Does Management Intensive Grazing Protect Groundwater by Denitrification?

Does Management Intensive Grazing Protect Groundwater by Denitrification?

Summary

Groundwater quality and gases were evaluated at four grazing paddocks and one conventionally cropped field. These results support the idea that denitrification actively mitigates groundwater nitrate contamination beneath management intensive grazing paddocks compared to conventional cropping. It is clear from the solute and dissolved gas composition of groundwater beneath the pasture that conditions were highly favorable for denitrification in groundwater under grazing and that a large portion of the nitrate that leached to groundwater was transformed to harmless N2 gas. Similar patterns were not evident at the corn study site due to the absence of a sufficient supply of DOC to fuel the denitrification reaction

Objectives/Performance Targets

The primary objective is to determine whether denitrification is higher in soil and groundwater under MIG than in annual cropping. We are focusing on coarse- and medium-textured soils, where nitrate loading potential is higher than fine-textured soils.

Accomplishments/Milestones

Results and Discussion/Milestones

Surface Soil Aeration
Soil gas samples collected from the rooting zone during dry periods showed that aerobic condition were generally present in the surface soil at both experimental sites under field capacity moisture conditions. The partial pressure of oxygen generally remained above 0.20 atm at all locations sampled and CO2 concentrations were generally less than 0.01 atm. Thus, low oxygen status conducive to denitrification was not detectable within the soil atmosphere samples collected from the macropores of the rooting zone at either study area. The fairly open architecture of soil pores in the coarse sandy soil, at both the pasture and corn study sites, appeared to allow rapid gas exchanges between the rooting zone and the general atmosphere levels. This suggests that that the air-filled pore space may also provide an important mechanism for exchange of atmospheric gases to and from the water table environment.

The apparent absence of reducing conditions within the rooting zone based on the biogenic gas composition of the soil atmosphere suggested that denitrified N gases (N2, N2O) encountered within groundwater generally would not be generated first within the rooting zone and then leached to groundwater. Thus, the occurrence of denitrified N gases in groundwater most likely primarily reflects the reduction of NO3- after leaching to groundwater.

However, the possibilities of leaching dissolved denitrified N gases from the unsaturated zone due to (1) sustained or transient denitrification progress within micropores of soil aggregates [Tiedje, 1988] or the poorly ventilated capillary fringe or (2) transient denitrification pulses associated with the infiltration of water within and beneath dung and urine patches can not be fully eliminated based on the existing soil gas data. These distinctions would require installation of soil moisture access tubes at selected locations for soil moisture depth profiles, collection of soil gas samples from the rooting zone during rainfall or snowmelt events, and installation of deeper soil gas wells for monitoring of biogenic gases within the unsaturated zone immediately within and above the capillary fringe.

Groundwater Chemistry
The dissolved solids (Fig. 2) in groundwater beneath both the pasture and corn study areas ranged from very dilute to moderately concentrated. The results for specific conductance, acid neutralizing capacity, dissolved inorganic carbon and pH suggest a comparable inorganic ionic composition between the sites. However, the slightly higher specific conductance and lower ANC and DIC concentrations at the corn study area are consistent with the idea the bicarbonate anions generated naturally by dissolution of carbonates (dolomite) have been partially depleted by strong acids associated with fertilizer salts (e.g., eq. [4], nitric acid generated by nitrification reactions).

[4a] NH4+ + Cl- + 2O2 = NO3- + 2H+ + H2O
[4b] 2H++ NO3- + Cl- + Ca2+ + 2HCO3- = Ca2+ + 2CO2+ 2H2O + NO3- + Cl-

Despite apparent similarities in inorganic chemical composition, dissolved organic carbon (DOC) and O2 concentrations in shallow groundwater revealed a strong biogeochemical contrast between the two study areas (Fig.3). DOC concentrations were significantly higher and ranged more broadly at the pasture site than at the corn site due to the leaching of dissolved organic matter from dung and urine patches. Consistent with greater availability of DOC as a potential electron donor and microbial food source, groundwater O2 concentrations were significantly lower beneath the pasture study area than the corn site. The percent O2 saturation (relative to air equilibration at field temperature) levels show that the differences were attributable to oxygen consumption rather the temperature dependence of O2 solubility. The absence of more complete depletion of O2 may reflect an exchange of gases to and from the general atmosphere through the air-filled pore space of the unsaturated zone.

Fueled by leaching of soluble organic matter, heterotrophic respiration in the water table environment will cause an increase in the concentration of dissolved carbon dioxide in groundwater. Accordingly, comparison of the dissolved CO2 concentrations beneath pasture and corn (Fig. 4) revealed somewhat higher concentrations of CO2 in groundwater beneath the pasture consistent with its higher DOC level. However, CO2 concentrations were apparently too high at many locations within both study areas to be explained solely by aerobic heterotrophic activity. Under strictly aerobic respiration, the depression of dissolved oxygen should have been accompanied by a roughly 1:1 stoichiometric increase in dissolved CO2 (i.e., if one mole of O2 consumed produces roughly one mole of CO2). Yet at both locations, the dissolved CO2 concentrations were substantially in excess of the apparent O2 deficit in groundwater (Fig. 3).

Under anaerobic conditions, the upper limit of CO2 production is no longer bounded by the solubility of dissolved O2 but is controlled by the availability of other electron acceptors (NO3-, Fe3+, Mn4+, SO42-, CO2). Thus, CO2 has the potential to accumulate to concentrations well above the saturation CO2 concentration corresponding to the consumption of O2 in air-equilibrated water. Despite the presence of measurable O2 at most locations, evidence of anaerobic conditions beneath the pasture (1) included an accumulation of CH4 produced by methanogens to concentrations well above its solubility under air-equilibration (Fig. 4), elevated concentrations of dissolved (2) Fe2+ and (3) Mn2+ reflecting the reduction of Fe and Mn oxides in the solids (data not shown), (4) an accumulation of N2O gas (an intermediate in denitrification) greatly above its solubility under air-equilibration (Fig. 4), and (5) total concentrations of N2 gas above the N2 concentration in air-saturated water at the apparent recharge temperature of the sample (Fig. 5). Evidence of anaerobic conditions beneath the corn was limited to elevated N2O and total N2 concentrations and was therefore less striking. It is important to note, however, that the accumulation of N2O is an ambiguous indicator of anaerobic conditions because N2O can be produced by both nitrification (an O2 requiring reaction) and denitrification reactions (Davidson et al. 2000).

At neither study area, was the accumulation of the products or consumption of reactants by anaerobic processes fully sufficient to explain the excess of CO2 in groundwater. Thus, it is probable that some of the apparent excess of CO2 was derived by the reaction of strong acids with bicarbonate salts in groundwater (e.g., reaction with nitric acid from nitrification reactions, see eq. [4], above).

The apparent co-occurrence of both aerobic and anaerobic indicators (albeit weakly at the corn study area) within groundwater at both locations deserves some consideration. One explanation is that the anaerobic processing of DOC in groundwater proceeds more rapidly at times than CO2 and O2 can be exchanged to and from the soil atmosphere within the aerobic unsaturated zone. In the pasture study area, pulses of soluble organic matter from dung and urine patches could periodically produce short term anaerobic responses which subside after fresh substrate has been consumed. Diffusive and advective pumping of O2 from the air-filled pore space of the unsaturated zone could then later resupply O2 and reset aerobic conditions. In the absence of apparent sources of pulses of soluble organic matter, the subtle combination of aerobic/anaerobic indicators is more difficult to explain at the corn study site. Further, study is needed to clarify the transient anaerobic/aerobic conditions and processes in the water table environment beneath pasture environments and to elucidate whether such conditions/processes are also happening beneath conventional crops.

The distributions of dissolved N (Fig. 6) and P (Fig. 7) provide an additional line of evidence supporting the importance of soluble organic matter from dung and urine patches in nutrient cycling beneath the pasture study area. At the corn study site high concentrations of dissolved N in groundwater were comprised almost exclusively of inorganic N as NO3-N, reflecting an excess of N from fertilizer salts in the N cycle. However, beneath the pasture, concentrations of NO3-N were much lower, DON comprised a much greater share of dissolved N (Fig.8), and there was a greater accumulation of NH4+ (The presence of high DOC and lower O2 concentrations at the pasture study area may inhibit the formation of NO3- from NH4+ by nitrification, allowing it to accumulate to higher levels in groundwater at the pasture site.). These patterns are consistent with a greater role of organic matter in the N cycle at the pasture site. Similarly, although the total dissolved concentrations of P in groundwater were somewhat similar at the two study areas, the distribution of P was dominated to a greater degree by dissolved organic P (DOP) at the pasture study area than at the corn study area (Fig. 8).

Groundwater Denitrification
Figure 9 illustrates the concentrations of N2 (XsN2) in excess of the amount contributed by equilibration with the general atmosphere during groundwater recharge:
[6] XsN2 = Total N2 – Recharge N2

Here “Recharge N2” represents the N2 concentration in air-saturated water at the apparent temperature of recharge (Tr)) based on the measured Ar concentration and “Total N2” represents the total N2 concentration measured in the groundwater sample (Fig. 5). To obtain “recharge N2”, the Henry’s Law constant for N2 at Tr was used to compute the concentration of dissolved N2 contributed by the partial pressure of N2 in air (~0.2095 x 0.965 atm) :
[7]

Based on the XsN2 measurements, the results in Fig. 9 suggest (1) that the dissolved concentration of denitrified N as N2 ranged from 0 to roughly 5 mg-N L-1 across both study areas and (2) that concentrations of denitrified N based on XsN2 were similar in groundwater at both study areas. However, at the pasture study area, concentrations of XsN2 were similar in magnitude to the measured NO3- concentrations. Thus, total NO3- concentrations reconstructed by eq. [2] were generally more than double the measured NO3- concentrations in groundwater beneath the pasture. In contrast, at the corn study area, concentrations of XsN2 were much smaller than measured NO3- concentrations and were a small contributor to total NO3-.

Only a small percentage of the NO3- leached to groundwater at the corn site was converted to N2 gas in the near water table environment (see denitrified N%, Fig. 9). Three factors potentially inhibit an efficient conversion of NO3- to N2 gas at the corn study site. First, the electron donation capacity of DOC is limited due to the relatively low concentration of DOC in groundwater. Second, the electron accepting capacity of the available NO3- may greatly exceed the electron donating capacity of the DOC present in groundwater. For example, assuming as an upper bound that each mole of DOC can potentially donate 4 moles of electrons, approximately 0.8 moles of NO3- could be reduced per mole of DOC in groundwater. On this basis 1 mg L-1 of DOC could potential reduce only 0.93 mg L-1 of NO3- to N2 gas; thus, comparison of the NO3- (Fig. 9) and DOC (Fig. 3) concentrations shows that the available NO3- probably greatly exceded the reducing power of DOC at the corn study site. Under such conditions, Chapin et al. (2002) suggest that N2O, an intermediate product of denitrifcation, will potentially accumulate to high levels, a condition which is consistent with the high N2O concentrations observed for the corn study in Fig. 4. (However, an alternate plausible explanation for the high N2O may be its accumulation as a byproduct of nitrification [Davidson et al. [2000]). Third, the oxygen levels in groundwater at the corn study site may inhibit denitrification.

In contrast, a much larger percentage of the NO3- leached to groundwater was converted to harmless N2 gas in the near water table environment at the Bestul’s north/east pasture study area (see denitrified N%, Fig. 9). Based on the same calculus used for the corn study area, the high concentrations of DOC present in groundwater beneath the pasture (Fig. 3) likely represent a sufficient supply of available electrons to reduce all NO3- (Fig. 9) to N2 gas. However, complete conversion of all NO3- is probably still somewhat inhibited by the residual dissolved O2 present in groundwater beneath the pasture.

Groundwater denitrification was readily quantified at two contrasting study areas based on the accumulation of excess N2 in groundwater. It is clear from the solute composition and dissolved gas composition of groundwater beneath the Bestul north/east pasture that conditions were highly favorable for denitrification in groundwater under management intensive grazing and that a large portion of nitrate that leached to groundwater was transformed to harmless N2 gas. Similar patterns were not evident at the corn study site due to the absence of a sufficient supply of DOC to fuel the denitrification reaction. Further, it is clear from comparing total NO3- concentrations (XsN2 + NO3-) at the two study areas that leaching to groundwater is much lower under managed intensive grazing than conventional cropping with corn. These results support the idea that denitrification actively mitigates groundwater nitrate contamination beneath management intensive grazing areas compared to conventional cropping.
Nitrogen Budgets and Groundwater Monitoring Wells
Nitrogen budgets were calculated for each study paddock and the cropped field. Calculations for inputs to the grazing paddocks included the number, duration, types of animal (adjusted for size, supplemental feed and loss in milk), addition of N (inorganic fertilizer and land spread manure), precipitation, and dry deposition. Grazing paddock nitrogen budget outputs included estimated N removal in forage by animals during grazing, and N losses to the atmosphere (ammonia and denitrification loss at 25% and 6% of the total N inputs, respectively). The amount of forage removal was based on the results of a previous study at the Bestul and Onan paddocks. That study estimated forage quantities before and after the paddocks were grazed and analyzed the N (crude protein) content in forage samples. The cropped field inputs included the addition of fertilizer, land spread manure, precipitation, and dry deposition and outputs from silage removal and N losses to the atmosphere.

Variability occurred between the sites with stocking densities, particularly when out-wintering animals was practiced in the study paddock (Fig. 10 ). Predictably, the out-wintered paddocks also had the greatest residual nitrogen. In 2003 the amount of nitrogen added to and removed from the Bestul paddocks was nearly balanced. At the Rambo field, in 2002 there was less N added to the land than was removed in the harvest, but in 2003 there were more than 112 kg/ha of N added than was removed by the crop. This was primarily due to a larger application of manure in 2003.

Average N concentrations in the groundwater monitoring wells showed a good relationship to the amount of residual N calculated in the N budget in 2002, with an r value of 0.74 in 2002. However, in 2003 the relationship between the residual N and groundwater concentrations was minimal. This difference may be due to the inconsistent movement of N to groundwater depending upon the amount of precipitation in a given year. During drier years with less groundwater recharge, residual N builds up in the soil, moving to groundwater during subsequent recharge events.

In central Wisconsin groundwater background concentrations are considered to be less then 1 mg L-1 for NO3 and 2 mg L-1 for Cl. Concentrations of NO3 in groundwater ranged from 0.5 to 47.8 mg L-1 from all of the wells at all of the sites. Site-by-site distribution of NO3 is shown in Fig. 11. Concentrations of NO3 were below the drinking water standard (10 mg L-1) in most of the samples collected from the Bestul north/east, Bestul west, and Breneman paddocks. However, the Breneman site had several outlier elevated concentrations.Greatest concentrations were measured in groundwater below the Rambo and Onan sites. A similar pattern was observed with Cl concentrations below the grazing paddocks. These concentrations ranged from 0.5 to 45 mg L-1. Background groundwater chloride concentrations in this region are generally near 2 mg L-1, demonstrating that there are impacts to groundwater from management intensive grazing. However, this large range also reveals the non-uniform nature of the contamination below these systems.
Soil In-field experiments
High variability was evident in data from field cores. There were few statistical differences among sampling locations and sampling dates that were consistent. In each pasture, higher NO3 concentrations were present under putative urine spots than other treatments on at least one sampling date (Fig. 12). Differences among sampling dates were present under corn (Rambo field) and NO3 concentrations tended to be smaller under corn than under pasture, at least on some dates. Results of DOC concentrations were more variable, with no differences among sampling times at any site and minor differences among treatments in two pastures (Fig. 13).

In the large-diameter soil cores, root mass was greater under pasture than under corn for the uppermost sampled horizon, but no differences were noted at deeper depths (Fig. 14). We had hypothesized that soil DOC would be related to visible horizon differences (mottling, old animal burrows, etc.), but we found no relationship. There were no difference in the depth distribution of soil DOC under corn and pasture (Fig. 15).
Intact soil core experiments
Soil water content varied during each experiment, but averaged 40 to 45%, 50%, and 60 to 65% water-filled pore space for the 0 to 15, 15 to ~40, and >~40 cm depth increments. Denitrification is increasingly likely when water-filled pore space exceeds 60%, as long as both NO3 and DOC are present. Although all soil samples were analyzed for DOC, the results showed no effect of treatment, sampling time, or depth in either experiment. This was unexpected, as others had noted rapid increases in DOC after urine application (Shand et al., 2000). A later report by the same group indicated that synthetic sheep urine caused much larger changes in soil solution nutrient composition, including DOC, than natural sheep urine (Shand et al., 2002).

The control cores showed no appreciable changes in soil pH, inorganic N content, or gas evolution during either experimental run (Fig. 16). The same was true for the dung application, except that trace gas evolution was twice or more than from the control cores. Methane evolution, in particular, was evident immediately after dung application. This can be expected, because dung contains high amounts of microbially available C substrates, and because oxygen levels are low due to high moisture content. Dung pats in Denmark released CH4 for 10 to 18 days in the field, but CH4 production from dung in pastures was estimated to be less than 4% that of stored manure (Holter, 1997). Yamulki et al. (1999) found that net release of CH4 from pastures was small relative to sources such as enteric fermentation in cattle. Greater CO2 evolution from dung was likely due to microbial respiration, whereas the initial release of CO2 with urine application was likely due to bicarbonates contained in urine.

Application of urine caused a rapid increase in soil pH in both the 0 to 15 and 15 to ~40cm depth increments. This is due to rapid hydrolysis of urea to ammonia, also evident in Fig. 16. As ammonium was oxidized to nitrate, soil pH declined because two moles of protons are produced per mole of NH4 oxidized. Thus, by the end of the month-long incubation, soil pH was 0.5 to 1 unit lower in the top two horizons than the control cores. This has been seen in field experiments (Vallis et al., 1982) and the changes that occur are related to the amount of urea + NH4 applied in urine and the pH buffering capacity of the soil. Increased soil pH enhances NH3 volatilization from urine spots (Fig. 16). Total NH3-N loss averaged 0.1, 2.5, and 18.8 mg N in the control, dung, and urine treatments in the second experiment, respectively. Ammonia loss from urine represented 13% of the maximum NH3 content of the 0 to 15 cm soil layer. Lower NH3 losses in the first experiment may be attributable to deeper infiltration of urine in the soil, as evidenced by higher NH4 contents below 15 cm in the first than the second experiment.

Soil nitrate contents increased beginning 7 to 10 days after urine application. This response is also typical of what has been observed in the field (Shand et al., 2002). There were no appreciable changes in NO3 contents in the dung or control cores, nor in the deepest depth of the urine cores. Even where NH4 was present in the deepest increment under urine in the first experiment, little NO3 accumulated. The low gas-filled pore space in this zone would have inhibited NH4 oxidation and enhanced conversion to N2O and other products during partial NH4 oxidation and reduction of NO3.

Nitrous oxide-N losses over 4 weeks were 0.5 mg N in the control, 1.8 mg N in the dung treatment, and 5.4 mg N with urine in the second experiment. Nitrous oxide loss comprised about 3% of the total accumulated NO3 at the end of the experiment, well within the ranges reported in the literature. However, incubations were not sufficiently long to determine total N loss as N2O, nor do we know how much NO or N2 was lost during denitrification. It is possible that these compounds comprised a significant proportion of total gaseous N loss.

No leaching of NO3 out of these cores occurred, because the base was sealed to permit a water table to be maintained. Under field conditions, NO3 formed in the upper part of the soil profile would be subject to uptake by plants and leaching to the water table in these permeable soils. Given the evidence in these experiments and in the literature for high denitrification rates under urine spots, we suspect that leaching of NO3 and DOC to the water table would result in rapid denitrification.

Impacts and Contributions/Outcomes

Groundwater denitrification was readily quantified at two contrasting study areas based on the accumulation of excess N2 in groundwater. It is clear from the solute composition and dissolved gas composition of groundwater beneath the Bestul north/east pasture that conditions were highly favorable for denitrification in groundwater under management intensive grazing and that a large portion of nitrate that leached to groundwater was transformed to harmless N2 gas. Similar patterns were not evident at the corn study site due to the absence of a sufficient supply of DOC to fuel the denitrification reaction. Further it is clear from comparing total NO3- concentrations (XsN2 + NO3-) at the two study areas that leaching to groundwater is much lower under managed intensive grazing than conventional cropping with corn. These results support the idea that denitrification actively mitigates ground water nitrate pollution beneath management intensive grazing areas compared to conventional cropping.

Residual N calculated from N budgets for the study areas and groundwater nitrate concentrations showed a fairly strong relationship to one another. Although management intensive grazing generally results in less nitrate contamination of underlying groundwater than some conventional agricultural practices, it is possible to overload the capacity for mitigation and/or denitrification of N by over-application of fertilizer or overloading the system with manure from practices such as out-wintering animals. These results substantiate the importance of creating and following nutrient management plans with management intensive grazing systems as well as other agricultural practices.

It is clear from the soil sampling and intact soil core experiments that soil inorganic N concentrations can be substantial in a pasture after urine is applied, but that little accumulation occurs under dung pats. Both ammonia and nitrous oxide losses were large after urine application to the intact cores. The intact cores allowed more precision, while maintaining some of the inherent soil variability present in the field. Patterns of inorganic N content were readily traceable over time with this method. These results support the field observations of active denitrification activity in shallow ground water under pasture.

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