1205

Artifi cial subsurface drainage is commonly used in midwestern agriculture and drainage losses of dissolved organic carbon (DOC) from such systems are an under-quantifi ed portion of the terrestrial carbon (C) cycle. Th e objectives of this study were to determine the eff ect of common agricultural management practices on DOC losses from subsurface tile drains and to assess patterns of loss as a function of year, time of year, and drainfl ow. Daily drainfl ow was collected across six water years (1999–2004) from a restored prairie grass system and cropping systems which include continuous corn (Zea mays L.) and corn-soybean [Glycine max (L.) Merr.] rotations fertilized with urea-ammonium-nitrate (UAN) or swine (Sus scrofa) manure lagoon effl uent. Th e DOC concentrations in tile drainfl ow were low, typically <2 mg L–1. Yearly DOC losses, which ranged from 1.78 to 8.61 kg ha–1, were not aff ected by management practices and were small compared to organic C inputs. Spring application of lagoon effl uent increased yearly fl ow-weighted (FW)-DOC concentrations relative to other cropping systems in three of the years and increased monthly FW-DOC concentrations when drainfl ow occurred within 1 mo of application. Drainfl ow was signifi cantly and positively correlated with DOC loss. Drainfl ow also aff ected DOC concentrations as greater 6-yr cumulative drainfl ow was associated with lower 6-yr FW-DOC concentrations and greater daily drainfl ow was associated with higher daily DOC concentrations. Our results indicate that lagoon effl uent application and fertilizer N rates do not aff ect long-term losses of DOC from tile drains and that drainfl ow is the main driver of DOC losses.

Dissolved Organic Carbon Losses from Tile Drained Agroecosystems

Matthew D. Ruark* University of Wisconsin–Madison

Sylvie M. Brouder and Ronald F. Turco Purdue University

Artificial subsurface tile drainage is a management practice for

improving crop production in poorly drained soils. Roughly 37%

of agricultural lands in the upper Midwest are tile drained, altering

the hydrologic pathways in these landscapes (Fausey et al., 1995).

Subsurface drainage decreases runoff losses of sediment, nutrients,

and chemicals, but increases infi ltration rates and leaching losses of

nutrients and chemicals (Skaggs et al., 1994). Tile drainage has been

shown to aff ect the cycling of organic C in managed agroecosystems

by altering the dominant water loss pathways (McTiernan et al.,

2001; Jacinthe et al., 2001). While many recent studies have focused

on tile drainage losses of nitrate (e.g., Huggins et al., 2001; Bakhsh

et al., 2002, 2005; Jaynes et al., 2001; Randall et al., 2003; Randall

and Vetsch, 2005; Kladivko et al., 1999, 2004) and pesticides (e.g.,

Kladivko et al., 1999, 2001), there is considerably less information

regarding losses of DOC. At the watershed-level, surface water fl ux

of DOC ranges from 10 to 100 kg ha–1 yr–1 (as reviewed by Hope

et al., 1994), with DOC fl uxes from small agricultural watersheds

in Indiana typically being <20 kg ha–1 yr–1 (Dalzell et al., 2007).

Th us, in watersheds dominated by tile-drained agriculture, it is

possible that tile drainage is an important contributor to the overall

DOC fl ux from the watershed, although the relative contribution

of tile drains to watershed scale DOC fl ux has not been quantifi ed

(Chantigny, 2003; McDowell, 2003).

Leaching losses of DOC from surface soils are usually small

compared to soil organic carbon (SOC) content or other C loss

pathways such as grain removal or gas fl ux (Brye et al., 2002). Mc-

Carty and Bremner (1992) reported that DOC concentrations

ranged between <0.3 to 2.9 mg L–1 from tile drains in central Iowa

and Beauchemin et al. (2003) reported average DOC concentra-

tions between 1.58 to 6.03 mg L–1 in tile drains in Quebec, Canada,

but DOC mass losses were not quantifi ed. Kovacic et al. (2000)

reported that average DOC concentrations in tile drainage ranged

between 2.6 and 3.6 mg L–1 in eastern Illinois, Owens et al. (2002)

determined that DOC concentrations in subsurface soil leachate

water (2.4 m depth) ranged from 0.5 to 3.2 mg L–1 in eastern Ohio,

and Brye et al. (2001) reported typical DOC concentrations be-

tween 5 and 20 mg L–1 in subsurface soil leachate (1.4 m depth)

in central Wisconsin. Th e DOC concentrations in Kovacic et al.

(2000), Owens et al. (2002), and Brye et al. (2001) translate to an-

Abbreviations: C, carbon; DOC, dissolved organic carbon; FW, fl ow-weighted; LOI, loss

on ignition; N, nitrogen; SOC, soil organic carbon; TDE, tile drain effi ciency; UAN, urea-

ammonium-nitrate; WQFS, water quality fi eld station.

M.D. Ruark, Dep. of Soil Science, Univ. of Wisconsin-Madison, 1525 Observatory Dr.,

Madison, WI, 53706. S.M. Brouder and R.F. Turco, Dep. of Agronomy, Purdue Univ., 915

W. State St., West Lafayette, IN 47907.

Copyright © 2009 by the American Society of Agronomy, Crop Science

Society of America, and Soil Science Society of America. All rights

reserved. No part of this periodical may be reproduced or transmitted

in any form or by any means, electronic or mechanical, including pho-

tocopying, recording, or any information storage and retrieval system,

without permission in writing from the publisher.

Published in J. Environ. Qual. 38:1205–1215 (2009).

doi:10.2134/jeq2008.0121

Received 6 Mar. 2008.

*Corresponding author (mdruark@wisc.edu).

© ASA, CSSA, SSSA

677 S. Segoe Rd., Madison, WI 53711 USA

TECHNICAL REPORTS: SURFACE WATER QUALITY

1206 Journal of Environmental Quality • Volume 38 • May–June 2009

nual DOC fl uxes between 2 and 48 kg ha–1 from surface soils.

Th ese DOC quantities may seem minimal in terms of system-

level mass balance, but the movement of DOC through the soil

profi le plays important roles in soil chemical and biological pro-

cesses (as reviewed by Kalbitz et al., 2000). Mineral weather-

ing (Raulund-Rasmussen et al., 1998), transport of metals and

organic pollutants (Temminghoff et al., 1997; Totsche et al.,

1997), and microbial activity in subsurface soils (Neff and As-

ner, 2001) can be infl uenced by leaching of DOC. In subsurface

soils, DOC is typically the limiting nutrient for denitrifi cation

and input of fresh DOC into this system will stimulate deni-

trifi cation (Sotomayor and Rice, 1996; Yeomans et al., 1992).

Furthermore, DOC plays an important role in the survivorship

of bacterial pathogens in soils (Jamieson et al., 2002) and seques-

tration of C in subsurface soil (Lorenz and Lal, 2005).

Vegetation type and amount of residue returned to the soil are

major determinants of surface soil DOC concentrations (often

determined as water-extractable organic C). In general, surface

soil DOC concentrations increase with application of organic

amendments such as crop residues and manures (Chantigny,

2003 and others therein). Siemens et al. (2003) reported DOC

concentrations between 15 and 25 mg L–1 and fl uxes between 63

and 87 kg ha–1 yr–1 at 90 cm of depth in soils where fertilizer was

provided largely as manure. Additions of manure can also increase

DOC losses and concentrations drainage waters (Bol et al., 1999;

Royer et al., 2007). In situ DOC concentrations may also be af-

fected by crop rotation as diff erences in water-extractable organic

C have been observed between legumes and gramineae crops

(Mazzarino et al., 1993; Campbell et al., 1999). However, several

studies have documented that surface soil DOC is rapidly decom-

posed or sorbed in soils and may not reach deep soil layers (Angers

et al., 2006; Franchini et al., 2001; McCarty and Bremner, 1992).

Likewise, there is some evidence that rate and source of nitrogen

(N) fertilizer impact DOC but, again, these impacts may be tran-

sient (Clay et al., 1995). At present, it remains unknown whether

management-related diff erences in in situ DOC concentrations

are suffi cient in magnitude and duration to result in diff erences

in either total quantity or varying temporal patterns of DOC loss

in drainage water. Th e objectives of this study were to determine

the eff ect of common agricultural management practices on DOC

losses from subsurface tile drains and to assess patterns of loss as a

function of year, time of year, and drainfl ow.

Materials and Methods

Site Description and Crop ManagementA 6-yr (1 Oct. 1998–31 Sept. 2004) fi eld experiment was

conducted at Purdue University’s Water Quality Field Station

(WQFS). Soils at the WQFS are predominantly Drummer silty

clay loam (fi ne-silty, mixed, superactive, mesic Typic Endoaquoll)

with a small area (<2%) of Raub silty clay loam (fi ne-silty, mixed,

superactive, mesic Aquic Argiudoll) with slopes ranging from 0

to 2%. Th e WQFS is arranged in a randomized complete block

design, with four blocks and 12 treatment plots per block. Eleven

treatments of common cropping systems and one treatment of

a perennial, mixed-vegetation, restored prairie grass (PG) system

dominated by big bluestem (Andropogon gerardii L.) were ran-

domly assigned within each block (Table 1). Th e cropping system

treatments encompassed factors of crop rotation [continuous corn

(CC), corn following soybean (CB), or soybean following corn

(BC)], source and timing of fertilizer application [spring-applied

or sidedressed UAN and spring- or fall-applied swine lagoon effl u-

ent (SLE and FLE, respectively)], and UAN fertilizer rate [low (L),

medium (M), and high (H)]. Low rates of UAN were applied as a

sidedress and medium and high rates were applied preplant. Sid-

edress applications of UAN were applied at a depth of 10 cm and

occurred between 5 June and 8 July. Preplant UAN was applied at

a depth of 15 to 20 cm between 14 April and 8 May all years ex-

cept 1999 (21 June) and 2002 (1 June). Th e medium N rates were

Purdue University recommendations for preplant applications to

continuous and rotated corn (179 kg ha–1 and 157 kg ha–1, respec-

tively) based on yield goal (Vitosh et al., 1995). Th e high and low

rates of N application correspond to 22 kg ha–1 above and below

Table 1. Description of agricultural management treatments, average yearly yield, and average soil organic C content.

Crop Code Previous crop Method Source Nitrogen rate Yield† SOC‡

—————kg ha–1————— g kg–1

Prairie grass PG 23.0ab

Soybean BCL Corn 3090 20.7ab

Soybean BCM Corn 2960 22.0ab

Soybean BCH Corn 3250 22.1ab

Corn CBL Soybean Sidedress UAN§ 135 8730bcd 20.6b

Corn CBM Soybean Preplant UAN 157 9170abc 22.2ab

Corn CBH Soybean Preplant UAN 179 9630a 22.0ab

Corn CCL Corn Preplant UAN 157 8530cd 20.9ab

Corn CCM Corn Preplant UAN 179 8670bcd 23.2ab

Corn CCH Corn Preplant UAN 201 9390ab 24.8a

Corn SLE Corn Spring injection Effl uent¶ 190 to 210 8230d 20.9ab

Corn FLE Corn Fall injection Effl uent 230 to 260 8140d 23.2ab

† 6-yr average corn and soybean yields (1998–2003). Corn yields followed by diff erent lowercase letters are signifi cantly diff erent (P ≤ 0.05).

‡ SOC, average soil organic carbon content (collected in the fall of 1999, 2000, 2003, and 2004). The SOC contents followed by diff erent lowercase letters

are signifi cantly diff erent (P ≤ 0.05).

§UAN, 28% urea-ammonium-nitrate.

¶ Effl uent, swine manure lagoon effl uent.

Ruark et al.: Dissolved Organic C Losses from Tile Drained Agroecosystems 1207

the Purdue University recommended rates, respectively. Lagoon

effl uent was injected at a depth of 15 to 20 cm. Fall lagoon effl u-

ent applications occurred between 26 October and 5 November

in every year and SLE applications occurred between 1 and 16

April, except in 2002 when excessive spring rainfall delayed ap-

plication until 1 June. Lagoon effl uent was applied at a rate of

72,500 L ha–1. Th is rate applied between 190 and 262 kg ha–1 of

total N (determined as Kjeldahl-N; Bremner, 1996) and between

300 and 630 kg ha–1 of total organic C [determined by loss on

ignition (LOI); Nelson and Sommers, 1996]. Analysis of total N

and total organic C was conducted by A&L Great Lakes Laborato-

ries Inc. (Fort Wayne, IN) on lagoon effl uent applied in the spring

and fall of 2003. Liquid starter fertilizer containing N and P was

applied with all corn plantings at a rate of 22 kg ha–1 N and 8 kg

ha–1 P placed 5 cm to the side and 5 cm below the seed.

For treatments involving corn-soybean rotations, eight plots

were assigned to each treatment allowing both crop species to

be present each experimental year. Corn was planted in 76 cm

row spacings and soybean was drill-seeded at 20 cm spacings,

in a north-south direction parallel to drainage tiles. In all years,

Pioneer corn seed variety 34G81 was planted at a rate of 66,720

seeds ha–1 and Pioneer soybean variety 93B45 was planted at a

rate of 518,700 seeds ha–1. All cropping system treatments re-

ceived insecticide application at planting, broadcast application

of herbicide after planting, and were disked and cultivated in the

spring. Additionally, all cropping system treatments were chisel-

plowed in the fall with the exception of FLE. To control plant

species composition, prairie grass plots were burned in 2000,

2001, 2002, and 2004 between 2 April and 1 May.

Daily rainfall totals were collected at the Purdue Agronomy

Center for Research and Education Weather Station, which was

located 3 km from the WQFS, and compiled by the Indiana State

Climate Offi ce. Soil samples (0–20 cm) were collected after har-

vest and before lagoon effl uent application in 1999, 2000, 2003,

and 2004 and analyzed by A&L Great Lakes Laboratories Inc.

(Fort Wayne, IN) for SOC content by LOI (Nelson and Som-

mers, 1996). Corn was harvested with a four-row harvester (White

3700) from the middle four rows (of 12) in each corn plot (3.1 by

48 m); soybean was harvested from a 3.0 by 48 m area from the

center of each plot using a John Deere 3300. Corn and soybean

yields were determined by weighing collected grain in a Parker

weigh wagon and analyzing subsamples for moisture with a Dick-

ey John GAC 3000 moisture tester (Dickey John Corp., Aurora,

IL). Corn and soybean yields were standardized to 15.5 and 13%

moisture content, respectively. Total aboveground crop biomass at

maturity was used to estimate crop residue return using a harvest

index (dry grain weight per total dry grain and biomass weight) of

0.53 for corn and 0.46 for soybean (Johnson et al., 2006).

Field Lysimeters and Drain Water CollectionAs described by Eigel et al. (1992), within every treatment plot

(10.8 by 48 m), a drainage lysimeter (10.8 by 24.4 m) was con-

structed as a bottomless clay box to allow the collection of drain-

fl ow from a hydrologically isolated area of soil. Bentonite slurry

was used to construct the walls of the clay box to a depth of 1.5 m.

Two plastic agricultural drain tiles (collection drain and compan-

ion drain), 0.1 m in diameter, were installed down the length of

every treatment plot at a depth of 0.9 m. Th e collection drain was

perforated within the lysimeter area and was nonperforated under

areas outside of the drainage lysimeter. Th e companion drain was

placed next to the collection drain and was nonperforated within

the lysimeter area but perforated outside the lysimeter area. Th e

paired drain tiles maintained similar soil water contents across

the entire plot area, simulating 10 m tile spacing. Th e companion

tiles drained into a nearby drainage ditch while the collection tiles

drained into eight instrument huts (six tile drains per hut). In each

instrument hut, stainless steel tipping buckets were positioned at

the end of the collection drains to measure hourly fl ow volumes.

Each tipping bucket was fi tted with a magnetic sensor switch

(Honeywell SR3C-A1, Sacramento, CA) to count the number of

tips. Tip counts were collected using a Campbell CR10 data logger

(Campbell Scientifi c, Logan, UT) and summarized every hour. A

collection bucket located next to each tipping bucket collected

about 10 mL of every other tip, up to 20 L of cumulative sample.

On days when drainfl ow occurred, fl ow-proportional subsamples

were manually retrieved from collection buckets; the remainder of

the sample was discarded.

Water AnalysisDrain water subsamples were recovered from the fi eld, fi l-

tered (Whatman #2) to remove particulates and stored frozen

(–4°C) until DOC analysis. Th e Whatman #2 fi lter (pore size

~8 μm) had been compared to 0.45 μm fi lters and no diff erences

in DOC concentrations were observed. Whatman #2 fi lters were

prerinsed with 100 mL of DI water to ensure no contamination

of DOC from the fi lter. Th e DOC concentrations were mea-

sured as nonpurgeable organic C using a Shimadzu TOC-VCSH

Total Organic Carbon Analyzer (Shimadzu Corp., Kyoto, Japan)

using a high sensitivity catalyst (platinum on quartz wool). For

all subsamples, DOC concentrations below the instrument de-

tection limit (0.05 mg L–1) were assigned values of 0.025 mg L–1,

halfway between zero and the detection limit.

Calculations and Statistical AnalysisTo best correspond with the daily subsample collection, hour-

ly drainfl ows were aggregated from noon to noon. Daily DOC

losses were calculated as the product of daily drainfl ow (L) and

daily DOC concentration (mg L–1). If a daily subsample was not

able to be collected on a day with drainfl ow because of fl ooding or

equipment malfunction, the DOC concentration from the closest

daily subsample collected within 5 d was used; if there were no

subsamples collected within 5 d, then the average yearly DOC

concentration for that drainage plot was used. To delineate 12-mo

periods on the natural water cycles, yearly data was calculated and

presented on a water year (1 October–30 September) basis. How-

ever, seasonal periods were delineated from the calendar year as:

winter (21 December–19 March), spring (20 March–20 June),

summer (21 June–21 September), and autumn (22 September–20

December). Daily drainfl ow and DOC losses were summed for

each individual tile across each of the six water years (1999–2004)

and on a yearly, seasonal, and monthly basis. Flow-weighted DOC

concentrations were calculated as the total loss divided by the total

1208 Journal of Environmental Quality • Volume 38 • May–June 2009

drainfl ow per plot on a 6-yr, yearly, seasonal, and monthly basis.

Tile drainage effi ciency (TDE) was estimated each year by divid-

ing average yearly drainfl ow per tile (mm) by the yearly rainfall.

Data loggers failed to measure hourly tips 13 times between

2000 and 2003, resulting in unmeasured hourly or daily drainfl ow.

Separate strategies were used to estimate the unmeasured hourly or

the unmeasured daily drainfl ow. For a given plot with unmeasured

hourly fl ow in a discrete drainage event, the observations that pre-

ceded and followed the break in the data record were fi t with an

interpolating, cubic spline curve (data function: CALL SPLINE)

in SAS (SAS Institute, 1999). Th e resulting function was used to

estimate the unmeasured hourly data. When a given drainage plot

had an entire day of unmeasured hourly data, a linear regression

model (Proc. REG) was determined using the daily drainfl ow and

the 6-yr cumulative drainfl ow of all drainage plots for which daily

data was collected. Th e resulting regression model was then used

to estimate unmeasured daily drainfl ow from each plot’s 6-yr cu-

mulative drainfl ow. For any day when six or fewer tiles had mea-

surable fl ow, unmeasured daily drainfl ow was assumed to be zero.

Two discrete drainage events (22 Jan. 1999 and 11 June 2004) had

unmeasured drainfl ow for all drainage plots and no estimations

could be performed. Th e unmeasured drainage event in 2004 was

believed to represent a large portion of the 2004 total drainfl ow

(based on the daily rainfall totals); subsequently, 2004 data were

excluded from some analyses. Overall, 20% of the 6-yr cumulative

drainfl ow was determined by using estimation strategies (spline

and regression methods) for each tile.

All statistical evaluations were conducted using SAS. Yearly and

6-yr cumulative drainfl ow and daily DOC concentrations were

analyzed using univariate statistics (Proc. UNIVARIATE). Agri-

cultural management treatment eff ects on 6-yr and yearly drain-

fl ow, DOC loss, and FW-DOC concentrations were analyzed

with ANOVA using the general linear model (Proc. GLM), with

ANOVA conducted separately for each year. Treatment eff ects on

average SOC content, average corn yield, and average soybean

yield were analyzed with ANOVA using Proc. GLM; eight and

three treatments were used in the ANOVA for average corn and

soybean yield, respectively. Tukey’s honest signifi cant diff erence

(HSD) was used to determine treatment diff erences for drainfl ow,

DOC loss, FW-DOC concentration, crop yield and SOC content.

Signifi cance was determined at the α = 0.05 level for all ANOVA

and Tukey’s HSD results, unless otherwise noted. Departures from

normality and nonconstant variance were evident with drainfl ow

and DOC losses. It was determined that neither transformation of

the data nor use of weighted least squares method aff ected the sig-

nifi cance of ANOVA results, therefore untransformed data were

reported. Preselected contrasts were conducted on yearly FW-

DOC concentrations (CONTRAST, Proc. GLM) to compare

sets of treatments. Th e set of continuous corn and corn-soybean

rotations receiving UAN (CCL, CCM, CCH, CBL, CBM, CBH,

BCL, BCM, BCH) was compared to each treatment not receiving

UAN (PG, SLE, or FLE). To evaluate eff ects of crop rotation, the

set of CC treatments (CCL, CCM, CCH) was compared to the

set of CB treatments (CBL, CBM, CBH), and the set of CB treat-

ments was compared to the set of BC treatments (BCL, BCM,

BCH). Likewise, the set of CC treatments was compared to the

set of all CB and BC treatments. Bonferroni adjustments were

applied to each signifi cance level (α = 0.05, 0.01, and 0.001) to

evaluate the statistical signifi cance of the contrast results. Correla-

tion analysis (Proc. CORR) was conducted among the following

variables per individual tile drainage plot: 6-yr cumulative drain-

fl ow, 6-yr cumulative DOC loss, 6-yr FW-DOC concentration,

6-yr average daily DOC concentration, average SOC content, and

estimated 6-yr cumulative crop residue. If drainage plots did not

have any yearly, seasonal, or monthly drainfl ow, the corresponding

FW-DOC concentrations was treated as missing data.

Regression analysis was used to characterize relationships

between drainfl ow and DOC fl ux. Th e drainfl ow-DOC fl ux

relationship was determined across two time scales: 6-yr and

daily. Th e relationship between 6-yr cumulative drainfl ow and

6-yr DOC fl ux was determined across all 48 tile drains, while

the relationship between daily drainfl ow and daily DOC fl ux

was determined among all tiles and all days with concentra-

tions greater than the minimum detection limit (0.5 mg L–1)

and fl ow rates >0.1 mm d–1. For each time scale, both variables

were log-transformed and fi t with a linear regression model

(SAS Proc. REG). Th e resulting expression was converted to

the simplifi ed expression based on non log-transformed data:

L = aQb [1]

where L = solute fl ux and Q = drainage fl ow rate. Th e exponent b

in Eq. [1] represents the slope in the linear relationship between

the log-transformed values of drainfl ow and DOC fl ux. Ninety-

fi ve percent confi dence intervals of the slope (b) were determined

with Proc. REG to test whether b was signifi cantly diff erent from

1. Slope values less than, greater than, and equal to one indicate

solute concentrations are decreasing with, increasing with, and

unaff ected by increasing drainfl ow, respectively.

ResultsYearly corn yields, averaged across all treatments, ranged from

7660 kg ha–1 (in 1998) to 9790 kg ha–1 (in 1999). Among crop

management treatments, CBH, CCH, and CBM had the highest

6-yr average corn yields (9630, 9390, and 9170 kg ha–1, respec-

tively) and were signifi cantly greater than the two lowest yielding

corn plots, SLE and FLE (8230 and 8140 kg ha–1, respectively)

(Table 1). Yearly soybean yields, averaged across all treatments,

ranged from 1310 kg ha–1 (in 1999) to 4090 kg ha–1 (in 2002).

Analysis of variance determined no eff ect of crop management

on 6-yr average soybean yield. Th e SOC content exhibited rela-

tively large variation within each management treatment, with

coeffi cients of variation (CV) ranging between 9.2 to 19.0%.

Th ere was a signifi cant eff ect of cropping system on SOC con-

tent, but the only signifi cant diff erence was between the treat-

ment with the greatest SOC content (CCL) and the treatment

with the lowest SOC content (CBL) (Table 1).

Drainfl ow and RainfallTh e 6-yr cumulative and yearly drainfl ows were not signifi cantly

diff erent among the experimental blocks or among agricultural

management treatments. Univariate statistics of 6-yr cumulative

Ruark et al.: Dissolved Organic C Losses from Tile Drained Agroecosystems 1209

drainfl ow exhibited large tile-to-tile variability (CV = 54%). An-

nual drainfl ow, across all drainage plots, was right skewed in 1999

through 2002, while 2003 and 2004 did not display severe skew-

ness. When averaged across all drainage plots, drainfl ow varied sub-

stantially among years; 2002 had nearly twice the drainfl ow as 1999

and more than three times the drainfl ow as 2003. Th e year-to-year

variability can be attributed to diff erences in rainfall patterns (Fig.

1) and total yearly rainfall (Table 2), although the average yearly

drainfl ow (1999–2003) was not signifi cantly correlated with year-

ly rainfall (r = 0.39). Yearly rainfall amounts ranged from 807 to

1131 mm, with a 6-yr average (966 mm) similar to the 30-yr aver-

age of 946 mm (1971–2000, Indiana State Climate Offi ce).

Year-to-year variations in rainfall and drainfl ow patterns were

evident when data were analyzed on a monthly time-step (Fig. 1).

For months when drainfl ow occurred, monthly rainfall and aver-

age monthly drainfl ow (calculated across all drainage plots) were

signifi cantly correlated, although the relationship was weak (r =

0.47). Mean tile drain effi ciency, calculated from 6-yr cumulative

drainfl ow and rainfall, was 21% (Table 2). However, yearly TDE

varied between 13 and 37%, with 2002 having a much greater

TDE than any other year. For some drainage plots, monthly TDEs

were >100% (average monthly drainfl ow + 1 standard deviation >

monthly rainfall; Fig. 1), which may be attributed to snowmelt or

drainfl ow originating from the previous month’s rainfall.

Dissolved Organic Carbon: Concentration,

Loss, Drainfl ow-Flux RelationshipsTh roughout the 6-yr experiment, a total of 8789 DOC samples

were collected; the number of samples per agricultural manage-

ment treatment ranged from 565 to 792. Collected daily DOC

concentrations ranged from below detection limit (<0.05 mg L–1)

to 22.6 mg L–1; 53 samples had DOC concentrations below the

instrument detection limit (0.05 mg L–1), 44 samples had DOC

concentrations >5 mg L–1, and seven samples had DOC concen-

trations >10 mg L–1. Th e population of daily DOC concentration

observations had a mean of 1.48 mg L–1, a median of 1.29 mg L–1

and an interquartile range of 0.88 to 1.88 mg L–1

Six-year cumulative DOC loss averaged 20.9 kg ha–1 across

all agricultural management treatments, with average yearly

losses ranging between 1.78 and 8.61 kg ha–1 (Table 3). Yearly

DOC losses were quite variable across replicates within treat-

ments, resulting in large standard deviations for yearly means.

Th e 6-yr cumulative and yearly DOC losses were not signifi -

cantly diff erent among experimental blocks or treatments. Th e

6-yr FW-DOC concentration, calculated from 6-yr DOC loss

and cumulative drainfl ow and averaged across all treatments,

was 1.84 mg L–1, with average yearly FW-DOC concentra-

tions ranging from 1.45 to 2.16 mg L–1 (Table 4). A signifi cant

treatment eff ect on FW-DOC concentration was observed in

Fig. 1. Monthly rainfall and average monthly drainfl ow during six water years (1999–2004). Average monthly drainfl ow (determined across all drainage plots) was reported when >0.1 mm. Error bars represent standard deviation.

Table 2. Descriptive statistics for yearly and 6-yr total subsurface drainfl ow from individual tile drains. Yearly and 6-yr rainfall totals and drainage effi ciencies are also reported.

Drainfl ow

Water year

Statistic 1999 2000 2001 2002 2003 2004 Total†

———————–––—————————————mm—————–––———————————————

Minimum 29 11 16 56 0 0 175

Maximum 648 771 873 2192 315 234 3903

Mean 226 148 176 423 126 96 1195

Median 219 120 136 393 133 102 1124

Interquartile range 145 71 95 178 77 66 665

Standard deviation 125 127 155 307 64 52 643

Skewness 0.89 3.0 2.7 4.2 0.29 0.23 1.6

Kurtosis 1.6 12 9.4 24 0.54 0.07 5.8

CV, % 55 86 88 72 51 55 54

Rainfall, mm 894 975 807 1131 998 989 5794

TDE‡, % 26 15 22 37 13 na§ 21

† Total, cumulative drainfl ow between 1 Oct. 1998 and 30 Sept. 2004.

‡ TDE, tile drainage effi ciency (mean drainfl ow/rainfall × 100).

§ na, data not available because of large periods of missing drainfl ow.

1210 Journal of Environmental Quality • Volume 38 • May–June 2009

1999 (P ≤ 0.05) and 2003 (P ≤ 0.1). In 1999, SLE had a sig-

nifi cantly greater FW-DOC concentration compared to CCH,

PG, FLE, and CBL (Tukey’s HSD). Th e 6-yr and yearly FW-

DOC concentrations were not signifi cantly diff erent between

prairie grass and crop rotations receiving UAN (Table 4).

Spring application of lagoon effl uent had a greater average 6-yr

FW-DOC concentration compared to crop rotations receiving

UAN (2.49 vs. 1.80 mg L–1); within years, contrast results for

this treatment comparison were signifi cant in 1999 (2.32 vs.

1.39 mg L–1), 2002 (2.84 vs. 2.09 mg L–1), and 2003 (2.50 vs.

1.51 mg L–1). No diff erences between fall application of lagoon

effl uent and crop rotations receiving UAN were found. Crop

rotation did not aff ect 6-yr or yearly FW-DOC concentrations,

but in 2003, within corn-soybean rotations, plots in soybean

had greater yearly FW-DOC concentration than plots in corn.

Th is diff erence may be attributed to the carryover eff ect of

more corn residue being returned to the soil at the beginning

of the water year compared to soybean residue; however, this

eff ect was not consistently observed.

Average seasonal DOC losses were 0.88 kg ha–1 for autumn,

1.09 kg ha–1 for winter, 1.14 kg ha–1 for spring, and 0.32 kg ha–1 for

summer. Th e seasonal DOC losses follow the same trend as aver-

age seasonal drainfl ow (38 mm for autumn, 69 mm for winter, 73

mm for spring, and 19 mm for summer). However, seasonal FW-

DOC concentrations were 2.08 mg L–1 for autumn, 1.65 mg L–1

for winter, 1.83 mg L–1 for spring, and 1.73 mg L–1 for summer,

indicating that DOC losses and DOC concentrations had dif-

ferent seasonal trends. Monthly FW-DOC concentrations show

periodic increases associated with the proximate timing of rainfall

to lagoon effl uent application (Fig. 2). Drainfl ow occurred within

1 mo of SLE application in 1999, 2001, and 2004 and FW-DOC

concentrations in these months appeared to increase relative to

other cropping system treatments in 1999 and 2001, but not in

2004 (Fig. 2). Likewise, drainfl ow occurred within 1 mo of FLE

application in 2002 and 2004. Th e FW-DOC concentrations for

FLE appeared to increase relative to other treatments in November

2002 and November 2004 (Fig. 2).

Among all 48 drainage plots, 6-yr cumulative drainfl ow was

signifi cantly and positively correlated with the 6-yr cumulative

DOC loss (r = 0.89), and signifi cantly and negatively corre-

lated with 6-yr FW-DOC concentration (r = –0.48) and aver-

age daily DOC concentration (r = –0.57) (Table 5). Th e 6-yr

FW-DOC concentrations and the average daily DOC concen-

trations were signifi cantly correlated (r = 0.90), but only daily

DOC concentrations were signifi cantly correlated with 6-yr

DOC losses. Th e SOC content was signifi cantly and positively

correlated with 6-yr FW-DOC concentration (r = 0.33) and

the average daily DOC concentrations (r = 0.47).

Th e 6-yr DOC fl ux was determined to be a linear function

of 6-yr cumulative drainage when log transformed (r2 = 0.84)

(Fig. 3). Th e slope (b = 0.76) of the 6-yr drainfl ow-DOC fl ux

relationship had a 95% confi dence limit between 0.66 and 0.86,

indicating b was signifi cantly <1 and lower FW-DOC concentra-

tions occurred in conjunction with high drainfl ow totals. Daily

DOC fl ux was also determined to be a linear function of daily

drainfl ow when log transformed (r2 = 0.90) (Fig. 4). Th e slope

(b = 1.10) of the daily drainfl ow-DOC fl ux relationship had a

95% confi dence limit between 1.09 and 1.11, indicating b was

signifi cantly >1 and that higher daily concentrations occurred in

conjunction with high daily drainfl ows.

Discussion

Treatment Eff ects on Dissolved Organic Carbon Loss

and Dissolved Organic Carbon ConcentrationTh e DOC losses via tile drainage were low when compared

to the available C in the soil and crop residue. Yearly DOC losses

to drainage represented between 0.001 to 0.006% of the organ-

ic C in the soil profi le (137.7 Mg ha–1; Hernandez-Ramirez et

al., 2009). Based on harvest indexes from Johnson et al. (2006)

and assuming a residue C concentration of 40%, the observed

DOC losses represented between 0.05 and 0.28% of corn residue

and between 0.10 and 0.52% of soybean residue (average yearly

residue returned ranged from 6800–8500 kg ha–1 for corn and

1560–4870 kg ha–1 for soybean, estimated from Table 1). Despite

large diff erences in crop biomass, the estimated input of crop resi-

due was not correlated with DOC loss or FW-DOC concentra-

tion indicating that within corn-based systems, relative crop pro-

ductivity was not an important indicator of DOC loss (Table 5).

Fresh inputs of corn residue may not be the main source of DOC,

as shown by Flessa et al. (2003), who determined that after 37 yr

of corn production, less than one-third of leached DOC originat-

ed from the corn biomass. Th e 1.78 to 8.61 kg ha–1 yr–1 drainage

losses of DOC were remarkably small when compared to gaseous

losses of C (as carbon dioxide), which have been reported to total

1.2 to 1.6 Mg ha–1 during a 94 d span (Parkin and Kaspar, 2004),

Table 3. Yearly and 6-yr dissolved organic carbon (DOC) loss in subsurface drainage as aff ected by agricultural management treatment and water year. Refer to Table 1 for description of treatments.

DOC Loss

Water year

Treatment 1999 2000 2001 2002 2003 2004 Total†

––––––––––––––––––kg ha–1––––––––––––––––––PG 2.98 2.43 3.81 4.01 1.69 1.19 16.1

BCL 3.11 1.65 2.66 7.35 2.68 1.62 19.1

BCM 3.72 2.12 3.67 8.91 3.41 1.85 23.7

BCH 2.98 1.97 2.68 7.82 2.65 2.05 20.2

CBL 2.58 2.38 1.86 8.63 1.45 1.96 18.9

CBM 2.94 2.61 2.20 11.27 1.29 2.53 22.8

CBH 3.27 1.50 2.58 8.35 1.61 1.88 19.2

CCL 2.68 3.50 4.82 14.67 1.96 1.92 29.6

CCM 2.75 1.72 2.90 8.60 1.30 1.35 18.6

CCH 2.17 1.69 1.89 7.91 1.61 1.65 16.9

SLE 4.06 2.60 3.39 8.00 2.12 1.60 21.8

FLE 2.95 2.64 3.24 7.86 1.76 1.81 20.3

Avg. 3.02 2.24 2.97 8.61 1.96 1.78 20.9

SD‡ 1.59 1.63 2.25 5.37 1.00 0.90 10.4

CV§, % 52.6 72.8 75.8 62.4 51.0 50.6 50.7

Variation df –––––––––––––––––––P > F–––––––––––––––––––Block 3 0.611 0.297 0.913 0.652 0.826 0.800 0.915

Treatment 11 0.975 0.927 0.899 0.640 0.064 0.898 0.956

† Total, cumulative DOC loss between 1 Oct. 1998 and 30 Sept. 2004.

‡ SD, standard deviation.

§ CV, coeffi cient of variation.

Ruark et al.: Dissolved Organic C Losses from Tile Drained Agroecosystems 1211

0.8 to 1.7 kg ha–1 during a 5-mo period (Rochette et al., 2000),

and 4.6 to 13 Mg ha–1 annually (Brye et al., 2002)

In general, the yearly FW-DOC concentrations determined in

this study agree with the FW-DOC concentrations reported by

Owens et al. (2002) (0.5–3.2 mg L–1) and Kovacic et al. (2000)

(2.1–3.6 mg L–1). Additionally, the daily DOC concentrations ob-

served in this study were similar to the range in DOC concentrations

reported by McCarty and Bremner (1992) (< 0.3 to 2.9 mg L–1)

and Beauchemin et al. (1–6 mg L–1), but were markedly lower than

concentrations reported by Brye et al. (2001). Brye et al. (2001)

reported that 60% of the DOC concentrations were between 5 and

20 mg L–1 whereas only 0.5% of our values were >5 mg L–1. Th e

dissimilarity of our results and those of Brye et al. (2001) is some-

what surprising given similarities in soil textures and SOC contents

between the fi eld sites. However, the diff erent results may be an ar-

tifact of study methodology. Th e equilibrium tension lysimeter used

by Brye et al. (2001) placed a tension on the soil slightly greater than

the matric potential, providing an opportunity to sample soil water

from smaller pores sizes that would not be collected using tile drains

(Kovacic et al., 2000), monolith lysimeters (Owens et al., 2002),

or drainage lysimeters (used in this study; McCarty and Bremner,

1992; and Beauchemin et al., 2003).

Our results suggest that annual DOC losses were small compared

to the amount of organic C applied in lagoon effl uent (300–630 kg

ha–1) and that application of lagoon effl uent does not appear to in-

crease yearly or 6-yr DOC losses from drainfl ow. However, applica-

tion of lagoon effl uent aff ected yearly and 6-yr FW-DOC concen-

trations, although the eff ect of increased FW-DOC concentrations

due to lagoon effl uent application was limited to spring applications

only. Regardless, both SLE and FLE applications did exhibit short-

term increases in monthly FW-DOC concentrations over other

treatments when drainfl ow occurred within 1 mo of application

(Fig. 2). Similar results were reported by Royer et al. (2007), who

demonstrated that fall application of manure increased DOC con-

centrations in drain water during the 2 mo following application.

Short-term spikes in DOC concentration following effl uent appli-

Table 4. Yearly and 6-yr fl ow-weighted dissolved organic carbon (FW-DOC) concentrations in subsurface drainfl ow as aff ected by agricultural management treatment and water year. Refer to Table 1 for descriptions.

FW-DOC Concentration

Water year

Treatment 1999 2000 2001 2002 2003 2004 Total†

–––––––––––––––––––mg L–1–––––––––––––––––––PG 1.28 1.51 1.69 1.72 1.56 1.87 1.53

BCL 1.39 1.57 1.89 1.81 1.80 1.66 1.69

BCM 1.63 1.90 1.97 2.29 2.01 2.11 2.00

BCH 1.45 1.66 1.92 2.03 1.89 2.08 1.85

CBL 1.12 1.78 1.47 1.99 1.15 1.66 1.61

CBM 1.35 1.46 1.77 2.36 1.20 1.92 1.86

CBH 1.40 1.27 1.77 1.94 1.36 1.74 1.66

CCL 1.45 2.11 2.05 2.10 1.68 2.16 1.94

CCM 1.42 1.38 1.68 2.30 1.29 2.27 1.84

CCH 1.33 1.46 1.79 2.00 1.48 2.02 1.73

SLE 2.32 2.21 2.24 2.84 2.50 2.23 2.49

FLE 1.27 1.67 1.73 2.50 2.02 2.68 1.90

Avg. 1.45 1.67 1.83 2.16 1.67 2.03 1.84

SD‡ 0.48 0.65 0.45 0.56 0.63 0.64 0.48

CV§, % 33.1 38.9 24.6 25.9 37.7 31.8 26.0

Variation df –––––––––––––––––––P > F–––––––––––––––––––

Block 3 0.020 0.031 0.076 0.058 0.244 0.123 0.055

Treatment 11 0.032 0.565 0.615 0.147 0.065 0.587 0.284

Contrasts¶

PG vs. UAN NS NS NS NS NS NS NS

SLE vs. UAN *** NS NS * * NS *

FLE vs. UAN NS NS NS NS NS NS NS

CC vs. CB NS NS NS NS NS NS NS

BC vs. CB NS NS NS NS * NS NS

CC vs. BC+CB NS NS NS NS NS NS NS

* Signifi cant at the 0.05 probability level (with Bonferroni adjustment).

*** Signifi cant at the 0.001 probability level (with Bonferroni adjustment).

† Total, FW-DOC concentrations calculated from the cumulative drainfl ow

and cumulative mass DOC loss between 1 Oct. 1998 and 30 Sept. 2004.

‡ SD, standard deviation.

§ CV, coeffi cient of variation.

¶ Contrast treatment combinations: UAN, all agricultural treatments in

continuous corn or corn-soybean rotations receiving UAN applications; PG,

restored prairie grass; CC, all continuous corn rotations (CCL, CCM, CCH, see

Table 1 for descriptions); CB, all corn cropping years within corn-soybean

rotations (CBL, CBM, and CBH, see Table 1 for descriptions); BC, all soybean

cropping years within soybean-corn rotations (BCL, BCM, and BCH, see Table

1 for descriptions).

Fig. 2. Average monthly fl ow-weighted dissolved organic carbon (FW-DOC) concentrations for selected agricultural management treatments. Treatments include: prairie grass (PG), corn-soybean, and continuous corn rotations with the Purdue University recommended N fertilizer application rates of UAN (BCM, CBM, and CCM), continuous corn with spring application of swine lagoon effl uent (SLE), and continuous corn with fall application of swine lagoon effl uent (FLE). Monthly FW-DOC concentrations were reported when >75% of drainage plots had fl ow. White arrows indicate the month of spring applied lagoon effl uent and black arrows indicate the month of fall applied lagoon effl uent. Error bars represent standard error.

1212 Journal of Environmental Quality • Volume 38 • May–June 2009

cation may, in part, refl ect the effl uent’s DOC moving through the

preferential fl ow pathways of macroporus soil. Both the nature and

drivers of DOC concentration dynamics in relation to lagoon effl u-

ent applications require further investigation.

In this study, restoration of prairie grass on tile drained soils did

not appear to aff ect drainfl ow, DOC losses, or FW-DOC concen-

trations. Th ese results were in contrast to Brye et al. (2001), who

observed large diff erences in 4-yr cumulative drainfl ow between a

chisel-plowed agroecosystem and a restored prairie (1575 and 461

mm, respectively); these drainage diff erences were associated with

a twofold diff erence in DOC loss (annual cropping system > prai-

rie). Additionally, the DOC concentrations in the agricultural soils

were typically higher than in the prairie soils (Brye et al., 2001).

Our results, however, are similar to those by Owens et al. (2002),

who reported similar DOC losses and concentrations in subsur-

face fl ow between pasturelands and corn-soybean rotations.

Univariate statistics of within-year and 6-yr variation in

drainfl ow and effi ciency (Table 2) indicate a large variability

in hydrologic properties among the relatively small drainage

plots. Yet previous studies have suggested this degree of varia-

tion in tile performance is typical. In a 36 plot (4446 m2 plots,

1.2 m depth, 28.5 m spacing), 6-yr study, Bakhsh and Kan-

war (2004) reported annual CV values for drainfl ow between

39 and 85%, which are similar to our CV values (51–88%).

Likewise, within years, the Bakhsh and Kanwar (2004) obser-

vations were also right-skewed around the mean although the

degree of skewness (0.89 and 2.3) was less than observed in this

study (0.23–4.2). Although plot-to-plot variability was high,

our average yearly TDEs were similar to TDEs in southern

Minnesota (36%; Jin and Sands, 2003) and southern Indiana

(13–16%; Kladivko et al., 2004).

Drainfl ow-Dissolved Organic Carbon Flux RelationshipsAmong tiles, the single greatest determinant of DOC loss

was drainfl ow. Th is relationship was observable and signifi cant

on a daily and 6-yr cumulative loss basis, the extremes of the

timescale examined in this study. Studies of nitrate fl ux from

individual tiles (Hofmann et al., 2004; Tomer et al., 2003)

Table 5. Correlation matrix of selected measured variables from individual drainage plots determined from 1 Oct. 1998 through 30 Sept. 2004. Variables are 6-yr cumulative totals or 6-yr averages per drainage plot.

Variable Unit nAverage daily

DOC concentrationCumulative

DOC loss FW-DOC concentration

–––––––––––––––––––––––––––r–––––––––––––––––––––––––––Average Daily DOC concentration mg L–1 48 1.00

Cumulative DOC loss kg ha–1 48 –0.29* 1.00

FW-DOC Concentration mg L–1 48 0.90*** –0.12 1.00

Cumulative drainfl ow L 48 –0.57*** 0.89*** –0.48***

Average SOC† mg kg–1 48 0.47*** 0.05 0.33*

Estimated cumulative residue‡ kg 44 0.22 0.12 0.22

* Signifi cant at the 0.05 probability level.

*** Signifi cant at the 0.001 probability level.

† SOC, soil organic carbon (0–20 cm, averaged across 1999, 2000, 2003, and 2004 soil samples).

‡ Estimated cumulative residue returned after harvest, using a harvest index of 0.53 and 0.46 for corn and soybean, respectively (Johnson et al., 2006).

Plots with prairie grass were not included because there were no estimates of residue return.

Fig. 4. Relationship between daily drainfl ow and daily dissolved organic carbon (DOC) fl ux presented on a log-log scale. Each data point represents daily drainfl ow and daily DOC fl ux from individual drainage plots from days when samples were collected and when daily drainfl ow was >0.1 mm d–1. Exponent (b) and coeffi cient (a) values are defi ned by Eq. [1] and r2 values were determined by linear regression on log-transformed drainfl ow and DOC fl ux values. Iso-lines for 0.5 and 3 mg L–1 are provided for orientation.

Fig. 3. Relationship between 6-yr cumulative drainfl ow and 6-yr dissolved organic carbon (DOC) fl ux presented on a log-log scale. Each data point represents total values from an individual drainage plot. Exponent (b) and coeffi cient (a) values are defi ned by Eq. [1] and r2 values were determined by linear regression on log-transformed drainfl ow and DOC fl ux values. Iso-lines for 1 and 3 mg L–1 are provided for orientation.

Ruark et al.: Dissolved Organic C Losses from Tile Drained Agroecosystems 1213

have also identifi ed discharge as a dominant factor determin-

ing annual and multi-year solute exports.

Th e diff erent regression coeffi cients for daily and 6-yr regres-

sion relationships provide some additional insight into the dy-

namics of DOC loss. Th rough the entire course of the study,

lower FW-DOC concentrations in high-fl owing tiles (b < 1, Fig.

3) suggest that either there are limits to the amount of relatively

soluble DOC, the preponderance of preferential fl ow pathways

allow newly added rainwater little time to equilibrate with the

soil matrix, or a combination of both. However, observations of

higher concentrations with greater drainfl ow on a daily time-step

indicate a diff erent soil DOC dynamic occurs at higher temporal

resolution. Hydrographs for drainage events typically follow a log-

normal distribution with peak rates occurring within a few hours

of initial fl ow (Jury, 1982). When associated chemographs have a

concentration peak near the peak fl ow rate followed by a decrease

in concentration as fl ow rates decrease, preferential fl ow path-

ways are thought to dominate soil solute transport (Lennartz et

al., 1999). In studies of nonadsorbed and adsorbed tracers and in

studies of diff erentially adsorbed pesticides, such parallel behavior

between hydrographs and chemographs have been observed when

rainfall occurs shortly after solute introduction to soil (Kung et

al., 2000; Kladivko et al., 1999, 2001). Th us, our results indicate

that high-volume, short-duration discharges may tend to occur in

conjunction with elevated DOC concentrations in soil macropo-

res. As noted previously, this interpretation is also supported by

our observations of elevated FW-DOC concentrations in SLE

and FLE treatments in the month following lagoon effl uent ap-

plication (Fig. 2). Evidence of preferential pathway transport of

DOC following application of DOC to the soil surface was also

provided by Gentry et al. (2000) and Royer and David (2005),

who observed high DOC concentrations in peak tile drain fl ow,

followed by a sharp decrease in DOC concentrations with decreas-

ing fl ow rates. Similar daily drainfl ow-solute fl ux behavior (b > 1)

has been observed for nitrate (Hofmann et al., 2004; Tomer et al.,

2003), but opposite trends have also been observed (Kung et al.,

2000; Kladivko et al., 2004). In forest soils, DOC exported during

macropore fl ow can be as much as 90 times greater than during

matrix fl ow (Kaiser et al., 2000), but the relative contributions of

preferential and matrix fl ow remain poorly understood for agri-

cultural soils. Further explanation of this phenomenon requires

detailed analyses of hydro- and chemograph behaviors.

Our data, along with results from Dalzell et al. (2007), suggests

that tile drains contribute toward the watershed level export of

DOC in watersheds dominated by tile drained agriculture. Dalzell

et al. (2007) examined watershed level DOC loss in a west-central

Indiana watershed (Big Pine Creek Watershed) and the authors

reported watershed level DOC loss as 19.5 and 14.1 kg ha–1 yr–1

in 2002 and 2003, respectively. Big Pine Creek Watershed is

dominated by agricultural land use (80%) and is similar in topo-

graphic relief, soil characteristics, and rainfall patterns to that of

the WQFS. If we assume that all agricultural soils in the Big Pine

Creek Watershed are tile drained, then, in 2002, tile drainage loss-

es of DOC (8.61 kg ha–1) represent 44% of the watershed level

DOC loss. Under the same assumptions, in 2003, tile drainage

losses of DOC (1.96 kg ha–1) represent 14% of the watershed level

DOC loss. Th erefore, there is evidence to suggest that while DOC

export from tile drains represents a very small portion of the terres-

trial carbon cycle, it represents an important DOC source in the

aquatic carbon cycle within agriculturally based watersheds.

ConclusionsIn general, our results indicate that common agricultural

management alternatives which have been consistently applied

for a decade have little eff ect on the drainage loss component

of the terrestrial carbon cycle. Annual FW-DOC concentrations

were similar among continuous corn and corn-soybean cropping

systems and between inorganic fertilized cropping systems and

restored prairie grass systems on tile drained fi elds. Th e only crop

management practice that aff ected DOC concentration was ap-

plication of lagoon effl uent, and then only for a short duration

and under specifi c climatic conditions. Drainfl ow was the main

driver of DOC fl uxes and also impacted DOC concentrations,

as large amounts of daily drainfl ow tended to increase DOC

concentrations, rather than dilute them. Th is relationship be-

tween drainfl ow and DOC concentration is important to con-

sider when determining watershed C budgets or modeling C

and N dynamics in subsurface soils. Concentrations of DOC

in tile drains of this study were typically <2 mg L–1; these values

are comparable to the results of Owens et al. (2002), but some-

what lower than in other eastern Corn Belt studies (Kovacic et

al., 2000; Royer and David, 2005). In addition, losses of DOC

from tile drains are relatively low compared to SOC content, soil

input from crop residue or lagoon effl uent, and reported carbon

dioxide fl uxes. While drainage losses represent a small portion of

the overall DOC fl ux from terrestrial systems, they can represent

large portions of watershed-level DOC fl ux. Th e ecological im-

plications DOC from tile drainage and other locations is largely

unknown, but recent work suggests that this DOC can be easily

used by bacteria (Waranoski, 2007), and therefore can play a

major role in the microbial ecology of river water.

AcknowledgmentsTh e authors would like to thank Brenda Hofmann for

her eff orts in collecting samples and maintaining the WQFS.

Th e authors acknowledge support from the United States

Department of Agriculture through the Consortium for

Agriculture Soils Mitigation of Greenhouse Gasses program

and through CSREES-NRI Watershed Processes and Water

Resources Grant Number 04-35102-14908.

ReferencesAngers, D.A., M.H. Chantigny, P. Rochette, and B. Gagnon. 2006. Dynamics

of soil water-extractable organic C following application of dairy cattle manures. Can. J. Soil Sci. 86:851–858.

Bakhsh, A., and R.S. Kanwar. 2004. Spatio-temporal analysis of subsurface drainage fl ow volumes. Trans. ASAE 47:1427–1436.

Bakhsh, A., R.S. Kanwar, T.B. Baily, C.A. Cambardella, D.L. Karlen, and T.S. Colvin. 2002. Cropping system eff ects on NO

3–N loss with subsurface

drainage water. Trans. ASAE 45:1789–1797.

Bakhsh, A., R.S. Kanwar, and D.L. Karlen. 2005. Eff ects of liquid swine manure applications on NO

3–N leaching losses to subsurface drainage

water from loamy soils in Iowa. Agric. Ecosyst. Environ. 109:118–128.

Beauchemin, S., R.R. Simard, M.A. Bolinder, M.C. Nolin, and D. Cluis.

1214 Journal of Environmental Quality • Volume 38 • May–June 2009

2003. Prediction of phosphorus concentration in tile-drainage water from the Montreal Lowlands soils. Can. J. Soil Sci. 83:73–87.

Bol, R., N.J. Ostle, C. Friedrich, W. Amelung, and I. Sanders. 1999. Th e infl uence of dung amendments on dissolved organic matter in grassland soil leachates—Preliminary results from a lysimeter study. Isot. Environ. Health Stud. 35:97–109.

Bremner, J.M. 1996. Nitrogen–Total. p. 1085–1121. In D.L. Sparks (ed.) Methods of soil analysis. Part 3. SSSA Book Ser. no 5. SSSA and ASA, Madison, WI.

Brye, K.R., S.T. Gower, J.M. Norman, and L.G. Bundy. 2002. Carbon budgets for a prairie and agroecosystems: Eff ects of land use and interannual variability. Ecol. Appl. 12:962–979.

Brye, K.R., J.M. Norman, L.G. Bundy, and S.T. Gower. 2001. Nitrogen and carbon leaching in agroecosystems and their role in denitrifi cation potential. J. Environ. Qual. 30:58–70.

Campbell, C.A., G.P. Lafond, V.O. Biederbeck, G. Wen, J. Schoenau, and D. Hahn. 1999. Seasonal trends in soil biochemical attributes: Eff ects of crop management on a Black Chernozem. Can. J. Soil Sci. 79:85–97.

Chantigny, M.H. 2003. Dissolved and water-extractable organic matter in soils: A review on the infl uence of land use and management practices. Geoderma 113:357–380.

Clay, D.E., S.A. Clay, Z. Liu, and S.S. Harper. 1995. Leaching of dissolved organic carbon in soil following anhydrous ammonia application. Biol. Fertil. Soils 19:10–14.

Dalzell, B.J., T.R. Filley, and J.M. Harbor. 2007. Th e role of hydrology in annual organic carbon loads and terrestrial organic matter export from a midwestern agricultural watershed. Geochim. Cosmochim. Acta 71:1448–1462.

Eigel, J.D., R.F. Turco, and E.J. Kladivko. 1992. Slurry trenching to isolate water quality research plots. Paper 922066. American Soc. of Agric. Engineers, St. Joseph, MI.

Fausey, N.R., L.C. Brown, H.W. Belcher, and R.S. Kanwar. 1995. Drainage and water quality in Great Lakes and Cornbelt States. J. Irrig. Drain. Div. Am. Soc. Civ. Eng. 121:283–288.

Flessa, H., B. Ludwig, B. Heil, and W. Merbach. 2003. Th e origin of soil organic C, dissolved organic C and respiration in a long-term maize experiment in Halle, Germany, determined by 13C natural abundance. J. Plant Nutr. Soil Sci. 163:157–163.

Franchini, J.C., F.J. Gonzalez-Vila, F. Cabrera, M. Miyazawa, and M.A. Paven. 2001. Rapid transformation of plant water-soluble organic compounds in relation to cation mobilization in an acid Oxisol. Plant Soil 231:55–63.

Gentry, L.E., M.B. David, K.M. Smith-Starks, and D.A. Kovacic. 2000. Nitrogen fertilizer and herbicide transport from tile drained fi elds. J. Environ. Qual. 29:232–240.

Hernandez-Ramirez, G., S.M. Brouder, D.R. Smith, and G.E. Van Scoyoc. 2009. Carbon and nitrogen dynamics in an eastern corn belt soil: Nitrogen source and rotation. Soil Sci. Soc. Am. J. 73:128–137.

Hofmann, B.S., S.M. Brouder, and R.F. Turco. 2004. Tile spacing impacts on Zea mays L. yield and drainage water nitrate load. Ecol. Eng. 23:251:267.

Hope, D., M.F. Billett, and M.S. Cresser. 1994. A review of the export of carbon in river water: Fluxes and processes. Environ. Pollut. 84:301–324.

Huggins, D.R., G.W. Randall, and M.P. Russelle. 2001. Subsurface drain losses of water and nitrate following conversion of perennials to row crops. Agron. J. 93:477–486.

Jacinthe, P.A., R. Lal, and J.M. Kimble. 2001. Organic carbon storage and dynamics in croplands and terrestrial deposits as infl uenced by subsurface tile drainage. Soil Sci. 166:322–335.

Jamieson, R.C., R.J. Gordon, K.E. Sharples, G.W. Stratton, and A. Madani. 2002. Movement and persistence of fecal bacteria in agricultural soils and subsurface drainage water: A review. Can. Biosys. Eng. 44: 1.1–1.9.

Jaynes, D.B., T.S. Colvin, D.L. Karlen, C.A. Cambardella, and D.W. Meek. 2001. Nitrate loss in subsurface drainage as aff ected by nitrogen fertilizer rate. J. Environ. Qual. 30:1305–1314.

Jin, C.X., and G.R. Sands. 2003. Th e long-term fi eld-scale hydrology of subsurface drainage systems in a cold climate. Trans. ASAE 46:1011–1021.

Johnson, J.M.F., R.R. Allmaras, and D.C. Reicosky. 2006. Estimating source carbon from crop residues, roots and rhizodeposits using the national grain-yield database. Agron. J. 98:622–636.

Jury, W. 1982. Simulation of solute transport using a transfer function model. Water Resour. Res. 18:363–368.

Kaiser, K., G. Guggenberger, and W. Zech. 2000. Organically bound nutrients in dissolved organic matter fractions in seepage and pore water of weakly developed forest soils. Acta Hydrochim. Hydrobiol. 28:411–419.

Kalbitz, K., S. Solinger, J.H. Park, B. Michalzik, and E. Matzner. 2000. Controls on the dynamics of dissolved organic carbon: A review. Soil Sci. 165:277–304.

Kladivko, E.J., L.C. Brown, and J.L. Baker. 2001. Pesticide transport to subsurface tile drains in humid regions of North America. Crit. Rev. Environ. Sci. Technol. 31:1–62.

Kladivko, E.J., J.R. Frankenberger, D.B. Jaynes, D.W. Meek, B.J. Jenkinson, and N.R. Fausey. 2004. Nitrate leaching to subsurface drains as aff ected by drain spacing and changes in crop production system. J. Environ. Qual. 33:1803–1813.

Kladivko, E.J., J. Grochulska, R.F. Turco, G.E. Van Scoyoc, and J.D. Eigel. 1999. Pesticide and nitrate transport into subsurface tile drains of diff erent spacings. J. Environ. Qual. 28:997–1004.

Kovacic, D.A., M.B. David, L.E. Gentry, K.M. Starks, and R.A. Cooke. 2000. Eff ectiveness of constructed wetlands in reducing nitrogen and phosphorus export from agricultural tile drainage. J. Environ. Qual. 29:1262–1274.

Kung, K.J.S., T.S. Steenhuis, E.J. Kladivko, T.J. Gish, G. Bubenzer, and C.S. Helling. 2000. Impact of preferential fl ow on the transport of adsorbing and non-adsorbing tracers. Soil Sci. Soc. Am. J. 64:1290–1296.

Lennartz, B., J. Michaelsen, W. Wichtmann, and P. Widmoser. 1999. Time variance analysis of preferential solute movement at a tile-drained fi eld site. Soil Sci. Soc. Am. J. 63:39–47.

Lorenz, K., and R. Lal. 2005. Th e depth distribution of soil organic carbon in relation to land use and management and the potential of carbon sequestration in subsoil horizons. Adv. Agron. 88:35–66.

Mazzarino, M.J., L. Szott, and M. Jimenez. 1993. Dynamics of soil total C and N, microbial biomass, and water-soluble C in tropical agroecosystems. Soil Biol. Biochem. 25:205–214.

McCarty, G.W., and J.M. Bremner. 1992. Availability of organic carbon for denitrifi cation of nitrate in subsoils. Biol. Fertil. Soils 14:219–222.

McDowell, W.H. 2003. Dissolved organic matter in soils–Future directions and unanswered questions. Geoderma 113:179–186.

McTiernan, K.B., S.C. Jarvis, D. Scholefi eld, and M.H.B. Hayes. 2001. Dissolved organic carbon from grazed grasslands under diff erent management regimes. Water Res. 35:2565–2569.

Neff , J.C., and G.P. Asner. 2001. Dissolved organic carbon in terrestrial ecosystems: Synthesis and a model. Ecosystems 4:29–48.

Nelson, D.W., and L.E. Sommers. 1996. Total carbon, organic carbon, and organic matter. p. 961–1010. In D.L. Sparks (ed.) Methods of soil analysis. Part 3. SSSA Book Ser. no 5. SSSA and ASA, Madison, WI.

Owens, L.B., G.C. Starr, and D.L. Lightell. 2002. Total organic carbon losses in subsurface fl ow under two management practices. J. Soil Water Conserv. 57:74–81.

Parkin, T.B., and T.C. Kaspar. 2004. Temporal variability of soil carbon dioxide fl ux: Eff ect of sampling frequency on cumulative carbon loss estimation. Soil Sci. Soc. Am. J. 68:1234–1241.

Randall, G.W., and J.A. Vetsch. 2005. Nitrate losses in subsurface drainage from a corn-soybean rotation as aff ected by fall and spring application of nitrogen and nitrapyrin. J. Environ. Qual. 34:590–597.

Randall, G.W., J.A. Vetsch, and J.R. Huff man. 2003. Nitrate losses in subsurface drainage from a corn-soybean rotation as aff ected by time of nitrogen application and use of nitrapyrin. J. Environ. Qual. 32:1764–1772.

Raulund-Rasmussen, K., O.K. Borggaard, H.C.B. Hansen, and M. Olsson. 1998. Eff ect of natural organic soil solutes on weathering rates of soil minerals. Eur. J. Soil Sci. 49:397–406.

Rochette, P., D.A. Angers, and D. Côté. 2000. Soil carbon and nitrogen dynamics following application of pig slurry for the 19th consecutive year: I. Carbon dioxide fl uxes and microbial biomass carbon. Soil Sci. Soc. Am. J. 64:1389–1395.

Royer, I., D.A. Angers, M.H. Chantigny, R.R. Simard, and D. Cluis. 2007. Dissolved organic carbon in runoff and tile-drain water under corn and forage fertilized with hog manure. J. Environ. Qual. 36:855–863.

Royer, T.V., and M.B. David. 2005. Export of dissolved organic carbon from agricultural streams in Illinois, USA. Aquat. Sci. 67:465–471.

SAS Institute.1999. SAS/STAT user’s guide. Vol. 2. Version 8. SAS Inst., Cary, NC.

Siemens, J., M. Haas, and M. Kaupenjohann. 2003. Dissolved organic matter induced denitrifi cation in subsoils and aquifers? Geoderma 113:253–271.

Skaggs, R.W., M.A. Breve, and J.W. Gilliam. 1994. Hydrologic and water quality impacts of agricultural drainage. Crit. Rev. Environ. Sci. Technol. 24:1–32.

Sotomayor, D., and C.W. Rice. 1996. Denitrifi cation in soil profi les beneath grassland and cultivated soils. Soil Sci. Soc. Am. J. 60:1822–1828.

Ruark et al.: Dissolved Organic C Losses from Tile Drained Agroecosystems 1215

Temminghoff , E.J.M., S.E.A.T.M. van der Zee, and F.A.M. de Haan. 1997. Copper mobility in a copper-contaminated sandy soil as aff ected by pH and solid and dissolved organic matter. Environ. Sci. Technol. 31:1109–1115.

Tomer, M.D., D.W. Meek, D.B. Jaynes, and J.L. Hatfi eld. 2003. Evaluation of nitrate nitrogen fl uxes from a tile-drained watershed in central Iowa. J. Environ. Qual. 32:642–653.

Totsche, K.U., J. Danzer, and I. Kögel-Knabner. 1997. Dissolved organic matter-enhanced retention of polycyclic aromatic hydrocarbons in soil miscible displacement experiments. J. Environ. Qual. 26:1090–1100.

Vitosh, M.L., J.W. Johnson, and D.B. Mengel. 1995. Tri-state fertilizer recommendations for corn, soybeans, wheat and alfalfa. Ext. Bull. E-2567. Michigan State Univ., East Lansing.

Waranoski, V.M. 2007. Environmental controls on the survival of E. coli in surface water. M.S. thesis. Purdue Univ., West Lafayette, IN.

Yeomans, J.C., J.M. Bremner, and G.W. McCarty. 1992. Denitrifi cation capacity and denitrifi cation potential of subsurface soils. Commun. Soil Sci. Plant Anal. 23:919–927.