Resources and the Environment – 'Water Quality and ...goal of zero discharge of pollutants into the nation’s water-ways. Yet, water quality remains poor in many locations ... water

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2nd Quarter 2007 • 22(2) CHOICES 79

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2nd Quarter 2007 • 22(2)

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Theme Overview: Agriculture and Water Quality in the Cornbelt: Overview of Issues and ApproachesBy Matthew J. Helmers, Thomas M. Isenhart, Catherine L. Kling, Thomas B. Moorman, William W. Simpkins,

and Mark Tomer

JEL Classification Code: Q25

More than three decades have elapsed since the passageof the Federal Water Pollution Control Act with its statedgoal of zero discharge of pollutants into the nation’s water-ways. Yet, water quality remains poor in many locationsand considerable loading of pollutants continues. This isparticularly true for agricultural sources of water pollutionand is typified by the Upper Mississippi River Basin,where more than 1,200 water bodies appear on the currentU.S. Environmental Protection Agency (EPA) listing ofimpaired waterways. Additionally, nitrate export from thisregion has been implicated as a significant cause of thehypoxic zone in the Gulf of Mexico, which covered nearly20,000 km2 in 1999 and more than 17,000 km2 in 2006(http://www.epa.gov/gmpo/nutrient/hypoxia_pressrelease.html). Although a substantial body of evidence on theeffectiveness of agricultural conservation practices onwater quality continues to be developed, the net effect ofthese programs and practices at the watershed scale isunclear. Increasingly, studies are being focused on thewatershed (or landscape) scale and complex interactionsbetween agricultural practices and inputs, the types andconfiguration of conservation practices on the landscape,and the resulting downstream water quality. While lowcost methods to reduce agricultural non-point source pol-lution exist, large changes in water quality in agriculturalregions are likely to be costly and met with resistance. Thisis because to achieve large changes in water quality, majoralterations to land use or installation of expensive struc-

tural practices may be required, and the costs are bornedirectly by producers and landowners, or by the taxpayer.

Given the potentially large cost for significantimprovements in water quality, it is critical to developtools that can support cost-effective design of conservationpolicy and/or voluntary implementation of watershedplans focused on water quality. The following set ofthemed papers related to water quality and agriculture dis-cuss these issues, with a specific focus on using integratedwater quality and economic models to support better pub-lic policy and watershed-based solutions to these prob-lems. The article following this one describes detailedfield-scale data collected as part of a Conservation EffectsAssessment Project supported by CSREES and ARS. Inaddition to assessing the effects of current conservationactivities on water quality in these watersheds, data areused to calibrate a water quality model and are being inte-grated with economic cost information to study the opti-mal placement of additional conservation activities in thewatershed. That article discusses the historical evolution ofconservation activities in the three watersheds, the currentwater quality challenges in the watersheds, and the rolethat the integrated models can play in solving the prob-lems.

In the third paper of the series, Secchi et al. employ amore aggregate unit of analysis (scale) for calibrating awatershed model and a biophysical carbon sequestrationmodel and integrating them with economic data coveringthe entire state of Iowa. The focus of their analysis is on

80 CHOICES 2nd Quarter 2007 • 22(2)

the potential unanticipated environ-mental effects of developing marketsin ecosystem services that focus on asingle service, such as carbon seques-tration.

The final paper in the setaddresses a different water qualityissue: drinking water and nitrate lev-els. Specifically, the paper by Burkartand Jha considers whether it wouldbe cost-effective for farmers to reducenitrogen applications at the farmlevel, thereby reducing nitrate con-centrations in the water supplies forresidential consumers, rather thancontinue to treat the water in a deni-trification plant prior to use.

In the remainder of this themeoverview, we attempt to provide thecasual reader with adequate back-ground information on agriculturalwater quality problems, as well as theinstitutional framework withinwhich these water quality problemsin agriculture are currently managed.This includes a brief primer on thekey pollutants, their sources, and therange of conservation methods thatcan attenuate their effects. It is alsonecessary to understand the funda-mentals of the policy environment,which differs markedly fromapproaches taken in other industries.Specifically, voluntary actions are thefocus of state and federal agencyefforts under the requirements thatthey have to develop and implementTotal Maximum Daily Loads(TMDLs). We briefly describe theTMDL process and note the range offederal and state conservation pro-grams that provide funding for vol-untary conservation efforts.

Agriculture and Water Quality PrimerProduction of food and fiber haveinevitable impacts on land and waterresources. Conservation practices are

intended to reduce those impacts,ideally with as little effect on the pro-ductive and ecosystem service capaci-ties of the land. The critical questionsfor planning and implementation ofeffective conservation systems arethen: What water quality pollutantsare of primary concern and whattypes of conservation practices willprovide benefits for various environ-mental impacts? Here, briefly, weprovide generic answers to thesequestions that are most pertinent toagricultural watersheds in the CornBelt generally, and Iowa and theUpper Mississippi Basin, specifically.Through this discussion, we empha-size key differences among specificpollutants, in terms of the hydrologicpathways from field to stream, andthe types of conservation practicesthat can minimize their transport toreceiving waters. The primary pollut-ants of concern in the Corn Beltinclude nitrate-nitrogen, phospho-rous and sediment, and pathogens.

Nitrates-NitrogenNitrate-nitrogen (NO3-N) is a keypollutant of concern for its potentialwidespread impact on both publichealth and ecosystem function.Nitrate-nitrogen is readily leachedthrough soils to groundwater andenters surface water systems directlyby groundwater flow and through thesubsurface drainage systems (tiledrains), which were installed acrosslarge areas of poorly drained Mid-western soils beginning about 100years ago. These drainage systemshave allowed the Midwest to becomethe highly productive agriculturalarea that it is today, while short-cir-cuiting the much slower, naturalgroundwater pathway to the stream.Concentrations of nitrate-nitrogen indrainage and stream water oftenexceed 10 mg NO3-N /L, resulting

in losses exceeding 20 kg N/ha insome years (Tomer et al., 2003).Regional nitrogen budgets for theMississippi River Basin have impli-cated tile-drained regions of the Mid-west as disproportionately contribut-ing to N loads to the Gulf (Burkartand James, 1999). Nitrogen fertilizeris commonly applied to corn, at ratesvarying from 100 to 200 kg/ha. Theefficiency of N uptake by the cropvaries because of environmental con-ditions. Nitrogen losses are mostprevalent in early Spring when cropsare not present or are too small toeffectively immobilize the availablenitrate.

The problem of nitrate-nitrogenexport is not solely caused by N fer-tilizer management or any other sin-gle factor, but rather it is a combina-tion of soil management practicesand physical, chemical, and biologi-cal characteristics of the soil, alongwith temperature and precipitationpatterns (Dinnes et al., 2002). As aresult, reducing nitrate loss is morethan a matter of reducing N-fertilizerrates and improving timing of appli-cations (Jaynes et al., 2004). Effectivepractices to control N losses includediversified crop rotations thatincrease use of forages and improvednitrogen management (includingimproved timing and rates of applica-tion, and use of nitrification inhibi-tors). Improved engineering of agingdrainage infrastructure, and use ofwetlands, cover crops, and denitrifi-cation walls or subsurface drainagebioreactors are other alternatives thathave been shown effective. Becausenitrate in extensively tiled areas istransported to streams primarily insubsurface drainage water, any filter-ing ability of riparian buffers andedge of field filter strips will bebypassed.


Phosphorous and SedimentSurface runoff is the dominant mech-anism that transports phosphorus,sediment, and pesticides and bacte-ria from agricultural fields, asopposed to the subsurface pathwaysof nitrate. Ecological impacts of Pand sediment include eutrophicationand sedimentation of receivingwaters. Phosphorus losses from agri-cultural fields may be only a fractionof those observed for N (< 1 kg/ha.yris commonly reported), but suchlosses can have major implicationsfor the ecological integrity of lakesand streams. Phosphorus runoff fromagricultural fields is largely controlledby soil P concentrations and crop res-idue cover (Sharpley et al., 2002).Residue cover encourages infiltrationand discourages erosion. To improvephosphorus management at water-shed scales, the use of “P indices” arebeing implemented that identify soilerodibility, soil P concentrations, res-idue management practices, andproximity to streams, to rank fieldsfor runoff P losses. These indices canbe used to target conservation prac-tices to control P losses (Birr andMulla, 2001) via reduced tillage, lim-ited manure or fertilizer applica-tions, terraces, vegetated filter strips,and/or riparian buffers. These prac-tices are known to reduce erosion andphosphorus. Watershed responses tothese conservation practices may beless than initially expected becausestreambank erosion, rather than agri-cultural fields, can contribute signifi-cant amounts of sediment and phos-phorus to streams and rivers. Thesesources may result from past manage-ment activities.

Sediment and nutrient lossesfrom agriculture, therefore, can resultin a legacy of impacts within water-sheds, necessitating a long-term com-mitment to their amelioration. For

example, elevated nitrate concentra-tions in groundwater have beenshown to remain for decades (Rod-vang and Simpkins, 2001). Also,phosphorus accumulations in sedi-ment may have a legacy, providing along-term, internal loading source ofmineral P to the water column(Christophoridis and Fytianos, 2006)and may ultimately affect groundwa-ter P concentrations (Burkart et al.,2004).

Bacterial Pathogens and Livestock ConcernsLivestock is an important economiccomponent of U.S. agriculture,accounting for over 60% of agricul-tural sales. Production estimates for2005 include 72.6 million hogs, 10.9million beef cows, 3.1 million milkcows, 150 million egg layers, and 131million broilers for the 12-stateNorth Central Region. In the Mid-west, swine are increasingly producedin concentrated animal feeding oper-ations (CAFOs) making manuremanagement increasingly important,both as a source of nutrients for sub-sequent crops and as a potential envi-ronmental problem. CAFOs are alsoimportant in poultry and beef pro-duction. Potential water qualityissues arising from manure applica-tion are nitrate leaching and loss intile drainage networks, and loss ofphosphorus and pathogens in over-land runoff. Conservation practicesseek to prevent accumulation ofexcess nutrients (nutrient manage-ment plans), reduce and/or treat run-off from feed lots, and mitigate run-off from manured fields (buffers,filter strips). Several studies suggestthat increasing CAFO size offers cer-tain economic advantages in produc-tion, but increases the amounts ofmanure applied to land near theCAFO, which increases the risk of

loss of excess nutrients (Kellogg et al.,2000).

Bacterial pathogens that threatenwater quality include Escherichia coliO157:H7, Salmonella, Enterococ-cus, Listeria, and Campylobacter.Pathogenic protozoa includeCryptosporidium and Giardia.Although these microorganismscause disease in humans, they arecommonly carried in livestock with-out visible symptoms. Because of thedifficulty and cost involved in screen-ing water samples for these patho-gens, public health and water supplyauthorities have long relied uponindicator bacteria. In the past, fecalcoliforms tests filled this function,but two indicators are now beingpromoted by U.S. EPA, Escherichiacoli and Enterococcus. Quick andreliable tests for both of these micro-organisms are now available and thepresence of these bacteria has beencorrelated with the presence of dis-ease-causing microorganisms. Mea-sured E. coli densities in stream watercan be evaluated against EPA’s cur-rent standards, but the identificationof the E. coli sources is more complexand important to developing effectivewatershed management strategies.Microbial source tracking is anemerging technology that allows thesource animal to be determined.Potential sources in most watershedsinclude wildlife, farm animals, andhumans.

Heterogeneity of Conservation PracticesThere is a wide range of conservationpractices used on agricultural landintended to provide water qualitybenefits, including engineered struc-tures, edge-of-field practices, in-fieldnutrient and crop residue manage-ment practices, and land retirement.Government programs since the


1930s have promoted installation ofconservation practices on agriculturallands. Much of the early focus ofconservation practices was specifi-cally on soil conservation, where thegoal was to preserve the soil and tomaintain its productivity.

Structural practices that havebeen used for controlling soil loss andthe formation of gullies include ter-races, grassed waterways, sedimentbasins, and grade stabilization struc-tures. Terraces are used to decreasethe length of the hill-slope to reducerill erosion and the formation of gul-lies. Many early conservation prac-tices were intended, in part, for waterconveyance to improve trafficability,and thereby maximize agriculturalproduction. In addition to structuralpractices, there are a variety of in-field management practices such ascontour farming and strip croppingand tillage management, such as con-servation tillage and no-till. Also, insome areas marginal lands that arehighly susceptible to soil loss havebeen taken out of agricultural pro-duction and converted back to peren-nial vegetation.

Over the past thirty years, therehas been an increased concern relatedto the overall water quality impacts ofagriculture, including nutrient, pesti-cide, and pathogen loss from agricul-tural lands. Some conservation prac-tices have been installed with anintended purpose of reducing theexport of these contaminants. Two ofthese are buffer systems (riparian orgrassed) and the reintroduction ofwetlands back into the landscape. Inaddition, relative to nutrient losses,there has been an emphasis on appro-priate nutrient management practiceswithin agricultural fields to reducethe application of excess nutrients.

We have also learned that someagricultural practices have effects thatwere not intended. Subsurface drain-

age was used historically to enhanceproductivity of poorly drained lands,but these production benefits are off-set by the environmental impacts ofincreased export of nitrate-nitrogenfrom these drainage networks. Sur-face inlets to subsurface drain systemsalso create a direct conduit for surfacewater to enter streams effectivelybypassing riparian buffers or wet-lands. Much of the agricultural land-scape has been altered throughstream straightening channelization.Stream straightening and subsurfacedrainage have significantly altered thehydrology of the landscape, whichhas led to significant streambank sta-bility problems in many areas. So,while many of the conservation prac-tices mentioned above may reducesoil loss from agricultural fields, ifthey do not significantly reduce waterflow in the streams, the stream poweris not reduced. As a result, ratherthan carrying sediment from fields,the streams may erode sediment fromthe streambed and streambanks.

While there is a wide range ofpractices that can be used on agricul-tural lands for providing water qual-ity benefits, many times the locationswithin the watershed where practicesare implemented have not been spe-cifically targeted to achieve the great-est reduction of contaminants indownstream water bodies. This islikely the result of the voluntaryenrollment in federal conservationprograms combined with ineffectivetargeting technology. Recentadvances in remote sensing and geo-graphic information systems offer anopportunity for dramatic improve-ments in our ability to target conser-vation practice installation in largewatersheds. With the limited amountof resources available for conserva-tion practices, there will likely beincreased importance on targetingimplementation to those areas where

there may be the greatest benefitfrom a water quality perspective. Oneprogram that has used targeting withsome effectiveness is the Conserva-tion Reserve Program (CRP), whichtargets land choices based on an envi-ronmental benefits index. While theeffects of CRP on soil quality, carbonstorage, and wildlife have beenassessed, the aggregate effects at thewatershed scale are less understood.

Finally, it is important to under-stand that water quality monitoringin the United States is done by a vari-ety of state and federal agencies,including USGS and USEPA, andmany municipal and commercialwater supply entities, but the greatmajority of streams and rivers are notroutinely monitored. Thus, in manycases, the actual level of pollutants issimply unknown.

The Policy Environment: TMDLs and Voluntary ImplementationVoluntary cost-share and incentiveprograms sponsored by USDA andStates are large in geographic scaleand fiscal commitment (over $4.5billion was spent in 2005 by USDA-funded programs alone). These pro-grams generally provide varyingincentives to farmers for the installa-tion of structural or managementpractices described above. The crite-ria for participant eligibility varyfrom program to program, and con-servation compliance provisionsrequire that landowners who farm onhighly erodible land undertake someconservation activities in order to beeligible for other government incen-tives or subsidies. In addition to thelargest program, the CRP, there is acost-share program entitled the Envi-ronmental Quality Incentive Pro-gram, which provides cost share toproducers willing to install variousconservation structures or practices


on their farms. Notably, the 2002Farm Bill contained a new programthe Conservation Security Program awatershed-based initiative intendedto compensate farmers for adoptingconservation practices. Like the CRP,which covers the full cost of retiringland from production, the programwas intended to cover the full cost ofadopting conservation practices(rather than less than 100% of thecost as traditional cost share pro-grams do), but the focus of the Con-servation Security Program is on landthat stays in production. However,funding constraints have preventedthe program as it was initially envis-aged from being fully implemented.

Ironically, while there are largeconservation programs funded andadministered through USDA, theprimary law that addresses nonpointsource agricultural pollution loadingsis under the auspices of the U.S.Environmental Protection Agency(EPA) via the Clean Water Act.Rather than assign standards andrequire that sources implementchanges in production or invest inabatement technology to meet thosestandards, as has been the norm forair and water quality problems stem-ming from point sources, the TotalMaximum Daily Load (TMDL)approach was adopted. Under theTMDL framework, states are respon-sible for compiling lists of water bod-ies not meeting their designated uses,which are then reported as “impairedwaters.” The sources of impairmentvary across locations. For example,Iowa has 213 water bodies on the listand pathogens (bacteria) are the lead-ing cause of listing, accounting forabout 20% of the impaired waterbodies, with sediment/turbidityaccounting for about 10%. Nation-ally, it has been estimated that 40%of rivers and estuaries fail to meetrecreational water quality standards

because of microbial pollution(Smith & Perdek, 2004).

Note that water bodies are viewedas impaired only if they do not meettheir “designated use.” Thus, twowater bodies can be equally contami-nated with only one being listed asimpaired if their designated uses aredifferent (e.g., boatable vs. swimma-ble). This is part of what makes theTMDL rules so difficult to interpretand why a simple indication ofwhether a water body is listed or notis not necessarily a good indication ofits level of water quality.

Once a water body has been iden-tified as not meeting its designateduse, the state is required to identifythe sources of the impairment andthe “maximum allowable daily load”of pollutants that would eliminatethe impairment. Finally, states are tosuggest reductions for the variouspollutant sources that would allowthe watershed to reach the TMDL.Importantly, there is no regulatoryauthority by the states or EPA torequire that these reductions occur.Thus, the institutional environmentin which nonpoint source water qual-ity reductions may occur is funda-mentally voluntary.

In the TMDL process, modelingand monitoring can play importantroles in allocating pollutant loads tovarious sources, such as helping todetermine the relative contributionsof row crops, CAFOs, and urbansources to loads of nutrients and bac-teria observed in large watersheds.Two models, the Soil and WaterAssessment Tool and the Hydrologi-cal Simulation Program-FORTRANmodels are most often used to sup-port TMDL assessments (Benham etal., 2006). These models combineGIS-based spatial data of watershedphysical features with information oncropping systems, animal densities,fertilizer and pesticide use, and point

sources. For non-point source pollut-ants, conservation practices are a keyto developing mitigation strategiesthat allow watersheds to meetTMDL goals. Since TMDLs may bedesigned to mitigate multiple pollut-ants (e.g., nitrate and bacteria), com-binations of conservation practicesmay be necessary to achieve the nec-essary improvements in water quality.

Final RemarksThe purpose of this overview is tointroduce readers to the set of waterquality problems associated withrow-crop agriculture and livestockoperations in the Corn Belt andUpper Mississippi River Basin. Theproblems are complex, with a greatmany individual decentralized deci-sion makers contributing, both posi-tively and negatively, to their solu-tions. Adding to these complexproblems are the ever-changingdemands on agriculture to supplyfood, feed, fiber, and fuel. Thesedemands are leading to new ques-tions and concerns related to agricul-ture and may allow for some solu-tions that are economically viableand environmentally beneficial.Some concerns are related to poten-tial use of marginal lands for rowcrop agricultural production andincreasing continuous corn acreage tosupply the bioeconomy. At the sametime, the bioeconomy, particularly ifcellulose biofuels become feasible,may provide opportunities for morediversified cropping systems thathave environmental benefits. Associ-ated with some of these issues is theincreasing importance of agribusinessthrough decisions such as siting ofCAFOs and ethanol plants. Sitingdecisions should consider the poten-tial environmental impacts of thesefacilities both from a water qualityand water quantity perspective.


In the remaining three papers ofthis water quality theme, the authorsdescribe how data and models can beused to characterize the problems,model the underlying biophysicaland economic processes, and ulti-mately (hopefully) contribute tosolutions. Given the policy environ-ment described above one of volun-tary-based action and a myriad ofconservation programs with diversegoals and ever-present funding con-straints we believe that models ofwater quality processes carefully inte-grated with economic models areessential, both to assess existing pro-grams, and more importantly, todesign and implement cost-effectiveapproaches to meeting society’s waterquality goals. These modeling effortswill be difficult and will appropri-ately come under a great deal of scru-tiny.

The complexity of the ecologyand the social issues (including a hostof topics not addressed here such asinternational trade agreements, ruralcommunity viability, rural-urbanconflicts, etc.) indicate a need foradditional research that considers thebreadth of the systems involved atscales that are appropriate. For exam-ple, much of our current knowledgeof the efficacy of conservation prac-tices is based on field scale researchwhich cannot be simply “scaled-up”to understand the workings at water-shed levels. While current researchefforts are beginning at this morechallenging scale, definitive resultswill be, in many cases, many yearsoff.

Before we leave the reader to diveinto the three following papers, wenote a final thorny point concerningthe potential for significant “legacy”problems possibly hiding in ground-water supplies. Over many decades ofagricultural activity, we have addednutrients and other effluents to

groundwater systems that haveundoubtedly not yet emerged at thesurface. When and where such pol-lutants will appear is not clear, but ifconservation programs are designedonly with current pollutant contribu-tions in mind, our efforts may wellfall short due to these legacy sources.

For More InformationBenham, B.L., Baffaut, C., Zeckoski,

R.W., Mankin, K.R., Pachepsky, Y.A., Sadeghi, A.M., Brannan, K.M., Soupir, M.L., & Habersack, M.J. (2006). Modeling bacteria fate and transport in watersheds to support TMDLs. Transactions of the American Society of Agricultural and Biological Engineers, 49(4), 987-1002.

Birr, A.S., & Mulla, D.J. (2001). Evaluation of the phosphorus index in watersheds at the regional scale. Journal of Environmental Quality, 30(6), 2018-25.

Burkart, M.R., & James, D.E. (1999). Agricultural nitrogen contributions to hypoxia in the Gulf of Mexico. Journal of Environmental Quality, 28, 850-59.

Burkart, M.R., Simpkins, W.W., Morrow, A.J., & Gannon, J.M. (2004). Occurrence of total dissolved phosphorus in surficial aquifers and aquitards in Iowa. American Journal of Water Resources Association, 40(3), 827-34.

Christophoridis, C., & Fytianos, K. (2006). Conditions affecting the release of phosphorus from surface lake sediments. Journal of Environmental Quality, 35(4), 1181-92.

Dinnes, D.L., Karlen, D.L., Jaynes, D.B., Kaspar, T.C., Hatfield, J.L.,

Colvin, T.S., & Cambardella, C.A. (2002). Nitrogen management strategies to reduce nitrate leaching in tile-drained Midwestern soils. Agronomy Journal, 94, 153-71.

Follet, R.F., & Hatfield, J.L. (ed.). (2001). Nitrogen in the environment: Sources, problems and management. Amsterdam: Elsevier Science.

Jaynes, D.B., Dinnes, D.L., Meek, D.W., Karlen, D.L., Cambardella, C.A., & Colvin, T.S. (2004). Using the late spring nitrate test to reduce nitrate loss within a watershed. Journal of Environmental Quality, 33(2), 669-77.

Kellogg, R.L., Lander, C.H., Moffitt, D.C., & Gollehon, N. (2000). Manure nutrients relative to the capacity of cropland and pastureland to assimilate nutrients: Spatial and temporal trends for the United States. USDA Economic Research Service Publication No. nps00-0579.

Kross, B.C., Hallberg, G.R., Bruner, D.R., Cherryholmes, K., Johnson, J.K. (1993). The nitrate contamination of private well water in Iowa. American Journal of Public Health, 83(2), 270-72.

Rodvang, S.J., & Simpkins, W.W. (2001). Agricultural contaminants in Quaternary aquitards: A review of occurrence and fate in North America. Hydrogeology Journal, 9(1), 44-59.

Sharpley, A.N., Kleinman, P.J.A., McDowell, R.W., Gitau, M., & Bryant, R.B. (2002). Modeling phosphorus transport in agricultural watersheds: Processes and possibilities. Journal of Soil and Water Conservation, 57(6), 425-39.


Smith, J.E., & Perdek, J.M. (2004). Assessment and management of watershed microbial contaminants. Critical Reviews in Environmental Science and Technology, 34, 109-39.

Tomer, M.D., Meek, D.W., Jaynes, D.B., & Hatfield, J.L. (2003). Evaluation of nitrate nitrogen fluxes from a tile-drained watershed in Central Iowa.

Journal of Environmental Quality, 32(2), 642-653.

Matthew J. Helmers ([email protected]) is Assistant Professor,Department of Ag/Biosystem Engi-neering, Thomas M. Isenhart ([email protected])Assistant Professor,Department of Natural ResourceEcology and Management. CatherineL. Kling ([email protected]) Profes-sor, Department of Economics and

Center for Agricultural and RuralDevelopment, and William Simpkins([email protected]) is Professor,Department of Geology and Atmo-spheric Sciences, Iowa State Univer-sity, Ames, IA. Thomas B. Moorman([email protected]) isMicrobiologist, and Mark Tomer([email protected]) isResearch Soil Scientist, USDA-ARS,National Soil Tilth Laboratory,Ames, IA.





2nd Quarter 2007 • 22(2)


A Tale of Three Watersheds: Nonpoint Source Pollution and Conservation Practices across IowaBy Keith E. Schilling, Mark D. Tomer, Philip W. Gassman, Cathy L. Kling, Thomas M. Isenhart, Thomas B.

Moorman, William W. Simpkins, and Calvin F. Wolter


Many conservation practices and implementation pro-grams exist to address nonpoint source (NPS) pollutionlosses from agricultural landscapes (Helmers et al., thisissue). In order to select the most appropriate practices andprograms for reducing NPS pollution in a specific regionwhile maintaining economic return for the landowner, theinteracting processes of agricultural management andwatershed hydrology need to be understood across broadspatial scales. On a nationwide basis, it is easy to see howNPS pollution in one part of the country might be differ-ent than those in another region of the country. For exam-ple, cotton growers in the South, dairy farmers in theNortheast, cattle ranchers in the West, and grain farmersin the Midwest all face unique challenges based on differ-ences in climate, soil types, and cropping patterns. Eachregion relies on a different set of conservation practicesand programs to address NPS pollution. To be effective,conservation systems must be based on an understandingof specific management impacts on water quality prob-lems, and therefore be targeted to reduce, intercept, and/ortreat contaminants moving via surface or sub-surface path-ways from working agricultural lands.

Within agricultural regions, one might expect greaterhomogeneity in biophysical features and cropping prac-tices and be tempted to think that one size fits all; i.e., thatone set of conservation prescriptions can be used toaddress the negative impacts of agriculture on aquatic andterrestrial integrity. If this generalization could be madeanywhere, certainly a state such as Iowa, dominated by itsvast extent of corn and soybean fields, would be suited fora limited set of conservation prescriptions. However, as

described in this tale of three watersheds, conservationpractices must instead be tailored to individual landownerobjectives and local landscape conditions in order to opti-mize their effectiveness.

The research described in this paper was conducted aspart of USDA’s Conservation Effects Assessment Project(CEAP) and its Watershed Assessment Studies (Mausbachet al., 2004). The objectives of the project are to evaluatethe effects of agricultural conservation practices on waterquality, with a focus on understanding how the suite ofconservation practices, the timing of these activities, andthe spatial distribution of these practices throughout awatershed influence their effectiveness. An additionalcomponent of the project is to evaluate social and eco-nomic factors influencing implementation and mainte-nance of practices.

Watershed DescriptionsTo evaluate the effects of watershed conservation practiceson water quality, and to assess the spatial distribution ofthese practices, we are focusing on three watersheds in dis-tinct landscape regions of Iowa (Figure 1). By studyingthree watersheds with differing physical characteristics andpossessing a unique set of pollutants, practices and pro-grams, we can better assess the effectiveness of conserva-tion activities and land management decisions.

LandformsThe Sny Magill Creek, Squaw Creek, and the South Forkof the Iowa River (South Fork) watersheds are representa-tive of three distinct landform regions of Iowa (Prior,


1991). In Northeast Iowa, Sny Mag-ill Creek is a third-order stream inClayton County that drains 35.6 mi2

of the Paleozoic Plateau landformregion before discharging directlyinto the Mississippi River. The land-scape of this region is characterizedby narrow, gently sloping uplandsthat break into steep slopes withabundant outcrops of sandstone andlimestone. The characteristic lime-stone bedrock of the area gives rise tokarst features (sinkholes, caves, andsprings) that are found throughoutthe Sny Magill Creek watershed (Fig-

ure 2). Nearly 80% of annual stream-flow is ‘baseflow’ attributable toground water discharge from thesesubsurface sources. This results in“cold water” conditions suitable forhighly popular trout fisheries.

The 18.3 mi2 Squaw Creekwatershed is located in South-CentralIowa in Jasper County in the South-ern Iowa Drift Plain. The landscapeof this region is characterized bysteeply rolling hills and a well-devel-oped stream network that developedon a landscape composed of geologi-cally old (>500,000 years) glacial till

(poorly sorted mixture of gravel,sand, silt, and clay) overlain by geo-logically recent (17,000 to 31,000year old) windblown silt (loess).Because of the sloping hillsides andpoor infiltration capacity of the soils,rainfall is primarily directed tostreams via overland runoff, and only55% of the stream discharge is attrib-utable to baseflow originating asground water.

The largest of the study water-sheds is the South Fork of the IowaRiver, which covers 301 mi2 withinHardin and Hamilton counties inCentral Iowa. The landscape is repre-sentative of the Des Moines Lobe,the dominant landform region ofNorth-Central Iowa. The terrain isyoung (about 12,000 years since gla-cial retreat), and thus much of thelandscape is dominated by low reliefand poor surface drainage. Prior tosettlement by Europeans, the land-scape was a complex of wetlands, andthe stream network was poorly devel-oped due to the relatively younglandscape. The geology of the DesMoines Lobe region consists largelyof glacial till deposits in morainesand flat to rolling uplands, clay andpeat in depressional “prairie pothole”areas, and sand and gravel deposits infloodplains of rivers and streams. Soilwetness is a major concern for landmanagement and agricultural pro-duction. Hydric soils (indicative ofsoil saturation on at least a seasonalbasis) occupy about 54% of thewatershed, and artificial tile drainage(Figure 2) was installed in thesehighly productive and nutrient richsoils to lower the water table andallow crops to be grown. Thus, about70% of the stream flow in the SouthFork watershed originates from sub-surface drainage (Green et al., 2006),with most tile discharge occurringduring spring and early summer.

Figure 1. Location of the three watersheds (and controls) and EPA Ecore-gions in Iowa.

Figure 2. Subsurface hydrologic features in Sny Magill and South Fork water-sheds. In the Sny Magill watershed (left photo), springs and caves dischargenatural drainage from unknown areas of karst terrain. In the South Fork water-shed, once-prevalent wetlands were converted to cropland with tile drainage.This 36-inch clay pipe (right photo) has discharged drainage collected fromabout 4,500 acres of cropland for nearly 100 years (note monitoring lines).


Physical features of the threewatersheds have a great influence onthe timing and magnitude of therouting of water to streams. Water-sheds draining older landscapes havegreater slope and greater stream den-sity (number of streams per squaremile) than younger landscapes (Fig-ure 3). For example, the Sny Magillwatershed has twice the average slopeas Squaw Creek, which has morethan twice the slope of the SouthFork watershed. Slopes in Sny Magillare further accentuated because of

the bedrock terrain and its proximityto the Mississippi River. The SquawCreek and Sny Magill watersheds alsohave nearly three times more streamsper square mile than the South Fork.The well-dissected landscape of theSny Magill watershed shows a greaterstream density; thus, rainfall can bequickly routed as overland runoff tosinkholes or streams. In the SouthFork watershed, where natural drain-age is poor, excess rainfall would col-lect in potholes or other surfacedepressions if not for prevalence of

subsurface tile drainage, which hasaccelerated the routing of rainfallwater off the land. Watersheds drain-ing the Des Moines Lobe may yieldas much water as those draining frac-tured carbonate bedrock (Schillingand Wolter, 2005).

Relation of Land Use to the Landform RegionDifferences in land cover among thethree watersheds can be traced largelyto their watershed morphologies andthe suitability of land for intensive

Figure 3. Land use, conservation practices, and other characteristics of the three watersheds.


row crop agriculture. Row crops inthe Sny Magill watershed, primarilyfound on narrow upland divides andbottomlands, only comprise 26% ofthe land area (Figure 3). Grasses andforest are widespread in the Sny Mag-ill watershed, located on steep terrainthat is difficult to cultivate. In theSquaw Creek watershed, slopes arenot as severe as in the Sny Magillwatershed, and row crops are foundon 81% of the land area. Grasses aredistributed around the watershed onhighly erodible land, a practiceencouraged by conservation pro-grams. The till plain of the DesMoines Lobe, represented by theSouth Fork watershed, is also heavilyutilized for row crop production,which occupies 85% of the water-shed area.

Animal AgricultureIn the early 1900s, most small farmsin Iowa had livestock, often includ-ing cattle, swine, and chickens. As aresult of changes in farm policy andeconomies of scale, all three water-sheds have experienced shifts in ani-mal agriculture that are representa-tive of changes across the largerlandform regions (Figure 1). Histori-cally, the Sny Magill watershed hadsignificant numbers of dairy cattleutilizing available grasslands for for-age. While still a significant industry,dairy cattle within Sny Magill havedecreased greatly, with a resultingshift in some grassland acreage to rowcrop agriculture (soybean acreageespecially increased in Iowa as pas-ture and hayland decreased). Live-stock is comparatively absent in theSquaw Creek watershed except forseveral cow-calf operations. Nowhereis the concentration of livestock moreapparent than in the South Forkwatershed, where most swine andchickens are raised in confined ani-mal feeding operations (CAFOs).

There are nearly 100 CAFOs (mostlyswine) in the watershed. Based oninventories reported for permittedfacilities, hogs and chickens in theSouth Fork watershed number 1,654and 2,880 per square mile, respec-tively, which are densities consider-ably greater than the other twowatersheds combined (Figure 3). Allthe reported chickens are housed inone large egg-producing facility,while swine facilities are abundantacross the central part of the water-shed. We estimate that about a quar-ter of the watershed receives manureapplications annually, assuming thisis applied prior to corn at a rateequivalent to that crop’s uptake ofnitrogen (about 190 lb N/ac). Usu-ally these applications are done byinjection, and carried out in the fallwhen soils tend to be dry and mosteasily trafficked by manure tankersand applicators.

NPS Pollutants and the LandscapeBecause of their different watershedcharacteristics, land use, and live-stock histories, non-point pollutantsources and transport vary greatlyamong the three watersheds (Figure3). Pollutants of particular concern inIowa are sediment, nutrients (nitrateand phosphorus), and fecal bacteria(E. coli). In Iowa, nitrate concentra-tions in streams relate to the amountof row crops in a watershed (Schilling& Libra, 2000), and nitrate-N con-centrations are highest in the SouthFork and Squaw Creek watersheds,with median concentrations of 14.2and 9.5 mg/L, respectively. Tiledrains contribute greatly to nitratelosses in the South Fork watershed.In the mid-1990s, the USGS foundstream nitrate concentrations in theSouth Fork watershed to be among

the highest observed in the UnitedStates (Becher et al., 2001).

In contrast, nitrate concentra-tions are considerably lower in theSny Magill watershed, averaging 3.3mg/L over 10 years, a value thatwould be the envy of most otherregions of Iowa. The smaller concen-tration results from the differences inland use (Figure 3). Fecal bacteriacounts are also highest in the SouthFork watershed; however, multiplesources of bacteria are suspectedbecause patterns do not always followthe distribution of livestock. Yet,research suggests that these bacteriallosses in runoff are greatest when thatrunoff occurs within several weeks ofmanure application. Fecal bacteriaconcentrations in Squaw Creek arealso elevated, which may be surpris-ing, given the lower livestock densi-ties. However, cattle with directaccess to the streams, wildlife, andinadequate septic systems may allcontribute to fecal contamination ofIowa streams.

Sediment loss is also a major con-cern in these watersheds (Figures 3and 4). The greatest annual sedimentloss per unit area was associated withthe Squaw Creek watershed (0.69tons/ac per year), whereas meanannual sediment loss from Sny Mag-ill and South Fork watersheds aver-aged 0.26 and 0.28 tons/ac, respec-tively. Considering that row cropscover only 26%of the land area in theSny Magill watershed, actual soil lossper acre of cropland may be substan-tially greater. Long-term sedimentmonitoring data in the Sny Magilland Squaw Creek watersheds indi-cates that sediment transport is veryflashy in both watersheds, with muchof the annual sediment loss trans-ported by runoff from a few intenserainfall events. In Squaw Creek, onaverage, about 40% of the water-


shed’s annual sediment loss occurs onthe single day of greatest runoff.

Sediment losses from watershedsresult from overland flow across thelandscape, causing sheet, rill, andgully erosion, as well as substantialcontributions from streambanks. Inthe South Fork watershed, sedimentlosses are actually about three timeshigher than typically measured in theDes Moines Lobe region. In thelower third of the watershed, landsbecome more highly erodible in anarea of hilly moraines near the DesMoines Lobe’s edge, and the rivererodes its banks as it meanders acrossan alluvial valley (Figure 4). In someIowa watersheds, streambank ero-sion can contribute more than half ofthe annual sediment load exportedfrom a watershed.

Phosphorus is strongly adsorbedto sediment, which was reflected inSquaw Creek having both the great-est average sediment yields and great-est median concentrations of total P(0.14 mg/L) among the three water-sheds. Squaw Creek’s median P con-centration is more than twice whatEPA has proposed as the standard forMidwest streams. Concentrations ofP in the South Fork, by comparison,had a median of 0.07 mg/L during

three years of weekly-biweekly sam-pling. Recent groundwater samplingfrom 24 wells located throughout theSouth Fork watershed has shownmedian and maximum total P con-centrations of 0.030 and 0.340 mg/L, respectively. These groundwater Pconcentrations are found in similarmaterials and landscapes in Iowa(Burkart et al., 2004), and suggestthat groundwater can also be a P con-tributor to streams.

Tailoring Conservation Practices to Watersheds and NPS PollutantsConservation practices used on rowcrop fields in the three watershedsreflect respective watershed charac-teristics and land use histories. Afield-by-field assessment of conserva-tion practices was conducted in eachwatershed to assess the variety anddistribution of practices. An analysisof tillage practices, terraces, and con-tour farming shows the degree towhich land managers have used theseconservation practices to reducenutrient and sediment losses fromthe three watersheds. Of the threetillage practices assessed (conven-tional tillage, mulch till, and no till),

mulch till was most widely utilized inall watersheds. Mulch tillage (>30%residue cover) was used on 62 to91% of all row crop fields, whereasno till was used on 8 to 16% of thefields. Conventional tillage (<30%residue cover) was rarely used in SnyMagill and Squaw Creek, but wasused on 30% of cropland in SouthFork. Erosion losses from crop fieldsin South Fork are not a major con-cern in the flat, till plain portion ofthe watershed. In areas of the North-ern United States, with relatively flatterrain and poorly drained soils,many producers still view conven-tional tillage as the most viable prac-tice because the exposed soil iswarmed faster in spring, often allow-ing earlier seeding and emergence ofthe crop.

In the Sny Magill watershed, con-tour farming, terraces, and otherengineered structures are prevalentpractices for reducing sediment lossesfrom the steeper slopes in that water-shed. Although row crop fields occu-pied only 26% of the land area inSny Magill, most are terraced (77%)and/or farmed using contour plant-ing (92%). Other engineered conser-vation practices are also used exten-sively throughout the Sny Magill

Figure 4. Eroding streambanks and erodible soils are possible contributors to sediment loads. In the South Fork water-shed, both cut-bank meanders and erodible soils are mostly found in the lower (eastern) part of the watershed.


watershed, including a total of over150 sediment basins and grade stabi-lization structures. Terraces are not ascommon in Squaw Creek (23%), buthalf the farm fields are planted oncontours. Terraces and contour farm-ing are not common in the SouthFork watershed; fields with terracesoccupy less than 10% of the water-shed’s cropland.

A Tale of Three Watersheds - revisitedIn this tale of three Iowa watersheds,significant differences in NPS pollut-ants and practices emerged in a stateconsidered by many to be uniformlyagricultural. Much of the differencescan be attributable to their uniquelandform history that has beenexploited uniquely for intensive rowcrop and livestock development. Inthe South Fork watershed, extensivewetlands on recently glaciated tillplains were drained by settlers, andagricultural development then inten-sified during the past century. Theland is well suited for crop and live-stock production, but subsurface tiledrainage increases losses of nitrate,and the rapid routing of tile dis-charge, combined with surface run-off, may enhance movement of bac-teria, P, and sediment. In SquawCreek, with steeper slopes in rowcrops, conservation practices such asreduced tillage and contour farmingmethods are more prevalent. How-ever, losses of nutrients, sediment,and fecal bacteria remain as majorconcerns in the watershed, possiblybecause hydrologically sensitive areasare used for row crops or grazing.Row crop acreage in the Sny Magillwatershed constitute only about 25%of the land area and most row cropfields have conservation tillage orstructural practices such as terraces.However, the steep slopes and karst

drainage in the watershed make SnyMagill watershed perhaps the mostvulnerable among the three streamsevaluated.

Apparent in this tale of threeIowa watersheds is that, in order toprovide the greatest return on thepublic’s investment in conservation,it is imperative that practices be tai-lored to the most local of landscapeconditions and landowner objec-tives. Targeting is needed to placespecific conservation practices on theland to either reduce pollutant con-centrations or attenuate their trans-port. No single practice can beviewed as the answer in all cases, anda one-size-fits-all approach is likelydoomed to failure, or at least doomedto provide little return on the publicinvestment. Recent advances inassessment technologies and recordkeeping are only now beginning toallow us to understand the distribu-tion of practices on the land andtheir impacts on water quality. Sig-nificant challenges remain to developbetter assessment, monitoring, andmodeling techniques to capture theinherent differences among ourwatersheds in order to design conser-vation practices and programs pro-viding greater water quality benefitsfor lower cost. The challenges are notonly in assessing resource needsagainst the mosaic of land use andterrain that occur within watersheds,but also to then develop better policyand planning tools that can helpachieve watershed-scale conservationgoals through implementation at theindividual farm scale.

For More InformationBecher, K.D., Kalkhoff, S.J.,

Schnoebelen, D.J., Barnes, K.K., & Miller, V.E. (2001). Water-Quality Assessment of the Eastern Iowa Basins – Nitrogen,

Phosphorus, Suspended Sediment, and Organic Carbon in Surface Water, 1996-98. U.S. Geological Survey Water-Resources Investigations Report 01-4175. Iowa City, IA.

Burkart, M.R., Simpkins, W.W., Morrow, A.J., & Gannon, J.M. (2004). Occurrence of total dissolved phosphorus in surficial aquifers and aquitards in Iowa. Journal of the American Water Resources Association, 40(3), 827-834.

Fields, C.L., Liu, H., Langel, R.J., Seigley, L.S., Wilton, T.F., Nalley, G.M., Schueller, M.D., Birmingham, M.W., Wunder, G., Polton, V., Sterner, V., Tial, J., & Palas, E. (2005). Sny Magill Nonpoint Source Pollution Monitoring Project: Final Report, Iowa Department of Natural Resources, Geological Survey Bureau Technical Information Series 48, Iowa City, IA.

Green, C.H., Tomer, M.D., DiLuzio, M., & Arnold, J.G. (2006). Hydrologic evaluation of the Soil and Water Assessment Tool for a large tile-drained watershed in Iowa. Transactions American Society of Agricultural and Biological Engineers, 49(2), 413-422.

Helmers, M.J., Isenhart, T.M., Kling, C.L., Mooreman, T.B., Simpkins, W.W., & Tomer, M. (2007). Agriculture and water quality in the Cornbelt: Overview of issues and approaches. Choices 22(2), this issue.

Mausbach, M.J., & Dedrick, A.R. (2004). The length we go: Measuring environmental benefits of conservation practices. Journal of Soil and Water Conservation, 59(5), 96A-103A.


Prior, J.C. (1991). Landforms of Iowa. Iowa City, IA: University of Iowa Press, 153 pp.

Schilling, K.E., Hubbard, T., Luzier, J., & Spooner, J. (2006). Walnut Creek watershed restoration and water quality monitoring project: final report. Iowa Department of Natural Resources, Geological Survey Bureau Technical Information Series 49, Iowa City, IA. Available online: http://www.igsb.uiowa.edu/.

Schilling, K.E., & Libra, R.D. (2000). The relationship of nitrate concentrations in streams to row crop land use in Iowa.

Journal of Environmental Quality, 29(6), 1846-1851.

Schilling, K.E., & Wolter, C.F. (2005). Estimation of streamflow, baseflow and nitrate-nitrogen loads in Iowa using multiple linear regression models. Journal of American Water Resources Association, 41(6), 1333-1346.

Keith E. Schilling ([email protected]) is Research Geologistand Calvin F. Wolter ([email protected]) is Geologist, IowaGeological Survey – Iowa Depart-ment of Natural Resources, Ames, IA.Mark D. Tomer ([email protected]) is Research Soil Scientist

and Thomas B. Moorman ([email protected]) is Research Leader,USDA-ARS, National Soil TilthLaboratory, Ames, IA. Philip W. Gas-sman ([email protected]) isAssociate Scientist, Center for Agri-cultural and Rural Development,Cathy L. Kling ([email protected])is Professor, Department of Econom-ics and Center for Agricultural andRural Development, Thomas M.Isenhart ([email protected]) isAssistant Professor, Department ofNatural Resource Ecology and Man-agement, and William W. Simpkins([email protected]) is Professor,Department of Geology and Atmo-spheric Sciences, Iowa State Univer-sity, Ames, IA.

Watershed Highlight 1: Historical and Human Dimension: Squaw Creek and Walnut Creek Paired Watershed Study.

Because Squaw Creek represents typical agricultural land management in Southern Iowa, the watershed was selected to be the control basin for a large land use experiment occurring in the neighboring Walnut Creek watershed (Figure 5). In the Walnut Creek watershed, large tracts of row-cropped land are being recon-structed to native prairie at the Neal Smith National Wildlife Refuge by the U.S. Fish and Wildlife Service. Before restoration began, land cover in both water-sheds was about 70% row crop. From 1992 to 2005, nearly 220 acres of prairie was planted each year, so that by 2005, native prairie occupied 23.5% of Walnut Creek watershed. Surface water samples collected in the treatment (Walnut Creek) and control (Squaw Creek) watersheds from 1995 to 2005 documented the effects of prairie restoration on water quality (Schilling et al., 2006). Stream nitrate concentrations were found to have decreased 1.2 mg/L over the 10-year project period at the Walnut Creek outlet, with nitrate concen-trations decreasing up to 3.4 mg/L over the same time period in one monitored subbasin with substantial landuse conversion. Interestingly, land use in the con-trol basin of Squaw Creek did not remain static during the same 10-year monitoring period. Row-crop land area increased 9.2% in Squaw Creek as lands previously enrolled in the Conservation Reserve Program as grass-land were converted back to row crop production in the late 1990s. Stream nitrate concentrations increased 1.9 mg/L at the Squaw Creek outlet, with annual nitrate in one monitored subbasin increasing nearly 12 mg/L in 10 years where substantial acres were converted to row crops. These results attest to the sensitivity of water quality parameters to changes in watershed management that are, in aggregate, the result of indi-vidual landowner decisions.

Figure 5. Extent of Prairie plantings in the Neal Smith National Wildlife Refuge within the Walnut Creek watershed.


Three separate projects were carried out spanning the time period of 1988 to 1999 to improve water quality in the Sny Magill Creek watershed. The cumulative adoption percentages and total levels of key BMPs implemented during the 1990s through the Sny Magill Hydrologic Unit Area (HUA) and Sny Magill Creek Watershed projects are listed for selected years in Table 1. The cumulative adoption of terraces in the watershed is also shown in Figure 6 for 1991, 1995, and 2005. A paired watershed approach was used to assess Sny Magill Creek water quality improvements from 1992 to 2001 (Fields et al., 2005). Analysis of Sny Magill stream flow and water quality data collected during 1991-2001 was performed using a pre/post statistical model.

The statistical results indicated that discharge at the watershed outlet increased by 8% over the 10-year period; this could partly be due to routing of runoff water captured by terraces into surface inlet drains (that are often installed just upslope of a terrace) and to the stream. The statistical analysis also showed that the BMPs installed during the 1990s resulted in a 42% decrease in turbidity but only a 7% decrease in total suspended solids (TSS). The TSS results imply that stream bed and bank erosion continued to contribute significant sediment loads to Sny Magill Creek, even after BMP installation reduced sediment delivery from upland areas. The increase in discharge may have further magnified the in-channel sedi-ment contributions. Overall, the TSS results suggest that a long lag time may occur before the full impacts of the installed BMPs are realized.

The statistical analysis also revealed that an increase in nitrate concentra-tions of 15% was found at the SMCW outlet. This indicates greater N leach-ing, which is consistent with increased infiltration of rainfall that naturally results when conservation practices successfully decrease sur-face runoff. However, the nitrate con-centration level still only slightly exceeded 3 mg/L at the end of the 10-year time period, which is quite low compared with the concentra-tions measured in most other Iowa stream systems, including the South Fork and Squaw Creek watersheds.

Table 1. Cumulative percentages of total BMP adoption that was cost shared by year (expressed as a percentage of the total amount implemented as given in the bottom line).

Year Terrace Subsurface tile Sediment basin Grade stabilization Field border Contouring

1992 28 22 28 92 16 11

1995 65 65 79 93 99 53

1998 95 94 98 100 100 100

2001 100 100 100 100 100 100

Total Units 269,585 ft 160,345 ft 61 total 90 total 26,700 ft 1,907 ac

Figure 6. Cumulative additions of terraces to specific land tracts in the Sny MagillCreek Watershed.

Watershed Highlight 2: Long-term Implications of BMP Implementation in the Sny Magill Creek Watershed.


The Conservation Title of the 1985 Farm Bill (Food Security Act) included provisions to reduce soil erosion on highly erod-ible land (HEL) through conservation practices such as Conservation Reserve Program (CRP) plantings and reduced tillage. Land enrolled in CRP was planted to perennial, non-harvested vegetation for at least a ten-year period in exchange for annual rental payments. Soil survey data, including slope, soil texture and depth are used to identify HEL. Those producers farming on HEL-dominated fields were to employ reduced tillage practices to remain eligible for USDA commodity pro-grams; this was known as the conservation compliance provision of the 1985 Farm Bill.

A one-time inventory of conservation practices in the South Fork watershed was conducted during 2005. We compared the distribution of no-tillage management and CRP plantings with the distribution of HEL, which occupies 12% of the watershed (Figure 7). Very little (2.4%) of the watershed’s cropland had been enrolled into CRP by producers, partly because this is some of the most productive rain-fed agricultural land in the U.S. While CRP has also been used to install buffers along streams and around livestock facilities, there has been apparent success in targeting of CRP towards HEL. That is, the proportion of HEL in the watershed in CRP is 4.6%, as opposed to only 2.2% of non-HEL (Table 2). The same is true of no-tillage practices that are highly effective in controlling erosion: although relatively few producers in this water-shed have implemented no-tillage, largely due to concerns about planting delays during wet, cool spring conditions, a greater proportion of HEL (11.3%) is under no-tillage than is non-HEL (6.7%). There is little comparative data to evaluate

whether these practices are better targeted towards HEL in this watershed than in other areas. However, targeting success may also be indicated if conventional tillage prac-tices that increase soil susceptibility to ero-sion have shifted away from HEL as these conservation practices were implemented. This does not appear to be the case, as con-ventional tillage occupies nearly the same proportion of HEL and non-HEL cropland, to within 2%.

It is important to note that in 1985, when current policies where initiated, most of this watershed was probably tilled conven-tionally. This inventory offers only a snap-shot of conservation practices. Current conservation policies, which have had a goal of controlling soil erosion from the most sensitive soils for 20 years, have encouraged better management on the most vulnerable lands in the watershed. Yet, the least desirable tillage practices apparently have not preferentially shifted away from HEL. This raises questions about social-behavioral responses to conserva-tion policies, which are made by individual producers, yet in sum determine the impact of those policies in each watershed.

Table 2. Comparative distributions of CRP and no-tillage conservation prac-tices on highly erodible and non-highly erodible lands in the South Forkwatershed, along with conventional tillage practices.

Management HEL (12%) Non- HEL (88%) Total (100%)

Conservation Reserve Program 4.6% 2.2% 2.4

No-tillage 11.3% 6.7% 7.2

Conventional tillage 28.0% 30.0% 29.8

Figure 7. Distribution of key conservation practices for erosion control inthe South Fork watershed, compared to the distribution of Highly ErodibleLand.

Watershed Highlight 3: Evaluating Targeting of Conservation Practices in the South Fork Watershed.





2nd Quarter 2007 • 22(2)


Privatizing Ecosystem Services: Water Quality Effects from a Carbon MarketBy Silvia Secchi, Manoj Jha, Lyubov Kurkalova, Hongli Feng, Philip Gassman, and Catherine Kling


With the specter of a new farm bill on the horizon, con-siderable discussion is occurring concerning the possibleredirection of conservation programming and financing.Notably, interest in the increased use of incentive systemsand market-like instruments continues to expand. Onesource of this interest lies in the desire to shift some of theburden of providing ecosystem services, such as protectingstream and river channels from erosion, maintainingbiodiversity, and providing clean water and air, to privatesector pockets. For example, in the fall of 2006, USDAand EPA announced a joint partnership to supportexpanded water quality credit trading for nutrients in theChesapeake Bay watershed, allowing farmers to receivecompensation for water quality improvements. Carbonmarkets, such as the active program in the EuropeanUnion, are also being discussed as a possible model forexpanded market-like programs in agricultural conserva-tion policy.

While the potential cost effectiveness of providingenvironmental goods from incentive-based methodsappears to be broadly understood, there is an additionalattribute that is less broadly acknowledged: due to numer-ous inter-linkages in natural ecosystems, the developmentof a market that provides one ecosystem service may sig-nificantly change the level of provision of other ecosystemservices. Thus, by developing the institutional structure tosupport and encourage the provision of one ecosystem ser-vice, changes, either positive or negative, in other servicesmay result.

The example we consider here is the case in which acarbon market that would allow U.S. farmers to receivepayment for sequestering carbon when they retire landfrom production is implemented. This could occur if the

United States were to unilaterally implement such a mar-ket, or if at some future time the United States chose tosign on to a Kyoto-like accord, where carbon sinks wereallowed to generate credits that could be traded to meetmandatory carbon reduction requirements. Under such ascenario, land retirement decisions would be driven by theprices paid for carbon and the amount of carbon that aparticular parcel could sequester (we abstract from theimportant question of measuring the carbon storagepotential of each parcel see the excellent work of Mooneyet al., 2004).

Removing a parcel of land from production willchange the suite of environmental benefits associated withthe parcel. In many cases, these effects are likely to be pos-itive for example, taking land out of active agriculturalproduction and placing it in perennial cover or forestedlands will usually yield reduced erosion and nutrient run-off relative to row crop agriculture. Indeed, the findings ofour study are consistent with this outcome. However, ifconservation practices are already in place on a workingland field, water quality improvements from retiring theland might be small and, in fact, could be negative. Thelatter could occur if land retirement results in the plantingof land cover that, on the whole, is not as effective in cap-turing nutrients and sediments as a working land systemthat already has effective conservation practices and man-agement in place.

In this paper, we consider the possible water qualityconsequences of a carbon trading policy that allows farm-ers to receive carbon credits from retiring their land fromagricultural production. To do so, we make a number ofsimplifying assumptions about the structure of the carbonmarket and the choices farmers make in response to the


existence of that market; many ofthese assumptions may not, in fact,represent how an actual marketmight be implemented. Rather thanview the results of this analysis asdefinitive, we present the findings inthe spirit of raising awareness of thepotential environmental conse-quences that can occur when a singleenvironmental benefit or target (car-bon sequestration) forms the basis ofenvironmental policy, as would bethe case if carbon trading marketsthat allowed land retirement to yieldcarbon credits were functioning withhigh carbon prices.

A Bit about Our Data and ModelsTo develop our models, we drawheavily from the National ResourceInventory to provide data on the landuse, cropping history, and farmingpractices in the state of Iowa. Thisinventory is the most comprehensivedata set on land use in the UnitedStates, and we use data on the 14,472physical points in Iowa that representcropland. Conceptually, our data andmodels are based on individual pro-

ducer and farm-level behavior, andwe treat an NRI point as a producerwith a farm size equal to the numberof acres represented by the point (theexpansion factor provided by the sur-vey). Figure 1 illustrates the 35watersheds corresponding to theUnited States Geological Survey 8-digit Hydrologic Cataloging Unitsthat are largely contained in the stateand are modeled in this study. Tocompute the amount of carbonsequestered when a land unit isretired from cropland, we rely onestimates from the EnvironmentalPolicy Integrated Climate Model ver-sion 3060. When land is retired fromcrop production, we assume thatannual grasses are planted and main-tained on the land, and we run a 30-year simulation to predict the carbonsequestration level associated withthis change.

In addition, we also rely on esti-mates from a watershed-based modelto assess the conservation policies.Unlike carbon sequestration, thedegree to which land retirementimproves in-stream water qualitydepends on critical interactions

between land uses in different loca-tions within a watershed. Thus, forotherwise identical tracts of land,more water quality improvementmay occur from retiring a piece ofland from production that is locateddownstream from numerous othercropped points relative to one that isnot. The potential filtering effect isjust one example of the physical pro-cesses that need to be captured toassess the in-stream water qualityeffects of land retirement.

So that we can capture these landuse interactions within a watershedsetting, we employ the Soil andWater Assessment Tool, a biophysicalwater quality model, to estimatechanges in nitrogen, phosphorous,and sediment loads from retiring aparticular set of parcels from produc-tion within a watershed. To estimatethe in-stream water quality conse-quences of the increase in land setaside, we have calibrated the waterquality model for each of the water-sheds identified in Figure 1 to base-line levels (Jha et al., 2005; Gassmanet al., 2005). By running the modelat the set-aside levels “after” the pol-icy, we can compute the changes inwater quality attributable to theincrease in land set-aside. The water-sheds studied correspond to 13 out-lets, at which the in-stream waterquality is measured. The water qual-ity measures of interest are sediment,nitrogen, and phosphorus.

Water Quality Effects of a Carbon MarketTo demonstrate the possible conse-quences of a carbon market that paysfarmers for the sequestration of car-bon in agricultural soils on waterquality, we consider a simple sce-nario. Suppose that through an activecarbon market, the price of carbon issuch that about 10% of Iowa crop-

Figure 1. Study area and watershed delineations.


land is retired; suppose further thatthe cost of retiring all land within thestate is about the same. While rentalrates for farmland do vary across thestate, they vary relatively little withrespect to productivity (see Secchi &Babcock, forthcoming), and this sim-plifying assumption allows us tofocus on environmental outcomes ofthe scenario without overly compli-cating the analysis. Under theseassumptions, the cropland that willbe removed from production will bethe land that produces the highestcarbon sequestration benefits per acreas this land will earn the highestreturn from carbon sequestrationcredits. In consequence, the landremoved from production may ormay not represent significant por-tions of the watersheds under consid-eration.

Based on this scenario, the landretired would be focused in the cen-tral part of Iowa, in the ecoregionknown as the Des Moines Lobe, a flatarea, with very productive agricultureand particularly suited for carbonsequestration. Figure 2 illustrates thequantity and location of the carbonthat the carbon simulation modelpredicts would be sequestered acrossthe state under this scenario. Approx-imately 2 million acres of land isremoved from production under thisscenario with about 2.7 million tonsof carbon being sequestered annually.

Does this land retirement,induced by a private market that paysfor ecosystem services, yield otherenvironmental benefits to the region?To answer this question, we estimatethe in-stream water quality effects ofthis land retirement using the waterquality model and present the per-centage reductions in three commonindicators in Figures 3-5.

Figure 3 reports the estimated in-stream sediment reductions from theretirement of this set of land parcels.

We find that for the two largestwatersheds whose sediment dynamicsis influenced by the presence of largereservoirs, the Des Moines and IowaRiver watersheds, there is only a littleimprovement in sediment. In con-trast, there are larger reductions insediment in the South-Western partof the state, likely because the land

that is retired from production thereis more erodible. In general, however,sediment reductions in percentageterms are lower than the reductionsin nutrients, because land that hashigh carbon sequestration potentialalso has good productivity levels andis, therefore, more heavily fertilized.

Figure 2. Carbon sequestration from carbon trading.

Figure 3. In-stream sediment changes from carbon trading (percentagereduction).


The retirement of land generallyimproves the total N level as seen inFigure 4. The reason for theimproved N levels in many cases, asmentioned above, is that the landtaken out of production is largelyprime agricultural land: heavily fertil-ized and reliant on tiled drainage sys-

tems. Some of the highest reductions,a total of over 10,000 tons annualaverage, are in the Des Moines Riverwatershed, which comprises largeparts of the Des Moines Lobe andincludes some of the most productiveland in the state where most of theacres of land retirement are located.

Finally, Figure 5 reports theresults for the in-stream phosphorouslevels predicted to occur as a result ofthe carbon trading program. Like thesediment results, the Western water-sheds show the highest improve-ments. This is not surprising giventhe sediment results, since phospho-rous typically moves with sediment.

The development of more mar-ket-like programs to provide ecosys-tem services from agriculture is aconcept with expanding interest. Inthis paper, we have identified anadditional issue associated with thisstrategy, changes in other environ-mental goods of interest. In the caseanalyzed here, these changes were allpositive; thus, the market-based sys-tem yields positive gains for otherecosystem services. By recognizingthat a system that pays for carbonsequestration via land retirementpotentially has effects on other envi-ronmental services, and that the spa-tial distribution of different environ-mental services is likely to differ,policy makers can incorporate theseeffects in planning and implementingmarkets for ecosystem services.

For More InformationFeng, H., Kurkalova, L.A., Kling,

C.L., & Gassman, P.W. (2007). Economic and environmental co-benefits of carbon sequestration in agricultural soils: Retiring agricultural land in the Upper Mississippi River Basin. Climatic Change, 80(1-2), 91-107.

Gassman P.W., Secchi, S., Kling, C.L., Jha, M., Kurkalova, L., & Feng, H. (2005). An Analysis of the 2004 Iowa Diffuse Pollution Needs Assessment using SWAT. Proceedings of the SWAT 2005 3rd International Conference, 11-15 July, Zurich, Switzerland.

Figure 4. In-stream nitrogen changes from carbon trading (percentagereduction).

Figure 5. In-stream phosphorous changes from carbon trading (percent-age reduction).


Jha, M., Gassman, P.W., Secchi, S., & Arnold, J.G. (2005). Nonpoint source needs assessment for Iowa part I: Configuration, calibration and validation of SWAT. Watershed Management to Meet Water Quality Standards and Emerging TMDL (Total Maximum Daily Load), Proceedings of the Third Conference 5-9 March 2005 Atlanta, Georgia. American Society of Agricultural Engineers, St. Joseph, Michigan. pp. 511-521.

Mooney, S., Antle, J.M., Capalbo, S.M., & Paustian, K.H. (2004). Influence of project scale and

carbon variability on the costs of measuring soil carbon credits. Environmental Management, 33(S1), S252-S263.

Plantinga, A.J., & Wu, J. (2003). Co-benefits from carbon sequestration in forests: Evaluating reductions in agricultural externalities from an afforestation policy in Wisconsin. Land Economics, 79(1), 74-85.

Secchi, S., & Babcock, B.A. (2007). Impact of corn prices on land set-aside: The case of Iowa. Forthcoming CARD working paper.

Silvia Secchi ([email protected]) isAssociate Scientist, Manoj Jha([email protected]) is Assistant Sci-entist, Hongli Feng ([email protected]) is Associate Scientist, and PhilipGassman ([email protected]) isAssociate Scientist, Center for Agri-cultural and Rural Development,Iowa State University, Ames, IA.Lyubov Kurkalova ([email protected]) is Associate Professor, School ofBusiness and Economics, NorthCarolina A&T State University,Greensboro, NC. Catherine Kling([email protected]) is Professor,Department of Economics and Cen-ter for Agricultural and Rural Devel-opment, Iowa State University, Ames,IA.

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2nd Quarter 2007 • 22(2)


Nitrate Reduction ApproachesBy Christopher Burkart and Manoj K. Jha


As noted in the overview to this set of papers, water qual-ity continues to be a growing concern. Nutrients appliedas commercial fertilizer and manure enter surface andground water, leading to several forms of water qualityimpairment. These impairments manifest themselves in anumber of ways. Excess phosphorus is responsible foralgae blooms, losses in water clarity, and even the presenceof toxic cyanobacteria in fresh water. Excess nitrogen isbelieved to be the limiting factor in low-oxygen dead zonesin several dozen locations around the globe. In somelocales, nitrate concentrations reach levels that are toxic toboth humans and aquatic animals. In the United States,local nitrate concentrations are largely uncontrolled. Theonly widely applied standard affects water used for humanconsumption. This is regulated by the Environmental Pro-tection Agency via National Primary Drinking Water Reg-ulations (EPA NPDWR). Similar requirements and guide-lines exist in Canada and Europe.

Several technologies can remove nitrates directly fromwater and are employed by municipal water works in orderto comply with drinking water standards during periods ofhigh nitrate concentrations in source water. These technol-ogies are costly to operate, suggesting an opportunity forcost savings via upland reductions of fertilizer application.This article explores possible tradeoffs in the context of anutrient-application-right trading scheme. Simulations ofboth water quality and economic effects in a test water-shed suggest that simple upland fertilizer reductions aremore costly than direct nitrate removal if the goal is com-pliance with drinking water standards. Other water qualitygoals merit consideration, but are difficult to model with-out objective standards and given the current nitrateremoval technology.

Watershed Background The area used for simulation is the Raccoon watershed,located in the state of Iowa in the United States. The Rac-coon River is the main stream for the watershed and drainsa large area containing an abundance of fertile soil. Thetotal area of the watershed is approximately 2.3 millionacres, 1.7 million of which are devoted to rotations of cornand soybean production. Nitrogen and phosphorus fertil-izer are applied at high levels on the corn crop and consti-tute the primary nonpoint nutrient pollutant source in thewatershed. Figure 1 shows a land-use map of the water-shed. The outlet of the watershed is near the capital city ofDes Moines, which along with other municipalities in thearea, uses the Raccoon River as a source of drinking water.The Des Moines Water Works is the supplier of drinkingwater and currently operates the world’s largest denitrifica-tion facility.

In-stream nitrate levels frequently exceed the maxi-mum allowed concentration of 10 milligrams per liter. Inthese instances, source water is run through the denitrifi-cation facility before being treated for use as drinkingwater. The facility uses an ion exchange process which pro-duces waste water with a high saline content in addition tothe nitrate removed. This waste water is currently dis-charged downstream at no cost to the facility. Downstreammunicipal water supplies are not adversely impacted bythis discharge, as they are able to meet their water needsfrom deeper ground water aquifers. For purposes ofNPDWR compliance this is not an issue, and the dis-charge is permitted by the EPA under the National Pollut-ant Discharge Elimination System.

The nitrate removal facility was constructed in 1990 ata cost of approximately $3 million. The scrubbers andmedia were the primary components of this large sunkcost, and would also be the bulk of the cost associated with


an expansion of the facility unlessanother removal technology wereemployed. Current processing vol-ume does not appear to requireexpansion in the near term, and therehas been no observed deterioration ofthe scrubber components. Operatingcosts of the facility are approximately$300 per million gallons of water,with a capacity of 10 million gallonsof water per day. In an average year,the facility runs approximately 50days.

Modeling ApproachWhile drinking water standards aregiven high importance due to theirdirect effects on human health, highnitrate levels cause other problems.However, control of ambient water

pollution in this watershed is stillbeing developed, and there are noexisting regulations outside of drink-ing water standards. Amelioratingproblems such as hypoxia and nitratetoxicity for aquatic animals wouldrequire both a lower threshold fornitrates and complete removal of thenitrate from the watershed. Meetingthe latter requirement with the tech-nology currently used for drinkingwater purposes is inappropriate as itreintroduces the nitrate to the envi-ronment. The analysis here proceedsin the framework of existing regula-tions and the technology currently inplace, but it is important to note thatthere are other impacts that meritconsideration: namely, the effects ofnitrate levels outside of drinkingwater considerations. Upland fertil-

izer reductions prevent nitrates fromentering waterways in the first place,and have positive effects beyond con-tributing to drinking water standardcompliance.

The goal of the modeling frame-work is to capture changes in waterquality generated by implementationof policy, as well as the associatedeconomic effects. This requires thecoupling of an economic model witha physical model. Nutrient applica-tion levels predicted by the economicmodel are used to supply land-useinputs to the physical model. Theoutput from the physical model inturn provides the water quality mea-sure of interest: nitrate concentra-tion over time. A hydrologic model isused to link the effects of upland fer-tilizer reductions to direct nitrateremoval at the outlet. The watershed-based Soil and Water AssessmentTool (SWAT) simulates the effects ofwatershed management on waterquality and water flow on a dailytime step. It is primarily used formodeling nonpoint source contribu-tions to nutrient and sediment loadswithin a watershed. The SWATimplementation employed uses datafrom the National Resources Inven-tory (NRI) to populate the watershedwith spatially detailed information. Apoint in the NRI effectively repre-sents a farm. Site-specific nutrientapplication data are generated by theeconomic model. The economicmodel predicts nitrogen fertilizerapplication rates based on prices ofcorn and fertilizer and a site-specificsoil characteristic. It also predictsyield, and thus returns to fertilizerapplication. Changes in nitrogen fer-tilizer prices, for example, via a tax onfertilizer or a cap on application, willcause a loss in returns for the farmer.This provides a measure of the costimposed by the policy. Data used toconstruct the model comes from a

Figure 1. Land use in the Raccoon watershed.


farm operator survey, the Agricul-tural Resource Management Survey,historical prices, and from a detailedsoil grid.

Policy Simulations Three scenarios are run through themodeling system described above.One is a baseline in which the eco-nomic model leaves prices and nitro-gen fertilizer applications unchanged,and the water quality model predictsthe associated nitrate concentrationsat the watershed outlet. The othertwo scenarios represent reductions infertilizer applications simulated bythe imposition of a nonpoint sourcetrading scheme. This scheme worksas follows: each farm is allocated fer-tilizer application permits for thetotal acreage it farms; for example, a100-acre farm might receive 12,000pounds worth of permits if the per-mit level is 120 pounds per acre. Afarm has three choices in using its

permits. One is to apply exactly thepermitted amount. Another is toapply less than permitted, and sellthe surplus permits to the thirdgroup, those who purchase permits inorder to apply at greater levels thaninitially permitted. Farmers maketheir choice of total applicationaccording to the model, taking intoaccount the prices they face, their soiltype, and the market price of a per-mit, which is determined by the dis-tribution of farmer types. The totalwatershed application is reduced aslong as the total permit allocation issmaller than the total amount origi-nally applied. For purposes of simula-tions, this is done at two levels of per-mit allocations. From a baselineaverage application rate of 135pounds per acre, one scenariorestricts the per-acre permit alloca-tion to approximately 120 poundsper acre and results in a simulated6% reduction in annual load ofnitrate at the watershed outlet. The

other restricts the allocation toapproximately 108 pounds per acreand results in an approximate 12%reduction in annual nitrate load.These reductions are the result of thetotal mass of nitrogen being appliedin the watershed being reduced.

Imposing the permit restrictionsbenefits those farmers who can sellexcess permits, but increases the costsof those who must purchase addi-tional permits. Since the totalamount of nitrogen application isbeing reduced, the net result is a lossfor farmers in the watershed as awhole. Loss or gain from the policyscenarios can be measured for indi-vidual farms and then aggregated tothe watershed level to gauge the costof the policy. Under the small reduc-tions, the total farm watershed loss isapproximately $161,000, and underthe larger reductions, losses areapproximately $700,000.

To compare the water qualitychanges resulting from the imple-

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Figure 2. Nitrate loads by monthly average.


mentation of these policies to opera-tion of the nitrate removal facility,the water-quality model is run on adaily time step and the nitrate con-centration for each day recorded. Thetrigger concentration for the nitrateremoval facility to run is 9mg/L (thelegal limit is 10mg/L). Under thebaseline scenario, that level wasexceeded 56 days of the year. Thesmall and large reduction scenariosreduced the number of run-days to51 and 48. Figure 2 shows a sum-mary of nitrate loads by monthlyaverage. Saving days of operation forthe nitrate removal facility impliescost savings and illustrates the short-ening in the number of run-daysrequired to maintain a safe level ofnitrate. The energy, labor, and rawmaterial costs of one run-day areapproximately $3,000. The lifetimeof the media used in the removal pro-cess is currently uncertain, making itdifficult to calculate the true cost ofoperation. The original media is stillin use and shows no sign of deteriora-tion after 14 years of use. As nitrateloads and water demand grow, theremay be a need for expansion in thefuture, involving significant capitalcosts and raising the cost of a day ofoperation. Such expansion may alsoinvolve a change in nitrate removaltechnology.

Trading nitrogen permitsbetween point and nonpoint sourcescan lower costs of reductions (Ran-dall & Taylor, 2000). This is usuallyconsidered in the context of a non-point source generating excess per-mits by purchasing upland reduc-tions. In that type of tradingarrangement, a trading ratio is estab-lished to equilibrate a pound ofupland reduction to a pound of pointsource discharge. Conceptually, thisapproach could work in reverse aswell: nonpoint sources could gener-ate permits for themselves by paying

for the removal system. While thesetrading opportunities are attractivepossibilities, a quick look at the dif-ference in costs in this case suggeststhat it would be much more efficientto simply run the nitrate removalfacility a few extra days rather thanimplement any restrictions on farmerapplication. Five run-days at $3,000per day is $15,000, far less than the$160,000 in losses that would beincurred by farmers. Eight run-daysof the nitrate removal facility are like-wise much less expensive than the$700,000 in losses associated withthe stricter cap-and-trade policy.

While the upland fertilizer reduc-tions examined here are more costlythan direct nitrate removal, this anal-ysis does not take into account otherpossibilities. There are also concernsbeyond drinking water standards,such as hypoxia and low-level nitratetoxicity (Camargo et al., 2005), thathave important impacts on ambientwater quality. Perhaps because drink-ing water issues pose the most imme-diate threat to human health, it is theonly form of existing pollution regu-lation that impacts this watershed. Asnew standards with broader impactsin mind are developed, such as TotalMaximum Daily Loads, this analysiscan be revisited, possibly with differ-ent conclusions. The upland reduc-tions have an effect on the ambientand downstream nitrate loads thatthe removal process does not andwould be more effective at meetingexpanded standards. Even if undermore comprehensive standardsupland reductions become more costeffective, there would be transactioncosts involved in any trading schemethat would also need to be consid-ered.

There are also combinations ofreduction strategies that could resultin superior reductions with similarcosts, even in the existing framework.

Coupling reductions with bufferstrips, grassed waterways, changes intillage, and application timing all cancontribute to reductions in nutrientloads to a watershed. In addition, amore complete comparison wouldrequire information on possible dete-rioration of the nitrate removalmedia and the associated replacementcosts, though these are at presentuncertain.

For More Information Atwood, J., McCarl, B., Chen, C.,

Edelman, B., Nayda, B., & Srinivasan, R. (2000). Assessing Regional Impacts of Change: Linking Economic and Environmental Models. Agricultural Systems, 63, 147-59.

Camargo, J., Alonso, A., & Salamanca, A. (2004). Nitrate toxicity to aquatic animals: A review with new data for freshwater invertebrates. Chemosphere, 58, 1255-67.

Jones, C. (2005). Nitrate and bacteria in the Raccoon River: Historical perspective and 2004 Summary, Des Moines Water Works. Available online: http://www.dmww.com/Laboratory/.

Nusser, S., & Goebel, J. The National Resources Inventory: a long-term multi-resource monitoring programme. Environmental and Ecological Statistics, 4(3), 181-204.

Randall, A., & Taylor, M. (2000). Incentive-based solutions to agricultural environmental problems: Recent developments in theory and practice. Journal of Agricultural and Applied Economics, 32, 221-34.

United States Environmental Protection Agency, National Primary Drinking Water Regulations (EPA NPDWR).


Current Drinking Water Rules. Available online: http://www.epa.gov/safewater/regs.html.

United States Geological Survey, National Water-Quality Assessment (USGS NAWQA). Nutrients in streams and rivers

across the Nation–1992-2001 National Water-Quality Assessment Program. Available online: http://water.usgs.gov/nawqa/.

Christopher Burkart ([email protected]) is Assistant Professor, Depart-ment of Marketing and Economics,University of West Florida, Pensa-cola, FL. Manoj K. Jha([email protected]) is Assistant Sci-entist, Center for Agricultural andRural Development, Iowa State Uni-versity, Ames, IA.



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