmental hazards because of their low toxicities and generally rapid biodegradation rates (Verschueren, 2001).


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  		Table 2. Volatile organic compound concentrations of leachate after spray irrigation at the landfill in Section 12, T6N R13W, Ottawa County Michigan, 2003–2006.

 
Leachate Application and Constituent Deposition
Between July and November 2003, approximately 32 cm of leachate were applied to the irrigation plots (Table 3 ). This initially high application rate was required to remove leachate from the landfill to prevent the rupture of the landfill cover. At the end of the 2003 irrigation season, soil solution conductivity and solute concentrations had increased significantly on irrigation plots. To reduce the potential for high solute leaching losses in subsequent years, we recommended timing leachate applications to match periods of greatest evapotranspirational demand (June–September). We also recommended reduced rates of leachate application to maintain soil solution conductivities at levels that would not inhibit plant growth (e.g., Adarve et al., 1998; Bowman et al., 2002). To avoid exceeding the average unsatisfied evapotranspirational demand during June, July, and August, we recommended that total leachate applications should not exceed 2.4 cm per irrigation date or a total of 9.6 cm in any one spray area during an irrigation season. To allow for adequate onsite leachate disposal, we suggested that each spray head should be rotated among several spray areas to distribute the leachate applied. The MDEQ adopted these recommendations, and rotation of each spray head among three spray areas effectively reduced total irrigation depths per spray area in 2004, 2005, and 2006 as compared with 2003 (Table 3). Although leachate application rates during 2003 exceeded the 20-cm maximum suggested by Watzinger et al. (2005) to avoid increased soil CH4 and N2O emissions, application rates in 2004, 2005, and 2006 were well below this level. Mean application rates between 2004 and 2006, ranging from 5.0 to 8.6 cm, still exceeded the 2.5 to 5.0 cm yr–1 suggested by Jones et al. (2006) as a sustainable rate of undiluted leachate application to a land-based system. Spray head rotation, by allowing a 3-wk resting period between successive leachate irrigations, may permit higher application rates than those suggested by Jones et al. (2006), but additional monitoring is required to determine if these higher rates are sustainable.


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  		Table 3. Estimated annual leachate constituent deposition{dagger} rates on irrigation plots at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, 2003–2006.

 
Annual constituent deposition for metals and SRP was below 1 kg ha–1 during the study period (Table 3). Annual total P, NO3–N, and SO4–S leachate deposition rates were at moderate levels (<30 kg ha–1), but Ca, Mg, NH4–N, TKN, TOC, and Cl were deposited at much higher rates (Table 3). In comparison, ambient wet plus dry atmospheric ionic deposition rates in southern Michigan range from around 2 kg ha–1 for Cl to 14 kg ha–1 for NO3–N + NH4–N (MacDonald et al., 1992). High deposition of leachate Cl (up to 2648 kg ha–1 in 2003) is of concern because it contributes to elevated soil solution EC, which may adversely affect plant growth (Cureton et al., 1991; Stephens et al., 2000). The high NH4–N deposition (up to 961 kg N ha–1 yr–1 in 2003) increased the risk of NO3–N losses to ground water and the potential for NO3–N movement off site. Nitrogen application rates observed in this study, however, were less than those reported by Bowman et al. (2002), who applied 1300 kg N ha–1 in leachate per year without adverse effects. As a result of spray area rotation, constituent deposition rates were greatly reduced in 2004, 2005, and 2006 as compared with 2003 (Table 3).

Initial Soil Properties
Surface and subsurface cover soils were slightly basic (pH 7.7–8.1) sandy loams, with the surface layer being substantially higher in organic carbon than the subsurface layer (Table 4). Both soil layers had available water capacities consistent with their textures and organic carbon contents, providing an average total available water capacity of 13 cm in the upper 50 cm of cover soil. Soil ECs were around 0.05 S m–1, representative of uncontaminated soils (Munshower, 1993). Soil physical and chemical properties on the plots assigned to irrigation treatments were very similar to and did not differ significantly from those on the plots assigned to be controls (Tables 4 and 5 ). Pre-irrigation soil metal concentrations were typical of concentrations in uncontaminated soils (Munshower, 1993), and no VOCs were detected.


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  		Table 4. Physical characteristics of surface and subsurface cover soils at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, May 2003.

 

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  		Table 5. Effects of leachate irrigation on surface soil (0–25 cm) properties at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, October 2003, 2004, and 2005.

 
Soil and Soil Solution Chemistry
Soil pH was not significantly affected by leachate irrigation, although pH tended to be somewhat higher on irrigation plots in 2004 and 2005 (Table 5). The slightly alkaline condition of the leachate (Table 1) and the elevated levels of Ca and Mg deposition (Table 3) may account for slight pH increases on the irrigated plots in our study. Winant et al. (1981) observed that pH increased from 4.7 to 5.8 in forest soils treated with pH 5.45 leachate (Menser et al., 1979a, 1979b) and concluded that the increased pH represented a beneficial effect on soil fertility. In contrast, laboratory treatment with more alkaline leachate (pH 10.3) increased soil pH to 10 and decreased soil hydraulic conductivity (Chan, 1982), demonstrating that the potential for adverse soil effects is dependent on leachate chemistry.

Electrical conductivity increased and then remained elevated in surface soils on irrigated plots as compared with control plots (Table 5), but soil ECs on irrigated plots decreased significantly after 2003. This was an intended result of the reduction in the amount of leachate applied beginning in 2004. Fall soil ECs on the irrigated plots remained below the 0.4 S m–1 threshold for adverse affects (Munshower, 1993; Jones et al., 2006), but mean soil solution conductivities exceeded this value in 2003 and approached it again in 2005 (Table 6 ). These data demonstrate the need for soil and soil solution monitoring to manage the application of leachate at rates that are not harmful to the vegetative cover.


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  		Table 6. Effects of leachate irrigation on volume-weighted mean annual soil solution concentrations at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, 2003–2006.

 
Leachate application had no statistically significant effects on soil metal concentrations during the study period, although there was a trend toward gradually increasing metal concentrations on the irrigated plots (Table 5). Godley et al. (2004b) and Thorneby et al. (2006) also reported minimal effects of leachate application on soil concentrations of metals and attributed this to the low rates of metal addition. Calace et al. (2001) observed that treatment with landfill leachate increased soil adsorption of Cu and Cd, so any increases in soil metal concentrations that take place over time may be the combined result of metal deposition along with an increased capacity to retain these metals. We sampled the entire upper 25 cm of soil, which may have masked more pronounced differences in metal concentrations between treatments that could have been present near the soil surface. For example, Winant et al. (1981) found increases in pH, Ca, Mg, K, Na, Sr, and Zn in the upper 5 cm of soils irrigated with landfill leachate, although Mn, Fe, and P concentrations were unaffected. In our study, metal concentrations in the upper 25 cm of cover soil remained near or below statewide default background levels on control and irrigation plots (Table 5). No VOCs from the applied leachate were detected in surface soils or soil solution samples in any year. Although volatilization losses from soil solution samples in suction cup lysimeters may reduce concentrations of VOCs with low vapor pressures (Wood et al., 1981; Weihermuller et al., 2007), the total absence of VOCs in soil and soil solution samples is consistent with substantial losses from post-application volatilization or biodegradation and is similar to observations by McBride et al. (1989a) and Thorneby et al. (2006).

Soil solution ECs (Fig. 1a ) and NO3–N concentrations (Fig. 1b) on irrigated plots followed a very pronounced seasonal progression, reaching peaks during irrigation seasons but declining as late fall rains diluted the soil solution and spring snowmelt leached solutes from the soil. Godley et al. (2004a, 2004b, 2005) reported similar short-lived effects of seasonal leachate irrigation when interspersed with periods of leaching during the winter and suggested that the resulting dilution of leachate constituents would reduce the impacts on receiving surface or ground waters. In our study, 2006 soil solution conductivities and NO3–N concentrations did not approach the high levels observed on irrigation plots in the three previous years due to the reduced leachate application rates combined with the diluting effects of high rainfall in July, September, and October. Leachate application produced elevated soil solution concentrations of Cl, NO3–N, and TOC, but average concentrations of these solutes on irrigation plots tended to decrease through time as leachate application rates were reduced (Table 6). Average concentrations of Ba, Cr, Cu, and Zn also increased in soil solution on irrigated plots but remained below drinking water criteria levels in all years (Table 6). Although average concentrations of NO3–N remained above drinking water criteria on irrigated plots, soil solution concentrations of NH4–N were below 0.3 mg L–1 on irrigated plots in all years (Table 6). Low NH4–N soil solution concentrations accompanied by increases in NO3–N as a result of nitrification also were observed by Robertson et al. (1995), Bowman et al. (2002), and Godley et al. (2004b, 2005) in studies of land-applied leachate. After application, NH4–N also can be lost from leachate via volatilization or through adsorption in the soil (Tyrrel et al., 2002). Effective removal of NH4–N is an important part of treatment because leachate toxicity is strongly related to the concentration of NH4–N (Kjeldsen et al., 2002; Ward et al., 2002). Concentrations of NH4–N, along with those of SRP and total P, tended to be higher on control plots in 2005 than in other years, possibly as a result of lower precipitation, drier soil conditions, and reduced plant uptake of these nutrients during that year.


Figure 1
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  		Fig. 1. Seasonal trends in soil solution (a) electrical conductivity and (b) NO3–N concentrations at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, June 2003 to April 2007. Error bars represent 1 SD.

 
Constituent Leaching Losses
Solute leaching estimates did not differ significantly between control and irrigated plots for SO4–S, Ca, Mg, SRP, total P, or Zn (Table 7 ). Leachate-irrigated plots had significantly elevated leaching losses of Cl, Cr, Cu, NH4–N, and NO3–N in all or most years, whereas As, Ba, and TOC leaching losses were greater on irrigated plots only in 2003–2004 (Table 7). Estimated annual net leaching losses of Cl and SO4–S from irrigation plots were substantially greater than annual leachate loading rates (Table 3). For SO4–S, this was probably the result of rapid mineralization of dissolved organic matter in leachate and from soil organic matter in response to higher moisture and nutrients on the irrigated plots. Chloride is a conservative ion, however, and can be used as a tracer to evaluate solute movements and leaching rates (Bowman et al., 2002).


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  		Table 7. Estimates of seasonal{dagger} (fall through spring) constituent leaching losses at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, 2003–2007.

 
During the 4-yr study period, our estimate of total net leaching losses of Cl from irrigation plots (irrigated – control, 6407 ± 3271 kg ha–1) exceeded total net Cl deposition (4347 ± 442 kg ha–1) by a factor of 1.47. In contrast, Bowman et al. (2002) found that Cl leaching estimates based on estimated solution fluxes and solution concentrations from conventional suction cup soil water samplers accounted for only 80 to 86% of Cl applied in leachate. Conventional suction cup soil water samplers allow only intermittent sampling, providing a discontinuous record of solution chemistry that may not accurately reflect fluctuations in solute concentrations over time (Bowman et al., 2002; Weihermuller et al., 2007). For example, in our study, subfreezing weather from late December through early March prevented routine collection of soil solution samples, and the mean seasonal volume-weighted soil solution concentrations we used to estimate leaching rates may be affected to some extent. Solute leaching estimates also are affected by uncertainty in the estimates of solution fluxes (Weihermuller et al., 2007), which may be overestimated by the water balance method (MacDonald et al., 1992). In our study, however, estimates of total net Cl leaching did not differ significantly from estimates of total net Cl deposition (P = 0.33 for ln-transformed data), indicating that the relative magnitudes of and trends in other solute leaching losses are not unreasonable. Leaching losses of most constituents from irrigation plots decreased through time, consistent with leaching losses of solutes declining as a result of the decreased loading rates.

Estimated total leaching losses of NO3–N (Table 7) amounted to approximately 53% of the NH4–N + NO3–N applied during the 4 yr of this study. Watzinger et al. (2006) also found that 44% of the nitrogen applied in landfill leachate, as measured by drainage from zero-tension lysimeters, was leached as NO3–N. In contrast, Bowman et al. (2002) reported much lower NO3–N leaching from spray-irrigated leachate (7–11% of N applied), largely as a result of management of their system to encourage denitrification. While maintaining higher soil moisture to increase denitrification reduces NO3–N leaching, it also directly encourages the production of greenhouse gasses (Watzinger et al., 2005) and may have adverse effects on vegetation (Gordon et al., 1989). In our study, the ultimate fate of NO3–N lost via leaching is not known. The landfill that we studied is separated from the Grand River on one side and adjacent uplands on the other by at least 100 m of floodplain forest, where additional plant uptake of nitrogen and reduction in NO3–N through denitrification can occur (Gilliam, 1994). Although leaching of NO3–N remains a concern given the potential for off-site movement to surrounding areas, estimated leaching of NO3–N during the last 3 yr of the study period decreased substantially as compared with the first year (Table 7) in response to reduced leachate applications.

Plant Biomass and Elemental Concentrations
Leachate irrigation significantly increased plant biomass (P = 0.001), with a 3-yr average of 473 ± 162 g m–2 on irrigated plots compared with 246 ± 126 g m–2 on control plots. Increased biomass on leachate-irrigated plots was consistent with the positive growth effects of added water and nutrients from leachate noted in several other studies (Cureton et al., 1991; Maurice et al., 1999; Revel et al., 1999). Except for three elements, most metal concentrations (dry weight basis) were below detection limits in vegetation from control and irrigated plots in all years. Concentrations of Ba were elevated in vegetation on irrigated plots in 2003 and 2004, whereas concentrations of Zn did not differ between treatments (2003 and 2004) or decreased with irrigation (2005) (Table 8 ). Although Ba is not an essential element for plants, toxicity may occur if it is supplied at levels in excess of Ca and Mg (Page et al., 1982), which was not the case in our study. We found slightly higher levels of Cu in vegetation on irrigated plots, but this effect was not significant (Table 8). As reported in other studies (Menser et al., 1979a, 1979b, 1983; Adarve et al., 1998), metal concentrations in leachate-treated vegetation were within ranges that should not pose a hazard to plants, humans, or wildlife (Williams and Schuman, 1987; Munshower, 1993).


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  		Table 8. Mean total metal{dagger} and nitrogen concentrations in plant tissues at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, Sept. 2003–2005.

 
Nitrogen concentrations of herbaceous vegetation tend to increase with the nitrogen concentration of the leachate applied, although significant increases were observed by Hernandez et al. (1999) only when leachate nitrogen concentrations exceeded 2400 mg NH4–N L–1. In our study, TKN concentrations (dry weight basis) of vegetation did not differ significantly between treatments (Table 8), similar to the response of five grass species treated with landfill leachate in a study by Menser et al. (1979b, 1983). Increased biomass on irrigated plots in our study did serve as a sink for some of the applied nitrogen. On average, at the end of the irrigation season there was an estimated 89 kg N ha–1 in plant biomass on irrigated plots, compared with 34 kg N ha–1 in plant biomass on control plots. Accounting for annual variation in biomass and TKN concentrations, the increased plant uptake of nitrogen on irrigated plots compared with that on the control plots represented an estimated 7.1% of the total nitrogen applied in 2003, 32.2% in 2004, and 10.6% in 2005. In 2004, the year of lowest nitrogen deposition (Table 3), this nitrogen uptake was substantially greater than the 5.8% average turf uptake of nitrogen reported by Bowman et al. (2002). Mean total biomass production in the latter study ranged from 212 g m–2 on control plots to 432 g m–2 on the most productive leachate-treated plots, comparable to the mean biomass values we found. In the Bowman et al. (2002) study, however, average annual nitrogen deposition rates were much higher (1300 kg N ha–1). The effectiveness of plant uptake as a sink for leachate nitrogen decreases as nitrogen deposition increases and may depend on the herbaceous species present on the irrigated site (Menser et al., 1979a, 1979b, 1983; Hernandez et al., 1999). For greater nutrient removals, irrigation of short-rotation tree crops over larger land areas has been proposed (Brander et al., 2004; Duggan, 2005; Zalesny et al., 2007). In our study, increased biomass on leachate-treated areas served as an effective erosion control measure and helped to maintain the integrity of the landfill cover. Excessive applications of high-conductivity leachate, however, can cause the vegetative cover to deteriorate (Hernandez et al., 1999; Bowman et al., 2002; Jones et al., 2006, Watzinger et al., 2006), reducing its effectiveness as part of the treatment system.

Water Balance and Leaching Rates
Precipitation amounts varied from year to year during the study (Fig. 2a ), with 2003, 2004, and 2006 having more precipitation during the growing season than in 2005, which was extremely dry. Heavy leachate applications in 2003 (Fig. 2a) provided much more water than plants needed for transpiration on the irrigated plots (Fig. 2b). In contrast, lower rates of leachate application in 2004 and 2005 allowed normal periods of seasonal drought when actual evapotranspiration fell below potential evapotranspiration on control and irrigated plots. On the control plots, soil-available water was reduced by evapotranspiration more rapidly than on irrigated plots in all years (Fig. 2c). Subsequent depletion of available water on irrigation plots during the summers of 2004 and 2005 suggests that more leachate could have been applied during these months, but observed effects on soil solution chemistry and constituent leaching rates indicate that higher leachate application rates would be undesirable. Estimated leaching losses of water from irrigation plots exceeded those on control plots in 2003 and 2006 (Fig. 2d). Elevated leaching losses in 2003 were the result of high leachate application rates (Table 3), whereas elevated leaching losses in 2006 were associated with abundant rain in the summer and fall (Fig. 2a). This illustrates the unpredictable risks of increased leaching losses associated with periods of high rainfall (McBride et al., 1989b). For example, elevated leaching losses of solutes from irrigated plots in 2006 (Table 7) were the combined result of the higher concentrations of these solutes in soil solution (Table 6) and the higher leaching losses of water resulting from abundant fall rains (Fig. 2a, 2d). Maintaining soils below field capacity during active irrigation periods is necessary to minimize surface runoff and immediate solute leaching losses (Godley et al., 2005) and to avoid the increased N2O production and reduced CH4 oxidation that has been observed in soils at higher leachate application rates (Watzinger et al., 2005). As suggested by McBride et al. (1989b) and reemphasized by Jones et al. (2006), balancing leachate irrigation rates with the possible variations in soil water storage, evapotranspiration, and precipitation is a necessary part of any leachate application system.


Figure 2
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  		Fig. 2. Estimates of (a) precipitation and leachate irrigation depths, (b) evapotranspiration, (c) soil available water, and (d) leaching losses of water on control and irrigated plots at the landfill in Section 12, T6N R13W, Ottawa County, Michigan, January 2003 to June 2007. Soil-available water and leaching losses of water are calculated for the upper 50 cm of landfill cover soil.

 

   Summary and Conclusions
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Methods
 Results and Discussion
 Summary and Conclusions
REFERENCES
 
We studied the effects of operational spray irrigation of municipal solid waste landfill leachate on a landfill cover soil and its vegetation using a replicated experimental design. Soil solution conductivity and nitrate concentration varied seasonally on irrigated plots, with the highest values observed in the late summer and fall at the end the irrigation seasons. Estimated leaching losses of nitrate and other solutes were elevated on irrigated plots, most dramatically during the first year of the study when leachate application rates exceeded 32 cm. To reduce the high solute leaching rates observed in 2003, leachate applications in 2004, 2005, and 2006 were timed to match periods of greatest evapotranspirational demand (June–September), and each of the spray heads on the landfill cover was sequentially rotated among three non-overlapping spray areas to further reduce localized deposition rates. Rotation of spray areas was effective; the total annual estimated leaching loss of water did not differ between control and irrigated plots in 2004 or 2005 and was only slightly greater on irrigation plots in 2006. This approach also substantially reduced the total deposition of leachate constituents on each spray area, greatly reducing the potential long-term impact of leachate application on soils and plants.

Leachate applications had no effect on soil metal concentrations, and soil ECs on irrigated plots in 2004, 2005, and 2006 were reduced compared with those observed in 2003. Total leaching losses of NO3–N from irrigated plots remained elevated above those of the control plots through 2006, but leaching losses of NO3–N and other solutes were greatly reduced as compared with 2003. Total plant biomass on irrigated plots was greater than on control plots, providing dense vegetation to help protect the landfill cover soil from erosion. Plant metal concentrations on irrigation plots remained in normal ranges that should not pose ecological or health concerns. Potential adverse impacts of increased soil EC remain a concern, and high leachate conductivity limits the amount of leachate that can be safely applied to the landfill cover without harming vegetation or soil properties. As concluded by other studies, onsite spray irrigation of municipal solid waste landfill leachate provides a convenient disposal method, but timing and rates of application need to be carefully planned to avoid adverse effects related to elevated nitrate leaching losses and increased soil EC. In our study, monitoring of leachate and soil solution chemistry allowed a timely adjustment of application logistics to avoid potential adverse effects to the soil–plant system from initially high leachate application rates.


   ACKNOWLEDGMENTS
 
This study was funded through the MDEQ and AWRI. AWRI and the Biology Department at Grand Valley State University also provided matching funds for administrative support and project management. Scott Lidgard of DLZ Michigan, Inc., was instrumental in facilitating all phases of the study, and Mark McCoy of EQ Industrial Services, Inc., assisted with onsite operations. AWRI staff, including Eric Andrews, Michelle Lelli, Eric Nemeth, Jim O'Keefe, Katherine Rieger, and Gail Smythe, assisted with sample collections and analyses. Sandy Gregg, Carol Smith, and other staff of the MDEQ Environmental Laboratory assisted with sample analyses. Alan Steinman and two anonymous reviewers provided perceptive and constructive comments. We thank all of these individuals for their unstinting efforts.


   NOTES
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ABSTRACT
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 Methods
 Results and Discussion
 Summary and Conclusions
 REFERENCES
 
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   REFERENCES
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 NOTES
 ABSTRACT
 INTRODUCTION
 Methods
 Results and Discussion
 Summary and Conclusions
 REFERENCES
 





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