U.S. Department of the Interior

U.S. Geological Survey

Ecological Effects on Streams from Forest 

Fertilization—Literature Review and Conceptual 

Framework for Future Study in the Western 

Cascades

By CHAUNCEY W. ANDERSON

Water-Resources Investigations Report 01–4047

Prepared in cooperation with

the BUREAU OF LAND MANAGEMENT

Portland, Oregon: 2002
U.S. DEPARTMENT OF THE INTERIOR

GALE A. NORTON, Secretary

U.S. GEOLOGICAL SURVEY

CHARLES G. GROAT, Director

The use of trade, product, or firm names in this publication is for

descriptive purposes only and does not imply endorsement by the

U.S. Government.

For additional information contact: Copies of this report can be

purchased from:

District Chief

U.S. Geological Survey USGS Information Services

10615 S.E. Cherry Blossom Drive Box 25286, Federal Center

Portland, OR 97216-3159 Denver, CO 80225-0046

E-mail: info-or@usgs.gov Telephone: 1-888-ASK-USGS

Internet: http://oregon.usgs.gov

Suggested citation:

Anderson, C.W., 2002, Ecological effects on streams from forest fertilization—Literature review and conceptual 

framework for future study in the western Cascades: U.S. Geological Survey Water-Resources Investigations 

Report 01–4047, 49 p.
iii

CONTENTS

Acknowledgments.................................................................................................................................................... v

Abstract .....................................................................................................................................................................1

Introduction...............................................................................................................................................................1

Literature Review..................................................................................................................................................... 2

Forest Processes................................................................................................................................................ 2

Water-Quality Effects on Streams from Forest Fertilization.......................................................................... 10

Immediate Nutrient Runoff ..................................................................................................................... 10

Longer-Term Nutrient Losses to Streams ............................................................................................... 10

Water-Quality Criteria............................................................................................................................. 11

Ecological Effects on Streams from Forest Fertilizations .............................................................................. 12

Stream Nutrient Dynamics.............................................................................................................................. 13

Hyporheic Processing ..................................................................................................................................... 15

Conceptual Model of Ecological Processing of Fertilizer Nitrogen ....................................................... 16

Suggested Approaches to Evaluate Ecological Effects of Forest Fertilization ....................................... 17

Little River Watershed ........................................................................................................................................... 20

Physiographic Setting ..................................................................................................................................... 20

Water-Quality Issues in the Little River Watershed....................................................................................... 22

Land uses in the Little River Watershed and Potential Effects on Water Quality.......................................... 23

Water-Quality Conditions............................................................................................................................... 24

Methods ................................................................................................................................................... 24

Quality Assurance ................................................................................................................................... 29

Environmental Data................................................................................................................................. 30

Framework for Fertilization Study in Wolf Creek and Little River Watershed .................................................... 35

Possible Study Approaches............................................................................................................................. 38

Summary ................................................................................................................................................................ 40

References Cited .................................................................................................................................................... 41

TABLES

1. Summary of studies of forest fertilization effects on stream-water quality......................................................... 4

2. Aquatic species of concern in the Little River watershed, Oregon.................................................................... 22

3. Sites sampled in the Little River watershed by the U.S. Geological Survey (USGS) in 1998 and 1999 ......... 25

4. Nutrient and field data in the Little River and tributaries from reconnaissance samplings 

during August and November, 1998.................................................................................................................. 26

5. Nutrient and field data for sites in the Little River and tributaries, August 1999.............................................. 27

6. Analytical methods and detection levels for nutrient analyses performed at the U.S. Geological Survey

National Water Quality Laboratory (NWQL) and Oregon State University Cooperative Chemical Analytical 

Laboratory (CCAL) ........................................................................................................................................... 29

7. Comparison of nutrient concentrations from standard reference samples analyzed at the

U.S. Geological Survey’s National Water Quality Laboratory and Oregon State University’s 

Cooperative Chemical Analytical Laboratory, August 1998 ............................................................................ 30

8. Algal biomass and chlorophyll a measured in the Little River watershed, August 1999 ................................. 35

9. Research components for different levels of investigation of effects of urea fertilization

on water quality and stream ecology ................................................................................................................. 39
iv

FIGURES

1. Map showing Little River watershed, Oregon, and proposed Bureau of Land Management 

fertilization units.............................................................................................................................................. 3

2. Forest cycling pathways representing major processes and fates of nitrogen fertilizers................................. 9

3. Schematic drawing of hypothetical transport pathways, dominant processes, and relative 

concentrations of nitrogen in response to urea fertilization in forested cathments of the western 

Cascades, Oregon, during fall/winter and late spring/summer ..................................................................... 18

4. Map showing sampling locations in Little River Watershed, Oregon 1998.................................................. 21

5. Graph showing morning and afternoon pH and dissolved oxygen saturation in the

Little River, July 28, 1998............................................................................................................................. 22

6.–9. Graphs showing:

6. Afternoon temperature in the main stem of the Little River, August 1999................................................... 32

7. Afternoon pH in the main stem of the Little River, August 1999 ................................................................. 32

8. Major ion chemistry in Little River and Wolf Creek .................................................................................... 33

9. Distribution of nutrient concentrations in the Little River and tributaries during August, 1999 .................. 34

 10. Concentrations of nitrate-nitrogen and total phosphorus, and daily 

maximum temperature and pH in the Little River Basin during August, 1998 ............................................ 36

CONVERSION FACTORS AND ABBREVIATIONS

Multiply By To obtain

acre 0.4047 hectare

pounds per acre per year [(lb/acre)/yr] 1.121 kilograms per hectare per year (kg/ha/yr)

Milligrams per liter (mg/L) is equivalent to parts per million; micrograms per liter (μg/L) is equivalent to parts per billion.

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as follows: °F=1.8 (°C) +32.

Specific conductance is expressed in microsiemens per centimeter at 25 degrees Celsius (μs/cm at 25°C).

Sea level: In this report, “sea level: refers to the National Geodetic Vertical Datum of 1929 (NGVD of 1929)—a geodetic datum 

derived from a general adjustment of the first-order level nets of both the United States and Canada, formerly called Sea Level

Datum of 1929.
v

ACKNOWLEDGMENTS

This project was conducted with the Roseburg Office of the Bureau of Land Management 

(BLM). Craig Kintop, silviculturalist, was instrumental in obtaining funding and for encouraging 

credible, objective science about the effects of fertilization on stream biota. BLM hydrologists 

Lowell Duell and Ed Rumbold provided local logistical support (including maps and stream 

access), helped with sampling, and provided ideas and interest. Dayne Barron and Anne Boeder 

were pivotal as managers of the project within the BLM. Many others provided insights into 

literature and processes of forest fertilization, and stream ecology, including especially Drs. Stan 

Gregory and Steve Wondzell at Oregon State University. Analytical data were provided by the 

Cooperative Chemical Analytical Laboratory (CCAL) established by memorandum of 

understanding number PNW-82–187 between theU.S. Forest Service and the Department of Forest 

Science, Oregon State University. Within the U.S. Geological Survey, Don Major, Bruce Bury, and 

Niels Luthold helped brainstorm about effects of fertilizer on biological species, especially 

amphibians, and acted as guides in the Wolf Creek Basin. Charlie Patton provided advice on 

analyses of nutrients at low concentrations. Frank Triska, John Duff, and Carol Kendall helped with 

ideas about how to look for the applied fertilizers and their effects, and Kurt Carpenter, Steve 

Hinkle, and Joe Rinella participated in many discussions of hydrology, nutrient dynamics, and 

stream ecology. 
1

Ecological Effects on Streams from Forest 

Fertilization—Literature Review and Conceptual 

Framework for Future Study

By Chauncey W. Anderson

Abstract

Fertilization of forests with urea-nitrogen has 

been studied numerous times for its effects on 

water quality. Stream nitrogen concentrations following fertilization are typically elevated during 

winter, including peaks in the tens-of-thousands 

of parts per billion range, with summer concentrations often returning to background or near-background levels. Despite these increases, waterquality criteria for nitrogen have rarely been 

exceeded. However, such criteria are targeted at 

fish toxicity or human health and are not relevant to 

concentrations that could cause ecological disturbances. Studies of the responses of stream biota to 

fertilization have been rare and have targeted either 

immediate, toxicity-based responses or used methods insensitive to ongoing ecological processes. 

This report reviews water-quality studies following 

forest fertilizations, emphasizing Cascade streams 

in the Pacific Northwest and documented biological responses in those streams. A conceptual model 

predicting potential ecological response to fertilization, which includes effects on algal growth and 

primary production, is presented. In this model, 

applied fertilizer nitrogen reaching streams is 

mostly exported during winter. However, some 

nitrogen retained in soils or stream and riparian 

areas may become available to aquatic biota during 

spring and summer. Biological responses may be 

minimal in small streams nearest to application 

because of light limitation, but may be elevated 

downstream where light is sufficient to allow algal 

growth. Ultimately, algal response could be greatest in downstream reaches, although ambient nutrient concentrations remain low due to uptake and 

benthic nutrient recycling. Ground-water flow 

paths and hyporheic processing could be critical in 

determining the fate of applied nitrogen. A framework is provided for testing this response in the 

Little River watershed, a tributary to the North 

Umpqua River, Oregon, at basic and intensive 

levels of investigation.

INTRODUCTION

Fertilization of public and private timberlands with nitrogen to boost forest productivity has 

been common in the Pacific Northwest and elsewhere since the late 1960’s (Fredriksen et al., 1975; 

Binkley et al., 1999), and more frequent use of fertilization is anticipated in the future (National 

Council of the Paper Industry for Air and Stream 

Improvement [NCASI], 1999). During 1990–98, 

over 850,000 acres, or approximately 5 percent of 

Oregon’s timberland, were fertilized (Oregon 

Department of Forestry, 1999), averaging about 

95,000 acres a year. Since 1992, most fertilization 

has occurred on private timberlands; however, 

applications averaging 16,000–36,000 acres per 

year continued to State and Federal lands from 

1997–99 (Oregon Department of Forestry, 1999). 

Regionally, over 120,000 acres of forest lands were 

fertilized each year in the Pacific Northwest during 
2

the late 1980’s, and in the southeastern United 

States over 850,000 acres of pine plantations were 

fertilized in 1996 alone (Binkley et al., 1999). Forest fertilization also is practiced in other parts of 

the world, including Japan, Australia, New 

Zealand, and Sweden. 

This report reviews literature on effects of 

forest fertilization on water quality, emphasizing 

Cascade streams in the Pacific Northwest and possible ecological effects on aquatic systems in those 

streams. Although the focus is on streams, the initial discussion describes interactions of fertilizers 

with soils to the extent that they influence nutrient 

transport to streams. A brief review of literature on 

processing of nutrients in underground near-stream 

(hyporheic and riparian) and in-stream (water and 

benthic) environments also is presented. Next, a 

conceptual framework for future evaluation of 

these effects is developed. Finally, an example 

study plan is provided for examining the possible operational fertilization of urea-nitrogen 

to selected areas of the Little River Adaptive 

Management Area (LRAMA) in southwestern Oregon (fig. 1). Water-quality issues there include 

occurrences of high pH due to excessive algal productivity and the degree to which nuisance algae 

are enhanced by forestry. Data from reconnaissance 

surveys in the LRAMA are provided to indicate 

stream water-quality and algal conditions prior to 

public timberland fertilization. The suitability of 

those areas for studying fertilization’s effects also 

is evaluated. 

In this report references are made to streams 

and watersheds of different sizes that are often 

nested within larger river basins. To avoid confusion, a consistent set of terminology proposed by 

McCammon (1994) is used. The term “river basin” 

is used to refer to the equivalent of a U.S. Geological Survey (USGS) third field accounting unit 

(Seaber et. al., 1987), generally the largest of the 

waterbodies, such as the North Umpqua River 

Basin, considered in the report. The term “watershed” refers to the equivalent of USGS fifth field 

cataloging units, a subunit of a river basin such as 

the Little River watershed. Successively smaller 

hydrologic units are referred to as “subwatersheds” 

(Wolf Creek subwatershed) and “drainages” (West 

Fork Wolf Creek drainage). 

LITERATURE REVIEW

The possibility of negative effects on streamwater quality from runoff of fertilizer nitrogen has 

long been recognized (Cole and Gessel, 1965). Forest fertilizer losses to streams and their effects on 

water quality have been studied often (Fredriksen 

et al., 1975; Moore, 1975; Bisson et al., 1992; Binkley and Brown, 1993a; Binkley et al., 1999). 

Water-quality criteria for nutrients have rarely been 

exceeded as a result of fertilization; however, few 

studies have examined the more subtle biological 

effects of fertilizer-nutrient inputs to streams. 

Table 1 provides an overview of data and findings 

of the relevant studies from the Pacific Northwest 

and several from other regions, and is referred to 

throughout this report. 

Forest Processes 

A large body of literature exists on forest fertilization, including proceedings from at least three 

conferences (Gessel et al., 1979; Lousier et al., 

1991; Chappell et al., 1992). However, most 

reports are directed at the efficacy of using fertilizers to enhance tree growth and nutrition (see also 

Haase and Rose, 1997), forest economics, soil processes, and fate of added nutrients in soils and 

trees. There are also over 25 reports worldwide on 

the effects of fertilization on water quality in 

receiving waters, and periodic reviews (Fredriksen 

et al., 1975; Moore, 1975; Bisson et al., 1992; Binkley and Brown, 1993a; Binkley et al., 1999). Of 

these reports, only three evaluated biological 

effects (Groman, 1972; Meehan et al., 1975; Stay et 

al., 1979), and none focussed on both soil processes 

and stream water or linkages between them (Binkley et al., 1999). 

Tree growth in the Pacific Northwest and 

many other locations is generally believed to be 

constrained by available nitrogen (Cole., 1979; 

Johnson, 1992; Fenn et al., 1998). For this reason, 

young (15–40-year-old) commercial forest stands 

are often fertilized with nitrogen, typically as urea 

[(NH2)2CO] pellets, although ammonium sulfate, 

ammonium nitrate, and various phosphate fertilizers also have been used (Klock, 1971; Tiedemann 

et al., 1978; Russel, 1979; Nason and Myrold, 

1992). Urea pellets (known as “prill”) consist of
3

Figure 1. Map showing Little River watershed, Oregon, and proposed Bureau of Land Management fertilization units.

123° 05' 122° 50'

43° 20'

123° 00' 122° 55' 122° 45' 122° 40'

43° 10'43° 05'43° 15'

0 1 2 3 4 MILES

0123 4 KILOMETERS

Cavitt

Fa

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Creek

Creek

Cr

Creek

Creek

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Blac W k hit N eegro

Emile

Creek

North Umpqua

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Little

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138

Glide
4

Table 1. Summary of studies of forest fertilization effects on stream-water quality [μg/L, micrograms per liter; vs, versus; %, percent; NH3, ammonia; NO3, nitrate; 1 kg/ha (kilogram per hectare), 0.89 lb/ac (pound per acre); 1kg Urea, 0.46 kg N; NR, 

nitrogen; NA, not applicable; d, day; ~, approximately, value is interpreted from report; <, less than; >, greater than; x, times; N, nitrogen; NO2, nitrite; Tot-N, total-N (s

invertebrates; chl a, chlorophyll a; mo, months; yr, year; mg/kg, milligrams per kilogram (parts per million); mg/ L, milligrams per liter; m, meter; km, kilometer; 

Geographical area and 

streams studied

Baseline 

nitrogen 

concentration

(μg/L)

Maximum 

posttreatment 

concentration 

(μg-N/L)

Estimated period 

and average 

magnitude of 

elevated 

concentration1,

vs control or 

baseline

% loss to 

streams

Biological 

components 

studied and results Sampling des

Location: Santiam Basin, Oregon; Application rate: 224 kg N/ha urea; Date: May 1969; Objectives: Effects of fertilization on water quality; Reference: MalueCrabtree Creek(control was upstream from study site)

NH3 <10 

NO3 <10 

TKN 400 

NH3 80 

NO3 250 

TKN 24,000

NH3 >100d, ~3x

NO3 >7 mo, 1.5–2x

TKN 2d, ~75x

NR NR

Also performed assays on urea pellets (N

NO3-N, 19.33 mg/kg; TKN, 440,000 m

during summer but peaked again in fall 

Location: Tahuya River, Kitsap Peninsula, WA.; Application rate: 227 kg N/ha; Date: October 1972; Objectives: Effects of fertilization on water Site 1, control(upstream)

NH3 10–80

NO3 0–200Urea 0–10

NH3 <10

NO3 470

Urea 50

NA NA NR

Water–quality data collected at all sites NO3 responded to flow. Conditions wer

Site 2, treatment (no buffer strip)

NH3 10–80

NO3 40–210

Urea 10–20

NH3 1,400

NO3 1,830Urea 27,000

NH3 25d, ~30–60x

NO3 ~7.5 mo, ~8x

Urea 6d, ~40x

NR NR

Lack of buffers contributed to increased site 3. NH3 peaks were also more imme

application. 

Site 3,treatment (buffer strip)

NH3 0–60

NO3 0–260

Urea 10–20

NH3 160

NO3 680Urea 4,300

NH3 2d, 10–40x

NO3 ~7.5 mo, ~3x

Urea 6d, ~40x

NR NR

NH3 peaks were delayed by more than a 

in fall.

Sites 4 & 5,downstream sites

NH3 0–80

NO3 0–350

Urea 0–30

NH3 60

NO3 470

Urea 40

NH3 ~31d, ~3–5x

NO3 ~31d, ~4x

Urea ~3d, ~2x

.45–1% NR

NH3 peaks were delayed by more than a 

in fall. Loss of nitrogen reported is for tupstream compared to control. 

Location: SE Alaska; Application rate: 210 kg urea-N/ha; Date: May 1970; Objectives: MCL, NH3 toxicity exceedances; Reference: Meehan et al., 1975Falls Creek, control NH3 ~20 

NO3 ~10 

NH3 ~100

NO3 ~200 NA NR Benthic inverts, algal biomass on Plexiglas 

slides. No significant differences found between treatment and control. High variability. No species data taken 

Application was to recently logged watemonth after application, weekly for secolow stream temperatures, average pH 6.summer. Phosphorus did not respond to 

Falls Creek, treatment NH3 ~20 

NO3 ~20 

NH3 1,280

NO3 ~1,600

NH3 ~ 1.5 mo, 10–20x

NO3 ~ 14 mo, 5–10x NR

Three Lakes Creek, control

NH3 ~20

NO3 ~20 

NH3 ~100

NO3 ~300 NA NR

Three Lakes Creek, treatment

NHc ~50 

NO3 ~10 

NH3 ~100

NO3 2,360

NH3 ~5d, ~3x

NO3 ~1.5 mo, >5x NR

Location: 6 locations in Pacific Northwest; Application rate: 224 kg urea-N/ha; Date: March-April, 1970–72; Objectives: not reported; Reference: Moore, 197Coyote Creek, South Umpqua Experimental Forest

NH3 5

NO3 2 Urea 6

NH3 48

NO3 177

Urea 1,390

NH3 ~5d, ~2x

NO3 ~2 mos, ~5–10x

Urea ~15d, ~10x 0.01% NR

After 3–6 weeks all loss of N was as NO

months, but NO3 had a second peak dur

92% of N lost during first year was duriwatershed area treated, old-growth mixe
5

Location: 6 locations in Pacific Northwest; Application rate: 224 kg urea-N/ha; Date: March-April, 1970–72; Objectives: not reported; Reference: Moore, 197Trapper Creek, Olympic National Forest

NH3 0

NO3 34Urea 8

NH3 10

NO3 121 

Urea 700

NR NR NR

<10% of watershed area treated, 40-yea

Jimmy-Come-Lately Cr., Olympic National Forest

NH3 0

NO3 5 

Urea 2

NH3 40

NO3 42

Urea 708

NO3 9 weeks NR NR

<10% of watershed area treated, 10-yea

Nelson Creek,Siuslaw River Basin

NH3 10

NO3 290

Urea <20

NH3 320

NO3 2100

Urea 8,600

NR NR NR

100% watershed area treated, young Do

Dollar Creek,McKenzie River Basin

NH3 30

NO3 60

Urea <20

NH3 490

NO3 130

Urea 44,400

NR NR NR

100% of watershed area treated, young 

Pat Creek,Yamhill River Basin

NH3 7

NO3 70

Urea 3

NH3 34

NO3 388

Urea 3,260

NR NR NR

63% of watershed area treated, 35-year-

Location: 25 Locations on 9 streams in Oakridge Ranger District, Willamette National Forest, Oregon; Application rate: 225 kg N/ha urea; Date: April 1976; selected chemical and biological aspects of streams; Reference: Stay et al., 1978, Stay et al., 1979Site 25, control NH3 5

NO3 5

TKN 87Urea ND

NH32 13

NO32 5

TKN2 63

Urea2 20

NA NR

Algal assays—no response; Periphyton—small increase in chl a but not biomass; Macroinverts—no direct change evident; Fish—no mortalities evident

By extending data collection through Juchanges from fertilization that were not Stay et al. (1978); these included small 

and differences in N-runoff between streSome increases were also found in specconcentrations. Algal assays using a greindicate colimitation by N and P. Stay ehelped minimize algal response to added tied to seasonal variability than to fertili

Treatments—24 sites (Ranges indicate reported concentrations from many sites)

NH3 5

NO3 5–10

TKN 47–100Urea ND

NH3 11

NO3 26

TKN 2,380Urea 8,000

NH3 no difference

NO3 ~1 yr, 1–3x

TKN <30d, <1–3xUrea <30d, ~1.5x

NR

Location: Vancouver Island, B.C.; Application rate: 200 kg N/ha; Date: November 1979; Objectives: Effects of fertilization on water quality in streams and doReference: Perrin et al., 1984 2 control streams NH3 <4

NO3 1–27

Urea <5

NH3 15

NO3 110

Urea 20

NA NA NR

Lower concentrations and longer transpbuffer strips than without buffer strips. Creduced nitrification resulting in longer 

baseline concentrations (relative to other 

Forest fertilization caused shift from N-lake, & algal blooms. 

12 sites on 10 streams draining 3 treatment watersheds entering a lake

NH3 <4

NO3 1–58

Urea <5

NH3 4,780

NO3 790

Urea 57,000

NH3 79–136d3NO3 4–84d3Urea 102–140d3

2.1–5.2% NR

Table 1. Summary of studies of forest fertilization effects on stream-water quality—Continued [μg/L, micrograms per liter; vs, versus; %, percent; NH3, ammonia; NO3, nitrate; 1 kg/ha (kilogram per hectare), 0.89 lb/ac (pound per acre); 1kg Urea, 0.46 kg N; NR, 

nitrogen; NA, not applicable; d, day; ~, approximately, value is interpreted from report; <, less than; >, greater than; x, times; N, nitrogen; NO2, nitrite; Tot-N, total-N (s

invertebrates; chl a, chlorophyll a; mo, months; yr, year; mg/kg, milligrams per kilogram (parts per million); mg/ L, milligrams per liter; m, meter; km, kilometer; 

Geographical area and 

streams studied

Baseline 

nitrogen 

concentration

(μg/L)

Maximum 

posttreatment 

concentration 

(μg-N/L)

Estimated period 

and average 

magnitude of 

elevated 

concentration1,

vs control or 

baseline

% loss to 

streams

Biological 

components 

studied and results Sampling des
6

Location: Vancouver Island, B.C.; Application rate: 224 kg N/ha; Date: September 1974; Objectives: Effects of fertilization on water quality; Reference: HethTC, control NH3 0–131

NO3 0–10

Urea 0–20

NH3 61

NO3 300

Urea 540

NA NA NR

Previously fertilized in 1967 at 96 kg N

16M, control NH3 0–93

NO3 4–109

Urea 0–20

NH3 22

NO3 89Urea 10

NA NA NR

Previously fertilized in 1967 at 258 kg N

TF1 (Lens Creek), treatment40-year-old plantation

NH3 0–79

NO3 7–177

Urea 0–30

NH3 540

NO3 2,700

Urea 14,000

NH3 13d

NO3 >14 mos

Urea 6d

5.9% NR

No buffer strips. 46% of watershed area Previously fertilized in 1968 and 1972 anitrate. Fall rains in 1975 caused increas

TF2, treatment NH3 0–80

NO3 28–151

Urea 0–220

NH3 1,900

NO3 9,300Urea 790

NH3 15d3,4

NO3 ~14 mos, ~9x

Urea 14d3 14.5% NR

No buffer strips. 80% of watershed area Previously fertilized in 1967 at 96 kg Nrains in 1975 caused increases in nitrate contributed to higher N-loss compared t

L, downstream site(Receives combined flow from both TF1 and TF2)

NH3 0–119

NO3 38–215

Urea 0–23

NH3 360

NO3 720Urea 160

NH3 33d3NO3 ~14 mos3Urea 5d3

NR NR

Located ~2 km downstream from TF1. Ndelayed until November 1974 after first 

Location: Western Washington; Application rate: 224 kg urea-N/ha; Date: July 1980; Objectives: determine water-quality effects of annual fertilizations; RefHook Creek,“control”

NH3 3

NO3 262

Tot-N 113

NH3 25

NO3 268Tot-N 488

NA

1.9–9%

 NR

“Heavily fertilized within 3 yrs before s

Willow Creek,Treatment—annual application.

NH3 6

NO3 96Tot-N 91

NH3 159

NO3 458Tot-N 8,597

NH3 40d, ~5x

NO3 77d, ~1.5x Tot-N 77d, ~3x

 NR

 “Heavily fertilized within 3 yrs before first yr of study and annually afterwards 

Needle Creek,“Control”

NH3 77 

NO3 1,270 Tot-N 874

NH3 1,580 

NO3 2,000 Tot-N 4,400

NA NR NR

“Heavily fertilized within 3 yrs before s

Gate Creek,Treatment—65 kg N/ha

NH3 10 

NO3 1,232 Tot-N 1,168

NH3 186

NO3 2,310 Tot-N 9,595

NH3 40d, ~1.5x

NO3 > 7 mos, ~2xTot-N > 7 mos,~3x

 NR NR

“Heavily fertilized within 3 yrs of study”of study and annually afterwards (treatmregarded as cause of high N-export thro

Debris Creek6, 7,Treatment—224 kg N/ha

NH3 5

NO3 211 Tot-N 105

NH3 630 

NO3 1,570 Tot-N 4,380

NA NR NR

“Relatively little past fertilization” plus study and annual treatment afterwards.

Eleven Creek6, 7,Treatment—224 kg N/ha

NH3 2

NO3 131 Tot-N 44

NH3 752 

NO3 1,680 Tot-N 37,553

NA NR NR

“Relatively little past fertilization” plus study, and annual treatment afterwards. application. High loss may also have be

Table 1. Summary of studies of forest fertilization effects on stream-water quality—Continued [μg/L, micrograms per liter; vs, versus; %, percent; NH3, ammonia; NO3, nitrate; 1 kg/ha (kilogram per hectare), 0.89 lb/ac (pound per acre); 1kg Urea, 0.46 kg N; NR, 

nitrogen; NA, not applicable; d, day; ~, approximately, value is interpreted from report; <, less than; >, greater than; x, times; N, nitrogen; NO2, nitrite; Tot-N, total-N (s

invertebrates; chl a, chlorophyll a; mo, months; yr, year; mg/kg, milligrams per kilogram (parts per million); mg/ L, milligrams per liter; m, meter; km, kilometer; 

Geographical area and 

streams studied

Baseline 

nitrogen 

concentration

(μg/L)

Maximum 

posttreatment 

concentration 

(μg-N/L)

Estimated period 

and average 

magnitude of 

elevated 

concentration1,

vs control or 

baseline

% loss to 

streams

Biological 

components 

studied and results Sampling des
7

Location: Western Washington; Application rate: various (see below); Date: various dates in 1988; Objectives: Drinking water criteria, dissolved N toxicity, an

Reference: Bisson, 1988. 

Forks Creek 7Treatment averaged 207 kgN/ha, on 1/88 and 2/88

NH3 <20

NO3 30

TKN 80

NH3 40 

NO3 50

TKN 160 

NH3 40d, ~2x

NO3 2d, ~1.5x

TKN >30d ~2x

 NR NR

Project also tested if fertilization jeopardhatchery or caused algal fouling of watedays after application.

Spring Creek 7, Treatment averaged 130 kgN/ha, on 2/88

NH3 <20 

NO3 1,000

TKN 70 

NH3 <20 

NO3 1,500TKN 180 

NH3 no increase

NO3 >30d, ~1.5xTKN >15d, ~2x

 NR NR

Tributary to Forks Creek

 Silver Lake Basin,Hemlock and Sucker Creeks, Treatment—92 kgN/ha each

NH3 20 

NO3 800

TKN 300 

NH3 200 

NO3 800 TKN 1,500 

NH3 >100d, 3-4x

NO3 no increaseTKN no increase NR NR

History of fertilizer application every ~5 after each application. Current fertilizatieutrophic with extensive macrophyte be

 Ryderwood, Pair 1 7

(Campbell Creek)

 Treatment—92 kg N/ha

NH3 30 

NO3 90 

TKN 100

NH3 275 

NO3 580 

TKN 2,000 

NH3 >100d, ~2-5x

NO3 100d, ~2x TKN >100d, ~3x

 NR NR

Tributaries to Cowlitz River. No buffer “control” watershed was recent clear cu

“control” was higher than in “treatment

 Ryderwood, Pair 27

(Arkansas Creek)

 Treatment—92 kg N/ha

NH3 20

NO3 200 

TKN 100 

NH3 150 

NO3 600 TKN 3,750 

NH3 >100d, ~3x

NO3 ~75d, ~2xTKN >100d, 2x

NR

 pHs averaged 6.5–7.0, increased ~0.3 units

Paired locations on Arkansas Creek (buupstream=control, downstream=treatme

Location: Fernow Exp. Forest, W. Virginia; Application rate: 336 kg N/ha as ammonium nitrate plus 224 kg P/ha as triple superphosphate8; Date: April 1976

responses in streams, and cumulative downstream effects, tracked from 3 to 10 years; Reference: Helvey et al., 1989; Edwards et al., 1991

North and South Facing Watersheds

NO3 ~500

Ca 2 mg/L

NO3 ~10,000

Ca 10 mg/L

NO3 >10 yrs, >5x

Ca >3 yrs, ~3xMg >3 yrs, ~3x

N9 23–27%P–<1%

NR

Specific conductance increased from ~2

remained high after 3 yrs, back to backg

remained ~40% higher than in control sbackground after 10 yrs except in one wwere around 5.0. No apparent changes i

Location: Western Washington; Application rate: 224 kg N/ha; Date: various in 1988–89; Objectives: Not reported; Reference: Bisson et al., 1992Louse Creek(western Cascades, second growth Douglas-fir)

NH3 ~30 

NO3 ~120 

TKN ~100

NH3 ~800 

NO3 ~1000 

TKN 80,000

NH3 >30d, 5-20x

NO3 >90d, 5-10xTKN ~4d, 10-100x

NR NR

Fertilized April 1989. Virtually all of wafter 90d, at onset of summer.

Ludwig Creek (Coast Range, second growth Douglas-fir)

NH3 ~20 

NO3 ~600

TKN ~200

NH3 ~400 

NO3 ~4,000 

TKN 50,000

NH3 >60d, 3-10x

NO3 >90d, 2-5x

TKN >7d, 10-100x

NR NR

Fertilized December 1988. Virtually all more protracted release of N from fertilRange may have higher N deposition ra

Table 1. Summary of studies of forest fertilization effects on stream-water quality—Continued [μg/L, micrograms per liter; vs, versus; %, percent; NH3, ammonia; NO3, nitrate; 1 kg/ha (kilogram per hectare), 0.89 lb/ac (pound per acre); 1kg Urea, 0.46 kg N; NR, 

nitrogen; NA, not applicable; d, day; ~, approximately, value is interpreted from report; <, less than; >, greater than; x, times; N, nitrogen; NO2, nitrite; Tot-N, total-N (s

invertebrates; chl a, chlorophyll a; mo, months; yr, year; mg/kg, milligrams per kilogram (parts per million); mg/ L, milligrams per liter; m, meter; km, kilometer; 

Geographical area and 

streams studied

Baseline 

nitrogen 

concentration

(μg/L)

Maximum 

posttreatment 

concentration 

(μg-N/L)

Estimated period 

and average 

magnitude of 

elevated 

concentration1,

vs control or 

baseline

% loss to 

streams

Biological 

components 

studied and results Sampling des
8

Location: Central Sweden; Application rate: 150 kg N/ha as ammonium nitrate; Date: August, 1986; Objectives: Simulate acidification from N-deposition, aninvertebrates, and fish; Reference: Göthe et al., 1993Orranstjärnbäcken NH3 5

NO3 15

Ca 1.8 mg/L

NH3 11,100

NO3 8,200 

Ca 3.1mg/L

NH3 >1 yr, 2-4x

NO3 2yrs, ~10x Ca 14d, ~2x,

NR

No long-term change in 

invertebrate species 

abundance, but there 

was an increase in drift 

in both streams, especially at furthest downstream stations. No 

mortality or density 

effects on fish.

Rödtjärnbäcken was fertilized with calceffects noted in either stream. 

Rödtjärnbäcken NH3 5 

NO3 10

Ca NR

NH3 15,400 

NO3 29,600 

Ca NR

NH3 >14d, >3-20x

NO3 2 yrs, ~6-7x 

Ca 14d, ~3x NR

1Concentrations expressed as a relative change in the active nutrient or ingredient, per liter.2Concentrations reported are averages rather than maximums.3Average concentrations not reported.4No streamflow during fertilization at TF2. Ammonia-N and nitrate-N concentrations had peaks attributed to fertilization in October and November 1974 after rain5Baseline concentrations calculated from Bisson (1982) by C.W. Anderson, USGS, 1999.6Debris Creek and Eleven Creek are paired treatment watersheds, with no control watershed.7“Control” concentrations are baseline concentrations in the same stream prior to fertilization.8Calcium phosphate.9N loss of 27% includes estimated loss in ground water.

Table 1. Summary of studies of forest fertilization effects on stream-water quality—Continued [μg/L, micrograms per liter; vs, versus; %, percent; NH3, ammonia; NO3, nitrate; 1 kg/ha (kilogram per hectare), 0.89 lb/ac (pound per acre); 1kg Urea, 0.46 kg N; NR, 

nitrogen; NA, not applicable; d, day; ~, approximately, value is interpreted from report; <, less than; >, greater than; x, times; N, nitrogen; NO2, nitrite; Tot-N, total-N (s

invertebrates; chl a, chlorophyll a; mo, months; yr, year; mg/kg, milligrams per kilogram (parts per million); mg/ L, milligrams per liter; m, meter; km, kilometer; 

Geographical area and 

streams studied

Baseline 

nitrogen 

concentration

(μg/L)

Maximum 

posttreatment 

concentration 

(μg-N/L)

Estimated period 

and average 

magnitude of 

elevated 

concentration1,

vs control or 

baseline

% loss to 

streams

Biological 

components 

studied and results Sampling des
9

46% N, and are usually applied at a rate of 224 kg 

N/ha (200 lb N/ac). This rate represents several 

decades of refinement, providing a reasonable economic tradeoff of tree growth to N-loss (Miller and 

Fight, 1979; Mika et al., 1992) through various processes (described below). 

After application, nitrogenous fertilizers 

undergo numerous reactions, and substantial literature has identified and quantified these reactions, 

including N-incorporation into trees and (or) Nlosses. Some of those reactions are summarized 

here, with an emphasis on urea fertilizers and processes that could affect stream water quality.

Reactions of applied urea-N include volatilization as ammonia, foliar uptake, hydrolysis to 

ammonium-N (NH4

+-N)1, rapid immobilization 

into soil organic fractions, ion exchange, mineralization, nitrification, plant (root) uptake, denitrification, and leaching to deeper soils and ground 

water (fig. 2). The fraction of NH3-N or urea-N not 

initially volatilized or taken up into foliage (including trees) is usually hydrolyzed to NH4

+-N. Subsequently, NH4

+-N can be immobilized by conversion to soil organic material or through ion 

exchange (Matzner et al., 1983; Edwards et al., 

1991; Moldan and Wright; 1998), taken up by tree 

roots, or converted to NO3-N (nitrified). Hydrolysis 

may temporarily increase soil pH, which can in turn 

enhance nitrification (Ochtere-Boateng, 1979). 

Nitrate is relatively mobile and can be rapidly 

leached into shallow or deeper ground-water systems, where it can be utilized by root systems or 

exit the forest through ground water or runoff processes. Mineralization is the conversion of soil 

organic nitrogen to inorganic N. 

The amount of urea-N entering trees, through 

foliar and plant uptake, ranges from 10 to 30% of 

that applied (Binkley, 1986; Nason and Myrold, 

1992). Thus, up to 70–90% is potentially retained 

in forest soils or other vegetation, volatilized, or 

passed from the forest ecosystem (Nason and 

Myrold, 1992), all of which are considered losses 

of applied fertilizer. These loss terms are highly

Figure 2. Forest cycling pathways representing major processes 

and fates of nitrogen fertilizers. (Modified from Nason and Myrold, 

1992.)

variable, but the largest are usually volatilization 

(<1–46%) (Craig and Wollum, 1982; Marshall, 

1986, 1991; Nason and Myrold, 1992) and immobilization (averaging about 50–60%) (Nason and 

Myrold, 1992). Much of the immobilized N 

apparently remains unavailable (Miller, 1986), 

so mineralization is likely small. The amount lost to 

streams, a topic explored in the following section, 

generally is less than 10% of applied N but can be 

larger (up to 27%) (table 1) under certain soil and 

moisture conditions. Denitrification is usually 

insignificant (Nason and Myrold, 1992), although in 

anoxic, saturated riparian soils, denitrification can 

be several orders of magnitude higher than in unsaturated upland soils (Cirmo and McDonnell, 1997). 

The relative extent of reactions listed above are 

highly dependent on the type of fertilizer applied. 

Nitrate-based fertilizers can have substantially different loss rates compared with the urea-based rates 

given above (Marshall, 1986, 1991). The organiccarbon content of soils and pore-waters is also a 

critical determinant of N-retention and transformations, both in upland areas as well as riparian 

regions (Cirmo and McDonnell, 1997; Dahm et. al, 

1998; Chestnut and McDowell, 2000).

1In this report, the term NH3-N is used in a general sense 

to represent the sum of free and ionized ammonia concentrations. 

Ionized ammonia, also known as ammonium, is referred to specifically as NH4

+-N when it is intended to mean only the ammonium 

(ionized) form. 

1. Volatilization

2. Foliar uptake

3. Hydrolysis

 4. Immobilization

5. Mineralization

6. Ion exchange

7. Nitrification

8. Plant uptake

9. Leaching to ground water

10. Denitrification

10 10 10

9

6

7

8

8

3

5

4

2 1

Soil

organic

N

NH3

NO3 NO2 N2O NO2

NH4

+

Ca2+, Mg+
10

The timing of urea application to forests is 

partly determined by efforts to minimize losses, 

especially volatilization and immobilization, the 

largest loss components. Volatilization is enhanced 

by higher temperatures, wind, and soil pH, and 

by low intensity rainfall. High intensity rainfall 

(though not so high as to significantly increase 

erosion and runoff) can reduce volatilization. 

Immobilization, which includes conversion of 

both fertilizer N and soil inorganic-N into organic 

forms, is minimized by cool, moist weather (Nason 

and Myrold, 1992). As a result, urea is typically 

applied during fall in the Pacific Northwest as 

weather conditions become cool and wet. (Mika et 

al., 1992). 

Water-Quality Effects on Streams from Forest 

Fertilization

Immediate Nutrient Runoff

All studies in which fertilizers were applied 

by helicopter reported some violation of stream 

buffers, direct applications to water, or inadvertent 

fertilization of control watersheds. Aerial application of fertilizer to forests is not exact, although 

recent advances in navigation using global positioning systems (GPS) should enhance the ability 

of pilots to minimize unintended application to 

areas not intended for fertilization. Actual application rates using helicopters, as measured on the 

ground, can vary by 20–60% from the targeted rate 

(Fredriksen et al., 1975; Binkley, 1986). Direct 

applications to water have invariably resulted in 

immediate and relatively high-concentration pulses 

of urea-N (often measured as organic- or KjeldahlN) and NH3-N (table 1). Peak urea-N concentrations usually exceed 1,000 μg/L (micrograms per 

liter) and have been reported as high as 80,000 

μg/L (Bisson et al., 1992). Peak NH3-N concentrations have been as high as 4,780 μg/L (Perrin,

et al., 1984). In the latter case, ammonia toxicity 

criteria concentrations (U.S. Environmental Protection Agency, 1986) might have been temporarily 

exceeded (although temperature and pH data 

required to make this assessment were not 

reported). With this exception, however, ammonia 

toxicity problems have not been reported, though 

ammonia toxicity is an obvious consideration and 

hence NH3-N almost always is monitored. 

Typically, pulses of urea-N and NH3-N 

decline in concentration and are short-lived following fertilization, usually lasting less than 1 month 

and often just a few days (table 1), depending on 

rainfall conditions. Maximum periods of elevation 

for urea and NH3-N have been more than 100 days 

(Malueg et al., 1972; Perrin et al., 1984). Extreme 

cases (several months or years) (Bisson, 1982; 

Edwards et al., 1991) were reported in instances 

where applications were in forests previously perturbed by forest management or excessive N-deposition. Reductions in urea and NH3-N are generally 

coincident with soil nitrification, when their supply 

available for runoff becomes greatly diminished, 

often to pretreatment or control levels (Bisson et 

al., 1992; Binkley et al., 1999). 

Longer-Term Nutrient Losses to Streams

Concentrations of nitrate-N (NO3-N) in 

streams typically remain low immediately after 

urea application, but increase rapidly during subsequent rainstorms as nitrification proceeds. Peak 

concentrations (table 1) have ranged from less than 

100 μg/L (for example, Fredriksen et al., 1975) to 

greater than 9,000 μg/L (Hetherington, 1985). In 

two studies, peak NO3-N concentrations exceeded 

EPA drinking water standards (10,000 μg/L); however, these studies were at the Fernow Experimental Forest in West Virginia (Helvey et al., 1989; 

Edwards et al., 1991) and in Sweden (Göthe et al., 

1993), where nitrogen-saturated soils associated 

with excess atmospheric nitrogen deposition are 

well-documented (Vitousek et al., 1997; Fenn et 

al., 1998). 

Whereas peak NO3-N concentrations occur 

chiefly during high runoff, ambient NO3-N concentrations in streams draining fertilized watersheds 

also are typically elevated as much as two- to tenfold, often for the entire winter and spring following a fall fertilization (Bisson et al., 1992). In most 

cases, however, NO3-N concentrations essentially 

return to background levels by summer due to 

uptake in soils and (possibly) uptake in stream 

water (Mulholland, 1992; Mulholland and Rosemond, 1992). Where sampling continued beyond 

the summer following fertilization into subsequent 

fall and winter periods (Malueg et al., 1972; 

Moore, 1971; Fredriksen et al., 1975; Meehan et 

al., 1975; Stay et al., 1979; Hetherington, 1985), a 
11

fall NO3-N peak, which is elevated relative to that 

in control streams, has been observed. This secondary peak indicates that applied fertilizer nitrogen 

remains available for leaching to streams beyond 

the spring and summer growing seasons. An 

extreme example of long-term availability was 

reported at Fernow Experimental Forest, where 

NO3-N remained elevated relative to control 

streams 10 years after fertilization with ammonium 

nitrate at 336 Kg N/ha (Edwards et al., 1991). 

Aside from the Fernow experiment and a long term 

watershed acidification study in Maine (Norton et 

al., 1994; Fry et al., 1995), no studies have followed fertilizer-nitrogen runoff for more than 1–

1.5 years.

The amount of applied nitrogen lost to 

streams varies from less than 0.5% to 14.5%, with 

the exception of Fernow Experimental Forest, 

where losses to streams were as high as 27% (table 

1). Most of the losses occur as NO3-N, extending 

over a protracted period, although in some cases 

immediate losses of urea-N or NH3-N accounted 

for as much as 50% of the loss of applied N 

(Moore, 1975; Perrin et al., 1984)

Reasons for differences among studies in 

losses of applied N and concentrations of NO3-N in 

streams vary, but often appear to be related to landmanagement history, nitrogen status of soils, or differences in the fertilizer applications (Bisson et al., 

1992). In a study originally intended to investigate 

the importance of repeated fertilizations, Bisson 

(1982) evaluated N-runoff to streams in six fertilized watersheds, four of which had been “heavily” 

fertilized (rates not reported) less than 3 years prior 

to the study. Each stream had NO3-N concentrations prior to the study fertilization higher than 

might be expected (100–1,200 μg/L) for forested 

streams in the Cascades, and even greater NO3-N 

concentrations after fertilization. “Extensive” prior 

fertilization was considered the cause of high Nexport (Bisson, 1982). One cause of this increase 

may be that fertilization stimulates growth of nitrifying bacteria in soils such that nitrification is 

increased if fertilizations are repeated within a few 

years (Bisson, 1982; Miegroet et al., 1990; 

Johnson, 1992). Enhancement of soil nitrification 

can also occur simply from large increases in atmospheric N-deposition (Fenn et al., 1998). Others 

have noted long-term increases in soil-N availability from single or multiple fertilizations (Binkley 

and Reid, 1985; Prescott et al., 1995; Norton et al., 

1994; Moldan and Wright, 1998) resulting from 

increases in N recycling, nitrification, or mineralization. In an experiment at a midwestern agricultural field, using isotopically enriched N, Wilkison 

et. al (2000) found applied fertilizer N in runoff for 

several years following fertilization. Additional 

fertilizations in succeeding years would be 

expected to increase leaching of N to ground water. 

From these studies it is apparent that a variety of factors related to the nitrogen status of soils 

determine a watershed’s response to urea fertilization, just as many factors are considered in predicting the response of tree growth from fertilization 

(Klinka, 1991; Carter, 1992). For watersheds, these 

factors include not just the N-pool but the amount 

of organic material, nitrification potential, cation 

exchange capacity (Mitchell et al., 1996; Edwards 

et al., 1991), tree types and stand ages, N-deposition rates, previous fertilization history, precipitation quantity and timing, and others. In watersheds 

in the eastern USA, with high N-deposition rates 

and N-saturated soils (Aber et al., 1989; Stoddard, 

1994; Fenn et al., 1998), it is not surprising that Nfertilization resulted in stream NO3-N concentrations that nearly violated water-quality standards. 

Such results are much less likely in the western 

Cascades, where N-limitation in forest soils, and 

probable uptake in the largely N-limited forest 

streams (see below), help maintain ambient stream 

NO3-N concentrations well below existing nutrient 

standards. 

Water-Quality Criteria

Comparison of stream nutrient concentrations resulting from forest fertilization with currently available criteria is arguably insufficient to 

evaluate fertilizations effects, particularly in 

streams in Cascade streams of the Pacific Northwest, where primary production is often limited by 

the supply of nitrogen in water (Triska et al., 1983; 

Gregory et al, 1987; Bothwell, 1992; Borchardt, 

1996). It is now well established that NO3-N concentrations resulting from forest fertilization rarely 

exceed the U.S. Environmental Protection 

Agency’s (EPA’s) 10 mg/L (milligram per liter) 

standard, and ammonia toxicity has rarely been 

observed (Bisson et al., 1992; Binkley et al., 1999), 

despite sometimes high concentrations (Göthe et 
12

al., 1993). However, the NO3-N standard is targeted towards human health protection in drinking 

water and is not intended to protect against ecosystem degradation. Furthermore, there are no criteria 

for phosphorus in streams, despite evidence that 

phosphorus concentrations exceeding 25 μg/L in 

lakes can produce eutrophic conditions (Welch, 

1992), and in streams similar concentrations may 

result in nuisance growth of periphyton (attached 

algae) (Dodds et al, 1997, 1998; Correll, 1998). 

The EPA has recently developed guidelines 

for the States to use in setting regional nutrient criteria (U.S. Environmental Protection Agency, 

2000a). The nutrient levels suggested to prevent 

nuisance conditions, however, are considerably 

lower than the existing criteria targeting drinking 

water and ammonia toxicity. For instance, upper 

limits for total nitrogen (TN), dissolved inorganic 

nitrogen (largely NO3-N and NH3-N), and total 

phosphorus could be approximately 650, 400, and 

38 μg/L, respectively (U.S. Environmental Protection Agency, 2000a). For the Cascades subecoregion, the suggested values are even lower— 

approximately 55, 5, and 9 μg/L, respectively 

[U.S. Environmental Protection Agency, 2000b]. 

These values are one to several orders of magnitude 

less than existing criteria and well below concentrations frequently observed following fertilization. 

However, few criteria have actually been set, and 

it will be left to States to do so, meaning that the 

establishment of criteria is likely to happen several 

years in the future, and will be variable among 

States and ecoregions. Thus, rather than using 

drinking-water criteria as a basis for decision making for forest streams, it might be more relevant at 

this point to increase attention on the ecological 

consequences of fertilization.

Ecological Effects on Streams from Forest 

Fertilizations

Many researchers have expressed the need 

for more direct investigation of fertilizer effects on 

stream ecology (Groman, 1972; Bisson, 1988; Binkley et al., 1999), though few studies (Groman, 

1972; Meehan et al., 1975; Stay et al., 1979; Göthe 

et al, 1993) actually examined any biological 

responses. These studies indicated that toxic effects 

to aquatic invertebrates and fish are unlikely. However, broad conclusions about the ecological effects 

of N-fertilization are tenuous because (1) conditions specific to individual studies may not be 

applicable in other settings, (2) sampling techniques may have been insensitive to the questions 

being asked, or (3) the understanding of hydrological and biological processes available at the times 

of the studies were limited and precluded examination of subtle, yet potentially important ecological 

processes. In some cases, changes in nutrient inputs 

were measured in ephemeral streams too small to 

support substantial increases in algal biomass, and 

(or) algal growth might have been limited more by 

light availability than nutrient concentrations. In 

these cases, added nutrients would be transported 

downstream to larger, less light-limited reaches, 

where complex processes of nutrient uptake by 

benthic organisms (periphyton or bacteria) and 

hyporheic processing might greatly reduce water 

column concentrations. Some streams, particularly 

those like the Little River in the western Cascades 

(fig. 1), might be especially susceptible to negative 

effects because of local physical and water-quality 

conditions or combined effects from other upstream 

land use. Results from the few studies with biological components are discussed below with respect to 

their applicability to the Little River watershed. 

Following spring fertilization of two previously clearcut watersheds (total 1,500 acres) in 

Alaska, Meehan et al. (1975) found no increase in 

periphyton biomass on Plexiglas slides, and no difference from natural variation in stream invertebrates despite five- to tenfold increases in stream 

NO3-N concentrations (table 1). Although these 

results indicate minimal biological effects from fertilization, biomass of periphyton growth alone may 

not adequately represent changes in algal community due to high variability and complicating 

effects from invertebrate grazing, scour, or changes 

in taxonomic composition. (Stevenson, 1996a). The 

authors did not report data on algal special composition, which can indicate differences in water quality (Lowe and Pan, 1996), so there is no way to 

assess whether changes in nutrient concentrations 

changed the algal communities. Also, artificial substrates such as Plexiglas slides are often poor indicators of natural periphyton because they tend to 

underestimate growth of green and blue-green 

algae (Cattaneo and Amireault, 1992), which are 

common in many Cascade streams. Neither was the 
13

potential for nutrient or light limitation reported. 

The first- and second-order study streams had very 

low flow (streams in one subbasin were dry during 

summer), so primary production in these ephemeral 

streams was likely insignificant anyway. Furthermore, no information was provided on downstream 

sites. Finally, fertilization took place in spring, to 

clearcuts, as opposed to fall applications to 10-40 

year-old stands, the current operational practice. As 

a result, nutrient dynamics in the forest floor and 

hydrologic conditions that contribute to nutrient 

runoff likely differed from those in locations like 

the Little River watershed.

After fertilization in the Cascade foothills of 

the Santiam River Basin, Oregon, Stay et al. (1979) 

found colimitation of Crabtree Creek water by N 

and P, using algal bioassays with the planktonic 

green algae Selenastrum capricornutum. There was 

no response to added N alone (table 1). Although 

valuable, laboratory demonstrations of colimitation 

in flasks by S. capricornutum, a planktonic green 

alga of European origin that lives in still or slow 

moving water, have limited transferability to Cascade streams dominated by periphyton. Reasons for 

this lack of transferability include the enhancing 

effect of moderate water velocities on periphyton 

metabolism (Stevenson, 1996b) and algal-grazer 

interactions (Steinman, 1996). In the same study, 

variation in populations of benthic invertebrates 

before and after treatment were indistinguishable 

due to high natural variability, and fish assays 

showed no mortality due to fertilization. The invertebrate studies, however, were designed to detect 

short-term changes from toxicity and could have 

been confounded by flow change. However, the 

study’s approach would have been insensitive to 

longer-term shifts in invertebrate assemblages. 

Finally, calculations of invertebrate diversity indicated already perturbed conditions prior to treatment, suggesting that sensitive invertebrate species 

may have already been eliminated, and remaining 

species may have been insensitive to additional 

effects from fertilization. 

Göthe et al. (1993), after fertilizing parts of 

two watersheds to evaluate stream acidification, 

found a temporary increase in drift of benthic 

invertebrates that was attributed to high concentrations of un-ionized ammonia. Drift was highest at 

the downstream stations, indicating cumulative 

upstream effects, and daytime drift was as large as 

nighttime drift, which may have rendered invertebrates more susceptible to daytime predation. No 

effects on fish density or mortality were noted. 

Though this study suggests some possible shortterm, toxicity-based effects for invertebrates, it did 

not address longer-term changes that could result 

from more indirect changes in habitat and food 

quality due to altered nutrient regimes. The authors 

also did not evaluate primary producers, which 

would have been directly affected by nutrient additions from fertilization.

Some investigators acknowledge that 

increases in stream primary production from fertilization may occur (Binkley et al., 1999), occasionally postulating that such increases could stimulate 

food web changes that enhance fish production 

(Fredriksen et al., 1975; Malueg et al., 1972; Harriman, 1978; Hetherington, 1985; Bisson et al., 

1992). In parts of an oligotrophic lake in British 

Columbia, nitrogen input from upstream forest fertilization with urea was sufficient to cause a temporary shift from nitrogen limitation to phosphorus 

limitation and an increase in plankton biomass 

(Perrin et al. 1984). Thus, the potential for ecological modification has long been recognized, and 

there is some evidence of its having occurred in 

certain instances. No investigators, however, have 

followed the movement of applied fertilizer nitrogen from the forest floor, through soil profiles and 

ground-water regions, to streams and into aquatic 

food webs, so the link between terrestrial and 

aquatic processes remains poorly understood 

(Binkley et al., 1999). 

Stream Nutrient Dynamics

The processing of nutrients in streams is 

complex, and mostly beyond the scope of this 

review. However, advances in the understanding of 

ecological processes and several recent conceptual 

developments are critical to understanding the possible effects of fertilization on streams, and potentially to investigating those effects in streams of the 

Cascade Mountains. These concepts include nutrient speciation and limitation, nutrient uptake by 

stream algae, hyporheic processing of nutrients, 

algal indicators of environmental changes, food 

web interactions between primary producers and 
14

higher trophic levels, and the use of stable isotopes 

as tracers of nutrient transfer in aquatic food webs.

Stream nitrogen budgets, though complex to 

measure, have been determined in several forested 

streams. In the Pacific Northwest, almost all have 

been in the Andrews Experimental Forest, in the 

Cascade Range of Oregon. In all cases, organic 

nitrogen has been an important component of both 

N-input and output in undisturbed streams. Sollins 

et al. (1980) found that nitrogen was biologically 

limiting in an old-growth watershed at the Andrews 

Experimental Forest, and that DON or NH3-N were 

the dominant forms of N in solution annually. In 

the same watershed, Triska et al. (1984) found that 

over 96% of the nitrogen leaving the outlet stream 

on an annual basis was as organic nitrogen, with 

77% as DON, whereas less than 4% was as NO3-N. 

Similarly, DON was the largest component and 

NO3-N the smallest (95% and 0.2%, respectively) 

of annual N-export in temperate, old-growth forests 

in Chile (Hedin et al., 1995). Overall, the relative 

amount of organic nitrogen (dissolved and particulate) output from subbasins in the Andrews Experimental Forest ranges from 28% to 85%, with the 

highest DON proportions occurring in fall (K. 

Vanderbilt, Oregon State University, unpub. data, 

1999). Interestingly, NH3-N export is roughly 

twice that of NO3-N in undisturbed watersheds in 

the Andrews Experimental Forest, whereas NO3-N 

export is greater in disturbed (though unfertilized) 

watersheds (K. Vanderbilt, Oregon State University, unpub. data, 1999). Wondzell and Swanson 

(1996) found that a conifer-dominated floodplain 

was the largest source of nitrogen to fourth-order 

McRae Creek, and more than 50% of that nitrogen 

was as DON. In contrast, inorganic nitrogen made 

up over 50% of the nitrogen entering McRae Creek 

through an alder-dominated gravel bar, with most 

of the remainder being as DON. 

One importance of N-speciation is related to 

the types of algae able to utilize that nitrogen, and 

the possible shifts in community structure that 

might occur if the amount or form of nitrogen 

changes. For instance, DON is usually assumed to 

be biologically unavailable (Sollins and McCorison, 1981). However, some diatom species are able 

to utilize DON for energy (heterotrophy) and (or) 

for nutrition, although this process is relatively 

inefficient because energy is expended in breaking 

down organic molecules to liberate energy and 

reduce N (Hellebust and Lewin, 1977). Inorganic 

nitrogen is more easily assimilated than DON by 

most algae, particularly green algae. NH3-N assimilation is the most energetically favorable but NO3-

N can readily be reduced as well. Where available 

nitrogen is scarce, algae that can fix elemental 

nitrogen (N2) from air or water have a competitive 

advantage and are often more abundant than in 

nitrogen-replete waters (Biggs et al., 1998). However, the importance of DON as a source of N to 

algae also may be elevated in these situations (Mulholland, 1992). N-fixing algae are typically bluegreen species but can also include certain diatom 

species containing cyanobacterial inclusions 

(Floener and Bothe, 1980). In the North Umpqua 

and Little River watersheds, the colonial bluegreen alga, Nostoc, and the diatom Epithemia 

Sorex, are commonly observed in nitrogen-poor 

environments (Anderson and Carpenter, 1998). 

Hence, it is reasonable to hypothesize that a 

fertilization-induced spring/summer shift of the 

predominant dissolved nitrogen form, to NO3-N 

from DON or NH3-N, could induce a shift in algal 

community structure. The algal community might 

change from heterotrophic and N-fixing species to 

a community dominated by non-N-fixing diatoms, 

non-heterocystous blue-greens, and possibly filamentous green algae. Such shifts in community 

were observed after NH4

+-N addition to a nitrogen 

limited, fifth-order stream (Lookout Creek) in the 

Andrews Experimental Forest, despite relatively 

unchanged algal biomass in the enriched sections 

(Lundberg, 1996). In that study, the changes in 

algal assemblages due to N enrichment were 

hypothesized to affect invertebrate grazers because 

certain species were known to prefer the epithemiacean (N-fixing) diatoms present prior to enrichment. Recent efforts to model benthic algal and 

invertebrate processes indicate that an increase in a 

limiting nutrient that causes a decrease in algal 

food quality can indirectly affect various invertebrate functional groups (McIntire et al., 1996).

A decrease in N concentrations in surface 

water during summer does not necessarily confirm 

that fertilizer N is not entering streams, nor does it 

confirm that benthic communities are unaffected. 

Concentrations of stream NO3-N could be reduced 

to low levels during summer because of increased 

plant uptake in upslope areas. Yet periphyton 

uptake can also increase nitrogen retention in 
15

streams (Kim et al., 1992; Peterson et al., 2001), 

and hence reduce dissolved-nitrogen concentrations. Thus, if dissolved-N input during summer is 

through the riverbed as ground water is important, 

subsequent algal uptake and spiraling (Newbold et 

al., 1981) is likely to reduce the ability to measure 

the added nitrogen in the water column (Peterson et 

al., 2001), even though it may be transported downstream by recycling in benthic layers or as sloughed 

algae. Additionally, added N in a hyporheic zone 

could increase heterotrophic metabolism (Mullholand et al., 1997, 1999; Storey et al., 1999) while 

obscuring the increased N input in streams. 

Hyporheic Processing

In recent years, hydrologic exchange and 

nutrient processing in hyporheic zones (subsurface 

and riparian near-shore environments) along 

streams have received increased attention. Extensive syntheses of various hyporheic processes and 

findings have been published by Cirmo and 

McDonnell (1997), Boulton et al. (1998), Dahm et 

al. (1998), and Storey et al. (1999). Hyporheic 

function can be critical in determining hydrologic 

flow paths and nutrient exchange, as well as transformations of carbon and nitrogen. These aspects 

are discussed briefly below with respect to forest 

fertilization and its effects on aquatic systems. 

Interactions of ground water with streams can 

follow several types of paths, including gaining, 

losing, parallel flow, and through flow (Woessner, 

2000), each of which may be intermittently present 

depending on stratigraphic and fluvial character of 

individual reaches or streams (Stanford and Ward, 

1993). These flow paths have potentially diverse 

implications for ecological processes. In gaining 

streams the inflowing water is derived from shallow to regional ground-water flow and will likely 

increase net stream solute transport. Benthic processes will be partly dependent on the quality of 

incoming hyporheic water, which in turn will be 

dependent on physical conditions in stream margins 

and on the quality of the ground-water source. In 

losing streams, benthic processes will be more predictably dependent on surface-water chemistry and 

physical conditions, and will in turn control 

hyporheic metabolic process (Boulton et. al, 1998). 

Parallel flow occurs when head gradients between 

surface and ground water are not distinct, with 

intermittent exchange occurring in both directions 

between stream and hyporheic zones depending on 

local variations in channel gradient and bed material. In these cases, water in hyporheic zones and 

streams may be of similar quality, and net solute 

transport may be relatively unaffected by the 

exchange processes. Through flow is likely to 

occur between bends in alluvial-stream reaches 

as a short-circuiting of ground-water flow from 

upgradient to downgradient. 

Transformations of solutes in hyporheic 

zones are often controlled by the hydraulic residence time and the amount of organic material in 

those zones (Boulton et. al, 1998; Wondzell and 

Swanson, 1996). Transformations of nitrogen are 

complex, depending also on redox conditions and 

available oxygen and organic carbon (Cirmo and 

McDonnell, 1997), and are seasonally variable 

(Wondzell and Swanson, 1996) (fig 2). During winter and high-flow periods, temperatures and storage 

time are typically low, so nitrification of incoming 

DON (including urea or NH3-N) from upland areas 

is likely to be minimized. Nitrogen entering the 

stream will be a combination of DON, NH3-N, and 

NO3-N (possibly nitrified in upland terrestrial 

soil). As temperatures rise and streamflows 

decrease during spring and summer, nitrification 

will increase such that the predominant form of N 

entering the stream will be nitrate. Along with 

these warm weather transformations will be losses 

of nitrogen resulting from microbial uptake (Boulton et. al, 1998; Wondzell and Swanson, 1996) and 

possibly denitrification (Cirmo and McDonnell, 

1997), so the overall load of nitrogen entering the 

stream will be reduced compared to that in winter 

and early spring. If the hyporheic zone contains relatively high amounts of DOC, microbial metabolism and hence nitrification rates may be further 

enhanced, and denitrification can be enhanced in 

anoxic environments. 

The form, timing, and relative amount of fertilizer-derived nitrogen entering streams will therefore depend partly on hyporheic zone processing, 

which itself will depend on subsurface flow paths 

and degree of saturation in upland and riparian 

areas, as well as hydraulic conductivity, stratigraphy, and amount of organic material in riparian 

areas. In regions as geologically complex as the 

Little River watershed and the Wolf Creek subwa-
16

tershed, heterogeneity may reduce the ability to 

make broad spatial generalizations. Nonetheless, if 

summer streamflow has an important ground-water 

component, fertilizer nitrogen could stimulate 

microbial nitrification and (or) denitrification in 

hyporheic zones and benthic algal growth at the 

interface between stream and hyporheic inputs. 

Conceptual Model of Ecological Processing of Fertilizer 

Nitrogen

There are a variety of possible ecological 

effects of fertilizer nitrogen in Cascade streams, 

depending on local conditions of hydrology, water 

quality, stream morphology, and aquatic biota. For 

the Little River watershed, a hypothetical process 

scenario following fall-winter urea fertilization is 

illustrated in figure 3, p. 18–19. The lettered 

descriptions below correspond to the respective lettered parts of the figure. Pie charts indicate hypothetical relative stream nutrient concentrations 

(areas of circles) and speciation (pie slices). 

A. Following fall fertilization, rains cause immediate overland and (or) subsurface runoff of 

organic and ammonium nitrogen to small 

streams draining fertilized stands, most of 

which is rapidly transported to larger streams 

(>4th order) such as Little River, and further 

downstream to successively larger rivers. 

Streams draining reference areas remain 

higher in organic nitrogen than NO3-N (Sollins and McCorrison, 1981), and have lower 

nitrogen concentrations overall. 

B. Winter rains and nitrification of applied nitrogen enhance longer-term inputs of NO3-N to 

small streams, which also is mostly transported out of the fertilized stands downstream to the Little River and farther. Some 

sequestration of nutrients during the spring 

may occur in stream biota, other organic 

material, or in hyporheic zones and shallow 

ground water. 

C. During late spring and early summer, nitrate 

concentrations in streams affected by, and 

downstream of, fertilization decline in conjunction with declines in stream discharge. 

Reasons for these reductions include uptake 

by plants and trees in forest soils, denitrification or sequestration in hyporheic environments, and uptake by periphyton in streams 

(Mulholland, 1992; Mulholland and Rosemond, 1992). Dissolved organic nitrogen 

(DON) remains proportionally high (Sollins 

and McCorrison, 1981; Triska et. al, 1984), 

especially in unfertilized basins.

D. Additional input of fertilizer nitrogen during 

summer is reduced but may not be altogether 

eliminated. Shallow ground-water flow from 

upslope may transport NO3-N or DON to 

periphyton or benthic bacteria through riparian and hyporheic zones. Nutrients could be 

transported from smaller streams in fertilized 

stands, where algal growth may be light limited, to higher order reaches where more 

open canopies allow greater algal growth 

(Gregory et al, 1987). Regional ground-water 

flow patterns (Harvey and Bencala, 1993) 

might discharge water and nutrients from fertilized areas well downstream of fertilized 

stands. 

E. Algal biomass in small, lower-order streams 

remains low during summer because of light 

limitation, though changes in species may 

occur if nitrogen input is increased. Biomass 

downstream may increase, or algal species 

may change, as light availability increases. 

Benthic uptake continues to keep water column nitrogen concentrations low. Nuisance 

algal growth occurs in some places where 

substrate and light conditions are favorable, 

enhancing overall primary production.

The N-transport and transformation processes postulated here would not be identical in all 

forests, even within the Pacific Northwest. Some 

factors might make certain streams more sensitive 

than others to nitrogen inputs, predisposing them to 

perturbation. For instance, coastal mountain 

regions of the Pacific Northwest often have higher 

ambient nitrate concentrations than Cascade 

streams, possibly because of differences in geology, weather patterns, and the predominance of red 

alder (Alnus rubra) (Brown et al., 1973; Miegroet 

and Cole, 1988). In contrast, relatively high concentrations of available phosphorus, often found in 

streams of young volcanic origin (Dillon and 

Kirchner, 1975) such as the Cascades, provide adequate phosphorus (10–40 μg/L) necessary to grow 

periphyton (Bothwell, 1988; Borchardt, 1996), 

making periphyton communities more dependent 

on nitrogen supplies. In low alkalinity streams, a 

given algal productivity or biomass could cause 
17

larger fluctuations in pH and higher daily maximum pH than in streams with high alkalinity (Teal 

and Kanwisher, 1966; Beyers, 1970). Also, more 

stable beds composed of relatively large, stable 

substrates (ranging from cobbles to bedrock) can 

often accumulate higher algal biomass, because the 

algal mats are able to withstand higher flows that 

would scour attached algae in streams with less stable surfaces (Peterson, 1996). 

Enhanced diel fluctuations of pH and DO 

resulting from increased primary production may 

be compounded, in streams dominated by bedrock 

or otherwise armored, if hydrologic exchange or 

discharge through hyporheic zones also is reduced. 

Because these zones are important areas for nutrient transformations and heterotrophic activity in 

streams (Mulholland et al., 1997 & 1999; Naegeli 

and Uehlinger, 1997; Boulton et al., 1998; Chafiq 

et al., 1999), heterotrophic respiration may be 

reduced in streams with smaller hyporheic zones 

(Mulholland et al., 1997 & 1999), potentially 

allowing higher pH maxima. Although maximum 

DO concentrations could be increased if hyporheic 

respiration is reduced, DO in small, high-gradient 

streams appears to be reaerated more rapidly by 

physical processes than does carbon dioxide 

(Guasch et al., 1998), so stream DO might not be 

altered to the same extent as pH. In a study of the 

North Umpqua River, DO was apparently controlled by temperature and reaeration, despite diel 

pH changes that were indicative of primary production (Anderson and Carpenter, 1998).

The influences of upstream land-management 

practices could play a vital role in determining 

stream response to N-application. Nutrient concentrations (including NO3-N) are known to be elevated by many forestry operations (Tamm et al., 

1974; Sollins and McCorison, 1981; Tiedemann et 

al., 1988; Adams and Stack, 1989; Binkley and 

Brown, 1993a, 1993b), with various subsequent 

effects on stream productivity (Gregory et al., 

1987). Other site-dependent factors that could 

influence algal production include soil characteristics, ground-water flow paths, light availability, 

water temperatures, and the amount of invertebrate 

grazing. 

Suggested Approaches to Evaluate Ecological Effects of 

Forest Fertilization

Care will be needed to differentiate the 

effects on streams of fertilization from those of 

other land-management practices and from natural 

variability. The susceptibility of streams to 

eutrophication and other ecological effects will 

depend on numerous watershed conditions, including hydrology, geology and stream morphology, 

geochemistry, nitrogen status of soils and streams, 

canopy cover and aspect, and history of land use. 

Many forested watersheds in the West have already 

been perturbed by historical resource management. 

As a result, added fertilizer N could contribute to 

cumulative stimulatory effects of forestry on primary production in streams (Norris et al., 1991). 

Forest management practices that can increase 

nutrient runoff to streams include road construction 

and logging (Fredriksen et al., 1975; Sollins and 

McCorison, 1981; Binkley and Brown, 1993a, 

1993b; Brown and Binkley, 1994), history of fire or 

fire prevention (Brown et al., 1973; Tiedemann et 

al., 1978; Norris et al., 1991) and fertilization history (Bisson et al., 1992). Additional conditions 

that could compound the effects of fertilizer nutrient inputs include (1) increased sediment transport, 

which can contribute nutrients and scour channels, 

(2) increased temperature, which can enhance algal 

growth (DeNicola, 1996) and accelerate invertebrate hatches, possibly decreasing grazing pressure 

on algae, and (3) increased light penetration, where 

small buffer strips might cause a shift from light to 

nutrient limitation. 

The occurrence of several or all of the above 

conditions can contribute to possible cumulative 

effects on streams and make differentiation of the 

effects from an individual practice difficult to discern. Thus, streams already experiencing increases 

in eutrophication, manifested as increases in nutrient concentrations, changes in algal growth or algal 

populations, elevated temperatures, increased light 

penetration, or exacerbation of diel DO and pH 

cycles resulting from elevated primary production, 

may be poor locations to examine effects from forest fertilization. 

Despite these complications, there are several 

ways to determine, more definitively than in the 

past, whether fertilizer nitrogen is transported to 

streams during growing seasons or contributes to 

benthic production. Potential methods include (1) 

measurements of nutrient processing and transport 

in hyporheic zones, (2) assessments of benthic 

metabolism either in chambers (Bott et al., 1997; 

Harvey et al., 1998) or by measurements of wholestream metabolism (Marzolf et al., 1994, 1998;
18

UPPER WATERSHED

Stream order 1-2

Low flow stream width 

about 1 meter

MID WATERSHED

Stream order 2-3

Low flow stream width 

1-3 meters

LOWER WATERSHED

Stream order 3-5

Low flow stream width 

3-6 meters

Stream order >5

NH3 NH3

NH3

NH3

NH3

DON

(Urea)

VOL

DON

DON

GW

DON

DON

GW

LPON

LEA

RO

TRAN

DIL

FPON

DON

DON

NH3

DON

DON

DON

RO

DIL

NO3

NO3

NO3

UNFERTILIZED

FERTILIZED

NIT

UP

UP

NH3

NO3

NO3

DON, NO3

NO3

NIT

LEA

LEA

LPON

PER

Reference

Treatment

PROCESS

DEN Denitrification

DIL Dilution

GW Ground-water exchange/transport

LEA Leachate

NIT Nitrification

PER Percolation

REC Recycling

RO Runoff

TRAN Transport

UP Uptake

VOL Volatilization

NITROGEN SPECIES

 Dissolved organic nitrogen

 Fine particulate organic nitrogen

 Large particulate organic nitrogen

Nitrogen gas

Ammonia nitrogen

Nitrate nitrogen 

NOTE: Size of pie chart is proportional to concentrations

FALL-WINTER SCENARIO

DON

FPON

LPON

N2

NH3

NO3

%NH3

A, B

%NO3

%DON

Figure 3. Hypothetical transport pathways, dominant processes, and relative concentrations of nitrogen in response to urea 

explained in the text.)
19

LOWER WATERSHED

Stream order 3-5

Low flow stream width 

3-6 meters

Stream order >5

Light 

Limited

Abundant 

Light

Moderate

Light

UPPER WATERSHED

Stream order 1-2

Low flow stream width 

about 1 meter

MID WATERSHED

Stream order 2-3

Low flow stream width 

1-3 meters

Periphyton growth

NH3

DON

DON

DON

DON

DON

LPON

FPON

REC

FPON

DON

DON

NO3

NO3

NO3

NO3

NO3

N2

UNFERTILIZED

FERTILIZED

NIT

NO3

NO3

NO3

NO3

DEN

GW

GW

UP

NH3

NIT NO3

UP

UP

UP

UP

C

D

E

Reference

Treatment

LATE SPRING-SUMMER SCENARIO

 fertilization in forested catchments of the western Cascades, Oregon, during fall/winter and late spring/summer. (Parts A-E are 
20

Parkhill and Gulliver, 1998), (3) experiments using 

nutrient-diffusing substrates to determine growthlimiting factors (Fairchild et al., 1985; Pringle 

and Triska, 1996; Tank and Webster, 1998), 

(4) measurements of cellular-nutrient content and 

sloughing algal material in transport to more accurately account for N inputs and outputs in streams, 

(5) use of stable isotopes of nitrogen, including possible additions of labelled 15N in urea fertilizer (Kahl 

et al., 1993), to track incorporation of applied N into 

different algal and food-web compartments (Harvey 

et al., 1998); and (6) assessments of benthic algal 

communities using autecological evaluations of 

water quality (Lowe and Pan, 1996; Pan et al., 1996; 

Anderson and Carpenter, 1998) and (or) biomass in 

cumulative assessments longitudinally and temporally in stream systems. In addition, in sensitive systems, continuous monitoring of DO and pH during 

selected periods might provide data to determine 

whether production is increasing and whether that 

production is having potentially deleterious effects 

on stream ecosystems. Most of these methods will 

require similar evaluations in control or untreated 

streams for comparison to determine the magnitude 

of response to fertilizer treatment. 

LITTLE RIVER WATERSHED

This section presents a framework for evaluating an operational application of urea fertilizer by 

the Bureau of Land Management (BLM) to selected 

stands in the Little River watershed, a nitrogen-limited tributary to the North Umpqua River, Oregon 

(fig. 4). The previous literature review is used as a 

basis to explore possible ecological effects of fertilization in the context of known geological and 

hydrological conditions in the basin. Urea-based fertilizer would be applied during late fall to individual 

15–40 year old stands on BLM lands, at a rate of 

200 pounds N/acre (224 kg N/ha). The study would 

examine water-quality and possible ecological 

effects of added fertilizer nitrogen, including 

changes in biomass of periphytic algae, dissolved 

oxygen (DO) and pH in streams, shifts in algal species and community composition, and changes in 

secondary grazer (macroinvertebrate) communities 

or food-web structure and function resulting from 

shifts in algal community composition and biomass. 

Data from reconnaissance samplings are included to 

provide an indication of water-quality and algal conditions prior to fertilization. Various portions of private timberlands in the watershed were fertilized in 

1998 or in 1999, but were not sampled prior to fertilization.

Physiographic Setting

The Little River watershed (fig. 1) is approximately 206 square miles in area, with elevations 

ranging from 730 to 5,275 feet above sea level (U.S. 

Forest Service and Bureau of Land Management, 

1995). Most of the watershed (83%), including 

almost all of the land east of Cavitt Creek, is located 

within the western Cascades geologic province 

(McFarland, 1983). Soils in this province tend to be 

high in nutrient content, particularly phosphorus, as 

they are derived from many layers of volcanic rock. 

In many places, streams have eroded deeply incised 

channels through the volcanic layers to surficial 

bedrock, and the potential for landslides is typically 

high on the steep slopes. Recharge and well yields in 

the province are typically low. In some places, 

including the eastern Wolf Creek subwatershed, 

earthflows have created a poorly defined drainage 

network where ponds and seeps are common, and 

where subsurface flow can exceed surface flow 

(U.S. Forest Service and Bureau of Land Management, 1995). Rocks of the Klamath Mountains comprise about 11% of the watershed, mostly in the 

southwest. Exposed rocks in this province include 

granitic and sedimentary rocks (composed of altered 

submarine volcanic flows, tuffs, flow-breccia, and 

agglomerates) and ultramafic rocks (including large 

outcroppings of peridotite and serpentinite). Well 

yields in this province tend to be somewhat higher 

than in the western Cascades (McFarland, 1983). 

The final 6% of the watershed, located in the 

extreme northwestern corner near the mouth of Little River, is made up of tertiary rocks from the Coast 

Range province, which are predominantly marine 

sedimentary rocks. 

The Wolf Creek subwatershed is located in 

the central portion of the watershed in what is 

known as the “Wolf Plateau Vicinity” (U.S. Forest 

Service and Bureau of Land Management, 1995), 

which also includes Negro Creek and White Creek. 

The vicinity is characterized by broad, gently sloping uplands formed from resistant ash-flow tuff, creating a rocky bluff along the northern, western, and
21

Figure 4. Sampling locations in Little River Watershed, Oregon 1998.

Cavitt

Creek

Creek

Creek

Creek

Wolf

We

s

t

Fo

rk

Negro

White

River

Little

Creek Emile

Creek

Egglestron

Jim Creek River

Cavitt

Creek

Little

10

18 17

16

11

15

30

7

6 31 2019

9

13

21

8

8

14 22

1

2

24 4

3

23 5

123° 00' 122° 55' 122° 50'

43° 10'

0 0 

43° 15'

E

Base from B1994-99. UTAmerican Da
22

southern edges. The plateau is dissected by drainage 

systems that have begun to cut through this resistant 

layer, creating steep gradients with numerous waterfalls that act as natural barriers to fish passage. The 

plateau is capped by a variety of volcanic deposits 

such as lava flows, mudflows, flow-breccias, and 

tuffs. Weathering of these deposits has generated 

fine, clay-rich, relatively impermeable soils. Massfailure processes are likely where channel incision 

has undermined adjacent banks, and during seasonal 

peak flows chronic sediment is delivered from these 

channels. Localized debris avalanches, slumps, and 

earthflows are found along steeper hillslopes (U.S. 

Forest Service and Bureau of Land Management, 

1995). 

Water-Quality Issues in the Little River Watershed 

Previous studies in the Umpqua Basin have 

indicated that several streams, including the South 

Umpqua River (Tanner and Anderson, 1996), North 

Umpqua River (Anderson and Carpenter, 1998), 

and Little River (Powell, 1995, 1998) experience 

nuisance growths of periphytic algae during summer low flow periods. In many locations, resultant 

photosynthesis has raised pH values higher than the 

State of Oregon standard of 8.5, with maximum 

values in the North Umpqua Basin reaching as high 

as 9.1 (Anderson and Carpenter, 1998; Powell, 

1995, 1998). Little River is listed on the State of 

Oregon’s 303(d) list of water-quality limited 

streams for pH, temperature, sedimentation, and 

habitat modification (Oregon Department of Environmental Quality, 1999). Longitudinal surveys in 

various streams in the basin have shown general 

increases in pH in a downstream direction, and 

there are substantial diel variations in pH and DO 

(fig. 5) characteristic of excessive primary production. Mechanisms that may account for nuisance 

algal growth and (or) exaggerated diel pH and DO 

cycles include increased primary production resulting from timber operations (Gregory et al., 1987; 

Powell, 1995) and lack of functioning hyporheic 

zones that could help to reduce pH through benthic 

respiration (Storey et al., 1999; Mullholand et al., 

1997, 1999). These effects are likely exacerbated 

when streams have low buffering capacity (alkalinity), and geochemical processes may be important 

locally in controlling stream pH.

Figure 5. Morning and afternoon pH and dissolved oxygen 

saturation in the Little River, July 28, 1998. (Source: Powell, 1998.)

Problems with water-quality and ecosystem 

processes in the Little River watershed extend 

beyond exceedances of State water-quality standards. There is a variety of aquatic species of concern in the Little River (table 2), including both 

anadromous fish and amphibians (John Raby, 

Bureau of Land Management, written commun., 

August 1999; Dr. R. Bruce Bury, U.S. Geological 

Survey, written commun., August 1999). These 

species could be affected by disruption of food supplies resulting from changed inputs of nutrients. To 

the extent that management activities on forest 

lands in the basin affect these sensitive species, 

public land agencies are mandated to strive to balance those activities with the needs of aquatic 

biota. Thus, it is important to understand the effects 

of forest management, including fertilization, on 

aquatic communities in the Little River watershed 

and in similar basins throughout the Pacific Northwest.

Table 2. Aquatic species of concern in the Little River watershed, 

Oregon

Species

Federal 

designation

State of Oregon 

designation

Umpqua River cutthroat trout Endangered Sensitive/Vulnerable

Oregon Coast coho salmon Threatened Sensitive/Critical

Oregon Coast steelhead trout Candidate Sensitive/Vulnerable

Pacific lamprey Species of 

Concern

Sensitive/Vulnerable

Red legged frog Sensitive1

1

listed on the National Forest Sensitive Species List

No Designation

7.5

9.5

8.0

8.5

9.0

pH

20 15 10 5 0

RIVER MILE

25

State standard for pH

90

110

95

100

105

DISSOLVED OXYGEN SATURATION, 

IN PERCENT

EXPLANATION

Afternoon pH

Morning pH

Afternoon DO (percent saturation)

Morning DO (percent saturation)
23

Land uses in the Little River Watershed and Potential 

Effects on Water Quality

Forestry is by far the dominant land use in the 

Little River watershed. Roughly 60 percent of the 

watershed’s area had been harvested for timber and 

reforested by 1995. The watershed’s 206 square 

miles (approximately 132,000 acres) is predominantly (>63%) Federal forest, of which 76% (63,575 

acres) is administered by the Forest Service and 24% 

(19,802 acres) is administered by the BLM. The 

watershed is sparsely settled (population approximately 1,200), and over 70% of the private lands are 

managed for commercial timber production. About 

78% of the Wolf Plateau Vicinity has been harvested, making it the most intensively logged vicinity in the Little River watershed. There are 960 miles 

of road in the watershed, the majority of which are 

used for forest management. Extensive road building 

and timber harvesting, especially prior to 1970 when 

techniques were poor, may have contributed to the 

Wolf Plateau Vicinity’s having the second highest 

frequency of land-management related landslides 

(5.2 per square mile) in the watershed (U.S. Forest 

Service and Bureau of Land Management, 1995). 

Fertilization continued on selected Forest Service 

lands until 1990, but no BLM lands have been fertilized in the watershed since 1975 

(C. Kinntop, Bureau of Land Management, written 

commun., 2000). Timber operations were recently 

identified as being among potential nutrient sources 

in the larger North Umpqua River Basin (Anderson 

and Carpenter, 1998). In that study, nutrient enrichment, excessive benthic algal growth, and maximum 

pH were evaluated in conjunction with timber and 

hydropower operations.

In 1994, the public land in the watershed was 

collectively designated as one of 10 Adaptive Management Areas (AMAs) under the President’s Northwest Forest Plan (U.S. Forest Service and Bureau of 

Land Management, 1994). The specific emphasis of 

the Little River Adaptive Management Area 

(LRAMA) is “the development and testing of 

approaches to integration of intensive timber production with restoration and maintenance of high quality 

riparian habitat” (U.S. Forest Service and Bureau of 

Land Management, 1995). As a result, evaluating 

management effects on water quality is the major 

emphasis of the LRAMA. 

A complicating aspect to the proposed study is 

the patchwork pattern of land ownership in the 

watershed (fig. 4), a common feature in forested 

lands of the western United States. The upper portions of the Little River watershed are located 

entirely on National Forest lands, whereas private 

timberlands intermix in alternating square-mile sections with BLM and Forest Service lands in the middle portions of the LRAMA, particularly in the Wolf, 

Negro, and Cavitt Creek subwatersheds. This pattern 

renders many of the perennial streams in these subwatersheds subject to influences from a mixture of 

upstream land uses and management practices, 

including previous fertilizations on much of the private timberlands during fall in 1998 and 1999. For 

this reason, care will be needed to differentiate fertilization effects to perennial streams on BLM lands 

and effects from other land uses. This problem 

necessitates including smaller scale research (reach 

and subwatershed), in addition to investigating larger 

streams, to minimize confounding effects from 

upstream treatments. 

In addition to forestry, there is a small amount 

of agriculture, a permanent population of about 

1,200, 2 camps that house a combined population of 

250–400 people at a time, and least 19 federally 

managed camping areas along the Little River (U.S. 

Forest Service and Bureau of Land Management, 

1995). Of the two camps, one has a small on-site 

wastewater treatment plant (WWTP), the other a 

large septic drainfield. Private residences in the 

watershed predominantly use septic systems, and 

recreation areas use either pit toilets or septic systems, some or all of which could provide nutrients to 

the river depending on their state of maintenance. 

With summertime streamflows dipping to less than 

30 ft3/s, the cumulative nonpoint nutrient inputs 

from these sources can potentially contribute to 

stream eutrophication. 

The two established camps, the Wolf Creek Job 

Corps Center and the Little River Christian Camp, are 

located just upstream from Wolf Creek at approximately river miles 12 and 13, respectively. The Job 

Corps Center houses between 230 and 275 people, and 

the Christian Camp averages about 90 during summer. 

The Job Corp Center’s WWTP discharges an average 

of 21,000 gallons per day. Coincidentally, both camps 

are located immediately upstream of Wolf Creek, the 

tributary with the largest number of forested stands 

planned for fertilization by the BLM and the most 
24

likely location in which to investigate fertilizer effects 

on streams. This fact, and possible effects from other 

upstream land uses such as recreation and forestry 

operations, complicates the assessment of the effects 

to the Little River from fertilization within and outside the Wolf Creek subwatershed due to the lack of 

an appropriate upstream reference. 

Water-Quality Conditions

Data were collected by the USGS and BLM 

during reconnaissance investigations in August and 

November 1998 (table 4), and August 1999 (table 5). 

The August surveys were to assess conditions during 

low flow prior to fertilization, and the November 

survey was intended to assess runoff during a fertilization to private timber land in the Wolf Creek subwatershed. Ultimately, the November samples were 

collected prior to fertilization during a storm. In the 

discussion below, emphasis is placed on data from 

the August 1999 survey because it is a more complete dataset and more quality assurance data were 

collected during that survey. Data from 1998 are referenced where they support or contradict findings 

from August 1999. 

Samples from all locations (fig. 4, table 3) 

were analyzed at the U.S. Geological Survey’s 

National Water Quality Laboratory (NWQL) in Denver, Colorado. A subset of samples from the August 

1999 survey also was submitted for nutrient analyses 

to the Central Chemical Analytical Laboratory 

(CCAL) in the Forest Sciences Laboratory at Oregon 

State University. This laboratory comparison was an 

attempt to assure the quality of CCAL for low-level 

nutrient analyses, particularly nitrogen species, 

because the NWQL does not have methods available 

to analyze organic nitrogen at the low concentrations 

routinely found in the Little River watershed (tables 

4 and 5). Organic nitrogen is a potentially important 

nitrogen species in the adjacent North Umpqua River 

Basin and typically constitutes the largest portion of 

nitrogen budgets in forested streams in the Cascades 

(Triska et al., 1984).

Methods

Discharge was measured using Price AA or 

pygmy current meters by established techniques 

(Rantz et al., 1982) wherever possible, although in 

some places water depths or velocities were too low 

to use meters. In these instances, discharge was measured by directing the entire flow of the stream into a 

bucket and measuring the volume of water filling the 

bucket in a known amount of time. Frequently this 

was done on the downstream ends of culverts under 

logging roads. Field parameters (temperature, specific conductance, pH, and dissolved oxygen) were 

measured in-place using Hydrolab® multiparameter 

instruments that were calibrated in the field according to manufacturer’s specifications. Where possible, field measurements were timed to document pH 

near its daily maximum (about 4:00–6:00 p.m.), and 

sampling during 1999 specifically included late 

afternoon measurements at each site for this purpose. 

Water chemistry samples for nutrients and 

major ions were taken by grab sampling in most 

cases. Grab sampling was considered representative 

where streamflows were low, there was little suspended material, and the streams were well mixed 

due to high gradients. Also, in many of the small 

catchments, the low streamflow would have prevented the use of larger depth- and width-integrating 

samplers. Samples from the main-stem Little River 

were collected using depth- and width-integrating 

techniques. Samples for whole-water (unfiltered) 

nutrients (total phosphorus [TP] and total organicplus-ammonium nitrogen [TKN]) were collected in 

acid-washed bottles. Those to be analyzed by the 

NWQL were immediately preserved with 0.2N 

H2SO4, whereas those being analyzed by CCAL 

were unpreserved. For filtered-water nutrients (dissolved organic plus ammonium nitrogen [DKN], 

NH3-N, nitrate plus nitrite nitrogen [predominantly 

NO3-N], nitrite nitrogen, soluble reactive phosphorus [SRP], and digested dissolved phosphorus 

[TDP]), water was subsampled on site by peristaltic 

pump with in-line, prewashed, disposable capsule 

filters (0.45 μm [micrometer] pore-size) into sample 

bottles and stored chilled and unpreserved. Samples 

for major chemistry were filtered into two polypropylene bottles; water analyzed for cations was preserved with nitric acid, and water analyzed for 

anions was unpreserved. Samples for alkalinity during August 1999 were filtered in the field and stored 

chilled until titrated at the U.S. Geological Survey’s 

Oregon District Laboratory (about 3–4 days). Based 

on good agreement in cation and anion balances, 

storage time did not appear to affect the alkalinity 

results. Analytical methods and detection levels for 

nutrient samples analyzed at the NWQL and CCAL 

are given in table 6. 
25

Table 3. Sites sampled in the Little River watershed by the U.S. Geological Survey (USGS) in 1998 and 1999×[ID, identification; BLM, Bureau of Land Management; USFS, U.S. Forest Service; Latitude/Longitude given in degrees, minutes, and seconds]

Map ID Site name

USGS 7-1/2 

minute 

topographic 

map

Range/

township/-

section 

Latitude/

longitude

Augus

1998

1 Wolf Creek at mouth Red Butte R2W/T27S-9 431358/1225700 X

2 Egglestron Creek at mouth Red Butte R2W/T27S-16 431341/1225648 X3 Unnamed tributary to Fork Wolf Creek at west bank near 

Wolf Creek Falls

Red Butte R2W/T27S-16 431314/1225650 X 

4 Wolf Creek above Egglestron ×Creek Red Butte R2W/T27S-16 431327/1225646 X 5 Unnamed tributary to Wolf Creek on east bank upstream of 

Egglestron Creek on BLM 16.0 road 

Red Butte R2W/T27S-16 431253/1225619 X 

6 West Fork Wolf Creek Red Butte R2W/T27S-20 431209/1225719 X7 Unnamed tributary to West Fork Wolf Creek at west bank 

on BLM 16.0 road

Red Butte R2W/T27S-20 431210/1225724 X

8 Cavitt Creek above Withrow Creek on road 25 Red Butte R2W/T28S-9 430914/1225614 X9 Unnamed tributary to Cavitt Creek below Evarts Creek Lane Mountain R3W/T27S-13 431312/1230021 X

10 Emile Creek at USFS/BLM boundary Mace Mountain R1W/T26S-31 431525/1225215 X

11 Emile Creek at mouth Old Fairview R2W/T27S -2 431455/1225403 X

13 Little River above Cavitt Creek at new bridge Lane Mountain R3W/T27S-11 431419/1230039 

14 Little River above Wolf Creek Red Butte R2W/T27S-9 431402/1225659 X

15 Little River above Job Corps at road 2701 bridge 

(Old Red Butte Road)

Red Butte R2W/T27S-3 431453/1225533 X

16 Little River at White Creek Campground Taft Mountain R1W/T27S-7 431340/1225137 X 17 Little River at Coolwater Campground Taft Mountain R1W/T27S-7 431350/1225214 

18 Negro Creek near mouth Taft Mountain R2W/T27S-12 431355/1225224 19 Wolf Creek near Red Butte below gravel quarry 

at BLM 14.1 road

Red Butte R2W/T27S-28 431159/1225630 

20 Wolf Creek at BLM road 16.0 Red Butte R2W/T27S-21 431214/1225643 21 Little River at Peel (USGS station ID 14318000) Glide R3W/T27S-2 431510/ 123013 X 

22 Little River below Wolf Creek Red Butte R2W/T27S-9 431358/1225704 

23 Wolf Creek above Wolf Creek Falls Red Butte R2W/T27S-16 431306/1225646 24 Reference tributary to Wolf Creek above Egglestron Creek Red Butte R2W/T27S-16 431337/1225650 30 Unnamed tributary to West Fork Wolf Creek in section 32

 at road 32.0

Red Butte R2W/T27S-32 431056/1225734 

31 Unnamed tributary to West Fork Wolf Creek, #2, 

at east bank at BLM road 16.0

Red Butte R2W/T27S-20 431212/1225709 
26

Table 4. Nutrient and field data in the Little River and tributaries from reconnaissance samplings during August and November, 1998[Sites are listed in downstream order. Map ID number refers to locations on figure 4. All nutrient samples were analyzed at the U.S. Geological Survey laboratory in Den

°C, degrees Celsius; (μS/cm), microsiemens per centimeter at 25 degrees Celsius; mg/L, milligrams per liter; μg/L, micrograms per liter; BP, Barometric pressure; mm H

oxygen; %, percent; μg/L, micrograms per liter; NH3, ammonia; NO2, nitrite; DKN, dissolved Kjeldahl nitrogen; TKN, total Kjeldahl nitrogen; NO3, nitrate + nitrite; TP

phosphorus; SRP, soluble reactive phosphorus; <, actual value is less than the indicated amount; USFS, U.S. Forest Service; BLM, Bureau of Land Management; --, no d

Station Name

Map ID 

number Date Time Flow (ft3/s)

Water 

temperature 

(×°C)

Specific

conductance 

(μS/cm)

BP 

(mm 

Hg)

DO

(mg/L)

DO 

saturation 

(%) pH

DKN

(μg/L 

as N)

TKN

(μg/L 

as N)

NH3

(μg/L 

as N)

Little River at White Creek Campground 

16 8/26/98 1835 8.8 16.4 88 -- 9.3 -- 8.2 -- 100 <2 

Little River above Negro Creek at Coolwater Campground

17 11/7/98 1500 -- -- 69 -- -- -- 7.6 -- 110 -- 

" 17 11/10/98 1500 120 6.5 64 725 11.7 100 7.7 <100 260 <2 

Negro Creek near mouth 18 11/7/98 1400 -- -- 85 -- -- -- 7.6 100 270 <2 

" 18 11/10/98 1420 18 6.4 80 723 11.2 96 7.8 120 280 2 

Emile Creek at USFS/BLM Boundary 10 8/26/98 1915 .99 14.9 56 -- 9.1 -- 7.5 -- <100 2 Emile Creek at mouth 11 8/25/98 1930 1.1 16.2 68 -- 9.2 -- 7.7 -- <100 <2 Little River above Christian Camp at USFS road 2701 bridge

15 8/26/98 1745 14 18.7 92 -- 9.3 -- 8.8 -- <100 <2 

Little River above Wolf Creek 14 8/24/98 1200 17 17.4 95 735 10.4 113 8.6 -- <100 <2 

" 14 11/9/98 1345 133 7.1 64 738 11.7 100 7.8 140 320 5 

Wolf Creek at BLM road 14.1 20(?) 11/9/98 1435 5.8 7.1 53 -- -- -- 7.6 <100 170 <2 West Fork Wolf Creek 6 8/27/98 1400 .64 12.8 129 -- 9.4 -- 7.9 -- <100 <2 Unnamed tributary to Wolf Creek east bank

5 8/25/98 1615 .01 14 90 -- 9 -- 7.4 -- <100 16 

Unnamed tributary to West Fork Wolf Creek 

7 8/27/98 1320 .09 13.1 101 -- 9.1 -- 7.6 -- <100 <2 

" 7 11/9/98 1735 -- -- -- -- -- -- -- <100 210 <2 " 7 11/10/98 1200 6.2 7.2 41 629 9.8 98 7.3 -- <100 -- 

Unnamed tributary to Wolf Creek west bank below Wolf Creek Falls

3 8/25/98 1415 .01 11.2 26 734 10.3 97 6.8 -- <100 12 

" 3 11/9/98 1120 .31 9.3 26 735 10.6 96 7.5 <100 <100 <2 

Wolf Creek aboveEgglestron Creek 

4 8/25/98 1435 2 13.8 120 -- 9.7 -- 8 -- <100 <2 

Egglestron Creek at mouth 

2 8/24/98 1600 .42 14.6 213 -- 9.4 -- 8.2 -- <100 <2 

" 2 11/9/98 1210 3.7 7.1 84 737 11.4 97 7.8 <100 190 <2 

Wolf Creek near Red Butte 

19 11/9/98 1230 4 6.2 91 -- -- -- 7.6 130 280 5 

Wolf Creek at mouth 1 8/24/98 1400 2.9 15.2 135 -- 9.9 -- 8.4 -- <100 <2 

" 1 11/7/98 1100 -- 74 -- -- -- 7.4 <100 <100 <2 " 1 11/9/98 1310 18 7.2 61 739 11.4 98 7.7 -- <100 -- 

Little River above Cavitt Creek at new bridge

13 8/26/98 1645 16 19.5 98 -- 9.3 -- 8.7 -- <100 <2 

Cavitt Creek above Withrow Creek 8 8/26/98 1407 2.3 12.7 138 699 9.5 98 8.2 -- <100 <2 Unnamed tributary to Cavitt Creek 9 8/27/98 1145 0 14.4 68 724 8.4 86 7 -- <100 <2 Little River at Peel 21 8/26/98 1545 26 18.5 111 740 10.1 111 8.4 -- <100 <2 
27

Table 5. Nutrient and field data for sites in the Little River and tributaries, August 1999[Sites are listed in downstream order. Map ID number refers to locations on figure 4. Sample types— f, field parameters (temperature, dissolved oxygen, pH, specific condAnalytical Laboratory at Oregon State University; U, U.S. Geological Survey Laboratory in Denver, Colorado; °C, degrees Celsius; mm Hg, millimeters of mercury; mg/μg/L, micrograms per liter; BP, Barometric pressure; ft3/s, cubic feet per second; (μS/cm), microsiemens per centimeter at 25 degrees Celsius; DO, dissolved oxygen; Alk

NO2, nitrite+nitrate; DKN, dissolved Kjeldahl nitrogen; TKN, total Kjeldahl nitrogen; NO3, nitrate + nitrite; TP, total phosphorus; TDP, total dissolved phosphorus; SRP, 

is estimated; <, actual value is less than the indicated amount; --, no data; USFS, U.S. Forest Service; BLM, Bureau of Land Management]

Station name

Map 

ID 

number

Sample 

type Dates Times

Water 

temperature 

(°C)

BP 

(mm 

Hg)

Flow 

(ft3/s)

Specific 

conductance 

(μS/cm)

DO 

(mg/L) pH 

Alk 

(mg/L)

NH3

(μg/L 

as N)

NO2

(μg/L 

as N)

DKN 

(μg/L 

as N)

T

(μ

as 

Little River at White Creek Campground 16 f 19990817 1705 17.0 727 -- 82 9.1 7.9 -- -- -- -- Little River at White Creek Campground 16 C 19990818 1500 -- -- -- -- -- -- -- <10 -- 50 4Little River at White Creek Campground 16 U 19990818 1500 17.3 741 16.5 80 9.5 8.2 39 3 1 102 E 7Little River at CoolwaterCampground 17 f 19990817 1725 16.7 727 -- 88 9.1 7.9 -- -- -- -- Little River at Coolwater Campground 17 f 19990818 1700 17.2 731 -- 86 9.3 8.2 -- -- -- -- Negro Creek near mouth 18 f 19990817 1755 14.5 727 -- 110 8.9 7.9 -- -- -- -- Negro Creek near mouth 18 C 19990818 1200 -- -- -- -- -- -- -- <10 40 3Negro Creek near mouth 18 U 19990818 1200 13.6 741 4.55 106 10.0 8.1 57 <2 <1 103 E 6Emile Creek at mouth 11 f 19990818 1705 17.7 729 -- 65 8.6 7.8 -- -- -- -- Little River below Emile Creek -- f 19990818 1705 18.8 729 -- 89 8.9 8.3 -- -- -- -- Little River at USFS road 2701 bridge 15 C 19990817 1530 -- -- -- -- -- -- -- <10 -- 40 6Little River at USFS road 2701 bridge 15 U 19990817 1530 19.4 737 25.5 86 9.1 8.5 40 <2 1 <100 <10Little River at USFS road 2701 bridge 15 f 19990817 1735 19.6 727 -- 88 8.6 8.3 -- -- -- -- Little River at USFS road 2701 bridge 15 f 19990818 1715 19.7 731 -- 86 9.1 8.6 -- -- -- -- Little River above Wolf Creek 14 C 19990816 1100 -- -- -- -- -- -- -- 38 -- 80 10Little River above Wolf Creek 14 U 19990816 1100 15.7 738 20.6 90 9.9 8.1 40 36 1 <100 11Little River above Wolf Creek 14 f 19990816 1850 19.1 735 -- 89 8.6 8.2 -- -- -- -- Little River above Wolf Creek 14 f 19990817 1730 20.1 735 (E) -- 87 8.9 8.6 -- -- -- -- Little River above Wolf Creek 14 f 19990817 1815 19.2 735 (E) -- 90 8.5 8.2 -- -- -- -- Little River above Wolf Creek 14 f 19990818 1730 19.9 735 (E) -- 91 8.7 8.6 -- -- -- -- Wolf Creek at mouth 1 C 19990817 1100 -- -- -- -- -- -- -- <10 -- 40 5Wolf Creek at mouth 1 U 19990817 1100 14.6 737 1.86 135 10.0 8.2 68 7 1 148 <10Wolf Creek at mouth 1 f 19990818 1735 16.4 735 (E) -- 141 9.1 8.2 -- -- -- -- Egglestron Creek at mouth 2 U 19990816 1620 14.8 738 .81 208 9.4 8.1 100 17 <1 <100 E 7Reference tributary to Wolf Creek above Egglestron Creek 24 C 19990816 1430 -- -- -- -- -- -- -- <10 -- 40 5Reference tributary to Wolf Creek above Egglestron Creek 24 U 19990816 1430 14.8 734 .02 20 9.0 7.4 9 13 <1 <100 E 8Wolf Creek above Egglestron Creek 4 U 19990816 1420 14.5 738 1.47 126 9.5 7.9 58 11 <1 <100 E 8
28

Wolf Creek above Wolf Creek Falls 23 U 19990816 1730 14.2 727 -- 131 9.5 8.1 65 11 <1 <100 26Unnamed tributary to West Fork Wolf Creek at BLM road 16.0 31 U 19990817 1340 12.6 714 .01 77 9.0 7.7 41 5 <1 <100 <10West Fork Wolf Creek 6 C 19990817 1530 -- -- -- -- -- -- -- <10 -- 30 4West Fork Wolf Creek 6 U 19990817 1530 13.5 714 .44 135 8.9 7.8 68 15 <1 105 <10Unnamed tributary to West Fork Wolf Creek 7 U 19990817 1720 13.8 714 .07 80 8.6 8.3 51 7 1 105 <10Wolf Creek at BLM road 14.1 19 C 19990818 1330 -- -- -- -- -- -- -- <10 -- 40 4Wolf Creek at BLM road 14.1 19 U 19990818 1330 12.7 698 .35 174 9.1 8.0 89 <2 1 <100 18Unnamed tributary to West Fork Wolf Creek in section 32 at road 32.0

30 U 19990818 1110 11.9 691 .18 56 8.7 7.5 35 2 <1 102 E 6

Little River below Wolf Creek 22 C 19990817 1130 -- -- -- -- -- -- -- <10 -- 6 9Little River below Wolf Creek 22 U 19990817 1130 18.3 737 24.3 91 10.0 8.5 42 6 2 <100 <10Little River below Wolf Creek 22 f 19990818 1745 19.6 -- -- 96 8.6 8.6 -- -- -- -- Little River at Cavitt Creek bridge 13 f 19990817 1835 21.1 -- -- 94 8.6 8.5 -- -- -- -- Little River at Cavitt Creek bridge 13 f 19990818 1800 19.4 739 -- 108 9.2 8.6 -- -- -- -- Little River at New Cavitt Creek bridge 13 f 19990818 1740 19.9 738 -- 90 9.5 8.8 -- -- -- -- Little River at Peel 21 f 19990818 1830 19.5 741 (E) -- 112 9.0 8.4 -- -- -- -- Little River at Peel 21 C 19990818 1000 -- -- -- -- -- -- -- <10 -- 12 10Little River at Peel 21 U 19990818 1000 18.7 741 32 109 9.0 8.0 51 10 1 E 78 21Little River Highway 27 bridge below Peel -- f 19990817 1900 20.3 -- -- 112 8.7 8.3 -- -- -- -- 

Table 5. Nutrient and field data for sites in the Little River and tributaries, August 1999[Sites are listed in downstream order. Map ID number refers to locations on figure 4. Sample types— f, field parameters (temperature, dissolved oxygen, pH, specific condAnalytical Laboratory at Oregon State University; U, U.S. Geological Survey Laboratory in Denver, Colorado; °C, degrees Celsius; mm Hg, millimeters of mercury; mg/μg/L, micrograms per liter; BP, Barometric pressure; ft3/s, cubic feet per second; (μS/cm), microsiemens per centimeter at 25 degrees Celsius; DO, dissolved oxygen; Alk

NO2, nitrite+nitrate; DKN, dissolved Kjeldahl nitrogen; TKN, total Kjeldahl nitrogen; NO3, nitrate + nitrite; TP, total phosphorus; TDP, total dissolved phosphorus; SRP, 

is estimated; <, actual value is less than the indicated amount; --, no data; USFS, U.S. Forest Service; BLM, Bureau of Land Management]

Station name

Map 

ID 

number

Sample 

type Dates Times

Water 

temperature 

(°C)

BP 

(mm 

Hg)

Flow 

(ft3/s)

Specific 

conductance 

(μS/cm)

DO 

(mg/L) pH 

Alk 

(mg/L)

NH3

(μg/L 

as N)

NO2

(μg/L 

as N)

DKN 

(μg/L 

as N)

T

(μ

as 
29

Table 6. Analytical methods and detection levels for nutrient analyses performed at the U.S. Geological Survey National Water Quality 

Laboratory (NWQL) and Oregon State University Cooperative Chemical Analytical Laboratory (CCAL)

[EPA, Environmental Protection Agency]

Constituent

NWQL CCAL

Method

Detection 

level 

(μg/L) Reference Method

Detection 

level 

(μg/L)

Total phosphorus (TP) EPA 365.1 8 EPA 424C 1

Total dissolved phosphorus (TDP) EPA 365.1 6 -- --

Soluble reactive phosphorus (SRP) I260689 1 Fishman, 1993 EPA 424F 1

Total organic + ammonium nitrogen (TKN) I451591 100 Patton and Truitt, 2000 Kjeldahl, Nessler finish 10

Dissolved organic + ammonium nitrogen (DKN) I261091 100 Patton and Truitt, 2000 Kjeldahl, Nessler finish 10

Ammonium nitrogen (NH4-N) I252589 2 Fishman, 1993 EPA 417F 10

Nitrate + nitrite nitrogen (NO3-N) I254691 5 Fishman, 1993 EPA 418F 1

Nitrite nitrogen (NO2-N) I254289 1 Fishman, 1993 -- --

Algal samples from 1999 were collected by 

scraping known areas from rocks into jars, homogenizing the sample in a blender, and subsampling 

the resulting slurry for algal biomass and chlorophyll a. Analysis of samples for ash free dry mass 

(AFDM) and chlorophyll a used standard methods 

(American Public Health Association, 1998). 

AFDM and chlorophyll a were analyzed in triplicate in the Oregon District Laboratory. 

Quality Assurance

Nutrient and major-ion sample results were 

checked for bias through the use of blank samples. 

Blanks were prepared in the field using the same 

equipment as environmental samples, with inorganic-free water obtained from the USGS’s Ocala 

Field Services warehouse. Laboratory blank and 

standard reference samples to check for bias and 

accuracy in low-level nutrient analyses were made 

in the Oregon District Laboratory, using Standard 

Reference Materials traceable to the National Institute of Standards and Technology (NIST-SRMs) in 

inorganic-free water. These quality-assurance samples were simultaneously submitted in triplicate to 

the NWQL and to CCAL. Precision was evaluated 

in field replicate samples that were submitted to 

both laboratories as well. 

Data from the August 1998 sampling indicated a possible contamination of dissolved organic 

nitrogen because dissolved Kjeldahl nitrogen 

(DKN) values were substantially higher than those 

for than TKN at more than half of the sites. TKN 

and DKN values less than 100 μg/L were expected 

on the basis of previous samplings in the watershed 

(Anderson and Carpenter, 1998) and previous data 

from the Oregon Department of Environmental 

Quality (D. Ades, Oregon Department of Environmental Quality, 1998, written commun.). Data for 

DKN from August 1998 were subsequently deleted 

from the database. 

Results of reference sample analysis (table 7) 

indicated that both laboratories performed comparably and well for phosphorus analysis but each had 

small bias from contamination of NH3-N, with the 

NWQL apparently having a somewhat higher bias 

(average ~12 μg/L) than CCAL (average ~6 μg/L). 

Bias in NH3-N measurement is not unusual because 

contamination is notoriously difficult to avoid 

(Holmes et al., 1999). Although new methods to 

cleanly sample and analyze for NH3-N have recently been developed (Holmes et al., 1999), they 

have not yet been adopted by the NWQL or CCAL. 

The largest difference was apparent in analysis of 

DKN. The NWQL’s reporting limit for this analysis 

(as for TKN) is 100 μg/L, although it will report values as estimates for detections between 50 and 100 

μg/L; whereas CCAL’s analytical detection limit for 

DKN and TKN is approximately 10 μg/L. For the 

DKN samples, NWQL analysis had a highly variable 

positive bias ranging from 46–110 μg N/L. 

Among environmental samples analyzed 

between the two laboratories, there was considerable disagreement for phosphorus analysis (table 

5), despite the favorable comparison of previous 

standard reference samples. Concentrations for TP 

from the NWQL were approximately half those 

reported from the CCAL, although results were 

comparable for SRP. Results for NO3-N and
30

Table 7. Comparison of nutrient concentrations from standard reference samples analyzed at the U.S. Geological Survey’s National Water 

Quality Laboratory and Oregon State University’s Cooperative Chemical Analytical Laboratory, August 1998

[Reference samples (R1, R2, R3) were prepared with only organic nitrogen and organic phosphorus; samples were also analyzed for ammonium nitrogen to 

check for decomposition of the organic nitrogen and to evaluate possible bias from contamination in low level ammonium data. NWQL, National Water 

Quality Laboratory; CCAL, Cooperative Chemical Analytical Laboratory; *, Replicate analyses performed by CCAL; (E) Concentrations are below the 

method reporting limit and are considered estimates; μg/L - micrograms per liter]

Sample 

Organic P (dissolved digested P) Ammonia-nitrogen

Organic-N + ammonia N

(dissolved Kjeldahl N)

Nominal 

value

(expected)

(μg/L)

NWQL

(reported) 

(μg/L)

CCAL

(reported) 

(μg/L)

Nominal 

value

(expected

(μg/L)

NWQL

(reported) 

(μg/L)

CCAL

(reported) 

(μg/L)

Nominal 

value

(expected)

(μg/L)

NWQL

(reported) 

(μg/L)

CCAL

(reported) 

(μg/L)

Blank 0 <4 <1 0 3 <10 0 (E) 71 <10

Blank 0 <4 *1/1 0 9 <10 0 (E) 88 <10

Blank 0 <4 <1 0 7 <10 0 (E) 95 <10

R1 21 21 22 0 13 0 43 (E) 89 40

R2 21 19 22 0 12 *4/7 43 134 40

R3 21 21 24 0 10 7 43 153 50

NH3-N in environmental samples were also comparable, although the small positive contamination 

from NH3-N was evident in some samples. On the 

basis of QA samples (table 7), data on organic nitrogen (TKN and DKN) from CCAL were considered 

more reliable than those from the NWQL; however 

differences in TP concentrations between the two 

labs remain unresolved. For most of the discussion 

here, data from the NWQL are used because samples were not submitted to CCAL from all sites; 

however reference is made to CCAL data where 

interpretation of nutrient status would be different. 

Environmental Data

Precipitation during 1999 in the Little River 

watershed was slightly higher than normal 

(Owenby and Ezell, 1992), and temperatures were 

near normal. Nonetheless, many tributaries located 

directly within the BLM’s proposed fertilization 

units were dry during sampling in August 1999 and 

an earlier reconnaissance trip in July 1999. The 

lack of water necessitated sampling farther downstream to find adequate water. However, as a result 

of relocating downstream from the fertilization 

units, samples contained water from additional tributaries that had entered the river from upstream 

subwatersheds containing mixed private and Federal (BLM) timberlands. In general, discharge was 

low in the Wolf Creek subwatershed, ranging from 

less than 0.01 ft3/s in some of the upland tributaries 

to 1.9 ft3/s at the mouth of Wolf Creek. These discharges were similar to those in August 1998, 

although Wolf Creek had 2.9 ft3/s at the mouth 

during 1998. Most tributary streams were covered 

by dense canopies of alder and low brush, and 

could often be straddled. An attempt was made to 

mass balance flows in Wolf Creek to evaluate possible ground-water discharge. Though most significant flows were measured, Wolf Creek upstream of 

the confluence with West Fork Wolf Creek, with 

visible flow, was not measured due to inaccessibility. It is therefore unclear if the approximately 

0.5 ft3/s gain from upstream to downstream 

resulted from unmeasured inflows of tributaries or 

ground water. Discharge in the Little River main 

stem increased from 16.5 ft3/s at White Creek 

Campground (the upstream border between BLMand USFS-managed lands) to 32 ft3/s at the USGS 

gaging station at Peel (table 5), a value equal to the 

average monthly discharge for August at the Peel 

gage during a previous period (1953–1987) when 

the station was in operation (Moffatt et al., 1990).

Where there was enough water to sample in 

the Wolf Creek subwatershed and its tributaries, 

field parameter data indicated few overt waterquality problems. Maximum temperatures were less 

than 16.5oC (degrees Celsius) at all sites and as low 

as 11.9oC at some sites. Maximum pH was as high 

as 8.4 at the mouth of Wolf Creek in 1998, and 8.3 

in West Fork Wolf Creek in 1999, but at all other 

sites ranged from 7.4 to 8.2. Dissolved oxygen 

(DO) concentrations were lowest (8.7 mg/L) at site 

#30, an unnamed tributary high in the West Fork 

Wolf Creek drainage; however, this investigation 
31

did not target the early morning period, when DO 

could be lowest if significantly affected by 

periphyton respiration. Therefore, the minimum 

DO conditions during August 1998 are unknown. 

Algal biomass was low and difficult to observe 

without magnification, at almost all locations in the 

Wolf Creek subwatershed except for two areas with 

open canopies. These were the mouth of Wolf 

Creek (site #1) and the upright wall of Wolf Creek 

Falls. 

Wolf Creek Falls is unique because of its 

physical structure, in which much of the water disperses and trickles thinly over a long, wide slab of 

rock. This slab has a western aspect and open canopy, so solar exposure is relatively good. As a 

result, the slab had a continuous and consistently 

thick film of healthy, bubbling filamentous green 

algae covering it as the water slowly poured over it, 

resembling a trickling-filter apparatus in a wastewater treatment plant. Due to air exposure, no grazing aquatic insects such as the stone-case building 

caddisfly Dicosmoecus, (observed elsewhere in the 

lower reaches of Wolf Creek and the Little River) 

colonized this mat. Thus the waterfall provides a 

naturally occurring habitat for filamentous algal 

growth with minimal grazing pressures. This phenomenon undoubtedly acts to reduce nutrient concentrations through uptake during certain times of 

the year, and may also increase pH in Wolf Creek 

upstream of the mouth such as previously observed 

(Powell, 1995). During mid-August 1999, the falls 

increased pH only about 0.1 units from top to bottom. 

The mouth of Wolf Creek (site #1) is covered 

by steps of bedrock with little alluvial material, and 

at this location solar exposure to the stream is the 

greatest in the subwatershed. There, luxuriant mats 

of filamentous green algae were observed, with 

individual strands exceeding 5 feet in length, during late August-early September 1999. 

In contrast to the Wolf Creek subwatershed, 

field parameter data in the main-stem Little River 

appear to reflect the effects of upstream land uses. 

During August 1999, daily maximum pH in the Little River (fig. 6) increased from 8.2 at White Creek 

Campground (site # 16) to 8.8 above Cavitt Creek 

(site # 13), with the largest increase (0.3 units) 

occurring in the 1.5 mile reach between the mouth 

of Emile Creek (river mile 14.9, site not shown on 

fig. 4) and the USFS road 2701 bridge (site #15). 

Similarly, during August 1998 pH was higher than 

the State standard of 8.5 at main-stem sites between 

the USFS road 2701 bridge and Cavitt Creek (table 

4). These patterns were similar to those observed 

by Powell (1998) in figure 4. Like pH, maximum 

water temperatures (fig. 7) did not meet the State 

standard (17.8oC) from below Emile Creek (river 

mile 14.9) to the Highway 27 bridge below Peel 

(river mile 3.2, not shown on fig. 4). Temperatures 

in the Little River during August 1998 did not meet 

the standard at the USFS road 2701 bridge, above 

Cavitt Creek, or at the Peel gage; temperature at the 

site above Wolf Creek may have met the standard 

because it was measured around noon rather than in 

late afternoon as most other main-stem sites were. 

Dissolved oxygen concentrations met the State 

standard of 8.0 mg/L at all stations, but were not 

investigated during the early morning to evaluate 

diel variation associated with periphyton metabolism.

Geologic or land-use differences in the area 

may be reflected in the water quality. In particular, 

the smaller streams draining the western slopes of 

the Wolf Creek subwatershed (sites 24, 7, 30, and 

31) had distinctly different chemical signatures, as 

measured by major-ion concentrations (fig. 8) and 

specific conductance (table 5), from other sites in 

Little River and Wolf Creek. Although the major 

anion at all sites was almost exclusively bicarbonate, and calcium and magnesium were the dominant 

cations at most sites, sodium was increased both in 

concentration and percent of total cations at these 

four sites and calcium concentrations were reduced. 

The same four sites also were the most dilute in the 

Wolf Creek subwatershed (specific conductances 

20–79 μS/cm [microsiemens per centimeter]). In 

contrast, the two sites draining the east side of the 

subwatershed (site 2—Egglestron Creek, and site 

20—Wolf Creek at the BLM 14.1 road) had the 

highest specific conductances (208 and 174 μS/cm, 

respectively) measured during the 1999 survey. 

Major ions were not measured during 1998, but 

specific conductances during August 1998 were 

similar to those during the August 1999 survey. 

The causes of the differences among sites, or their 

possible effects on nutrient retention in soils and 

(or) transport in the streams, have not been 

expressly investigated.
32

 

Figure 6. Afternoon pH in the main stem of the Little River, August 1999.

8.0

8.2

8.4

8.6

8.8

pH

Below Emile CreekRoad 2701 bridgeLittle River Christian CampJob Corps CenterAbove Wolf CreekBelow Wolf CreekAbove Cavitt Creek (new bridge)Below Cavitt CreekGaging station (Peel)Highway 27 bridge

State standard

20 16 14 12 10 8 6 4 2 18

RIVER MILE

Figure 7. Afternoon temperature in the main stem of the Little River, August 1999. 

20 16 14 12 10 8 6 4 2 18

17

18

19

20

21

21.5

TEMPERATURE, IN DEGREES CELSIUS

Road 2701 bridgeLittle River Christian CampJob Corps CenterAbove Wolf CreekBelow Wolf CreekAbove Cavitt Creek (new bridge)Below Cavitt CreekGaging station (Peel)

State standard

Below Emile CreekHighway 27 bridge

Nutrient concentrations in the Little River 

watershed and Wolf Creek subwatershed were 

typically low during August 1999, consistent 

with nitrogen limitation as previously observed 

(Anderson and Carpenter, 1998) (fig. 9). Median 

concentrations for all dissolved inorganic nitrogen 

species were less than 20 μg/L, and TP concentrations (median 20 μg/L) were typically high enough 

to saturate algal growth (Bothwell, 1989). Nitrate-N 

and NH3-N concentrations were slightly higher in 

streams of the Wolf Creek subwatershed than in the main stem Little River, possibly 

reflecting uptake of nitrogen in Little River 

and (or) light limitation in Wolf Creek. In 

contrast SRP concentrations, and to a lesser 

extent TP concentrations, were somewhat higher

 in Little River than in Wolf Creek. Organic plus 

ammonium nitrogen, measured as TKN or DKN, 

was generally less than 100 μg/L (using data from 

CCAL).
33

Figure 8. Major ion chemistry in Little River and Wolf Creek. (Sites that are the most chemically distinct are the reference tributary 

to Wolf Creek above Egglestron Creek [site 24], an unnamed tributary east of the West Fork of Wolf Creek on road 16.1 [site 30], an 

unnamed tributary west of the West Fork of Wolf Creek on road 16.1 [7], and an unnamed tributary to West Fork Wolf Creek in 

section 32 at road 32.0 [site 31]. See figure 4 for locations.)

31 24

7

30 PERCENT

PERCENT

SULFATE PLUS CHLORIDE

MAGNESIUM

100

80

60

40

20

0

100

80

60

40

20

0

100

80

60

40

20

0

CALCIUM PLUS MAGNESIUM

SULFATE

100

80

60

40

20

0

100

80

60

40

20

0

100

80

60

40

20

0

CARBONATE PLUS BICARBONATE

SODIUM PLUS POTASSIUM

100

80

60

40

20

0

100

80

60

40

20

0

100

80

60

40

20

0

100

20

40

60

80

0

CALCIUM CHLORIDE, FLUORIDE, NITRATE PLUS NITRITE

PERCENT

 Typically N:P ratios <7 (on an atomic weight 

basis) are considered indicative of nitrogen limitation, whereas N:P >10 could indicate phosphorus 

limitation (Wetzel, 1983; Hillebrand and Sommer, 

1999). For this report, ratios of total nitrogen tototal phosphorus (TN:TP) during 1999 depend on the 

source of the data (NWQL or CCAL) and the treatment of censored data (that is, nondetections). 

Using NWQL data only, and taking nondetected 

concentrations at the value of their respective 

reporting limits (an approach that overestimates 

nitrogen in nondetected samples), the median 

TN:TP was 9.3 (range 5.1–24). In contrast the 

median TN:TP was 2.3 (range 1.1–4.6) using data 

only from CCAL, but this source accounts for only 

12 of 17 samples because not all sites had samples 

submitted to CCAL. By substituting NWQL 

organic nitrogen data with CCAL data where available, the median TN:TP is 5.9 (range 2.6–15.5). 

Thus, in a few cases, phosphorus limitation might 

be indicated using censored data from the NWQL, 

but using data from CCAL, with less positive bias 

in TKN and higher TP concentrations, nitrogen 

limitation is almost universally indicated. Similar 

results are obtained for the ratio of dissolved inorganic nitrogen to soluble reactive 
34

Figure 9. Distribution of nutrient concentrations in the Little River 

and tributaries during August, 1999. (All data are from the U.S. 

Geological Survey’s National Water Quality Laboratory (NWQL) in 

Denver, Colorado, except dissolved Kjeldahl nitrogen (DKN) and 

total Kjeldahl nitrogen (TKN), which include data from both NWQL 

and Cooperative Chemical Analytical Laboratory (CCAL) at Oregon 

State University.)

phosphorus (DIN:SRP) (table 5). Similar conclusions can also be drawn from the August 1998 data. 

Although median TP concentrations during 1998 

(~12 μg/L) were lower than from 1999, N-limitation 

is supported by a median DIN:SRP ratio of ~1.

Little River watershed nitrogen concentrations (tables 4 and 5) are lower than the national 

median flow-weighted NO3-N and total N (NO3-N 

+ TKN) concentrations (87 and 260 μg/L, respectively) reported by Clark et al. (2000) for relatively 

undeveloped streams. In contrast, Little River 

watershed phosphorus concentrations are roughly 

equivalent to national median flow-weighted concentrations in those streams (TP=22 μg/L, 

SRP = 10 μg/L). These comparisons are made with 

caution, however, because data are not available to 

determine flow-weighted median nutrient concentrations for the Little River watershed. Also, data 

for the Little River watershed are from summer, 

when benthic uptake and the lack of particulate 

matter probably cause concentrations to be lower 

than flow-weighted median concentrations would 

be. Even so, during the stormflows sampled in 

November 1998, nitrate and total N concentrations 

were equivalent to or lower than those reported by 

Clark et al. (2000), and TP was only slightly elevated. Together these comparisons support the suggestion that streams in the Little River watershed 

are strongly nitrogen limited, indicating that algal 

growth could be stimulated by inputs of fertilizer 

nitrogen to streams, even in small quantities. 

Periphytic algal biomass (reported as ash

free dry mass, or AFDM) was generally low in 

streams of the Wolf Creek subwatershed, but relatively abundant in Little River (table 8). In many 

cases, rocks in the smaller Wolf Creek streams 

lacked visible algal growth, though chlorophyll a

analysis indicated nominal growth. Algal biomass 

within the Wolf Creek subwatershed was visibly 

heaviest on the rock wall of Wolf Creek Falls, 

which was not sampled during this survey. The 

highest measured biomass in the subwatershed, as 

AFDM, was at the mouth of Wolf Creek, although 

chlorophyll a in Egglestron Creek was similar to 

that at the mouth of Wolf Creek. Algal biomass in 

Little River was highest below Wolf Creek and at 

the Peel gaging station, and chlorophyll a was also 

highest below Wolf Creek. During the August survey, algal nuisance conditions (mats with filaments 

several feet long) were not observed; however nuisance growths of green algae, with filaments up to 

several feet long, were observed during a brief 

inspection in early September 1999 at the mouth of 

Wolf Creek, and in isolated mats in Little River 

above and below Wolf Creek. Macroinvertebrate 

grazers (a case building caddisfly of the genus 

Dicosmoecus) were abundant in Little River and 

the lower reaches of Wolf Creek in early August, 

and likely contributed to keeping algal biomass 

low. Their subsequent emergence in the warm 

waters sampled during mid-August may have 

allowed algal growth to accelerate afterwards. 

Spatial patterns of water quality in the watershed (fig. 10) generally followed the conceptual 

model proposed previously (fig. 3). For instance, 

NO3-N concentrations during the 1998 survey were 

generally higher at upstream tributary sites where 

dense riparian shading prevented light penetration, 

somewhat lower at downstream tributary sites 

where more light was available, and lowest in the 

main stem where light is not limiting and algal 

uptake apparently reduces nutrient concentrations.

  	 	



















  	 


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(2) (4)

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35

.

Table 8. Algal biomass and chlorophyll a measured in the Little River watershed, August 1999

[AFDM, ash free dry mass; chl a, chlorophyll a; g/m2, grams per meter squared; mg/m2, milligrams per meter squared] 

Sampling site Site number

AFDM

g/m2

Chl a

mg/m2

Unnamed tributary to West Fork Wolf Creek in 

section 32 at road 32.0 30 2.2 2.6

Wolf Creek at USFS road 14.0 19 3.7 9.4

Wolf Creek above falls 23 2.6 7.8

Wolf Creek below falls 33 4.0 14

Wolf Creek above Egglestron Creek 4 6.5 19

Wolf Creek at mouth 1 17 24

West Fork Wolf Creek 6 2.1 11

Unnamed tributary to West Fork Wolf Creek 7 2.3 5.5

Reference tributary to Wolf Creek above 

Egglestron Creek 24 4.9 7.0

Unnamed tributary to West Fork Wolf Creek, 

number 2 31 2.8 2.9

Egglestron at mouth 2 3.3 27

Negro Creek at mouth 18 2.7 11

Little River at White Creek Campground 16 14 21

Little River above Job Corp 15 16 35

Little River above Wolf Creek 14 18 60

Little River below Wolf Creek 22 20 68

Little River at Peel 21 22 57

The one main-stem site at which NO3-N was greater 

than 10 μg/L was just upstream of Wolf Creek, illustrating possible NO3-N inputs from the youth camp, 

Job Corps Center, or individual streamside residences. In contrast, total-phosphorus concentrations 

were moderate at most sites except the smallest tributaries sampled and in Little River at sites below 

Wolf Creek, where concentrations were again 

reduced by uptake. Daily maximum temperature and 

pH were each typically higher in the Little River 

than in the tributaries, except at the most upstream 

sites surveyed in Little River. Maximum pH was 

lower overall at upstream, shaded sites and somewhat higher at downstream sites. Patterns of chlorophyll a from 1999 followed those of pH. 

Chlorophyll a and pH were highest in the main stem 

of the Little River below Emile Creek, where an 

open canopy likely indicates that algal growth is not 

limited by light availability, and nutrient sources are 

apparently adequate to support relatively high primary production. During August 1999, NO3-N was 

low (<10 μg/L) throughout the upper main stem Little River, but it was relatively high (> 40 μg/L) at 

the mouth of the shaded but heavily timbered (and 

possibly recently fertilized) Negro Creek subwatershed. However, associations between the waterquality parameters in figure 10 and stream order or 

elevation were not statistically significant, indicating that a variety of processes and conditions must 

be considered in the Little River watershed for studies of forest fertilization to be conclusive. 

FRAMEWORK FOR FERTILIZATION STUDY IN 

WOLF CREEK AND LITTLE RIVER WATERSHED

In keeping with its mission under the President’s Northwest Forest Plan (U.S. Forest Service 

and Bureau of Land Management, 1994), the BLM 

plans to study the effects of forest fertilization with 

urea-N on water quality and stream ecology in the 

LRAMA to determine if fertilization adversely 

affects water quality and stream biota. Specifically, 

the objectives of the study are to determine: 

 1. Effects of fertilizer nutrient inputs on the 

aquatic ecosystem, including algae and 

higher trophic levels, such as macroinvertebrates and (or) amphibians, 
36

Figure 10. Concentrations of nitrate-nitrogen and total phosphorus, and daily maximum temperature and pH in the Little River Basin during

August, 1998. (Sampling sites are given in table 3 and figure 5, and water quality data are provided in tables 4 and 5.)

Main stem 

 <10

 10-25

 >25

Tributaries

 <10

 10-25

 >25

Nitrate-N (μg/L)

Maximum temperature (Degrees Celsius)

Main stem

 <12

 12-16.8

 >16.8

Tributaries

 <12

 12-16.8

 >16.8

Total P (μg/L)

Main stem

 <10

 10-25

 >25

Tributaries

 <10

 10-25

 >25

Maximum pH

Main stem

 <8

 8-8.5

 >8.5

Tributaries

 <8

 8-8.5

 >8.5

 2. Relations in the Little River watershed 

among nutrient inputs (natural and anthropogenic), watershed characteristics, and 

water quality, particularly pH, DO, nutrients, and temperature, and 

 3. Downstream cumulative effects, both spatial and temporal, of forest fertilization on 

water quality and aquatic-biological systems.

An investigation into ecological effects from 

forest fertilization in the LRAMA will require care 

and detail to distinguish effects of fertilization 

from natural processes and other land uses. Ancillary factors include the history of timber production on alternating federal and private lands, 

mixed-age stands (including recent clearcuts), 

recent forest fertilizations on some private lands, 

and residential and recreational land adjacent to the 

Little River. Existing extremes of pH and temperature that already suggest possible degradation from 

natural conditions may confound interpretation of 

water-quality data. 

An obvious approach would be to used paired 

drainages, or sets of pairs, to provide controls and 

treatments. Where possible, this approach is suggested for fertilization studies. Unfortunately, few, 

if any, drainages in the Little River watershed’s 

study area are small enough to contain somewhat 

homogeneous upstream land uses (especially without private timberland in the subbasins) and yet 

large enough to avoid drying in summer. The best 

choices for control streams include one small drainage in the Wolf Creek subwatershed with predominantly old-growth trees, and several possible 

locations in tributaries to Little River well 

upstream of Wolf Creek. Furthermore, benthic 

growth in the small tributary streams and immediately below the areas to be fertilized is most likely 

limited by light rather than nutrients (fig. 3). Thus, 
37

fertilizer nutrients are expected to be transported 

downstream to lower reaches of Wolf Creek and 

Little River, where investigation of cumulative

biological effects would be most confounded. This 

lack of suitable subbasin pairs necessitates an alternative approach, with emphasis on comparison of 

conditions before and after fertilization and on longitudinal changes in study streams. Longitudinal 

evaluations involving an upstream-downstream 

comparison need to take into account the possible 

inputs from upstream private timber lands with 

unknown histories (fig. 4) and the likely changes in 

stream function along an elevational or streamorder gradient (fig. 3). 

Adherence to water-quality criteria for nutrients is not perceived as an effective benchmark to 

determine fertilization’s effects on water quality in 

this case, with the possible exception of pulsed 

ammonia toxicity during rainstorms immediately 

following fertilization. Rather, the focus will primarily be on biological endpoints such as measures 

of biomass (AFDM and chlorophyll a), changes in 

algal community structure (autecology, species 

diversity, and dominant species types) and function 

(primary production, nutrient sequestration and 

uptake), secondary effects on water quality (DO 

and pH), and possibly secondary interactions with 

higher trophic levels (macroinvertebrate grazing or 

amphibian abundance). Nutrient processes will be 

investigated to provide insights into relevant ecological processes, evaluate transport, and make 

comparisons among stations. 

Ideally, incorporation of fertilizer-nutrient 

into biological tissues could be traced using unique 

signatures of naturally occurring 15N (herein 

termed “natural-15N”), the stable isotope of nitrogen (Lajtha and Michener, 1994; Kendall and 

McDonnell, 1998). Natural isotopes of oxygen 

(

18O) can also be used in conjunction with 15N to 

determine hydrologic flow paths and water sources 

(Kendall and McDonnell, 1998). However, urea 

fertilizer typically has a natural-15N signature that 

generally cannot be differentiated from natural-15N 

found in forest soils, with δ15N values near 0 ‰2. It 

is possible that volatilization, nitrification, and 

uptake processes in forest floors and along hydrologic pathways would cause fractionation of the 

urea’s 15N to a heavier fraction, producing a traceable signal (Udy and Dennison, 1997). However, 

several algal and moss samples from throughout the 

watershed, including downstream of Federal lands, 

private lands fertilized in fall 1998, and potential 

septic or WWTP influences in Little River, did not 

indicate a strong enough gradient (δ15N>3 ‰) to 

identify sources or processes. Thus it is unlikely 

that natural-15N alone can adequately be used to 

trace urea-N transport from forest fertilization 

through the forest floor and into aquatic biota. 

Nonetheless, additional reconnaissance of natural15N, alone or in conjunction with 18O, and testing 

of the urea to be applied, is warranted because of 

the great advantage that this technique would provide if a distinct fertilizer signature were available. 

There have been many studies of N movement in forest floors or into trees using fertilizers 

artificially enriched in 15N (herein referred to as 

“labelled-15N”) to ensure a distinct tracer (Marshall 

and McMullan 1976; Nason et al., 1988; Preston et 

al., 1990; Fry et al., 1995; Downs et al., 1996; Jordan et al., 1997), but no studies to date have traced 

the movement of these isotopes from the forest 

floor to streams or aquatic biota. Although this 

technique is likely to be the most definitive way to 

trace the effect of urea-N into streams, it would 

also be expensive for even one of the fertilizer units 

in the BLM’s proposed fertilization. A coarse cost 

estimate was made for an upper elevation drainage 

basin with minimal upstream influence from private land and a flowing stream in 1999 (above site 

30, in section 32 of the USGS Red Butte 7-1/2 

minute topographic map). Using basic costs from 

Fry et al. (1995), and scaling the effort to the 0.2-

square-mile (52 ha) fertilization unit immediately 

upstream of site 30, with an application rate of 224 

kg/ha, yields an estimate of $75,000, including purchase of the labelled urea and logistical costs of 

mixing the labelled urea with nonlabelled urea 

prior to application. Furthermore, application of 

labelled-15N to a limited area could prove to be 

highly useful near the area of application, but its 

signature might be diluted below detection in 

2 15N enrichment is measured in a relative sense compared 

to a known reference material. The delta value, expressed as δ15N 

and with units of ‰ (parts per thousand), is determined as 

δ = (Rx/Rs-1)*1000, where R is the ratio of the heavy to light 

isotope in the standard (S) and the sample (X). A positive δ value 

means that the isotopic ratio of the sample is higher than in the 

standard, and a negative δ value means that the sample is isotopically lighter than the standard. 
38

downstream reaches most likely to respond biologically to increased nutrients. Thus, the use in this 

study of labelled urea to trace the movement of 

nitrogen into streams and ecological compartments, 

while an attractive method, could be too expensive 

for practical use at the drainage-basin scale. More 

detailed cost estimates of this method for specific 

locations would be warranted prior to final decisions about its use. 

Possible Study Approaches

The level of investigation at Little River, and 

therefore the degree to which the study would 

address its objectives, will greatly depend on available resources. Table 9 indicates two possible levels of research and associated activities to address 

fertilizer impacts. Although these approaches are 

targeted to the Little River watershed and some of 

its specific complications, the concepts could apply 

broadly to other investigations of the effects of forest fertilization on aquatic systems. 

A basic investigation would examine study 

streams for gross biological responses (table 9) 

before and after fertilization. The focus would be 

on the Wolf Creek subwatershed, although a few 

sites in the Little River would be sampled as well. 

The relative loading of nutrients to streams from 

fertilized and unfertilized areas would be determined. In order to separate confounding upstream 

influences in the Little River watershed, relations 

between catchment scale characteristics such as 

upstream forest-and land-management history, 

slope, riparian vegetation, geology, surface and 

subsurface hydrology, stream morphology and 

water-quality constituents would also be considered. Cumulative effects downstream would be 

evaluated at a relatively coarse level. 

A variety of sampling activities would be utilized for the above approach including standard 

water-quality analyses, plus periphytic algal biomass and species data, prior to and following fertilization. Synoptic surveys would provide 

“snapshots” (Salvia et. al, 1999) of summer lowflow conditions before and after fertilization. 

Monthly sampling at a few sites during summer 

would provide data on algal biomass and succession, and on the variability of nutrients and algae. 

Using these data, gross summer-nutrient loading 

could be estimated, major sources of water and 

nutrients would be defined, and estimates could be 

made of fertilizer effects on algal growth. Majorion data would help evaluate geological influences 

on water quality and quantity, and possibly indicate 

a catchment response to fertilizer through ion 

exchange. Recording monitors for temperature, pH, 

dissolved oxygen, and specific conductance would 

be used to define diel variability in those parameters, and the timing and magnitude of their seasonal 

maxima. The validity of the assumption that primary production is nitrogen limited would be tested 

in an assay using nutrient diffusing substrates. A 

reconnaissance of possible ground-water inputs 

would be done by sampling seeps and mass balancing streamflows. Potential signatures of different 

nutrient sources using natural-15N levels in water 

and algal tissues would be assessed at a few locations. Selection of sites longitudinally within the 

Wolf-Creek subwatershed and the Little River will 

allow differentiation of runoff from upland fertilized stands compared with unfertilized stands in 

mixed-use forested areas, as well as generalized 

cumulative effects downstream. This analyses 

would be aided with broad characterization of 

upstream land uses from existing GIS data layers. 

If there is a large biotic response to fertilization (objective 1, page 35), the basic approach 

above might successfully detect it. However, with 

the variability of forest management history in 

Wolf Creek, and of upstream nutrient sources to 

Little River, it is likely that subtle effects on biota 

or subtle cumulative changes in water quality 

would not be attributable specifically to any one 

cause. Nor would this effort generate information 

about the relative retention or loss of applied ureaN or its downstream transport, or define potentially 

important transport processes (through riparian or 

hyporheic zones, benthic recycling, or spiraling 

of nutrients). Relations among nutrients, riparian 

characteristics, aquatic habitat, or other water-quality parameters such as pH could also be tenuous.

Effects of fertilizer-N on higher trophic levels and questions about cumulative impacts could 

be better addressed with a more extensive level of 

study (table 9). This could include expanded efforts 

to evaluate (1) the status of water quality and nutrient sources in the Little River above and below 
39

Table 9. Research components for different levels of investigation of effects of urea fertilization on water quality 

and stream ecology

[Research objectives are given on page 35. esp, especially; DO, dissolved oxygen; SC, specific conductance; GIS, Geographic Information System]

Research 

level Objectives addressed Approach Activities

Basic

Gross responses in nutrient 

concentrations and algal biomass 

Relations in Little River 

watershed among watershed 

characteristics and water quality 

(esp. pH, DO, nutrients inputs, 

and temperature)

Basic analysis of water quality and algal 

growth, primarily within Wolf Creek subbasin 

but including some sites in Little River

Determine gross loading of nutrients to streams 

and major sources

Determine possible differential nutrient 

sources or hydrologic flow paths using 

streamflow and geochemical data and 

reconnaissance of ground-water inputs

Evaluate land use and relate to water quality.

Determine gross cumulative effects on Little 

River

• Late summer synoptic surveys before and after 

fertilization. Include nutrients, major ions, daily 

maximum pH and DO, algal biomass and species 

composition. Evaluate stable isotopes as possible 

indicators of differences in N sources. Include a few 

sites in Little River to evaluate influences of upstream 

nutrient sources

• Monthly sampling for nutrients, algal biomass, algal 

nutrient content (including algal slough/drift), and 

field parameters at select sites

• Verification of limiting nutrient using nutrient 

diffusing substrates

• Recording monitors for continuous measurement of 

flow and field parameters (pH/DO/SC/temp) during 

selected seasons before and after fertilization

• Reconnaissance of ground water in basin, including 

nutrients, pH, and stable isotopes (natural-18O or 

natural-15N) 

• Characterize upstream land use with GIS data layers, 

including fertilization and harvest history in watershed

Extensive

Previous objectives, with 

additional detail, plus

Cumulative downstream effects 

(spatial and temporal), of forest 

fertilization on 

• Water quality, and

• Aquatic biological systems

Basic level, plus:

More intensive evaluation of cumulative 

effects on Little River, including multiple 

upstream land uses

Evaluation of nutrient retention, transport, and 

fate, including percentage of fertilizer N in 

different compartments

Relation of fertilization effects on stream biota 

with riparian condition

More intensive evaluation of ground water and 

hyporheic processes

Evaluation of effects of fertilization on higher 

food webs

Evaluation of longer term effects from 

fertilization

Experimental tracer of urea with 15N in 

selected fertilization units

Level 2, and:

• Expanded network of sites in Little River and Wolf 

Creek, and use of tracers (Little River only) for septic 

waste 

• Possible use of isotopically labeled urea-N in selected 

areas

• Sample for 15N (naturally occurring or artificially 

labeled) in algae, sediments, and biota from hyporheic 

zones, streambeds

• Nutrient sampling during and immediately after 

fertilization, and measuring effective fertilization rate 

on ground

• Establish piezometer network near selected stream 

reaches to determine flow paths, transient storage, and 

nutrient transformation/retention in hyporheic zones

• Measure primary production in treatment and 

reference streams

• Macroinvertebrate (and possibly aquatic amphibian) 

sampling in conjunction with algal sampling

• In-depth GIS analysis, including influences of 

confounding upstream land uses, measuring and 

mapping riparian conditions

• Nutrient and algal sampling in second summer 

following fertilization

Wolf Creek, (2) nutrient transport and retention, (3) 

the relative importance of riparian buffers and vegetation in modifying stream response to fertilization, (4) ground-water inputs, including regional 

flow and local, hyporheic transformations, (5) effect 

of altered nutrient regimes on higher trophic levels, 

(6) long-term (> 1 year) contributions of fertilizerN to streams, and (7) possible use of urea with 

labelled-15N. 

Study elements for this comprehensive 

approach would involve sampling at more sites 

in Wolf Creek and in the Little River in order to 

evaluate possible septic or sewage inputs to Little 

River and reference sites outside of Wolf Creek 
40

subwatershed. It would also include an assessment 

of major ions, and possibly 18O to evaluate source 

water based on potential geochemical differences 

among sites. To investigate the potential for longterm changes in nutrient regimes, sampling would 

be extended into fall and possibly an additional 

summer of the second year following fertilization. 

In addition to mass balances on nutrients and major 

ions, inputs from possible septic or sewage sources 

to the Little River could be traced by using either 

natural-15N, caffeine, or other methods. Retention 

of nutrients in streams would be assessed by analyzing benthic and drifting algae for tissue-nutrient 

content and calculating the amount of N and P 

retained or transported in biomass. Communitylevel effects on algae would be determined by 

multivariate analysis of periphyton species data and 

by measuring metabolic rates prior to and following fertilization. Higher trophic levels would be 

investigated by sampling of macroinvertebrates in 

conjunction with algae in summers prior to and following fertilization to investigate effects on secondary consumers. If reconnaissance data indicate 

that natural-15N will prove useful in following fertilizer N movement, or if urea with labelled-15N is 

used, macroinvertebrates would also be sampled 

for 15N levels. GIS data layers, including mapping 

of riparian areas and (if possible) private-timberland management and fertilization history, would 

be generated. These data would be used to help 

define the importance of various influences on 

water quality at downstream stations. Depending 

on available resources this study would also 

include installation of test wells near selected 

stream sites to investigate ground-water exchange 

with streams and localized nutrient dynamics in 

hyporheic zones. 

Together, these approaches would help to 

more definitively determine changes in algal communities among sites affected and not affected by 

fertilization, secondary effects on macroinvertebrates, and relations between nutrient concentrations (in water and algal tissues) and water-quality 

parameters. They would also provide a better 

opportunity to observe cumulative effects in downstream reaches, including Little River, and differentiate them from effects of other land uses. 

Ideally, models would eventually be constructed to 

evaluate the effects of fertilizer N inputs on stream 

biota and water quality; however, modeling periphytic systems is still relatively imprecise. Spreadsheet models of ground-water-N input and stream 

dynamics have been developed for intensively studied streams (Peterson et al., 2001), but they do not 

model primary production or its resulting affects on 

DO and pH. The Oregon Department of Environmental Quality (2000) recently developed a model 

for use in setting TMDLs in periphytic streams. 

This model predicts DO and pH as a function of 

nutrient concentration, and may work reasonably 

well for streams where point sources have been 

reduced, but has not yet been tested for systems 

with diffuse nutrient sources. 

SUMMARY

Forests in the Pacific Northwest and elsewhere have long been fertilized to increase timber 

productivity, with over 120,000 acres per year 

being fertilized in the Pacific Northwest in the late 

1980’s. Recent (1990–98) fertilization levels in 

Oregon have averaged approximately 95,000 acres 

annually. A review of literature on water-quality 

effects from fertilization of forests with nitrogen 

indicates that applied nitrogen does indeed run off 

to streams, in amounts ranging from less than 1% to 

as much as 27% of applied nitrogen. The amount of 

applied nitrogen lost to streams depends on many 

factors, including the amount and form of fertilizer 

applied, timing of application (usually fall), 

weather during and after application, degree to 

which the application was able to avoid direct input 

to streams, width of riparian buffers, nitrogen status of soils in the watershed, hydrologic processes 

in the watershed (including ground-water residence 

time), and history of forestry or other land-use 

practices in the watershed. Invariably there have 

been high-concentration pulses of nitrogen, usually 

as urea (or total Kjeldahl nitrogen) and NH3-N 

(ammonia-nitrogen), during runoff immediately 

following applications, with subsequent decreases 

in concentrations. Subsequent increases in NO3-N 

(nitrate-nitrogen) concentrations can be more prolonged, often for the duration of the winter and 

spring. Summer NO3-N concentrations are frequently low, often resembling background, but usually have been elevated during the following fall in 

streams draining treated watersheds. 
41

Despite these increases, water-quality criteria 

for nutrients have almost always been met, except 

in rare instances such as where soils were already 

nitrogen saturated. However, water-quality criteria 

for nutrients are targeted towards human health (for 

NO3-N) or aquatic toxicity (for NH3-N), and are 

not set at levels relevant to ecologic processes in 

most forested aquatic ecosystems. Biological processes following fertilization have rarely been studied, and most were completed prior to the 

development of key concepts of nutrient processing 

and ecological dynamics in streams. In several 

cases, techniques were not sensitive to potential 

processes in the streams studied. Meanwhile, many 

forest streams continue to indicate breakdown of 

ecological systems, from eutrophication to potential food-web alterations and loss of sensitive species. Thus, key questions about the ecological 

effects of forest practices remain unresolved. For 

these reasons, new approaches to evaluation of forest management practices, such as fertilization, are 

necessary. 

In Cascade streams of the Pacific Northwest, 

productivity in mountainous streams, like forests, 

is typically nitrogen limited. Increases in nitrogen 

inputs to streams can potentially increase primary 

production, and possibly alter successional patterns, community dynamics, and trophic structure 

of benthic communities. Nutrient inputs have long 

been linked to occurrences of nuisance algal 

growth in many streams, with secondary effects on 

water quality (DO and pH) from algal metabolism. 

These situations are increasingly frequent in forested systems. 

Pathways for nitrogen input to streams from 

upland disturbances include direct runoff, groundwater inputs, and hyporheic flow. Instream pathways for nitrogen processing, besides classical 

transport, include hyporheic retention and processing by microbial communities, uptake by benthic 

algae, and downstream transport by boundary layer 

recycling or transport of sloughed, particulate 

forms of algae. All of these processes can be 

extremely efficient and represent significant portions of the nitrogen budget of a stream. Yet most 

are ignored by standard approaches to water sample 

collection. Thus, the actual amount of nitrogen 

entering streams and contributing to ecological processes from upland sources (such as fertilization) 

may have been underestimated in some previous 

studies. 

The Little River watershed, in southwestern 

Oregon, has been designated as one of 10 Adaptive 

Management Areas (AMA’s) under the President’s 

Northwest Forest Plan. Forest land ownership in 

the watershed is predominantly Federal but private 

timberland also constitutes much of the watershed 

and is interspersed among many Federal tracts. 

Currently, water quality in the Little River during 

summers does not meet State standards for temperature or pH in some locations, and in many locations nuisance algal conditions are common. 

Nutrient concentrations are typically low and 

streams are generally nitrogen limited. 

To accompany a proposed operational fertilization of Federal (Bureau of Land Management) 

timberlands in portions of the watershed, a multilevel framework for investigation of water quality 

and ecological processes is suggested. The studies 

would focus primarily on biological endpoints but 

also would include hydrologic components and 

nutrient-data collection to help understand ecological processes. The different levels of study would 

help, to varying degrees, define the effects, if any, 

of fertilizer-nutrient inputs on aquatic ecosystems 

and processes, relations between nutrient inputs, 

watershed characteristics, and water quality, and 

finally, downstream cumulative effects on both 

water quality and aquatic-biological systems.

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U.S. Department of the Interior

U.S. Geological Survey

Ecological Effects on Streams from Forest Fertilization—

Literature Review and Conceptual Framework for 

Future Study in the Western Cascades

Prepared in cooperation with

the BUREAU OF LAND MANAGEMENT

Water-Resources Investigations Report 01–4047
Cover photographs: 

Upper: Little River above Wolf Creek (photograph by John C. Risley, U.S. Geological Survey).

Lower: West Fork Wolf Creek (photograph by Chauncey W. Anderson, U.S. Geological Survey).