Effects of Flow Modification on a Cattail

Wetland at the Mouth of Irondequoit Creek

near Rochester, New York

Water Levels, Wetland Biota, Sediment, and Water Quality

In cooperation with the

MONROE COUNTY DEPARTMENT OF HEALTH

Water-Resources Investigations Report 03-4224

U.S. Department of the Interior

U.S. Geological Survey
This page has been left blank intentionally.
Effects of Flow Modification on a Cattail 

Wetland at the Mouth of Irondequoit 

Creek near Rochester, New York

Water Levels, Wetland Biota, Sediment, 

and Water Quality

By WILLIAM F. COON

U.S. Department of the Interior

U.S. Geological Survey

In cooperation with the Monroe County Department of Health

Water-Resources Investigations Report 03-4224
U.S. Department of the Interior

Gale A. Norton, Secretary

U.S. Geological Survey

Charles G. Groat, Director

U.S. Geological Survey, Ithaca, New York: 2004

For more information write to:

U.S. Geological Survey

30 Brown Road

Ithaca, New York 14850

World Wide Web: http://ny.usgs.gov

For sale by U.S. Geological Survey, Information Services

Box 25286, Denver Federal Center

Denver, CO 80225

For more information about the USGS and its products:

Telephone: 1-888-ASK-USGS

World Wide Web: http://www.usgs.gov/

Any 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.

Although this report is in the public domain, permission must be secured from the individual copyright owners to reproduce any copyrighted materials contained within this report.
iii

CONTENTS

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

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

Irondequoit Creek basin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3

Previous studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

Irondequoit Creek basin studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

Ellison Park wetland studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

Purpose and scope . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

Study area . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

Study design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

Flow-control structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

Additional flow modifications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11

Ecological monitoring program. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

Water-surface-elevation and discharge measurements. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

Plant surveys . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

1991 and 1996 protocols . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

1999 and 2001 protocols . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

Fish surveys . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

Sampling methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

Species classification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

Age and growth-rate calculation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

Trophic classification. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Contaminant-related stress indicators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Benthic-macroinvertebrate surveys . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Bird surveys . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Sediment studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

Sedimentation rate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

Sediment quality. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

Surface-water data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

Stream discharge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

Water quality. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

Quality assurance and quality control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

Constituent loads and removal efficiency . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

Statistical analyses of load data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

Effects of flow modification. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

Water-surface elevations and discharge. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

Plants. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

Species richness and community composition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

Cattail productivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28

Chemical composition of cattail tissues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30

Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

Species richness and distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32
iv

Resident- and nonresident-fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

Growth rates of selected species . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

Trophic status of fish community . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

Contaminant-related stress indicators in brown bullhead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

Turtles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

Benthic macroinvertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

Community indicators of water quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38

Species indicators of water quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

Fish-stomach-contents data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

Birds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

Sediment characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

Sedimentation rates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

Sediment quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

Constituent loads and removal efficiencies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

Bias in water-quality data. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

Criteria for evaluation of flow-control structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

Total phosphorus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

Orthophosphate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

Ammonia-plus-organic nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

Ammonia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

Nitrate-plus-nitrite nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

Total and volatile suspended solids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67

Chloride and sulfate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 70

Concentrations of other constituents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 70

Water-quality improvement of Irondequoit Bay . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74

Annual total phosphorus load . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74

Trophic status of the Bay. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74

Suggestions for future monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74

Summary and conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76

References cited . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77

Reports of biological studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 81

FIGURES

1. Map showing location of study area and principal geographic features of Irondequoit 

Creek basin, Monroe County, N.Y. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3

2. Map showing locations of streamflow- and water-quality-monitoring sites in Ellison Park 

wetland, Monroe County, N.Y.. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

3. Graphs showing: A. Monthly precipitation (1990-2001) and normal mean monthly 

precipitation (1961-90) recorded at Greater Rochester International Airport, Rochester, N.Y. B. 

Monthly mean discharges of Irondequoit Creek above Blossom Road and at Empire Boulevard, 

Monroe County, N.Y., (1990-2001), with monthly maximum, mean, and minimum discharges

for the period of record at Blossom Road (1981-2001) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

4. Graph showing daily mean water-surface elevation of Ellison Park wetland, Monroe

County, N.Y., as measured at wetland outlet (Empire Boulevard), 1990-2001. . . . . . . . . . . . . . . . . . . . . . 9

5. Schematic diagram and photograph of flow-control structure at the Narrows in Ellison

Park wetland, Monroe County, N.Y., 2001. A. Schematic diagram. B. Downstream-facing 

view of flow-control structure with H-piles extending above water surface . . . . . . . . . . . . . . . . . . . . . 10
v

6-8. Maps of Ellison Park wetland, Monroe County, N.Y., showing:

6. Locations of flow modifications to Irondequoit Creek and the Millrace, 1997-98 . . . . . . . . . . . . . . . . . . . 11

7. Locations of data-collection sites for plant and fish surveys, 1991-2001. . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

8. Locations of data-collection sites for sedimentation-rate and sediment-quality studies, 

1991-2002 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

9. Graph showing peak discharges recorded at Irondequoit Creek above Blossom Road,

Rochester, N.Y., 1990-2001. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

10. Graphs showing cattail (Typha glauca) biomass measured during five surveys along 

permanent transects in Ellison Park wetland, Monroe County, N.Y., 1991-2001 . . . . . . . . . . . . . . . . . . . . . . 30

11. Graphs showing thickness of accumulated sediment at sedimentation-rate-measurement sites 

in Ellison Park wetland, Monroe County, N.Y., 1991-2002: A. Upstream from flow-control structure. 

B. Downstream from flow-control structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

12. Box plots showing concentrations of 16 constituents in and discharges associated with 

surface-water samples collected during pre- and postcontrol periods from inflow (Blossom 

Road) and outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, N.Y., 

1990-2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

13-21. Graphs showing monthly loads and flow-weighted concentrations, and wetland’s removal

efficiency for selected constituents, Ellison Park wetland, Monroe County, N.Y., 1990-2001:

13. Total phosphorus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63

14. Orthophosphate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64

15. Ammonia-plus-organic nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 65

16. Ammonia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 66

17. Nitrate-plus-nitrite nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68

18. Total suspended solids. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69

19. Volatile suspended solids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71

20. Chloride . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72

21. Sulfate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73

22. Graph showing trophic status of Irondequoit Bay, 1971-2001, as function of measured summer 

mean concentrations of chlorophyll_a and computed concentration of potential phosphorus . . . . . . 75

TABLES

1. Components of ecological-monitoring program for Ellison Park wetland study, Monroe County, 

N.Y., 1991-2002, by year . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

2. Sampling methods for fish surveys in Ellison Park wetland, Monroe County, N.Y., 1991-2001 . . . . . . . . . 16

3. Performance ratings for selected standard reference samples analyzed by Monroe County 

Environmental Health Laboratory, 1990-2001 (results of U.S. Geological Survey’s analytical 

evaluation program.) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

4. Explanatory variables used in regression equations developed to compute loads for selected 

constituents in water from Irondequoit Creek, Monroe County, N.Y., 1990-2001, at sites above 

Blossom Road (inflow to wetland) and Empire Boulevard (outflow from wetland) . . . . . . . . . . . . . . . . . . . 22

5. Regression statistics for equations used to compute constituent loads for Irondequoit Creek 

sampling sites above Blossom Road and at Empire Boulevard, Monroe County, N.Y., 1990-2001 . . . . . 24

6. Factors addressed in assessment of hydrologic effects of flow-control structure, Ellison Park 

wetland, Monroe County, N.Y., 1997-2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

7. Measurements obtained in plant surveys along three permanent transects in Ellison Park 

wetland, Monroe County, N.Y., 1991-2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27

8. Cattail biomass in above- and below-ground tissues along three permanent transects in 

Ellison Park wetland, Monroe County, N.Y., 1991-2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28

9. Above- and below-ground Typha spp. biomass reported in the literature, and average values

for each of three permanent transects sampled during four surveys in Ellison Park wetland, 

Monroe County, N.Y., 1991-2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29
vi

10. Mean concentrations of selected elements in above- and below-ground cattail tissues collected 

along three permanent transects in Ellison Park wetland, Monroe County, N.Y., 1991-2001 . . . . . . . . . . 31

11. Fish species found in Ellison Park wetland, Monroe County, N.Y., 1991-2001 . . . . . . . . . . . . . . . . . . . . . . . . . 33

12. Results of fish surveys in Ellison Park wetland, Monroe County, N.Y., 1991-2001. . . . . . . . . . . . . . . . . . . . . . 34

13. Growth rates of selected fish species in Ellison Park wetland, Monroe County, N.Y., 1991-2001 . . . . . . 35

14. Trophic status of fish community in Ellison Park wetland, Monroe County, N.Y., 1991-2001 . . . . . . . . . . . 36

15. Indicators of contaminant-related stress found in brown bullheads in Ellison Park wetland, 

Monroe Country, N.Y., 1991-2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

16. Turtle species found in Ellison Park wetland, Monroe County, N.Y., 1991-2001 . . . . . . . . . . . . . . . . . . . . . . . 37

17. Benthic macroinvertebrates found in Ellison Park wetland, Monroe County, N.Y., 

September-October 1996 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38

18. Physical characteristics of sampling sites and benthic-macroinvertebrate measures of 

ecological health in southern and northern areas of Ellison Park wetland, Monroe County, N.Y., 

September-October, 1996 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

19. Bird species observed in or near Ellison Park wetland, Monroe County, N.Y., 1980-2002, and 

designation of breeding status in wetland (possible, probable, or confirmed) . . . . . . . . . . . . . . . . . . . . . . . . 42

20. Size distribution of sediment samples collected for chemical analyses from Ellison Park wetland, 

Monroe County, N.Y., 1994-2001, by percentage of particles finer than index-particle size. . . . . . . . . . . . 45

21. Sediment thickness and net sedimentation rates in Ellison Park wetland, Monroe County, N.Y., 

1991-2001. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

22. Chemical concentrations in sediment samples from Ellison Park wetland, Monroe County, N.Y., 

1994-2001, and sediment-quality guidelines for selected chemicals: A. major and trace 

elements. B. Polycyclic aromatic hydrocarbons. C. Organochlorine and organophosphate 

compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48

23. Mean difference (bias) between concentrations of selected constituents in water samples 

collected by automatic sampler (Auto) and equal-width-increment method (EWI), and 

comparative statistics for concentrations of constituents in quality-assurance/quality-control 

samples (QA/QC) and all automatically collected samples at the inflow (Blossom Road) and 

outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, N.Y., 1990-2001 . . . . . . . . . . . . . . . . 53

24. Concentrations and summary statistics of selected constituents, Ellison Park wetland, Monroe 

County, N.Y., 1990-2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

25. Average annual inflow and outflow loads of selected constituents, and wetland’s removal 

efficiency for each constituent during pre- and postcontrol periods, Ellison Park wetland, 

Monroe County, N.Y., 1991-2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

26. Statistics (p values) related to tests of significant difference between the monthly constituent loads 

and flow-weighted concentrations, as computed for inflow (Blossom Road) and outflow 

(Empire Boulevard) of Ellison Park wetland, Monroe County, N.Y., 1990-2001 . . . . . . . . . . . . . . . . . . . . . . . . 62

APPENDIX: Monthly and annual inflow and outflow loads, associated errors, and removal efficiencies for 

selected constituents, Ellison Park wetland, Monroe County, N.Y., 1996-2001: 

A. Phosphorus compounds. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 82

B. Nitrogen compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84

C. Suspended solids. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87

D. Chloride and sulfate. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89
vii

Conversion Factors and Datum

INCH-POUND TO INTERNATIONAL SYSTEM (SI) UNITS

Multiply By To obtain

Length

inch (in.) 2.54 centimeter (cm)

inch (in.) 25.4 millimeter (mm)

foot (ft) 0.3048 meter (m)

mile (mi) 1.609 kilometer (km)

Area

acre 4,047 square meter (m2

)

square foot (ft2) 0.09290 square meter (m2)

square mile (mi2

) 2.590 square kilometer (km2)

Volume

cubic foot (ft3) 0.02832 cubic meter (m3)

acre-foot (acre-ft) 1,233 cubic meter (m3)

Flow rate

cubic foot per second (ft3/s) 0.02832 cubic meter per second (m3/s)

Mass

ounce, avoirdupois (oz) 28.35 gram (g)

pound, avoirdupois (lb) 0.4536 kilogram (kg)

ton, short (2,000 lb) 0.9072 megagram (Mg)

 INTERNATIONAL SYSTEM (SI) TO INCH-POUND UNITS

Multiply By To obtain

Length

centimeter (cm) 0.3937 inch (in.)

millimeter (mm) 0.03937 inch (in.)

meter (m) 3.281 foot (ft)

kilometer (km) 0.6214 mile (mi) 

Area

hectare (ha) 2.471 acre

square meter (m2

) 10.76 square foot (ft2)

square kilometer (km2) 0.3861 square mile (mi2)

Volume

cubic meter (m3) 35.31 cubic foot (ft3) 

cubic meter (m3

) 0.0008107 acre-foot (acre-ft) 

Flow rate

cubic meter per second (m3/s) 35.31 cubic foot per second

Mass

gram (g) 0.03527 ounce, avoirdupois (oz)

kilogram (kg) 2.205 pound, avoirdupois (lb)

gram per square meter (g/m2) .0002049 pound, per square foot (lb/ft2)

megagram (Mg) 1.102 ton, short (2,000 lb)
viii

Specific conductance is given in microsiemens per centimeter at 25 degrees Celsius 

(μS/cm at 25 ̊C).

Temperature in degrees Celsius ( ̊C) may be converted to degrees Fahrenheit ( ̊F) as follows:

 ̊F = (1.8 x  ̊C) + 32

Time: h, hour; min, minute; s, second

Concentrations of chemical constituents in water are given either in milligrams per liter 

(mg/L) or micrograms per liter (μg/L).

Vertical coordinate information is referenced to the National Geodetic Vertical Datum of 

1929 (NGVD of 29).

Horizontal coordinate information is referenced to the North American Datum of 1927 

(NAD 27).
Effects of Flow Modification on a Cattail Wetland at the 

Mouth of Irondequoit Creek near Rochester, New York— 

Water Levels, Wetland Biota, Sediment, and Water Quality

by William F. Coon

ABSTRACT

An 11-year (1990-2001) study of the Ellison Park wetland, 

a 423-acre, predominantly cattail (Typha glauca) wetland at the 

mouth of Irondequoit Creek, was conducted to document the 

effects that flow modifications, including installation of a flowcontrol structure (FCS) in 1997 and increased diversion of 

stormflows to the backwater areas of the wetland, would have 

on the wetland’s ability to decrease chemical loads transported 

by Irondequoit Creek into Irondequoit Bay on Lake Ontario. 

The FCS was designed to raise the water-surface elevation and 

thereby increase the dispersal and detention of stormflows in 

the upstream half of the wetland; this was expected to promote 

sedimentation and microbial utilization of nutrients, and 

thereby decrease the loads of certain constituents, primarily 

phosphorus, that would otherwise be carried into Irondequoit 

Bay. An ecological monitoring program was established to 

document changes in the wetland’s water levels, biota, 

sedimentation rates, and chemical quality of water and sediment 

that might be attributable to the flow modifications. 

Water-level increases during storms were mostly confined 

to the wetland area, within about 5,000 ft upstream from the 

FCS. Backwater at a point of local concern, about 13,000 ft 

upstream, was due to local debris jams or constriction of flow 

by bridges and was not attributable to the FCS.

Plant surveys documented species richness, concentrations 

of nutrients and metals in cattail tissues, and cattail 

productivity. Results indicated that observed differences among 

survey periods and between the areas upstream and downstream 

from the FCS were due to seasonal changes in water 

levels—either during the current year or at the end of the 

previous year’s growing season—that reflected the watersurface elevation of Lake Ontario, rather than water-level 

control by the FCS. Results showed no adverse effects from the 

naturally high water levels that prevail annually during the 

spring and summer in the wetland, nor from the short-duration 

increases in water levels that result from FCS operation. 

Fish surveys documented the use of the wetland by 44 

species, of which 25 to 29 species were found in any given year. 

Community composition was relatively consistent during the 

study, but seasonal and year-to-year variations in dominant 

resident and nonresident species were noted, and probably 

reflected natural or regional population patterns in Lake Ontario 

and Irondequoit Bay. The FCS allowed fish passage at all water 

levels and had no discernible adverse effect on the fish 

community.

Bird surveys documented the use of the wetland by more 

than 90 species for breeding, feeding, and migration. Groundnesting birds were unaffected by the FCS. Seasonally high 

water levels, rather than short-duration increases caused by the 

FCS, might have caused the scarcity or absence of certain 

wetland species by limiting the extent of breeding habitat for 

some species and the exposure of mud flats that attracted other 

species. Some noticeably scarce or absent species also were rare 

or absent elsewhere along the south-central shore of Lake 

Ontario.

Benthic-macroinvertebrate studies were of minimal use 

for evaluating the effect of the FCS because no surveys were 

conducted after FCS installation. The precontrol results allowed 

assessment of the ecological quality of the wetland on the basis 

of biotic indices, and generally indicated moderately to severely 

impaired conditions. Differences between the 

macroinvertebrate communities in the southern part of the 

wetland and those in the northern part were attributed to habitat 

differences, such as substrate composition, water depth, and 

density of submerged aquatic vegetation.

Sedimentation rates in the areas upstream and downstream 

from the FCS increased after the flow modifications, more in 

the area upstream from the FCS than in the downstream area. 

The concurrent downstream increase and the dynamic patterns 

of deposition and scour indicated that although the FCS and the 

other flow modifications undoubtedly were major factors in the 

postcontrol upstream increase in sedimentation rates, other 

factors, such as the magnitude, frequency, and the timing 

(season) of peak flows, might also have contributed.

Periodic analyses of sediment samples from three longterm depositional sites in the wetland documented the 

concentrations of major and trace elements, polycyclic aromatic 

hydrocarbons, and organochlorine and organophosphate 

compounds. The concentrations of most constituents showed no 

substantial fluctuation or consistent upward or downward trend 
2 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

during the years sampled, nor did they identify any change after 

FCS installation. Comparison of the measured concentrations 

with sediment-quality guidelines that are used to assess the 

ecological quality of substrate environments indicated that the 

wetland was moderately to severely impaired—an assessment 

consistent with the benthic-macroinvertebrate biotic indices. 

During the precontrol period (1990-96), the wetland was a 

sink for particulate constituents (removal efficiencies for total 

phosphorus and total suspended solids were 28 and 47 percent, 

respectively), but had little effect on conservative constituents 

(chloride and sulfate). The wetland was a source of 

orthophosphate and ammonia (removal efficiencies were -38 

and -84 percent, respectively). 

During the postcontrol period (1997-2001), the wetland 

continued to be a sink for particulate constituents (removal 

efficiencies for total phosphorus and total suspended solids 

were 45 and 52 percent, respectively); the exportation of 

orthophosphate by the wetland decreased (by 7 percent), 

whereas that of ammonia increased (by about 70 percent). The 

outflow loads of orthophosphate and ammonia represented 

about 15 and 2.3 percent of total phosphorus and total nitrogen 

loads, respectively. Changes in the loads of conservative 

constituents were negligible, and the overall removal 

efficiencies for other constituents during the precontrol period 

differed from those of the postcontrol period by no more than 

5.4 percent. 

Statistical analyses of monthly inflow and outflow loads 

indicated significant differences between inflow and outflow 

loads of most constituents during the pre- and postcontrol 

periods. Load data were adjusted to remove the effects of 

dissimilar hydrologic conditions that prevailed during the preand postcontrol periods, and to isolate the water-qualityimprovement effect that could be attributed solely to the FCS. 

Results indicated that the FCS contributed significantly to the 

decrease in total phosphorus loads, and slightly to a decrease in 

ammonia-plus-organic nitrogen loads, but had little or no 

significant effect on loads of other constituents. 

Introduction

Storm runoff from urban areas causes local flooding and 

carries large amounts of contaminants and sediment to 

downstream receiving water bodies (O’Brien and Gere, 1983; 

U.S. Environmental Protection Agency, 1983). Runoffmanagement practices such as stormflow-detention basins and 

erosion-control measures are routinely implemented to mitigate 

these adverse effects (Natural Resources Conservation Service, 

1986). One practice that has been considered a cost-effective 

means of decreasing chemical loads in urban runoff is 

stormwater detention within wetlands, which have the potential 

to attenuate peak discharges, detain stormflows, and improve 

stormflow quality by facilitating the retention or removal of 

sediment and nutrients (Kadlec and Knight, 1996; Coon and 

others, 2000; U.S. Environmental Protection Agency, 1993; 

Strecker and others, 1992). This latter function can be achieved 

through several processes, including sedimentation (settling), 

adsorption, chemical precipitation, filtration, biochemical 

interactions and microbially mediated transformations, nutrient 

assimilation, volatilization and aerosol formation, infiltration 

(Strecker and others, 1992), ion exchange, biodegradation (U.S. 

Environmental Protection Agency, 1993), and long-term 

storage of chemicals through accretion and burial (Mitsch and 

Gosselink, 1986). These processes can cause a wetland to act as 

either a temporary or permanent sink for certain constituents. 

Conversely, other wetland processes, such as leaching, 

decomposition, dissolution, diffusion from sediments, and, 

again, microbially mediated transformations, can cause a 

wetland to act as a source of certain constituents and to export 

them to receiving waters downstream (Kadlec and Knight, 

1996; Strecker and others, 1992; Mitsch and Gosselink, 1986; 

U.S. Environmental Protection Agency, 1993). 

The water-quality-improvement function of natural and 

constructed wetlands has been studied since the 1960’s; 

wetlands are receiving wide use as a cost-effective and 

environmentally beneficial means of mitigating the deleterious 

effects of contaminated runoff and wastewater on surface-water 

and ground-water systems (Kadlec and Knight, 1996; Strecker 

and others, 1992; U.S. Environmental Protection Agency, 1993; 

Livingston, 1989; Gadbois, 1989). Of the processes listed 

above, microbial activity—the assimilation and transformation 

of nutrients by microbes—has been identified by many 

researchers as the principal mechanism by which water quality 

can be improved through detention in a wetland (Mitsch and 

Gosselink, 1986; Johnston, 1991; Hickok and others, 1977). 

The effectiveness of this mechanism might depend on 

macrophyte density—that is, the amount of surface area to 

which bacteria, fungi, and epiphytic algae can adhere (Kadlec 

and Knight, 1996). Physical processes, such as adsorption and 

sedimentation, also can play major roles in water-quality 

improvement (Johnston, 1991). 

The processes of water-quality improvement in a wetland 

are interrelated, and the relative importance of any one 

mechanism can vary from wetland to wetland. The broad range 

in the effectiveness of wetlands in treating stormwater results 

from (1) their diversity in terms of vegetation, depth and 

duration of inundation, climatic factors, soils, and hydraulic 

characteristics; and (2) differences in water sources (surface 

water, ground water, or atmospheric deposition) and water 

chemistry. The net effect of a wetland on water quality depends 

on the balance between chemical uptake and retention on the 

one hand, and chemical release and export on the other. 

Chemical retention or export can be affected by many factors, 

including:

-size and water-storage capacity of the wetland; 

-rate and velocity of flow; 

-detention time of stormwater; 

-pattern and degree of flow dispersal or mixing (circulation); 

-presence or absence of pools for sediment accumulation; 

-frequency, duration, and depth of inundation; 

-types and density of vegetation; 
Introduction 3

Allen Cr eek

Thomas

Creek

Erosion-control

project

EXPLANATION

Lake

Ontario

City of Rochester

Irondequoit

Bay

Irondequoit Creek

MONROE COUNTY

ONTARIO COUNTY

WAYNE COUNTY

LIVINGSTON COUNTY

Erie (Ba rge) Ca nal

MONROE COUNTY

Gen ese

e

R

vi er

1 2 3 4 5 MILES

0 1 2 3 4 5 KILOMETERS

0

Base from U.S. Geological Survey 

State base map 1:500,000

43o

43o

07'

30"

43o

15'

77o37'30" 77o30' 77o22'30"

Ellison Park

wetland

(Area shown

in Figure 2.)

LOCATION MAP

Irondequoit

Creek Basin NEW

YORK

Greater

Rochester

International

Airport

-water pH, temperature, and dissolved-oxygen concentration; 

-turbidity or amount of light penetration; 

-season; 

-types of maintenance practices used (removal of sediment or 

vegetation); and 

-rates of chemical exchange between water and sediment 

(Livingston, 1989; Strecker and others, 1992; Kadlec and 

Knight, 1996; Adamus and Stockwell, 1983). 

A wetland’s structure and ability to improve water quality 

depend largely on its hydrologic characteristics, which, in turn 

(1) determine or modify the composition, species richness, and 

distribution of wetland flora; (2) affect the level of productivity 

through volume and circulation of flow and through duration 

and frequency of inundation; and (3) affect (a) rates of 

sedimentation, resuspension, and erosion; (b) rates of 

organic-matter accumulation and export; and (c) the 

input, output, transformation, and cycling of nutrients 

and other constituents (Mitsch and Gosselink, 1986). 

The use of natural wetlands for treatment of 

stormwater is less common than use of constructed 

wetlands because State and Federal regulations 

restrict the intentional use of natural wetlands for 

stormwater treatment (United States, 1977; New York 

State, 1980). The unintentional use of a wetland for 

this purpose can be unavoidable, however, depending 

on its location within its basin and its proximity to 

urbanized areas. When this occurs, the hydrologic, 

biological, and chemical characteristics of the wetland 

will be affected by the increased inflows and chemical 

loads (U.S. Environmental Protection Agency, 1993), 

and a new equilibrium condition will become 

established. One such natural wetland that reflects 

several decades of increasing upstream development 

is the Ellison Park wetland, at the mouth of 

Irondequoit Creek on Irondequoit Bay, near 

Rochester, N.Y. (fig. 1).

Irondequoit Creek Basin

The Irondequoit Creek basin, which 

encompasses parts of Monroe and Ontario Counties, 

has been the subject of many water-quality studies in 

response to public concern over the sedimentation and 

eutrophication of Irondequoit Bay (fig. 1). 

Irondequoit Creek is the bay’s main tributary and its 

principal source of (1) nutrients that, in the past, 

supported algal blooms in the bay (Pixley, 1982); (2) 

dissolved chloride, which can alter the bay’s chemical 

and thermal stratification and interfere with seasonal 

mixing of the bay’s waters (Diment and others, 1974; 

Bubeck and Burton, 1989); and (3) sediment to which 

heavy metals and organic compounds can adhere 

(Schroeder, 1985). 

Until the spring of 1978, effluent from 14 wastewatertreatment facilities within the basin and combined-sewer 

overflows (CSOs) from the city of Rochester was discharged to 

the creek. Since 1978, the wastewater-treatment effluent has 

been diverted out of the basin by a wastewater-interceptor 

system. In addition, a bedrock tunnel system that permits 

subsurface storage of stormwater has decreased the frequency 

of CSO discharges to the creek—from discharging during 

nearly every storm to discharging during only about one storm 

in 10 years (Michael Schifano, Monroe County Pure Waters, 

oral commun., 1999). Wastewater and combined-sewer water 

are now treated at a tertiary-treatment plant at the north end of 

Rochester and discharged into Lake Ontario. 

Monroe County also has implemented several practices to 

decrease nonpoint-source chemical loads in Irondequoit Creek, 

including (1) onsite controls of erosion and storm runoff at new 

Figure 1. Location of study area and principal geographic features 

of Irondequoit Creek basin, Monroe County, N.Y.
4 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

43o9'

77o32' 77o31'

43o10'

Base from scanned U.S. Geological Survey

Rochester East, NY 1:24,000, 1978

EXPLANATION

Wetland boundary

Streamflow-measurement and water-

 quality monitoring site

0423205010

1000 2000 FEET

0 250 500 METERS

0

The Narrows

and flow-control

structure

Town boundary

p

Irondequoit

Brighton

Penfield

E L

L

I

S

O

N

P

K

Millrace

I r o n d e q u o i t

NYS Route 590

B a y

Blossom Road

Browncroft

Empire

Boulevard

Boulevard

Irondequoit Creek

0423205010

0423205025

p

p

developments, with expanded use of multipurpose 

stormflow-detention basins; (2) conversion of dry detention 

basins in older developments to wet basins to promote the 

removal of nutrients in stormwater through increased 

biological uptake; and (3) rip-rap reinforcement of erodable 

banks along Irondequoit Creek and one of its tributaries, 

Allen Creek (fig. 1). In addition, the County has directly 

treated Irondequoit Bay through (1) application of alum to 

bottom sediments to inhibit the release of phosphorus to the 

water; and (2) injection of oxygen into the Bay’s 

hypolimnion to maintain aerobic conditions. These 

measures, along with the diversion of wastewater out of the 

basin, have decreased nutrient loads to Irondequoit Bay 

(Coon and others, 2000) and have improved its eutrophic 

state (Monroe County Health Department, written commun., 

2001), but further improvement has been slowed by the 

continued inflow of contaminants from nonpoint sources in 

the basin—primarily nutrients, metals, and organic 

compounds washed from impervious surfaces by rain and 

snowmelt (R.S. Burton, Monroe County Environmental 

Health Laboratory, written commun., 1996).

Previous Studies

Streamflow and water quality of Irondequoit Creek and 

its major tributaries, Thomas and Allen Creeks (fig. 1), have 

been extensively monitored by Monroe County and the U.S. 

Geological Survey (USGS). Irondequoit Creek basin was one 

of 28 basins that was included in the Nationwide Urban 

Runoff Program (NURP) during 1979-81. In addition, local 

projects have studied specific water-resource issues with the 

intent of extrapolating small-scale solutions to address 

basinwide concerns.

Irondequoit Creek Basin Studies

The hydrologic characteristics of the Irondequoit Creek 

basin were documented during 1979-81 as part of the NURP 

study (O’Brien and Gere, 1983; Zarriello and others, 1985; 

Kappel and others, 1986). One goal of the Irondequoit Creek 

NURP study was to assess the effect of storm runoff and its 

associated nutrients and contaminants on the quality of water in 

Irondequoit Bay. Kappel and others (1986) estimated annual 

loads of selected constituents that entered the bay (during 1980-

81): 214 tons of ammonia-plus-organic nitrogen, 20.2 tons of 

total phosphorus, and 19,100 tons of total suspended solids. 

O’Brien and Gere (1983) estimated that 15.7 tons of 

phosphorus were entering Irondequoit Bay annually (1980-81) 

from external sources and that, given a zero net phosphorus 

release from the bottom sediments, the annual external loading 

would need to be decreased to about 5.6 tons to maintain the bay 

in a trophic state appropriate for recreational usage. They also 

indicated that the 3-month period that included the major 

seasonal snowmelt and spring runoff accounted for 50 to 75 

percent of the annual phosphorus loads entering the bay. 

A unique component of the Irondequoit Creek NURP 

study’s monitoring program (Kappel and others, 1986) was the 

inclusion of an in-stream natural wetland (Ellison Park) (fig. 2). 

Several NURP projects considered wetlands to be promising 

mechanisms for treatment of urban runoff, but these projects 

did not include development of performance or design criteria 

(U.S. Environmental Protection Agency, 1983). 

Historical (12,000 to 2,000 years before present), preEuropean-settlement, and recent sedimentation rates in and 

upstream from Irondequoit Bay were estimated by Young 

(1992; 1996) and Schroeder (1985). Young (1992) used 

radiocarbon-dating techniques to estimate postglacial 

sedimentation rates from two sediment cores taken at the outlet 

Figure 2. Locations of streamflow- and water-quality 

monitoring sites in Ellison Park wetland, Monroe County, N.Y. 

(Location is shown in fig. 1)
Introduction 5

of Irondequoit Bay (fig. 1) and one core taken from along 

Irondequoit Creek near Browncroft Boulevard (fig. 2); his 

results indicated the average historical sedimentation rate at the 

mouth of Irondequoit Creek to be 1 ft per 99 years, or 0.01 ft/yr 

(3 mm/yr). This rate is consistent with the rate of rise in the 

water-surface elevation, and consequent deepening, of the 

depositional pool along the southern shore of Lake Ontario in 

response to a differential rate of postglacial isostatic rebound of 

the Lake Ontario basin. The rate of glacial rebound at the 

northeastern end (outlet) of Lake Ontario is 1.25 ft per century 

and, at the southwestern end, is 0.25 ft per century (Clark and 

Persoage, 1970; Larsen, 1985). Because the lake outlet has been 

rising faster than the rest of the lake, the water level along the 

southern shore has been rising at a “differential” rate of about 

1 ft per century. The rate of sediment accumulation in the 

vicinity of Irondequoit Bay, as computed by Young (1992), 

approximates this rate of water-level rise and the resulting 

deepening of the depositional pool in the area that encompasses 

the Ellison Park wetland and Irondequoit Bay. 

Young (1996) estimated sedimentation rates for the past 

100 to 1,200 years from the radiocarbon dates of 10 organicmatter samples from 6 excavation sites in Ellison Park between 

Blossom Road and Browncroft Boulevard (fig. 2). Results 

indicated an average sedimentation rate of 14.5 in. per century 

(0.012 ft/yr or 3.7 mm/yr) prior to European settlement of the 

Irondequoit Creek basin. This rate agrees closely with the 

historical sedimentation rate estimated by Young (1992) and 

the rate of isostatic rebound. Three samples from the postsettlement period indicate that sedimentation rates tripled—to 

an average rate of 43 in. per century or 0.036 ft/yr (11 mm/

yr)—between the early 1800’s and 1950, presumably as a result 

of deforestation and land-use changes that accompanied 

settlement of the area. From a third to half of this rate is 

assumed to be a result of the southwestward tilting of the Lake 

Ontario basin in response to isostatic rebound; the remainder is 

attributed to increased erosion resulting from development 

within the Irondequoit Creek basin (Young, 1996). 

Recent (1953-80) sedimentation rates estimated by 

Schroeder (1985) were based on lead-210 and cesium-137 

radioisotopic analyses of four bottom-sediment cores taken 

from Irondequoit Bay in 1980; results indicated yearly 

sediment-accumulation rates of 0.10 to 0.36 g/cm2 on a dryweight basis. The calculated thickness of sediment deposited 

between 1964 (the year of peak cesium-137 concentration in the 

atmosphere) and 1980 (when the samples were collected) (data 

from R.A. Schroeder, U.S. Geological Survey, written 

commun., 1995), indicates a sedimentation rate of 0.0092 to 

0.012 ft/yr (2.8 to 3.7 mm/yr), which is consistent with the 

historical rate estimated by Young (1992) but much less than 

the post-settlement rate estimated by Young (1996).

Ellison Park Wetland Studies

The Irondequoit Creek NURP study (O’Brien and Gere, 

1983) identified the Ellison Park wetland as the most costeffective means of decreasing nutrient loads entering 

Irondequoit Bay. Kappel and others (1986) estimated that the 

wetland decreased the phosphorus loads carried by Irondequoit 

Creek to Irondequoit Bay by 10 percent and theorized that an 

additional 15-percent decrease could be achieved, primarily 

through the deposition and mineralization of particulate 

phosphorus associated with the silt-clay sediment fraction, if 

flows through the wetland were regulated to increase water 

dispersal and detention time. Similarly, O’Brien and Gere 

(1983) estimated that the wetland could remove 24 percent of 

the creek’s phosphorus load if a flow-control structure were 

installed at the Narrows, a natural constriction midway through 

the wetland between Browncroft and Empire Boulevards 

(fig. 2).

An 11-year study of the wetland was begun in 1990 by the 

USGS in cooperation with the Monroe County Department of 

Health to evaluate the potential of the wetland to improve the 

water quality of Irondequoit Bay. During 1997, in response to 

the recommendation of O Brien and Gere (1983), Monroe 

County installed a flow-control structure (FCS) at the Narrows 

(fig. 2) to cause within-bank flows, especially the early spring 

first-flush and snowmelt flows, to overtop the banks of the main 

channel and disperse into the cattail-covered backwater areas of 

the wetland where natural detention, sedimentation, and 

biological uptake of nutrients could occur. This dispersal of 

stormflows over a larger area and for longer periods than 

occurred naturally would increase contact of the nutrient-laden 

water with plant life and microbes (bacteria, fungi, algae) and, 

theoretically, would decrease the loads of nutrients entering 

Irondequoit Bay and thereby improve its water quality. 

Throughout the study, streamflow and water quality at the 

inflow and outflow of the wetland were monitored, and the flora 

and fauna in the wetland were surveyed periodically. (See 

“Reports of Biological Studies” section, at the end.) The study 

consisted of two phases: the first (1990-96) entailed an 

assessment of wetland functions and documentation of the 

wetland’s ecological status; the second (1997-2001, the subject 

of this report) entailed an assessment of changes in wetland 

functions or ecological status that might have resulted from the 

operation of the FCS.

The first phase of the study is summarized by Coon (1997) 

and Coon and others (2000). Coon (1997) described the 

hydrologic, sedimentological, and biological characteristics of 

the wetland and concluded the following: 

1. Water levels in the wetland were controlled by the surface 

elevation of Lake Ontario. 

2. Bankfull and lower flows of Irondequoit Creek did not 

disperse through the wetland, but were mostly confined 

within the banks of the main channel and usually passed 

through the wetland in less than 3.5 hours. Dispersal of 

stormflows occurred only when flows exceeded the capacity 

of the channel (overbank flows), which on average occurred 

twice a year. 

3. Dispersed water moved into the cattail-covered backwater 

areas of the wetland, where it was detained for 3 to 15 hours 

or more (the maximum detention time was not measured). 
6 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

4. Dispersal of flows also occurred during backwater conditions 

that resulted from high water levels on Lake Ontario, usually 

from April through July. 

5. The channel sediments in the wetland ranged from sand, 

which was found in the main channel of Irondequoit Creek, 

to silt and clay with about 4-percent organic matter (by 

weight), which was found in the backwater areas. 

6. Sediment accumulation in the wetland was sporadic, but an 

annual sedimentation rate of 0.006 to 0.016 ft/yr (1.8 to 

4.9 mm/yr) was estimated. 

7. The wetland, which was dominated by cattails (Typha 

glauca), was highly productive; estimated average total 

biomass was 5,230 g/m2. 

8. The wetland supported diverse fish and bird communities; at 

least 37 fish species and 28 possible wetland breeding-bird 

species were identified.

Coon and others (2000) assessed the wetland’s effects on 

water quality of Irondequoit Creek and calculated inflow and 

outflow loads and removal efficiencies for selected constituents 

for 1990-96. The results indicated that the Ellison Park wetland 

was (1) a sink for total phosphorus and total suspended 

solids—28 and 47 percent removal efficiencies, respectively; 

and (2) a source of orthophosphate and ammonia— -38 and -84 

percent removal efficiencies, respectively. The wetland 

facilitated a slight decrease in total nitrogen loads and 

significant decreases in concentrations of zinc, iron, and lead, 

but had a negligible effect on conservative constituents such as 

chloride and sulfate.

Purpose and Scope

This report presents the results of the second phase of the 

study (1997-2001), which entailed an assessment of the wetland 

water levels, biota, sedimentation rates, and water and sediment 

quality during the first 4.5 years after installation of the FCS, and 

comparison of these data with those collected during the 

precontrol period (1990-96). The increased flooding potential 

was evaluated from discharge measurements and water-surface 

profiles. The plants, fish, benthic-macroinvertebrate, and bird 

communities were evaluated through biological surveys. 

Sedimentation rates were measured, and sediment samples 

were analyzed for major and trace elements, polycyclic 

aromatic hydrocarbons, and organochlorine and 

organophosphate compounds. Concentrations of selected 

constituents were measured in inflow and outflow water 

samples, and loads of these constituents, as well as the 

wetland’s removal efficiency for each, were calculated. This 

report (1) summarizes the design of the FCS and the ecologicalmonitoring program; (2) describes the methods used for each 

component of the study—water-level and discharge 

measurements, ecological surveys, sediment studies, waterquality analyses, estimation of monthly chemical inflow and 

outflow loads, and calculation of the wetland’s removal 

efficiency for selected constituents; (3) presents results of each 

of these study components; and (4) assesses the changes in 

wetland functions and characteristics that might be attributable 

to the FCS. Monthly and annual inflow and outflow loads and 

the estimated removal-efficiency values for phosphorus 

compounds, nitrogen compounds, suspended sediment, 

chloride, and sulfate are presented in an appendix.

Acknowledgments

This project required the combined efforts of many people. 

The Monroe County Environmental Health Laboratory 

(MCEHL) assisted in the maintenance and operation of the 

streamflow- and water-quality-monitoring sites, analyzed water 

samples, participated in several components of the ecologicalmonitoring programs, and operated and maintained the FCS at 

the Narrows. Richard Burton, Anna Madden, Charles Knauf, 

and Gary Brown (MCEHL laboratory and project 

administrators) provided guidance and suggestions throughout 

the project. The FCS was designed by Mark Ballerstein of the 

Monroe County Department of Environmental Services, and 

was installed by C.P. Ward, Inc., under the oversight of William 

Keihl, Harold Ford, and Randall Williamson. Access to the 

Narrows for installation and operation of the FCS was granted 

by the Town of Brighton. 

The plant surveys were conducted by Dr. John Bernard, 

Department of Biology, Ithaca College; Dr. Franz Seischab, 

Department of Biology, Rochester Institute of Technology; and 

Dr. John Hunter (Department of Biological Sciences) and Dr. 

Mark Noll (Department of Earth Sciences), State University of 

New York (SUNY) College at Brockport. Field and laboratory 

assistance for the plant surveys was provided by Nicole Perry, 

Jamie Brennan, Duane Hanselman, Tracy Cohen, Jessie 

DiLorenzo, Amy Gardner, Heather Halbritter, Jamie Kirby, Kit 

Sheehan, Frank Evans, Nicholas Byerle, Sarah Lewis, and 

Laura Sontag-Brown. 

The fish surveys were conducted by Dr. Neil Ringler, 

Daniel Miller, Darran Crabtree, and Patricia Thompson 

(Department of Environmental and Forest Biology), SUNY 

College of Environmental Science and Forestry at Syracuse.

Benthic macroinvertebrate work was conducted by Dr. 

James Haynes (Department of Biological Sciences) and Dr. 

James McNamara (Department of Mathematics), SUNY 

College at Brockport. 

Bird surveys were conducted by volunteers Robert 

McKinney and Robert Spahn, members of the Genesee 

Ornithological Society, Rochester, N.Y.

Data on (1) the wetland’s ground-water hydrology and 

chemistry; (2) the movement, deposition, and generation of 

phosphorus in the wetland; and (3) evapotranspiration by the 

cattails and the effect of this process on surface water were 

collected by Cornell University graduate students Michael 

Traynor and Oliver Pierson, under the supervision of Dr. 

Rebecca Schneider, Department of Natural Resources, Cornell 

University, Ithaca, N.Y.
Study Area 7

STUDY AREA

The Ellison Park wetland occupies 423 acres at the mouth 

of Irondequoit Creek, which drains 151 mi2 and flows 

northward into Irondequoit Bay (fig. 1). On the basis of satellite 

imagery (New York State, 2000), only 25 percent of the 

Irondequoit Creek basin is classified as “developed,” but taxparcel data classify about 54 percent of the basin as urban or 

suburban (Christopher Sciacca, Monroe County Planning 

Board, written commun., 1998); most of this developed area is 

north of the New York State Erie (Barge) Canal (fig. 1). The 

upstream (southern) part of the basin is dominated by forest and 

farmland but is becoming increasingly developed. 

Average annual precipitation in the Rochester area is 32.4 

in. (168 years of record), including the amount derived from an 

average annual snowfall of about 90 in. (55 years of record), as 

recorded by the National Weather Service station at the Greater 

Rochester International Airport (National Climatic Data Center, 

1996). Monthly precipitation and the resultant monthly 

discharges during 1990-2001 (Northeast Regional Climate 

Center, Ithaca, N.Y., written commun.) are plotted with 30-year 

(1961-90) mean monthly precipitation values in figure 3A and 

with monthly maximum, mean, and minimum discharges for 

the period of record (1981-2001) for Irondequoit Creek above 

Blossom Road in figure 3B. 

The wetland lies along the township boundaries of 

Irondequoit, Brighton, and Penfield in Monroe County and is 

mostly within Ellison Park, a county park (fig. 2). The wetland, 

which forms a transition zone between the riparian environment 

of the creek and the lacustrine environment of Irondequoit Bay 

at and below an elevation of 250 ft (DeGaspari and Bannister, 

1983), represents less than 0.5 percent of the creek’s total 

drainage area. Bounded by steep valley sides on the east and 

west, the wetland covers the entire valley floor from its inlet 

above Blossom Road to Irondequoit Bay at Empire Boulevard 

(fig. 2). The area between Browncroft and Empire Boulevards 

is dominated by cattails and is divided by a constriction at the 

Narrows into a southern and a northern segment (263 acres and 

160 acres, respectively, DeGaspari and Bannister, 1983). The 

width of the wetland ranges from 200 ft at the Narrows to 

2,500 ft in the southern and northern parts.

The wetland is classified as a “palustrine persistent 

emergent” wetland by the U.S. Fish and Wildlife Service 

(Cowardin and others, 1979); that is, a nontidal marsh 

characterized by erect, rooted, herbaceous hydrophytes that 

may be temporarily or permanently flooded at the base and 

normally remain standing until at least the beginning of the next 

growing season. Cattails, primarily Typha glauca, cover about 

265 acres, or 63 percent of the wetland area between 

Browncroft and Empire Boulevards. (Acreages were digitized 

from a USGS 7.5-minute topographic map, Rochester East, 

N.Y. quadrangle.) Plant species richness is greatest on the 

banks of the main channel of Irondequoit Creek and along the 

wetland margins, where soil wetness varies greatly with slight 

differences in elevation. Open water, including channels, 

covers about 99 acres, or 23 percent of the wetland. A forested 

wetland area between Browncroft Boulevard and Blossom 

Road covers about 25 acres, or 6 percent of the wetland, and 

maintained grass areas of the county park below the wetlanddemarcation elevation of 250 ft (National Geodetic Vertical 

Datum, NGVD) cover the rest of the wetland (about 34 acres, or 

8 percent of the wetland).

The wetland contains hydric soil, defined by the U.S. Soil 

Conservation Service (1985) as soil that, in its undrained 

condition, is saturated, flooded, or ponded long enough during 

the growing season to develop anaerobic conditions that favor 

the growth and regeneration of hydrophytic vegetation. The 

wetland soil is a mineral soil, formed through the deposition of 

alluvium (flood-plain deposits) from the Irondequoit Creek 

basin, that contains partly decomposed organic material. 

Organic matter in soil samples from the cattail-covered areas of 

the wetland represented from 35 to 40 percent of the sample 

material (by weight) on average, but represented as much as 

79 percent of some samples (Hunter and others, 2000; Noll and 

others, 2002).

Wetland conditions in Ellison Park are dependent on the 

surface elevation of Lake Ontario, which is maintained within a 

narrow range by control structures on the St. Lawrence Seaway. 

Stage data from a streamflow-monitoring station at Empire 

Boulevard (fig. 2) indicated that water-surface elevations at the 

mouth of the wetland fluctuated less than 4.5 ft during 1990-

2001 (fig. 4). Water levels in Lake Ontario usually rise during 

the spring-runoff period and are maintained through the 

summer for navigational and recreational purposes and for 

hydroelectric-power generation on the St. Lawrence River near 

Massena, N.Y. Water levels are lowered in December and 

January to increase the storage capacity for spring runoff (Paul 

Yu, U.S. Army Corps of Engineers, oral commun., 1997). A 

water-surface elevation of 246.9 ft at Empire Boulevard results 

in the inundation of most of the cattail-covered area throughout 

the wetland. A drop of less than 0.5 ft causes water levels in the 

southern wetland area to fall below land surface, leaving only 

the northernmost part of the northern wetland area inundated. 

Water levels throughout the wetland range from at least 2 ft 

above land surface to 1.5 ft below. Flat, vegetated areas of the 

wetland can be briefly inundated by runoff from large storms 

and are seasonally inundated for periods of 4 to 7 months. The 

duration of seasonal inundation (hydroperiod) varies from year 

to year and increases from south to north; in some years, the 

water level rises only enough to inundate the southern wetland 

area for several days, whereas in other years, high water levels 

can persist from January through August (fig. 4; Coon and 

others, 2000). 

Two channels convey water through the wetland—the 

main channel of Irondequoit Creek, which carries most of the 

flow, and a smaller channel, locally referred to as the Millrace 

(fig. 2), which formerly was a saw-mill raceway on the east 

bank of the creek just downstream from Blossom Road (F.W. 

Pugsley, Town of Pittsford Historian, written commun., 1942); 

currently it is a diversion from the main channel to the eastern 

part of the southern wetland area north of Browncroft 

Boulevard. The main channel in the Blossom Road vicinity is 

about 70 ft wide and 10 to 12 ft deep at bankfull stage, whereas 
8 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

the Millrace is 20 to 25 ft wide and 5 to 6 ft deep. The mainchannel dimensions decrease between Browncroft Boulevard 

and the Narrows to a bankfull width of 50 to 60 ft and depth of 

5 to 8 ft. The extension of the Millrace channel into the southern 

wetland area widens to about 50 ft but is only about 3 ft deep. 

The depth of Irondequoit Creek downstream from the Narrows 

gradually decreases to 3 or 4 ft. 

STUDY DESIGN 

The FCS was installed in 1997 at the Narrows, where the 

valley-bottom width is about 200 ft and the channel width is 

about 60 ft. The Narrows divides the main cattail-covered area 

of the wetland into a 144-acre southern (upstream) area between 

the Narrows and Browncroft Boulevard and a 160-acre northern 

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

8

2

4

6

PRECIPITATION, IN INCHES

Monthly precipitation

30-year normal mean

monthly precipitation (1961-90) 

A. PRECIPITATION

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

500

100

200

300

400

DISCHARGE, IN CUBIC FEET PER SECOND

Mean

Maximum

Minimum

Blossom Road (wetland inflow)

Empire Boulevard (wetland outflow)

B. DISCHARGE

EXPLANATION

Monthly mean discharge for this study (1991-2001) Monthly discharge at Blossom Road for period of record (1981-2001)

Figure 3 . A. Monthly precipitation (1990-2001) and normal mean monthly precipitation (1961-90) recorded at Greater Rochester 

International Airport, Rochester, N.Y. B. Monthly mean discharges of Irondequoit Creek above Blossom Road and at Empire 

Boulevard, Monroe County, N.Y. (1990-2001), with monthly maximum, mean, and minimum discharges for period of record at Blossom 

Road (1981-2001) (Locations are shown in figs. 1 and 2.)
Study Design 9

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

245

250

246

247

248

249

WATER-SURFACE ELEVATION,

IN FEET ABOVE NGVD OF 1929

Elevation above which most of

the cattail-covered area of the 

southern (upstream) part of the 

wetland is inundated.

Elevation above which most of 

the cattail-covered area of the 

northern (downstream) part of 

the wetland is inundated.

(downstream) area between the Narrows and the wetland outlet 

(Empire Boulevard) (fig. 2). As explained in the previous 

section, water depths in the northern area are greater, and 

periods of inundation persist longer than in the southern area. 

Comparison of periodic measurements of water depth in the 

cattail-covered areas with water-surface elevations at the 

wetland outlet indicates that much of the cattail mat in the 

northern area floats when water levels are high, whereas much 

of the cattail mat in the southern area is anchored to the 

underlying sediment. 

The entire wetland has been subjected to large chemical 

loads from nonpoint-source urban runoff for decades and, until 

1978, received effluent from wastewater-treatment plants and 

combined-sewer overflows. The northern area also received 

chemical loads from two other sources—a landfill (currently 

closed to household wastes but used for composting operations) 

and a combined-sewer overflow (whose overflow frequency 

was decreased from almost every storm in the 1970’s to less 

than once in 10 years since the mid-1980’s). Both sources are 

on the western side of the wetland and close to the Narrows.

These differences were considered minor when the two 

wetland areas were assessed for suitability for data 

comparisons. Other nearby wetlands along Lake Ontario were 

considered as control sites for this study but were rejected 

because of substantial differences, despite their similar 

locations in the landscape. 

Flow-Control Structure

The FCS, which was designed and operated by Monroe 

County, consisted of two parts—a semipermanent base control 

made of stop logs and sheet metal, and a dynamic control made 

of hinged gates and removable stop logs (fig. 5). The base 

control was designed to (1) approximate the channel crosssection shape at a point about 100 ft downstream from the 

Narrows where an erosion-resistant till creates a shallow, gravel 

riffle in the creek during low-water periods, and (2) have 

minimal effect on flood levels upstream from the FCS. This 

latter criterion was based on the results of hydraulic analyses of 

high flows and water-surface elevations with a onedimensional, gradually varied, steady-flow model (Shearman, 

1990). The dynamic control was designed and operated to have 

minimal effect on flooding in the day-use area of the park 

between Blossom Road and Browncroft Boulevard and to have 

no effect on flood levels at and above Blossom Road, about 2.4 

mi upstream (fig. 2). 

Construction of the base control entailed driving nine H 

piles, spaced about 6 ft apart, into the channel bottom to the 

point of resistance. Foot-square stop logs were slid into the 

channels of adjacent piles to an elevation of 244 ft in the four 

center bays and 246 ft in the side bays, which created a 2-ft deep 

notch in the center of the structure. The bottom elevation of the 

notch was below the thalweg elevation of the riffle downstream 

to ensure that the notch would be submerged at all times and 

would permit unobstructed passage of fish and small boats 

during low-water periods. 

The dynamic control, which permitted water-level 

manipulation to an elevation of 247 ft, was built upon the base 

control and consisted of removable stop logs above the side 

sections of the FCS and hinged gates in the “notch” bays 

(fig. 5). Operational needs and public concerns warranted 

modifications to this design, which included (1) driving sheetmetal plates into the channel bottom along the upstream face of 

the stop logs to prevent undercutting of the structure; (2) cutting 

a center H-pile to increase the maximum opening between piles 

Figure 4. Daily mean water-surface elevation of Ellison Park wetland, Monroe County, N.Y., as measured at wetland outlet (Empire 

Boulevard), 1990-2001. (Location is shown in fig. 2.)
10 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

80 90 100 110 120 130 140 150 160 170 180 190 200 210

DISTANCE FROM WEST END OF CROSS SECTION, IN FEET

238

258

240

242

244

246

248

250

252

254

256

ELEVATION, IN FEET ABOVE NGVD OF 1929

Precontrol channel cross section

Base-control cross section

Side bays at 246 feet

Center "notch" bays at 244 feet

Water surface

Stop logs and 

sheet-metal

plates

H-pile

Canoe portage (rock on

erosion-resistant fabric)

Removable stop log

Hinged gate

Large boulder riprap

Gabions

EXPLANATION

Figure 5. Schematic diagram and photograph of flow-control structure at the Narrows in Ellison Park wetland, Monroe County, 

N.Y., 2001. A. Schematic diagram. B. Downstream-facing view of flow-control structure with H-piles extending above water 

surface.
Study Design 11

to 11 ft to facilitate passage of canoes and small boats; (3) 

laying large boulder riprap or gabions on the streambanks, and 

erosion-resistant fabric and rock on the west-bank overflow 

area to control erosion; and (4) constructing a canoe portage 

around the FCS along the western bank. 

The purpose of the FCS was to increase the dispersal and 

short-term detention of stormflows with recurrence intervals 

between 1 and 2 years—the annual peak flow. Under natural 

conditions, stormflows of this magnitude are confined between 

the elevated banks of the creek channel, and little water flows 

into the cattail-covered areas of the wetland (Coon, 1997). 

Manipulation of the FCS could cause the water-surface 

elevation associated with these frequent high flows to rise 

above bank level and thereby increase the dispersal of flows 

into the cattail-covered areas. It also could cause larger 

flows—those that would ordinarily overtop the banks—to 

disperse more widely, and to be detained longer, than would 

occur without the FCS. Detention time of stormflows in the 

southern part of the wetland was expected to increase by as 

much as 0.7 days (17 h) for small peak flows (500 ft3

/s); from 

2.5 to 4.8 days (60 to 115 h) for high peak flows (1,500 ft3/s); 

and from 2.6 to 4.7 days (62 to 113 h) for extremely high peak 

flows (4,000 ft3/s) (Kappel and others, 1986). The actual period 

of inundation depends not only on the magnitude of flow, but 

on the antecedent water levels in Irondequoit Bay and Lake 

Ontario.

The FCS was operated in accordance with weather 

forecasts and water levels at the Blossom Road streamflowmonitoring station. The gates in the “notch” bays adjacent to the 

center bay and the removable stop logs were left in their 

maximum-detention positions most of the time. In anticipation 

of stormflows, the Monroe County Environmental Health 

Laboratory (MCEHL) would raise the gates in the center bay. If 

stormflows remained within the creek’s banks at Blossom 

Road, no adjustments were made to the FCS, but if weather 

forecasts predicted additional rain that might cause flooding at 

Blossom Road, the removable stop logs were removed, and the 

gates were lowered to obtain the base-control configuration.

The FCS was designed and operated to permit free passage 

of fish under all flow conditions and downstream water levels. 

During the period of seasonally high water levels (late spring 

and summer), the FCS was submerged, and water flowed over 

its top. During low-water periods (late fall and winter), water 

flowed through the 11-ft-wide notch in the center of the FCS at 

all times to a depth of at least 2 ft. 

The base control was installed near the end of February 

1997, and the gates in the center notch were installed in early 

September 1997. The FCS was fully operational by September 

15, 1997. An operation record was maintained, in which 

changes in stop-log placement and position of gates were noted. 

Additional Flow Modifications

Natural flow detention that occurs in the southeastern part 

of the wetland, and the potential water-quality benefits that 

might be realized if water from Irondequoit Creek were diverted 

to this cattail-covered backwater area, have been described by 

Coon (1997) and Coon and others (2000). To maximize these 

benefits, the flow patterns in the creek and Millrace were 

modified by Monroe County between February 1997 and the 

6

Day-use area

43o9'

77o32' 77o31'

43o10'

Base from scanned U.S. Geological Survey

Rochester East, NY 1:24,000, 1978

EXPLANATION

Wetland boundary

Streamflow-measurement and water-

 quality monitoring site 0423205010

1000 2000 FEET

0 250 500 METERS

0

The Narrows

and flow-control

structure

A B

C

Millrace

I r o n d e q u o i t

NYS Route 590

B a y

E

D

F

-A,B,C Streambank notch

D Channel widening

E Diversion channel

Culvert

Direction of flow through streambank notch

F

Letter indicates location and type

of hydrologic modification

Blossom Road

Browncroft

Empire

Boulevard

Boulevard

Irondequoit Creek

0423205010

0423205025

p

p

p

Figure 6. Locations of flow modifications to Irondequoit Creek and 

the Millrace, Ellison Park wetland, Monroe County, N.Y.,1997-98.
12 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

summer of 1998 to (1) increase dispersal of flows along the 

Millrace and in the wetland, (2) decrease flooding between 

Blossom Road and Browncroft Boulevard (the day-use area of 

the park), and (3) divert as much as 50 percent of bankfull 

stormflows from the creek to the Millrace, which drains directly 

into the southeastern part of the wetland (fig. 6). First, the bank 

of the creek was notched upstream from Browncroft Boulevard 

(point A in fig. 6) to increase diversion of within-bank flows to 

the wetland area bounded by Browncroft Boulevard to the 

north, the creek to the west, and the Millrace to the east. A 

second notch was made at the Millrace end of this wetland area 

(point B) to allow water that exceeded the storage capacity of 

this area to pass into the Millrace. A third notch was constructed 

in the east bank of Irondequoit Creek within the wetland proper, 

about halfway between Browncroft Boulevard and the 

Narrows, to permit movement of water from the creek directly 

into the backwater area of the wetland (point C). Additional 

modifications to the Millrace included (1) widening the channel 

immediately downstream from the point of diversion from 

Irondequoit Creek (point D), and (2) constructing a top-of-bank 

diversion channel from the Millrace to a wooded wetland area 

east of the Millrace and south of Browncroft Boulevard (point 

E). Finally, the top of bank was lowered and a third culvert was 

added to the existing two culverts at the point of diversion from 

Irondequoit Creek to the Millrace just downstream from 

Blossom Road to increase flow to the Millrace (point F).

Ecological Monitoring Program

The ecological-monitoring program was designed to 

identify any effects that the FCS might have on the wetland’s 

ecology. The program was based on the advice of a technical 

oversight committee that included representatives of the USGS, 

U.S. Army Corps of Engineers, U.S. Fish and Wildlife Service, 

New York State Department of Environmental Conservation, 

county and town agencies, and researchers from biology and 

geology departments of local universities. The program 

included periodic surveys of plants, fish, benthic 

macroinvertebrates, and birds. The plant surveys were designed 

to (1) identify plant species, (2) map plant communities, (3) 

measure cattail density and biomass, and (4) analyze above- and 

below-ground cattail tissues along permanent transects across 

the wetland. Fish surveys were designed to (1) identify fish 

species, (2) describe community composition, (3) record use of 

the wetland by resident and nonresident fish, (4) document 

species distribution within the wetland, and (5) note symptoms 

of contaminant-related stress. An unplanned consequence of 

using trapnets during the fish surveys was the capture of turtles, 

which were counted and identified by species. The benthicmacroinvertebrate and bird surveys were designed to (1) 

document species richness, and (2) confirm the presence or 

absence of species. Each of these studies included assessment 

of any identifiable or probable effect that the FCS had or might 

have on the particular biological community that was surveyed. 

The monitoring program also entailed (1) annual measurement 

of sediment accumulation at permanent measurement points; 

and (2) periodic collection and analysis of sediment samples for 

major and trace elements and organic compounds, including 

pesticides, polychlorinated biphenyls, and polycyclic aromatic 

hydrocarbons. Baseline surveys and measurements were made 

during 1991-96 to document the precontrol differences between 

the southern and northern wetland areas. These results were 

compared with those obtained from similar surveys and 

measurements conducted during the postcontrol period to 

identify changes that might be attributable to the FCS. The 

timing and frequency of the components of the ecologicalmonitoring program during the study period are given in 

table 1.

METHODS

The Ellison Park wetland project consisted of several 

studies to assess changes in the wetland after 1997 that might be 

attributable to the FCS. These were: (1) studies to assess the 

effects of the FCS on flooding extent and high-water levels; (2) 

an ecological-monitoring program that consisted of periodic 

surveys of the plant, fish, benthic-macroinvertebrate, and bird 

communities; (3) calculation of sedimentation rates, and 

chemical analyses of sediment samples; (4) measurement of 

discharge and water quality at the inflow and outflow of the 

wetland; (5) a quality-assurance and quality-control program to 

assess the accuracy and reliability of the water-quality data, 

which were provided by the MCEHL; and (6) calculation of the 

loads of selected chemical constituents in wetland inflow and 

outflow to obtain removal-efficiency values. The methods used 

in each of these study components are described below. 

Water-Surface-Elevation and Discharge 

Measurements

Two studies were conducted during the postcontrol period 

to assess the effect of the FCS on flooding and high-water 

levels, especially in the day-use area of Ellison Park in the 

vicinity of Blossom Road (fig. 6). The first study entailed use of 

WSPRO, a water-surface-profile model of one-dimensional, 

gradually varied, steady flow in open channels (Shearman, 

1990). The model was calibrated from measured water-surface 

elevations and discharges along Irondequoit Creek and the 

Millrace from the Narrows to Blossom Road, and was then used 

to predict the water-surface profiles that would result from the 

FCS during high flows. Water-surface elevations were 

measured during major storms with a weighted steel tape from 

surveyed measuring points on six bridges that span Irondequoit 

Creek and the Millrace. These elevations were compared with 

the water-surface profiles generated by WSPRO to confirm the 

extent of backwater from the FCS. 

The second study entailed making high-flow 

measurements at Blossom Road, about 13,000 ft upstream from 
Methods 13

the FCS, and checking against the stage-to-discharge relation 

for the USGS streamflow-monitoring station above Blossom 

Road to identify any measurable backwater that could be 

attributed to the FCS at that point. The stage-to-discharge 

relation was based on 15 years (1981-96) of discharge 

measurements. Discharge was measured by standard USGS 

current-meter methods as described by Rantz and others (1982).

Plant Surveys

Five wetland-plant surveys were conducted. The first two 

(1991 and 1996) preceded the FCS installation and were done 

by a research team from the Departments of Biology at Ithaca 

College, Ithaca, N.Y., and Rochester Institute of Technology, 

Rochester, N.Y. (Bernard and Seischab, 1991; 1997; Bernard 

and others, 1998). The last three surveys (1999-2001) were 

done by the State University of New York (SUNY) College at 

Brockport (Hunter and others, 2000; Noll and others, 2002). 

The survey design was based on the hypothesis that the 

FCS might cause prolonged periods of higher water levels in the 

southern wetland area than typically occurred prior to FCS 

installation and could thereby (1) decrease the density and 

growth of cattails and decrease plant-species richness; or (2) 

alter the chemical composition of cattail tissue upstream from 

the FCS through preferential deposition of sediment. The 

surveys consisted of (1) identification of individual plant 

species, (2) mapping of plant communities, (3) measurements 

of cattail density and above- and below-ground biomass, and 

(4) chemical analyses of above- and below-ground cattail 

tissues. The 1991 plant identification and mapping, and 

measurements of cattail density, were conducted on June 5 and 

July 5; cattail tissues for biomass calculations and chemical 

analyses were collected from August 15 to September 10. All 

components of the 1996 study were conducted from July 25 to 

August 25. All work for the 1999, 2000, and 2001 surveys was 

done during August 14-18 of each year. The 2000 survey was 

done gratuitously by the SUNY Brockport researchers and 

included only plant identification and measurements of cattail 

density and biomass. 

Two sampling transects (fig. 7) were established—one 

across the southern part of the wetland, and one across the 

northern part—along which 3.3 x 3.3-ft(1-m2) permanent plots 

were established every 33 ft (10 m); the northeastern corner of 

each plot was marked with a 1.5-in. PVC pipe. A third transect 

was established in 1996 in a cove in the southeastern part of the 

wetland that had been identified in 1991 as an area of unique 

 

Table 1. Components of ecological-monitoring program for Ellison Park wetland study, Monroe County, N.Y., 

1991-2002, by year

Component 1991 1992 1993 1994 1995 1996 1997a

 a Flow-control structure was installed February 1997; gates in center notch were installed September 1997.

1998 1999 2000 2001 2002

Plant survey

Identify plant species, map plant 

communities, measure cattail density and 

biomass, analyze above- and belowground cattail tissues along permanent 

transects across wetland 

X- - - -X- -X Xb

 b Chemical analyses of cattail tissues and calculation of chemical standing stocks were not done.

X -

Fish survey

Document species diversity and 

community composition, use of wetland 

by resident and nonresident fish, species 

distribution, and symptoms of 

contaminant-related stress

X- - - -X- -X-XBenthic-macroinvertebrate survey

Document species diversity and 

distribution, confirm presence or absence 

of selected species 

X - - - -X- - - - - -

Bird survey

Document species diversity, confirm 

presence or absence of selected species

X - - - -X- - - - -X

Sedimentation measurement

Measure accumulation at permanent 

measurement points 

- XXXXXXXXXXX

Sediment-quality analysis

Periodic collection and analysis of samples 

for major and trace elements, and organic 

compounds, including pesticides, polychlorinated biphenyls, and polycyclic 

aromatic hyrdocarbons

- - -X- -X-X-X-
14 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Empire Boulevard

Northern

Area

Southern

Area

Irondequoit

Bay

Irondequoit

Browncroft

Boulevard

Millrace

The Narrows

and flow-control

structure

Cove

site

EXPLANATION

Wetland boundary

Plant-survey transect

Gill net

Trap net

Minnow trap

0 500 1000 FEET

0 125 250 METERS

Base from New York State Plane NAD 83 Orthoimagery at URL

http://www.nysgis.state.ny.us/gateway/mg/monroe_download.htm, 2002

43o09'30"

43o10'

43o10'30"

77o32' 77o31'30" 77o31'

Creek

Figure 7. Locations of data-collection sites for plant and fish surveys in Ellison Park wetland, Monroe County, N.Y., 1991-2001
Methods 15

plant-species richness. Species nomenclature was in accordance 

with Gleason and Cronquist (1991), except for Typha glauca, 

which in this study was treated as a species, whereas Gleason 

and Cronquist (1991) treat it as a hybrid. The height of cattail 

shoots, cattail density, percentage of plot surface covered by 

cattails, and water depth were measured at each plot. 

Additional plots were established in 1999 along the 

elevated banks (levees) of the main channel of Irondequoit 

Creek, where a relatively large wetness gradient—from 

saturated to dry conditions—extends along a relatively narrow 

strip of land and permits the growth of a more diverse plant 

community than is found in the low-lying, cattail-dominated 

areas nearby, where wetness conditions are uniform, at least 

seasonally. Aquatic plots in the channels adjacent to the levee 

plots were established to identify submerged and floating 

plants. These plots were surveyed during 1999 and 2001 to 

document species richness.

Methods for biomass collection and processing, and for 

chemical analysis of the cattail-tissue samples varied, 

depending on the contractor. Every effort was made to 

minimize these differences to ensure comparability of the 

analytical results.

1991 and 1996 Protocols

Sample collection and biomass calculation. During the 

1991 survey, samples of above-ground tissue were harvested 

from a 3.3 x 3.3-ft (1-m2) area adjacent to each permanent plot, 

and samples of below-ground tissue were collected from a 0.82 

x 0.82-ft (0.25 x 0.25-m) area within each of the adjacent plots 

to a depth of about 12 in. (30 cm). Each sample was washed free 

of sediment and debris, then air-dried and weighed, and the 

biomass per unit area was computed. Unusually high water 

levels and hazardous walking conditions in the field, as well as 

a need to decrease the time needed to process the large volume 

of plant material, prompted modification of the above 

procedures for biomass collection across the northern transect 

during the 1991 survey. This modification entailed collection of 

a representative sample of above-ground biomass (2 or 3 cattail 

shoots) adjacent to each permanent plot, and use of their 

rhizomes and roots as the below-ground biomass sample. After 

drying and weighing, the above-ground biomass per unit area 

was computed as the product of the average weight per shoot 

and the number of shoots in the adjacent permanent plot. The 

below-ground biomass per unit area was calculated similarly 

from the average rhizome-plus-root weight per shoot. This 

method of calculation was reported to give reliable results for a 

Carex sp. (sedge) wetland in New York (Bernard and Gorham, 

1978). 

During the 1996 survey, five representative sites were 

randomly chosen along each transect for collection of biomass 

samples. As in the 1991 survey, samples of above-ground tissue 

were harvested from a 3.3 x 3.3-ft (1-m2) area adjacent to the 

transect of permanent plots, but the samples of below-ground 

tissue were collected within these areas from a larger 1.6 x 1.6-

ft(0.5 x 0.5-m) area to a depth of about 12 in. (30 cm). Samples 

were processed as described for the 1991 survey. 

Preparation for chemical analysis. A small subsample of 

each above-ground sample and each below-ground sample was 

processed for chemical analysis before drying. These 

subsamples were washed in a solution of detergent and 0.4-

percent hydrochloric acid for 30 s, rinsed twice in distilled 

water, oven-dried at 70 ̊C, and ground in a Wiley mill with a 40-

mesh screen. The 1991 samples were composited according to 

soil wetness at the permanent plots, and at least three composite 

samples from each of three “wetness” environments (dry, 

shallow-water, and deep-water areas) along each transect were 

processed. Each of the five 1996 samples was processed 

individually. All samples were analyzed for metals (copper, 

nickel, chromium, cobalt, molybdenum, zinc, aluminum, iron, 

boron, manganese, and lead) and nutrients (nitrogen, 

phosphorus, potassium, calcium, magnesium, and sodium) at 

the Soil and Plant Testing Laboratory in the Soils, Crops and 

Atmospheric Sciences Department at Cornell University, 

Ithaca, N.Y. 

1999 and 2001 Protocols

Sample collection and biomass calculation. Biomass 

samples were harvested from randomly selected sites between 

every third and fourth permanent plot along each transect. 

Above- and below-ground biomass samples were collected 

from a 0.82 x 0.82-ft (0.25 x 0.25-m) area at each site; belowground samples were collected to a depth of about 12 in. 

(30 cm) and rinsed free of sediment and debris in the field. All 

samples were stored at 4°C to minimize weight loss through 

respiration before processing. Samples were then heated to 

65°C and dried to constant weight; biomass per unit area was 

computed as the product of mass per cattail shoot and density of 

cattail shoots.

Preparation for chemical analysis. Cattail tissues were 

ground and sieved through a 60-mesh screen, and 1-g samples 

of dried material were ashed in a muffle furnace at 360 to 

400° C for 24 to 36 h. After cooling, the samples were removed 

and weighed for the loss of carbon on ignition. Each sample was 

dissolved in 5 mL of 6M nitric acid and 5 mL of concentrated 

hydrogen peroxide, then diluted with 10 mL deionized water. 

The samples were stored in plastic bottles and refrigerated. 

Analysis was done by inductively coupled plasma–atomic 

emission spectroscopy—the 1999 samples at XRAL 

Laboratories, Toronto, Canada, and the 2001 samples at Copper 

State Analytical Laboratories, Tucson, Ariz. Sample 

preparation for nitrogen-content analyses entailed grinding at 

least 2 g of each sample to pass through a 60-mesh screen. The 

1999 samples were analyzed at the Soil and Plant Tissue 

Testing Laboratory, University of Massachusetts, Amherst, 

Mass.; the 2001 samples were analyzed at the Agricultural 

Analytical Services Laboratory, Pennsylvania State University, 

University Park, Pa. Percent nitrogen was calculated by the 

Kjeldahl method (Issac and Johnson, 1976).
16 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Fish Surveys

Fish surveys, conducted during 1991-92, 1996, 1999, and 

2001, were done by the Department of Environmental and 

Forest Biology, SUNY College of Environmental Science and 

Forestry, at Syracuse (Miller and Ringler, 1992; Crabtree and 

Ringler, 1998; Thompson and others, 2000; Crabtree and 

Ringler, 2002). The same contractor was chosen for each survey 

to ensure procedural uniformity. The survey design was based 

on the hypotheses that the FCS might (1) act as a physical 

barrier to fish passage and thereby cause a decrease in use of the 

southern wetland area by nonresident fish species; or (2) cause 

a change in nutrient and sediment distribution and accumulation 

rates (an increase in the southern wetland area and a decrease in 

the northern wetland area) and thereby alter the growth rates of 

selected fish species or the trophic status of the fish community. 

The surveys were designed to identify fish species and to 

document abundance, diversity, and the distribution of species 

within the wetland.

Sampling Methods 

Three fish-sampling methods were used during most of the 

surveys—South Dakota trapnets (1.5-m x 1.5-m opening with 

1-cm mesh), Gee wire minnow traps (44-cm long x 9-cm 

diameter with 0.5-cm mesh), and a gill net (50-m x 2-m with 5 

panels of mesh sizes 1.2, 2.5, 3.5, 5.1, and 6.3 cm) (table 2). 

During the 2001 survey, only trapnets were used, and during 

September 1991, when low water prohibited the use of trapnets, 

seining and gill nets were used. Other techniques— 

electrofishing in 1996 and video monitoring of fish passage 

through the FCS in 2001—were used to test their utility in 

identifying any effect the FCS might have on the fish 

community. The same sampling effort was expended upstream 

and downstream from the FCS during each survey. Generally, 

one gill net, two trapnets, and 20 minnow traps were deployed 

in each area for 1, 4, and 6 consecutive nights, respectively, 

during each of three seasons (late spring, midsummer, and early 

fall) during a given survey year (fig. 7). Captured fish were 

identified and tallied; a random subset of black crappie, 

bluegill, largemouth bass, northern pike, and pumpkinseed 

were measured for total length and weighed, and a scale sample 

was collected for calculation of growth rates for these species. 

Species Classification

Fish were classified as resident species (those likely to 

spend all of their lives within a marsh setting) or nonresident 

species (those that migrate into a marsh environment for only 

part of their life cycle). This classification was based on a 

correspondence analysis by Jude and Pappas (1992), which 

separated species along an environmental gradient or according 

to habitat type. Any species that was recognized by Jude and 

Pappas (1992) as a marsh species was classified as a resident 

species of the wetland; all others were classified as nonresident 

species. 

Age and Growth-Rate Calculation

The age of five species—black crappie, bluegill, 

largemouth bass, northern pike, and pumpkinseed—was 

estimated from scale samples. Measurements were made from 

the center of the scale (focus) to the margin and to every 

apparent true annulus; the sum of the annuli corresponded to the 

age of the fish. The length of an individual at an estimated age 

was adjusted for the growth of the fish before scale formation 

by the Fraser-Lee method (Bagenal and Tesch, 1978) to obtain 

the mean length for each age (cohort) within a species. This 

calculation was based on the total length of the fish, the radius 

of the scale, and the radius of an individual annulus. The mass 

of an individual at an estimated age was calculated from the age 

of the fish and the measured mass or, if only length was 

measured, from a linear regression of mass and length. The 

mean mass was the average mass of all fish of a specific age. 

The instantaneous growth rate for a species was calculated as 

Table 2. Sampling methods for fish surveys in Ellison Park wetland, Monroe County, N.Y., 1991-2001

[Number, number of devices or techniques applied. DS (US), downstream (upstream) from flow-control structure.

Sampling days, number of sampling days during given survey (generally divided equally among three sampling periods each year).]

Sampling

method

1991-92 1996 1999 2001

Number Sampling days Number Sampling

days

Number Sampling 

days

Number Sampling 

DS US 7/91, 5/92 9/91 days a

 a Extremely low water levels during September 1991 made seining and increased use of gill nets necessary.

DS US DS US DS US 

South Dakota trap nets 2 2 8 0 2 2 12 2 2 12 2 2b

 b Extremely low water levels prevented use of trap nets in wetland area upstream from flow-control structure 

12

Gee minnow traps 20 20 12 6 20 20 18 20 20 18 0 0 0

Gill nets 1 1 2 6 1 1 3 1 1 2 0 0 0

20-foot bag seine hauls 7 0 0 1 0 0 0 0 0 0 0 0 0

Electrofishing sites 0 0 0 0 9 9 1 0 0 0 0 0 0

Video monitoring - - - - - - - - - - - - 3

.
Methods 17

the difference between the mean mass of a cohort for year (n) 

and that of a cohort for year (n-1) for each species (Bagenal and 

Tesch, 1978).

Trophic Classification

Fish species were classified as insectivore, invertivore, 

omnivore, or piscivore on the basis of Plafkin and others (1989). 

Percentages of species in each trophic group were computed for 

each season and for each wetland area (upstream and 

downstream from the flow-control structure).

Contaminant-Related Stress Indicators

Brown bullhead have been used as biological indicators of 

stress associated with contaminated sediments (Hickey and 

others, 1990; Hirethota and Ringler, 1993). Sediment from 

highly urbanized areas can contain high concentrations of 

potentially toxic chemicals, such as metals and organic 

compounds, and can accumulate in depositional areas such as 

wetlands and pose a hazard to bottom-feeding fish. Brown 

bullhead were examined for signs of contaminant-related stress, 

including barbel deformities, melanotic pigmentation, lesions, 

ulcers, and tumors.

Benthic-Macroinvertebrate Surveys

Benthic-macroinvertebrate surveys were conducted during 

1991-92 by the Department of Environmental and Forest 

Biology, SUNY College of Environmental Science and 

Forestry, at Syracuse (Miller and Ringler, 1992), and during 

1996 by the Department of Biological Sciences, SUNY College 

at Brockport (Haynes and others, 1998). No survey was 

conducted after installation of the FCS. 

The design of the 1991-92 survey was to provide a 

“snapshot” of the benthic-macroinvertebrate community 

through identification and enumeration of individuals found in 

the stomach contents of fish captured during July or September 

1991 or May 1992 (Miller and Ringler, 1992). Stomachs were 

removed from formalin-preserved fish, and the contents were 

flushed, identified, and enumerated. Insects were keyed to the 

lowest possible taxonomic unit, which generally was family. 

The design of the 1996 survey was more comprehensive 

than the 1991-92 survey and, in anticipation of similar surveys 

after FCS installation, was based on the hypothesis that the FCS 

might cause a change in the distribution and accumulation rates 

of sediment and sediment-associated pollutants (an increase in 

the southern wetland area and a decrease in the northern 

wetland area) and thereby alter the composition and species 

abundance of the benthic-macroinvertebrate community. The 

survey was designed to identify benthic-macroinvertebrate 

species and to document abundant, present, and absent species 

and the distribution of species within the wetland. A survey of 

the macroinvertebrate communities was conducted from late 

September through October 1996 at creek-channel and 

backwater-channel (wetland) sampling sites in the southern part 

of the wetland, and at two similar sites in the northern part. 

Three microhabitats—channel bottom, vegetation, and pelagic 

(open-water) zone—were sampled at each location (Haynes and 

others, 1998). The channel bottom at the creek site in the 

southern part of the wetland was mostly sand and was sampled 

by the kick method (Bode and others, 1990); the channel 

bottoms of the other three sites were muddy and were sampled 

with an Ekman dredge. Submergent and emergent vegetation 

were sampled with D-ring dip nets. The open-water areas at 

these sites were sampled by multiplate samplers (Bode and 

other, 1990), which were deployed during a 5-week period 

beginning September 25. 

Five replicate samples were collected in each of the three 

microhabitats at each site. Samples were preserved in the field 

in 4-percent formalin solution. After 24 hours, samples were 

transferred to 70-percent ethanol containing rose bengal dye. 

Macroinvertebrates were separated from debris and identified 

to the lowest practicable taxonomic unit. Chironomid 

mouthpart deformities were identified through criteria of 

Warwick (1990, 1989). Physical characteristics of each 

sampling site, including water temperature, specific 

conductance, dissolved-oxygen concentration, pH, flow 

velocity, and water depth, were measured, as were percent 

canopy cover, amount of vegetation, and substrate composition, 

to define the comparability among data collected at the two 

creek sites and the two wetland sites. Community and species 

indicators of water quality were identified, and water-quality 

indices were calculated. 

Bluegill (Lepomis macrochirus) and pumpkinseed (L. 

gibbosus) were collected by electrofishing. Stomach contents 

were removed from 20 fish of each species and preserved in 

formalin. The macroinvertebrates contained therein were 

identified and compared with those identified from the stomach 

contents of fish captured during 1991-92.

Bird Surveys

The New York State Department of Environmental 

Conservation and the Federation of New York State Bird Clubs 

had conducted a statewide survey during 1980-85 that included 

the Ellison Park wetland to identify bird species and map the 

extent of their breeding ranges (Andrle and Carroll, 1988). 

Birds were identified by sight, sound, and nest or egg 

identification and were classified as possible, probable, or 

confirmed breeders on the basis of breeding evidence. Possible

breeders were those species observed in breeding season in 

possible nesting habitat or whose breeding calls were heard. 

Probable breeders were those species that exhibited breeding 

activities, including breeding calls heard on more than one date 

in the same place, territorial behavior, courtship displays, 

visitation of probable nesting site, and nest building or 

excavation of a nest hole. Confirmed breeders were those 

species for which definite evidence of breeding was found, such 

as a used nest, an identifiable nest with eggs, fledglings, and 

adult carrying food for young. 
18 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

The design of the bird surveys that were conducted during 

the Ellison Park wetland study (1990-2001) was based on the 

hypotheses that the FCS might (1) cause changes in the timing, 

duration, and depth of high-water levels during the breeding 

seasons of ground-nesting birds, and (2) increase sedimentation 

rates that would prematurely fill open-water areas. The surveys 

were designed to identify species and document their presence 

or absence. Two precontrol bird surveys and one postcontrol 

survey were conducted. Miller and Ringler (1992) identified 

bird species by sight or call on July 9-15 and September 23-29, 

1991, and May 19-21 and 28-30, 1992, while canoeing in the 

wetland for the 1991-92 fish survey, but did not identify the 

breeding status of these species. McKinney (1996) conducted a 

survey of the breeding birds of the wetland by walking the 

wetland perimeter and making nine canoe trips into the wetland 

during June 10 to July 22, 1996. Spahn (2002) visited the 

wetland five times during May 17 to July 11, 2002—four trips 

on foot and one by canoe. The 1996 and 2002 surveys used bird 

calls and recordings of bird songs to elicit responses, and 

documented evidence of breeding by the same criteria as used 

by Andrle and Carroll (1988).

Sediment Studies

Two main ecological concerns regarding installation and 

operation of the FCS were (1) the probable increase in the 

deposition of fine-grained sediment (silt and clay) in the 

southern wetland area and the accelerated filling-in of openwater areas in the wetland, and (2) an increase in the 

concentration of metals and organic compounds that are 

associated with fine-grained sediments that could adversely 

affect the wetland biota, including benthic macroinvertebrates, 

amphibians, reptiles, bottom-feeding fish, and the predators of 

these animals. To address these concerns, the USGS made 

annual measurements of sediment accumulation (1992-2002) 

and periodically collected sediment samples (during 1994, 

1997, 1999, and 2001) for chemical analyses.

Sedimentation Rate

Measurement of sedimentation rates entailed anchoring 

1-ft-diameter plastic disks to the cattail mat in depositional 

environments and adjacent to open-water areas of the wetland 

(fig. 8) to provide reference surfaces from which sediment 

depths could be measured. Measurements were made annually 

at 6 or 7 points on each disk; the average of these values was 

used as the sediment thickness. Annual inspections coincided 

with seasonally low water levels from October through 

November to ensure that the disks were not submerged at the 

time of inspection. Measurements were adjusted for thickness 

of any organic matter that had accumulated on the disk, 

primarily duckweed and decaying cattails. Sedimentation rates 

(mm/yr) were computed as the net sediment thickness divided 

by the durations of the precontrol and postcontrol periods (in 

years). Use of a length measurement (mm) rather than a weight 

measurement by volume (g/cm3) or area (g/m2) permitted 

comparison of current sedimentation rates with previously 

published historical and near-recent rates. Alternative methods 

were considered, including surveying along permanent 

transects (Ritchie and McHenry, 1985), using sediment traps 

similar to those used in lake-sedimentation studies, using 

radioactive-isotopic dating of sediment cores (Ritchie and 

McHenry, 1990), and using a dendrogeomorphic technique to 

relate tree age to sediment depth over the tree’s root collar and 

initial-growth root zone (Hupp and Morris, 1990), but these 

methods were not used because (1) the surveying technique is 

imprecise in wetland environments, (2) sediment traps tend to 

measure total sediment influx rather than net deposition or a 

sedimentation rate that represents the natural deposition and 

removal processes in a wetland, and (3) radio-isotopic coring 

and dendrogeomorphic measurements provide only long-term 

sedimentation rates.

Sediment Quality

Samples of fine-grained sediment were collected at three 

sampling sites in the Ellison Park wetland during the fall of 

1994, 1997, 1999, and 2001 (fig. 8). Two of the sites were 

upstream from the FCS, and one was downstream; all were 

within open-water, long-term depositional areas and distant 

from channels that might carry high-velocity, erosive flows. 

Samples were carefully scooped under shallow water from the 

top 2 to 3 in. of sediment, packed in ice, and shipped to the 

USGS National Water Quality Laboratory in Arvada, Colo., by 

methods described by Shelton and Capel (1994) and Radtke 

(1997). No samples contained particle sizes greater than sand 

(2 mm and greater). All of each sample was analyzed for (1) 

total organic carbon, (2) polycyclic aromatic hydrocarbons by 

gas chromatography with a mass-spectrometric detection, and 

(3) organochlorine compounds by gas chromatography with 

electron-capture detection as described by Wershaw and others 

(1987); and (4) organophosphate compounds by gas 

chromatography with flame-photometric detection as described 

by Jha and Wydoski (2003). The clay fraction (particle size 

< 63 μm) was analyzed for major and trace elements by 

procedures listed by Timme (1995) and described by Fishman 

and Friedman (1989). Separate samples were collected and 

submitted to the USGS Sediment Laboratory in Iowa City, Iowa 

for particle-size analyses. Concentrations of selected 

constituents were compared with those published as sedimentquality guidelines for assessment of the probability of adverse 

effects to aquatic life.
Methods 19

Empire Boulevard

Northern

Area

Southern

Area

Irondequoit

Bay

Irondequoit

Browncroft

Boulevard

Millrace

The Narrows

and flow-control

structure

0 500 1000 FEET

0 125 250 METERS

Base from New York State Plane NAD 83 Orthoimagery at URL

http://www.nysgis.state.ny.us/gateway/mg/monroe_download.htm, 2002

43o09'30"

43o10'

43o10'30"

77o32' 77o31'30" 77o31'

Creek

EXPLANATION

Wetland boundary

Sedimentation–rate measurement site and site identifier

Sediment–quality sampling site and site identifier

SR1

SQ1

SR5

SQ3

SR4

SR8

SR6

SR9

SR3

SR13

SR12

SR2

SR7 SR1

SR10 SQ2 SQ1

Figure 8. Locations of data-collection sites for sedimentation-rate and sediment-quality studies

in Ellison Park wetland, Monroe County, N.Y., 1991-2002
20 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Surface-Water Data

Stream discharge and water-quality were monitored at the 

USGS streamflow-monitoring stations at the inlet (Blossom 

Road) and outlet (Empire Boulevard) of the wetland (fig. 2) 

throughout the study period. Each monitoring station included 

an electronic data logger, water-stage and temperature sensors, 

and an automatic water sampler. Sensors made measurements 

every 15 min, and the measured values were stored in the data 

loggers. 

Stream Discharge

Discharge at the upstream site was computed from a stageto-discharge relation by standard USGS procedures for gaging 

streamflow (Carter and Davidian, 1968; Rantz and others, 

1982), measuring stage and discharge (Buchanan and Somers, 

1968 and 1969, respectively), and analyzing stage-to-discharge 

relations (Kennedy, 1984). Computation of discharge at the 

downstream site was complicated by backwater conditions 

caused by fluctuating water levels in Irondequoit Bay and Lake 

Ontario; therefore, water velocity was measured with an 

acoustic velocity meter (AVM), and the recorded value was 

correlated with the mean velocity in the channel as calculated 

from discharge measurements made by current meter from the 

Empire Boulevard bridge or by acoustic Doppler current 

profiler (ADCP) from a boat. Discharge was computed from the 

velocity relation and from the relation between stage and flow 

area. Standard USGS procedures for measuring discharge with 

an ADCP (Lipscomb, 1995), and gaging streamflow with an 

AVM and analyzing stage-and-velocity-to-discharge relations 

(Laenen, 1985) were followed.

Water Quality

Automated samplers that extracted water samples from the 

channel (near the centroid of flow) hourly and stored them in 

refrigerated bottles were maintained by the Monroe County 

Environmental Health Laboratory (MCEHL). Samples were 

retrieved twice weekly and delivered to MCEHL. Sampling 

periods were divided on the basis of three flow 

conditions—base or steady flow, and the rising and falling 

phases of a storm hydrograph. If base- or steady-flow 

conditions prevailed during the entire sampling cycle (3 to 4 

days), equal volumes of water from all samples collected during 

that cycle were composited for a single analysis. During storms, 

samples collected during the rising phase were composited 

separately from those collected during the falling phase. Equal 

volumes of water from all samples collected during a given 

phase were composited for a single analysis. Laboratory 

analyses were done by MCEHL, which participated in the 

USGS quality-assurance program for cooperating analytical 

laboratories. Analytical procedures are described in American 

Public Health Association and others (1995). Samples were 

analyzed for phosphorus (total and orthophosphate), nitrogen 

(total, ammonia-plus-organic, nitrate-plus-nitrite, and 

ammonia), chloride, sulfate, and suspended solids (total and 

volatile). Samples collected during stormflows, and 

periodically during low-flow periods, were also analyzed for 

total organic carbon, biochemical and chemical oxygen 

demand, alkalinity, specific conductance, and trace metals 

(zinc, lead, copper, and cadmium). 

Quality Assurance and Quality Control

Quality-assurance and quality-control protocol (QA/QC) 

for the water-sampling and analysis program conducted by 

MCEHL included comparison of constituent concentrations 

measured in depth-integrated cross-sectional samples collected 

by the equal-width-increment (EWI) method (Shelton, 1994) 

with those measured concurrently in samples collected with 

automatic samplers (Auto). Statistical analyses (paired t-test; α

= 0.05) of the mean of the differences between paired 

data—(EWI and Auto)—identified any significant systematic 

bias in the concentrations in the automatically collected water 

samples. 

MCEHL also participated in the semiannual USGS 

Standard Reference Sample (SRS) program for cooperating 

analytical laboratories, which is coordinated by the USGS 

Office of Water Quality. Participating laboratories receive 

(1) reference samples, which are analyzed for trace elements, 

major ions, and nutrients; and (2) low-ionic strength samples 

that simulate precipitation. The analytical results from all 

participating laboratories are transmitted to USGS, and a “most 

probable value” (MPV) is statistically calculated for each 

constituent. MCEHL results were consistently rated 

“satisfactory” to “good” (within 1.50 to 0.51 standard 

deviations of the MPV) for trace elements; and “good” to 

“excellent” (within 1.00 to 0.00 standard deviations of the 

MPV) for major ions, nutrients, and mercury, and for low-ionicstrength samples (table 3). 

Constituent Loads and Removal Efficiency

Monthly and annual loads of phosphorus compounds, 

nitrogen compounds, suspended solids, chloride, and sulfate at 

the two water-quality-monitoring stations were calculated by 

the USGS program ESTIMATOR (Gregory Baier, Timothy 

Cohn, and Edward Gilroy, U.S. Geological Survey, written 

commun., 1995), which is based on a log-linear regression 

equation that relates nutrient concentrations to surrogate 

variables of discharge, time (year and decimal-date), and season 

(Cohn and others, 1989; 1992). Nine variables were considered 

for each equation: a constant; a quadratic fit to the logarithm of 

discharge (log flow and square of log flow); a quadratic fit to 

time (decimal time and square of decimal time); and two 

sinusoidal (first- and second-order Fourier) functions to account 

for the effects of annual seasonality (Cohn and others, 1992). 

Pre- and postcontrol loads were computed from data 

collected during each respective period. The loads for October 
Table 3. Performance ratings for selected standard reference samples analyzed by Monroe County Environmental Health Laborato

1990-2001 (results of U.S. Geological Survey’s analytical evaluation program) [Rating scale: 4, excellent; 3, good; 2, satisfactory; 1, marginal; 0, unsatisfactory. n, number of samples analyzed. bdl, concentration below detec

limit of analytical equipment used by laboratory.] 

Date

of

analyses

Overall

weighted

rating n

Nutrientsa

 a NH3, ammonia; NH3+Org N, ammonia-plus-organic nitrogen; NOx, nitrate plus nitrite; TP, total phosphorus; PO4, orthophosphate.

Major ionsb and physical properties

 b Alk, alkalinity; Cl, chloride; SO4, sulfate; SC, specific conductance. 

Trac

 c Cd, cadmium; Cu, copper; Pb, lead; Zn, zinc.

Ratingd

 d Overall rating for constituent category.

n NH3

NH3+

Org N NOx TP PO4 Rating4 n Alk Cl SO4 TP pH SC Rating4 n 

May 1990 3.3 51 3.3 11 3 3.5 2 4 4 3.6 13 4 3 3 4 4 3 3.1 15 

Nov. 1990 3.1 59 3.2 22 2 3.5 4 4 3.5 3.5 12 4 4 3 3 4 2.5 15 Apr. 1991 3.0 56 3.6 14 4 3.8 4 3.8 4 3.0 13 4 4 3 4 4 0 2.3 18 Oct. 1992 3.1 55 3.7 18 3.8 3 4 3.8 4 3.7 13 4 4 4 4 4 3 2.0 15 

Apr. 1993 3.0 59 3.8 18 4 3.2 4 3.8 3.8 2.7 13 3 0 0 4 4 4 2.6 28 Oct. 1993 3.0 97 3.8 - 4 3.5 4 3.8 3.5 3.5 13 4 4 4 4 4 3 2.5 18 

Oct. 1994 2.6 75 3.7 10 3 4 4 4 3.5 3.2 13 4 4 4 3 2 2 2.2 40 May 1995 2.9 57 3.3 10 3.5 3 3.5 3.5 3 3.4 13 4 4 4 4 0 3 2.5 22 Oct. 1995 3.1 54 3.4 10 3.5 2 4 3.5 4 3.2 13 4 3 4 2 4 3 2.7 19 Apr. 1996 2.6 68 3.8 5 4 4 3 4 4 3.2 13 4 1 0 4 4 3 2.3 38 Sept. 1996 3.0 76 3.4 9 3.5 2.5 4 4 3.5 3.0 13 4 3 4 4 3 3 2.8 43 Apr. 1997 2.9 85 3.6 10 3.5 3.5 3 4 4 3.2 14 4 3 4 4 4 4 2.5 35 Sept. 1997 2.7 102 3.7 10 3.5 3.5 3.5 4 4 3.2 13 3 4 3 3 4 4 1.7 17 Apr. 1998 2.7 78 3.8 10 4 4 3.5 3.5 4 2.6 13 3 4 3 bdl 4 3 2.4 20 Sept. 1998 2.6 72 3.6 10 3.5 3.5 3.5 3.5 4 2.8 11 bdl 4 4 3 4 2 2.4 20 Mar. 1999 2.9 76 3.1 8 4 4 3 3.5 1.5 2.9 13 3 4 3 0 4 2 2.3 22 Oct. 1999 2.5 57 3.5 10 3.5 3.5 3.5 4 3 2.2 13 0 3 4 4 0 4 2.1 22 Mar. 2000 2.9 59 3.7 10 3.5 3.5 3.5 4 4 3.1 14 4 4 4 4 2 3 2.3 23 Oct. 2000 2.7 59 3.7 10 4 4.5 4.5 3.5 4 2.7 14 4 4 3 4 4 0 2.0 23 Apr. 2001 2.8 59 3.8 10 3.5 4 3.5 4 4 3.3 14 4 4 4 4 4 4 1.7 23 Sept. 2001 2.6 57 3.5 10 3.5 3.5 3 3.5 4 3.1 14 4 4 4 4 4 4 2.1 22 Average 2.9 3.6 3.6 3.5 3.6 3.8 3.7 3.1 3.8 3.4 3.3 3.7 3.3 2.9 2.3 
22 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

1990 through September 1996 were those computed by Coon 

(2000). The loads for October 1996 through February 1997, 

when the base control was installed, were computed from the 

data collected from October 1990 through February 1997, and 

were appended to the loads calculated by Coon (2000) to 

complete the precontrol data set. Postcontrol loads were 

calculated from data collected from March 1997 through 

September 2001. 

Each dataset was used to develop a load equation for each 

constituent at each sampling site. For a given constituent, if the 

coefficient for a variable was significantly different from zero 

(p < 0.05) for either the inflow- or outflow-load equation, that 

variable was included in both equations to maintain 

comparability of the results (table 4). The postcontrol equations 

used the same variables as the precontrol equations. The 

constituent loads were computed from continuous water-quality 

data (composited hourly samples); typically at least two 

samples were provided per week, but more were obtained 

during storms, when samples were composited during the rising 

and falling phases. The discharge associated with a given water 

sample was computed as the mean of the 15-min discharges 

indicated by the stage-to-discharge relation at the Blossom 

Road monitoring station and the stage-and-velocity-todischarge relation at the Empire Boulevard site for the period 

covered by the hourly water samples that were composited into 

the analyzed sample. 

Daily loads were computed by the load equation from the 

daily mean discharges recorded at each site. From these data, 

total monthly, annual, and period-of-record loads were 

computed for each constituent. (See appendix.) The precision of 

the estimated loads can be described in terms of a confidence 

interval, which is based on the estimated mean and the standard 

error of prediction calculated by the equation. At a 95-percent 

confidence interval (α = 0.05), the confidence limits are the 

estimated load ±1.96 x the standard error of prediction. (See 

appendix.)

Potential problems in calculating loads from regression 

equations include (1) multiply censored water-quality data 

(which concentrations are found to be below an analytical 

detection limit that can change with improvements in analytical 

measurement capabilities), (2) logarithm-retransformation bias, 

and (3) lack of fit between predicted and observed values. The 

ESTIMATOR program handles multiply censored data through 

an adjusted maximum-likelihood estimation procedure (Cohn, 

1988; 1995). Loads computed from the logarithm of constituent 

concentrations can show bias when the loads are transformed 

back into nonlogarithmic units. The ESTIMATOR program 

corrects this bias, which tends to underestimate the loads, 

through a minimum-variance unbiased estimator (Cohn, 1988; 

1995; Cohn and others, 1992). Misspecification of a regression 

equation, which is likely to result when linear regression is used 

to describe the relation among time-series data, can be indicated 

by serial correlation (nonindependence) of the residuals. The 

ESTIMATOR program minimizes serial correlation of 

residuals and the effect of seasonality by including sinusoidal 

functions of time (sin2πT, cos2πT, sin4πT, cos4πT, where T is 

time, in years) (Cohn and others, 1992). Lack of fit of a 

regression equation can be indicated by a low coefficient of 

determination and by nonnormal distribution and serial 

correlation of the residuals. Cohn and others (1992) point out 

that (1) load estimators based on log-linear equations, in 

general, appear to be relatively insensitive to violations of the 

assumptions of linear regression—that is, nonnormality and 

independence of the residuals; (2) a minimum variance 

unbiased estimator, in particular, which is based on the 

assumptions of a log-linear equation, provides reliable 

estimates of loads despite these violations; and (3) the variances 

of annual or monthly load estimates based on infrequent 

sampling appear to be well described by log-linear equation 

theory. Equation validity or goodness of fit was evaluated 

through regression diagnostic statistics—standard error of 

prediction (in appendix), coefficient of determination, standard 

Table 4. Explanatory variables used in regression equations developed to compute loads for selected constituents in 

water from Irondequoit Creek, Monroe County, N.Y., 1990-2001, at sites above Blossom Road (inflow to wetland) and

Empire Boulevard (outflow from wetland) 

[T, time, in years. Coefficient for specified variable was found to be significantly different from zero (p-value less than 0.05) for one or

both of the inflow-load and outflow-load equations. If, for a given constituent, a variable was significant for only one of the equations,

it was nevertheless used in both equations to maintain comparability of results.]

Constituent

Constant

Log 

flow

Log flow 

squared

Decimal 

time

Decimal 

time 

squared

Seasonality function 

Sin(2πT) Cos(2πT) Sin(4πT) Cos(4πT)

Total phosphorus, as P X XX X X X X - -

Orthophosphate, as P X X X X X XXXX

Ammonia plus organic nitrogen, as N X X - X - XXXX

Ammonia nitrogen, as N X X - X - XXXX

Nitrate plus nitrite, as N X X X X X XXXX

Total suspended solids XX- X X X X - -

Volatile suspended solids XX- X X X X - -

Chloride X X X X - XXXX

Sulfate X X X X X ----
Effects of flow modification 23

error of regression, serial correlation of residuals, and 

probability plot-correlation coefficient (a measure of the 

normality of the residuals)—with consideration given to the 

number of total and censored observations included in the 

development of each equation (table 5). 

The USGS and MCEHL collected an extensive 

streamflow- and water-quality data set. As a result, the chemical 

loads could have been computed directly from the 

concentration data by means of a spreadsheet, but the 

ESTIMATOR program was chosen because it provided easier, 

more efficient computation. The validity of using the 

ESTIMATOR-computed loads rather than manually computed 

loads was tested through a statistical comparison of manually 

computed monthly 1997-98 loads with the ESTIMATOR loads. 

The Wilcoxon signed-ranks test in SAS (SAS Institute, Inc., 

1989; 1990) was used to identify statistically significant 

differences between the values obtained by each method. Of 20 

comparisons of 1997-98 loads of phosphorus compounds, 

nitrogen compounds, suspended solids, chloride, and sulfate at 

the inflow and outflow sites, 17 indicated no significant 

differences (p = 0.1180 to 0.9779) between the two 

computation methods; the remaining three indicated significant 

differences for ammonia, ammonia-plus-organic nitrogen, and 

total nitrogen loads at the outflow site only; the corresponding 

p values were 0.0010, 0.0064, and 0.0396. These results 

indicated that use of the ESTIMATOR program to compute 

loads was a valid alternative to manual computation of loads.

The wetland’s effectiveness in decreasing chemical loads 

can be evaluated in terms of removal efficiency—the 

percentage of an inflow load that is retained in, or exported 

from, the wetland. Removal efficiency for a given constituent 

was computed as the difference between inflow and outflow 

load divided by the inflow load and expressed as a percentage. 

Monthly and annual removal efficiencies were computed for 

each constituent. (See appendix.) A positive value indicated a 

net retention of a constituent in the wetland, and a negative 

value indicated the generation of a constituent within the 

wetland (or, possibly, its adjacent drainage area) and a net 

export of a constituent.

Statistical Analyses of Load Data

The computed loads were statistically analyzed through 

SAS (SAS Institute, Inc., 1989; 1990) to provide an objective 

basis for assessing the effectiveness of the FCS in decreasing 

chemical loads. Nonparametric tests were used because most of 

the monthly load data did not show normal distributions. Three 

approaches were followed: 

 (1) Statistically significant differences between the monthly 

inflow and outflow loads were identified by the Wilcoxon 

signed-ranks test on the ranked differences between the 

paired (inflow and outflow) loads. This test indicated 

whether the median of the differences between inflow and 

outflow loads was significantly different from zero. 

 (2) Three major high flows during the postcontrol period 

exceeded all of the peak flows that occurred during the 

precontrol period. This dissimilarity could greatly affect 

the interpretation of the FCS’s ability to decrease 

chemical loads; therefore, a second set of data was 

generated. The estimated monthly loads (in tons) were 

divided by the total monthly volume of water (in ft3) 

passing the respective streamflow-monitoring sites, and 

the resulting values were converted to monthly flowweighted concentrations (in mg/L). Wilcoxon signedranks tests were performed on the ranked differences 

between the paired (inflow and outflow) flow-weighted 

concentrations. 

 (3) The precontrol data reflected the effect of the wetland on 

chemical loads, whereas the postcontrol data reflected the 

combined effects of the wetland and the FCS. The 

wetland’s effect was removed from a given analysis by 

adding or subtracting, as necessary, the mean precontrol 

difference—inflow load or flow-weighted concentration 

for a given constituent minus its outflow load or flowweighted concentration, as described above—from the 

differences between the postcontrol inflow and outflow 

loads or flow-weighted concentrations. Theoretically, the 

result would represent the change in chemical load that 

could be attributed solely to the FCS. These adjustments 

were made to the load- and the flow-weightedconcentration differences, and the Wilcoxon signed-ranks 

tests were again performed.

EFFECTS OF FLOW MODIFICATION

Throughout the following discussion of the effects of flow 

modifications on the Ellison Park wetland, references are made 

to the precontrol period (or before FCS installation) and the 

postcontrol period (or after FCS installation). In all instances, 

these references are meant to include not only the installation of 

the FCS, but also the changes in flow patterns that resulted from 

the bank notches and diversions from Irondequoit Creek to the 

Millrace upstream from the wetland, as described in the section 

“Additional Flow Modifications” (p. 11). Although these 

additional flow modifications were made over a period from 

February 1997 through the summer of 1998, the pre- and 

postcontrol periods have been defined by the month of the FCS 

installation (February 1997) for comparison purposes.

The FCS was expected to have several measurable effects. 

First, it was designed to attenuate the discharge of many, if not 

all, peak flows passing through the wetland and to cause some 

amount of backwater upstream. It also was expected to increase 

the dispersal and detention of stormflows in the upstream part 

of the wetland and thereby affect the loads of chemical 

constituents carried by Irondequoit Creek into Irondequoit Bay. 

The FCS, as well as the other flow modifications, could increase 

the sedimentation rate of fine-grained sediments, which in turn 
Table 5. Regression statistics for equations used to compute constituent loads for Irondequoit Creek sampling sites above Blossom Road an

Empire Boulevard, Monroe County, N.Y., 1990-2001 Constituent Precontrol period (October 1990 through February 1997) Postcontrol period (March 1997 

Number of observations Coefficient of deter- minationa a Coefficient of determination (R2) - the percentage of the total variance in the computed loads that is accounted for by the regression equation. Values close t

Standard error of regressionb b Standard error of regression - a measure of the dispersion of the data around the regression line. Low values are desired. Serial correlation of residualsc c Serial correlation of residuals - a measure of the dependence or correlation in time sequence between residuals. The serial correlation coefficient ranges from 

1.00 (strong serial correlation). PPCCd d Probability plot correlation coefficient - a measure of the likelihood that the residuals are normally distributed. A value greater than 0.97 implies that the dat

from a population with a normal distribution. Number of observations Coefficient of deter- mination Stand

erro

regres

Total Censored Total Censored 

Blossom Road Total phosphorus, as P 1080 0 0.748 0.664 0.314 0.988 603 0 0.812 0.5

Orthophosphate, as P 1081 3 .764 .477 .508 .989 631 2 .891 .34Ammonia-plus-organic nitrogen, as N 1075 2 .851 .381 .382 .989 601 5 .860 .4Ammonia nitrogen, as N 1044 473 .574 .855 .432 .953 625 338 .709 .8Nitrate plus nitrite, as N 1065 0 .939 .214 .635 .950 591 0 .960 .1Total suspended solids 324 4 .664 .838 .197 .982 191 0 .708 .94Volatile suspended solids 320 23 .647 .724 .271 .986 189 33 .704 .84Chloride 1086 0 .928 .196 .578 .974 595 0 .937 .2Sulfate 1074 1 .853 .174 .177 .936 595 0 .891 .1

Empire Boulevard Total phosphorus, as P 940 0 .836 .445 .404 .977 572 0 .843 .44

Orthophosphate, as P 941 1 .871 .320 .535 .983 670 1 .903 .3Ammonia-plus-organic nitrogen, as N 928 0 .876 .309 .468 .981 574 5 .872 .3Ammonia nitrogen, as N 911 99 .523 .700 .575 .986 660 29 .772 .5Nitrate plus nitrite, as N 929 0 .958 .186 .624 .991 599 0 .970 .1Total suspended solids 243 3 .733 .633 .224 .984 159 1 .691 .8Volatile suspended solids 242 29 .752 .504 .292 .987 159 27 .681 .7Chloride 944 0 .930 .200 .572 .978 602 0 .941 .2Sulfate 928 0 .874 .163 .246 .975 600 0 .908 .14
Effects of flow modification 25

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002

0

3,500

500

1,000

1,500

2,000

2,500

3,000

DISCHARGE, IN CUBIC FEET PER SECOND

Flow-control structure

installed during 

late February 1997

could (1) accelerate the infilling of ponded areas in the upstream 

part of the wetland, and (2) increase the concentrations of 

sediment-associated chemicals in depositional sites in that area. 

The flow modifications also could cause changes in the wetland 

plant, fish, benthic-macroinvertebrate, and bird communities. 

The interpretation of results took into account the effects 

of precipitation and streamflow differences between the preand postcontrol periods, as follows. Precipitation quantities and 

monthly mean streamflows (fig. 3) were more extreme during 

the postcontrol period than during the precontrol period. 

Monthly precipitation for July 1992, October 1995, and June 

1996 (the precontrol period) greatly exceeded the long-term 

average, whereas two prolonged wet periods occurred during 

the postcontrol period June through August 1998 and April 

through September 2000. One- to 3-month periods of belowaverage monthly precipitation occurred throughout the 

study prolonged dry periods of at least 4 months occurred 

from May through August of 1991 and 1993 (4 months), and 

from January through June of 1995 (6 months) of the precontrol 

period. Dry periods occurred from April through July of 1997 

and 2001 (4 months), and from September 1998 through July 

1999 (11 months) of the postcontrol period. 

Peak discharges, which included annual peaks as well as 

flows that exceeded a base discharge of 900 ft3/s, were much 

lower during the precontrol period than during the postcontrol 

period (fig. 9). The maximum flow during the period of record 

at the Blossom Road streamflow-monitoring station before 

installation of the FCS (1982 through February 1997) was 

1,730 ft3/s. Of the nine peak flows that occurred during the 4.5 

years after the FCS installation, four exceeded the precontrol 

maximum flow, and one of these, which occurred within a year 

of the FCS installation, was the peak of record (3,300 ft3/s on 

January 8, 1998—almost twice the previous maximum peak). 

These peak flows greatly affected the chemical loads that 

entered the wetland during 1998 and 2000.

The seasonal periods of wetland inundation during the 

precontrol period were comparable to those of the postcontrol 

period, in that both periods contained persistently inundated 

(wet) and noninundated (dry) growing seasons (fig. 4). The 

prolonged wet condition of 1992-93 did not recur during the 

postcontrol period, but 1997 and 1998 had two successive wet 

periods that, at least during the growing season, were comparable 

to the 1992-93 wet period. Dry, low-water years occurred during 

each period; 1995 during the precontrol period, and 1999, which 

was the drier of the two, during the postcontrol period.

Water-Surface Elevations and Discharge

Water-surface elevations during the postcontrol period 

were compared with precontrol water-surface profiles and those 

predicted by the WSPRO model. Measurements were made 

during five storms with peak flows that ranged from 962 to 

3,300 ft3/s, the latter of which was the period-of-record (21 

years) peak flow at the upstream (Blossom Road) streamflowmonitoring site (table 6). Flooding in Ellison Park occurs when 

discharge exceeds about 900 ft3/s. The measured water-surface 

elevations were lower and the measured falls in water-surface 

elevations from Blossom Road to the Narrows were greater 

than those predicted by the WSPRO model. These results 

reflect the conservative assumptions and parameter values used 

in the model. 

Figure 9. Peak discharges recorded at Irondequoit Creek above Blossom Road, Rochester, N.Y., 1990-2001
26 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

The discharge measurements made at Blossom Road 

during the five storms were compared with the established 

stage-to-discharge relation at the streamflow-monitoring 

station above Blossom Road. Measured discharges ranged from 

311 to 2,910 ft3/s (table 6) and generally were rated as “good,” 

or within 5 percent of the true discharge. Except for the 

measurement on July 9, 1998, all measured discharges were 

within 5 percent of the discharges indicated by the stage-todischarge relation or had a positive departure from the relation; 

that is, a greater value than expected from the stage-todischarge relation. All measurements made on January 8-9, 

1998 (the period-of-record peak flow) showed positive 

departures that are attributed to the uncertainty of the stage-todischarge relation, which had been extrapolated beyond the 

previous highest measurement of 1,400 ft3/s. The measurement 

on July 9, 1998 indicated 0.68 ft of backwater at the 

streamflow-monitoring station, but this was a result of debris 

caught on the branches of a fallen tree that was overhanging the 

channel about 200 ft downstream. The four negative departures 

from the stage-to-discharge relation (table 6), which indicated 

less flow than expected and might have indicated backwater 

conditions, were within the 5-percent accuracy range of the 

measurements.

The FCS caused from 0.9 to 1.5 ft of backwater in the 

wetland immediately upstream of the FCS, depending on 

discharge, the gate and stop-log configuration of the FCS, and 

the amount of debris that had caught on the FCS. Extremely 

high flows were able to pass the FCS on the low, western bank 

at the Narrows; this decreased the magnitude of backwater that 

might have resulted from these flows. The results of these 

studies indicated that (1) the water levels at Blossom Road were 

controlled by the bridge-opening constrictions of Browncroft 

Boulevard over Irondequoit Creek and the Millrace, about 

5,000 ft upstream from the FCS, and (2) backwater from the 

FCS did not extend as far upstream as Blossom Road (about 

13,000 ft upstream from the FCS).

Plants

Three botanical characteristics of the wetland upstream and 

downstream from the FCS were evaluated for possible effects of 

the FCS; these were (1) species richness (number of species) and 

community composition, (2) cattail productivity, and (3) 

chemical composition of cattail tissue. Five surveys were 

conducted, mainly during mid-August of the summers of 1991, 

1996, and 1999-2001. The 2000 survey did not include chemical 

analyses of cattail tissue. Water levels in the wetland were high 

during the 1991 and 2000 surveys; low during the 1999 survey; 

and intermediate during the 1996 and 2001 surveys. Observed 

changes in species richness and cattail-growth characteristics 

appeared to be related to depth of inundation.

Species Richness and Community Composition

Typha glauca was the dominant species within the 

wetland, although 59 other plant species in 10 distinct plant 

communities were identified in 1991 (Bernard and others, 

1998). Species richness within the cattail-dominated areas was 

low; generally the only other major species found among the 

cattails were floating aquatic plants—primarily duckweeds 

(Spirodela polyrhiza, Lemna minor, Wolffia sp.). Species 

richness was greatest in areas of wetness transition, such as 

along the wetland perimeter, where 34 species were identified 

in 1991 (Bernard and Seischab, 1991), and on the elevated 

banks or levees adjacent to the main channel of Irondequoit 

Creek, where 28 species were found in 1991 (Bernard and 

Seischab, 1991), and 22 species were found in selected 1 x 2-m 

plots during 2001 (Noll and others, 2002). These areas differed 

only slightly in elevation but represented a large wetness 

gradient over a relatively short distance that created a variety of 

habitats within a small area. Many of the levee plots were 

dominated by Typha glauca, where an average of 8.2 plant 

Table 6. Factors addressed in assessment of hydrologic effects of flow-control structure, Ellison Park wetland,

Monroe County, N.Y., 1997-2001

Date

Peak flow

(cubic feet

per second)

Water-surface

profile

measured

Backwater at

Blossom Road

indicated

Discharge

measured

Discharge

(cubic feet

per second)

Departure from

stage-discharge

relation 

(percent)

September 29, 1997 962 yes no yes 311 -0.2

January 8, 1998 3,300 yes no yes 2,900 +7.4

2,910 +5.4

January 9, 1998 - no - yes 2,600 +8.3

July 9, 1998 1,830 yes no yes 814 a

July 10, 1998 no - yes 421 -3.4

January 24, 1999 1,900 no - yes 692 -2.9

February 25, 2000 782 no - yes 746 -0.6

May 13, 2000 2,420 yes no no - -

March 23, 2001 1,160 yes no no - -

a

a 0.68 feet of backwater from debris caught on tree overhanging channel about 200 feet downstream from gaging station.
Effects of flow modification 27

species per plot were identified; in contrast, the cattaildominated areas along the permanent transects contained only 

from 1.67 to 2.85 species per plot (table 7; Noll and others, 

2002). Species richness was high in the cove area of the 

wetland, where 23 species were identified in 1999 (Hunter and 

others, 2000), and an average of 3.78 to 7.88 species per plot 

were found (table 7). Deep-water areas throughout the wetland 

were found to contain 19 species of submerged or floating 

aquatic plants in 1991 (Bernard and Seischab, 1991), whereas 

selected 1 x 2-m channel plots adjacent to the levee plots 

contained only 8 species and an average of 2.1 species per plot 

in 2001 (Noll and others, 2002). Species richness also was low 

in Phragmites australis communities, where only 3 other plant 

species were found (Bernard and Seischab, 1991).

Species richness and distribution in the wetland were 

inversely related to depth of inundation of the cattail mat. 

During wet years, such as 1991 and 2000 (fig. 4), when average 

water depths at the permanent transects exceeded 0.25 ft, 6 to 7 

plant species were identified along the northern transect, and 10 

to 14 species were identified along the southern transect 

(table 7). In contrast, during a dry year, such as 1999, when 

average water levels were more than 0.65 ft below the top of the 

cattail mat, 11 and 20 species were found along the northern and 

southern transects, respectively. The increase in species 

numbers during 1999 might have resulted from the germination 

of seeds from annual or short-lived perennial plants—seeds that 

were dormant in the cattail mat awaiting optimum (drier than 

normal) conditions for germination and growth (Noll and 

a Water depth estimated from measured water-surface elevation at mouth of wetland.

Table 7. Measurements obtained in plant survey along three permanent transects in Ellison Park wetland, 

Monroe County, N.Y., 1991-2001

[FCS, flow-control structure. >, greater than; --, not measured. Transect locations are shown in fig. 7.]

Data category and sampling area

Year 

1991 1996 1999 2000 2001

Survey Dates

Transect downstream from FCS July 5 August 21 August 16 August 16 August 14

Transect upstream from FCS June 5 August 19 August 17 August 17 August 15

Cove site upstream from FCS -- August 23 August 18 -- August 16

Number of plots

Transect downstream from FCS 26 20 30 30 30

Transect upstream from FCS 31 31 29 29 29

Cove site upstream from FCS 09809

Approximate water-surface elevation at mouth of wetland (feet, North American Vertical Datum, 1988) 

Transect downstream from FCS 247.22 246.96 246.15 247.45 246.51

Transect upstream from FCS 247.86 246.94 246.11 247.51 246.45

Cove site upstream from FCS -- 246.90 246.22 -- 246.39

Average water depth at transect (feet; negative number indicates depth to ground water) 

Transect downstream from FCS .33 .0 a - .66 a .82 .0

Transect upstream from FCS .36 - .66 a - 1.31 a .26 > - .98

Cove site upstream from FCS -- -- -- -- 1

Number of species

Transect downstream from FCS 7 4 11 6 7

Transect upstream from FCS 10 9 20 14 16

Cove site upstream from FCS -- 12 23 -- 21

Average number of species per plot

Transect downstream from FCS 2.85 1.75 2.47 1.67 2.03

Transect upstream from FCS 2.45 1.81 2.48 2.07 2.31

Cove site upstream from FCS -- 3.78 7.88 -- 6.67

Average number of cattail shoots per 3.3-square-foot (1-square-meter) plot

Transect downstream from FCS 30.6 ± 1.4 38.3 ± 2.7 37.2 ± 1.7 42.8 ± 1.7 35.3 ± 1.7

Transect upstream from FCS 27.4 ± 2.2 34.8 ± 2.2 33.2 ± 1.8 34.4 ± 2.1 30.7 ± 1.7

Cove site upstream from FCS -- 31 ± 4 35 ± 4 -- 25 ± 3

Average height of cattail shoots (feet)

Transect downstream from FCS 10.0 ± 0.06 9.7 ± 0.23 8.7 ± 0.20 9.4 ± 0.20 8.2 ± 0.30

Transect upstream from FCS 7.1 ± 0.23 9.5 ± 0.16 8.5 ± 0.20 9.6 ± 0.36 8.2 ± 0.23

Cove site upstream from FCS -- 9.1 ± 0.43 9.2 ± 0.23 -- 8.9 ± 0.26
28 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

others, 2002). These plants were able to complete their life 

cycle during a single dry growing season. In every survey, 

fewer species were found along the northern transect than along 

the southern transect, where the slightly higher land-surface 

elevation allowed drier conditions than would concurrently 

prevail in the northern wetland area. Conditions along the cove 

transect were drier than along either of the other two transects 

and, thus, showed the greatest species richness during the three 

years in which all three transects were surveyed (1996, 1999, 

and 2001; table 7).

Cattail Productivity

Average height, density, and biomass of Typha glauca

were used as measures of cattail productivity in the wetland 

(tables 7, 8). Cattail height generally averaged 8 to 9 ft by the 

mid-August sampling season but could exceed 10.5 ft locally. 

Cattail density ranged from 25 to 43 shoots per 1-m2 (3.3 x 3.3-

ft2) plot and averaged 34 shoots/m2 among all plots in all five 

survey years. The average weight of an individual cattail shoot 

(above-ground material) ranged from 25.7 to 90.5 g; the 

average weight of below-ground material ranged from 33.7 to 

108 g per shoot (table 8). The measured weights were greatest 

during the 1991 and 1996 surveys; those measured after the 

1997 FCS installation were less than in 1991 and 1996 along the 

upstream and downstream transects but were comparable along 

the cove transect.

Average Typha glauca biomass values were computed 

from the measured values obtained from the samples collected 

along a given transect during a given year (table 8). These 

values were compared with those reported in the literature, 

which reflect seasonal differences within a site, as well as a 

wide range of climatic and ecological differences among sites. 

Above-ground biomass of T. glauca reported in other studies 

ranges from 758 to 2,320 g/m2 (table 9; Van der Valk and 

Davis, 1978; Bernard and Fitz, 1979; Andrews and Pratt, 1978); 

shoot biomass of other Typha species range from 378 to 

2,338 g/m2

 (Whigham and others, 1978; Mason and Bryant, 

1975; McNaughton, 1966; Van der Valk and Davis, 1978; Boyd 

and Hess, 1970; Boyd, 1970, 1971; Bray, 1960; Bray and 

Table 8. Cattail biomass in above- and below-ground tissues along the three permanent transects in 

Ellison Park wetland, Monroe County, N.Y., 1991-2001

[do, ditto; --, not measured; g, grams; kg/m2, kilograms per square meter. Locations of transects are shown in fig.7.]

Data category and sampling area 

Year 

1991 1996 1999 2000 2001

Survey dates

Transect downstream from FCS August 15 -

September 10

July 25 -

August 25

August 16 August 16 August 14

Transect upstream from FCS do do August 17 August 17 August 15

Cove site upstream from FCS -- do August 18 -- August 16

Number of plots sampled

Transect downstream from FCS 26 5 10 10 9

Transect upstream from FCS 315999

Cove site upstream from FCS 05303

Average above-ground biomass per unit area (kg/m2)

Transect downstream from FCS 2.4 3.5 1.3 ± 0.17 1.5 ± 0.20 1.0 ± 0.14

Transect upstream from FCS 2.1 2.5 1.1 ± 0.14 1.3 ± 0.22 1.1 ± 0.14

Cove site upstream from FCS -- 1.0 0.9 ± 0.23 -- 1.0 ± 0.28

Average below-ground biomass per unit area (kg/m2)

Transect downstream from FCS 3.3 3.6 2.8 ± 0.79 1.6 ± 0.19 1.3 ± 0.19

Transect upstream from FCS 2.7 2.5 1.9 ± 0.31 1.7 ± 0.23 1.5 ± 0.19

Cove site upstream from FCS -- 1.0 1.7 ± 0.56 -- 1.4 ± 0.31

Average total biomass per unit area (kg/m2)

Transect downstream from FCS 5.7 7.1 4.1 ± 0.81 3.1± 0.78 2.3 ± 0.66

Transect upstream from FCS 4.8 5.0 3.0 ± 0.39 3.0 ± 0.92 2.6 ± 0.65

Cove site upstream from FCS -- 2.0 2.6 ± 0.52 -- 2.4 ± 1.2

Average above-ground biomass per shoot (g) 

Transect downstream from FCS 77.4 90.5 34.9 35.0 28.3

Transect upstream from FCS 79.3 72.0 33.0 37.8 35.8

Cove site upstream from FCS -- 32.3 25.7 -- 40.0

Average below-ground biomass per shoot (g) 

Transect downstream from FCS 108 94.5 75.3 37.4 36.8

Transect upstream from FCS 102 71.4 57.2 49.4 48.8

Cove site upstream from FCS -- 33.7 48.6 -- 56.0
Effects of flow modification 29

others, 1959). Average above-ground biomass in the Ellison 

Park wetland ranged from 900 to 3,500 g/m2. Below-ground 

biomass of T. glauca reported in other studies ranges from 

1,167 to 3,100 g/m2 (Van der Valk and Davis, 1978; Bray, 

1960; Andrews and Pratt, 1978); for other Typha species, 

below-ground biomass ranges from 556 to 5,053 g/m2

(Whigham and others, 1978; McNaughton, 1966). Average 

below-ground biomass in the Ellison Park wetland ranged from 

1,000 to 3,600 g/m2. Total biomass, which is infrequently 

reported in other studies, ranges from 972 to 4,720 g/m2

(McNaughton, 1966; Andrews and Pratt, 1978; Bray, 1960). 

The average total biomass in the Ellison Park wetland ranged 

from 2,000 to 7,100 g/m2 (table 9). On the basis of these 

comparisons, the Ellison Park wetland would be considered a 

highly productive ecosystem.

Above-ground biomass generally represented from 42 to 

50 percent of the total biomass, except during the dry 1999 

season, when it represented only 32 to 37 percent. The biomass 

values were highest during 1991 and 1996 and either exceeded 

or were close to the high end of the range of values reported in 

the literature (table 9); values measured from 1999 through 

2001 generally were within the range of reported values. 

Biomass was greatest along the northern transect during 1996 

and smallest in the cove area during that year. Biomass was 

greater during the precontrol period than during the postcontrol 

period; this decline was observed upstream and downstream 

from the FCS, but not in the cove area (fig. 10). Within-season 

variations in cattail biomass have been documented in many of 

the studies referenced above. Annual variation in wetland-plant 

biomass also has been reported; for example, 47- and 81-

percent changes in the total biomasses of two Typha spp. 

communities in Oregon were found over a 1-yr period 

(McNaughton, 1966), and a 37-percent change in above-ground 

biomass of Phragmites australis was found in Norfolk, 

England, over a 3-yr period (Boar, 1996). Average total 

biomass in the Ellison Park wetland varied among survey 

years—by 68 percent along the northern transect, 48 percent 

along the southern transect, and 23 percent along the cove 

transect. These local fluctuations could result from the 

differences in sampling dates, water depth, or other 

environmental factors.

Table 9. Above- and below-ground Typha spp. biomass reported in the literature, and average values for

each of three permanent transects sampled during four surveys in Ellison Park wetland, Monroe County, N.Y., 

1991-2001 

Biomass component

Value

(grams per square meter) Location Reference

Above-ground biomass 

Typha glauca 758 to 2,106 Iowa Van der Valk and Davis, 1978

1,361 New York Bernard and Fitz, 1979

2,320 Minnesota Andrews and Pratt, 1978

Other Typha species 378 to 1,336 central USA McNaughton, 1966

626 to 2,338 Mid-Atlantic coast, USA Whigham and others, 1978

428 to 2,252 southeastern USA Boyd and Hess, 1970

530 to 1,132 South Carolina Boyd, 1971

684 South Carolina Boyd, 1970

1,118 England Mason and Bryant, 1975

1,070 England Bray and others, 1959

1,731 New England Bray and others, 1959

1,400 Minnesota Bray, 1960

Ellison Park Typha glauca 900 to 3,500 New York This study

Below-ground biomass

Typha glauca 1,167 to 1,450 Iowa Van der Valk and Davis, 1978

2,960 Minnesota Bray, 1960

2,400 to 3,100 Minnesota Andrews and Pratt, 1978

Other Typha species 556 to 2,646 central USA McNaughton, 1966

1,800 to 5,053 Mid-Atlantic coast, USA Whigham and others, 1978

Ellison Park Typha glauca 1000 to 3,600 New York This study

Total biomass

Typha spp. 1,289 to 3,405 Oregon McNaughton, 1966 

972 to 3,982 central USA McNaughton, 1966 

4,230 to 4,720 Minnesota Andrews and Pratt, 1978;

Bray, 1960

Ellison Park Typha glauca 2,000 to 7,100 New York This study
30 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

1990 2002 1990 1992 1994 1996 1998 2000

0

4

1

2

3

1990 2002 1990 1992 1994 1996 1998 2000

0

4

1

2

3

BIOMASS, IN KILOGRAMS PER SQUARE METER

EXPLANATION

Downstream from flow-control structure

Upstream from flow-control structure

Cove site

1990 2002 1990 1992 1994 1996 1998 2000

0

8

2

4

6

A. ABOVE-GROUND BIOMASS

B. BELOW-GROUND BIOMASS

C. TOTAL BIOMASS

Growth rates of emergent macrophytes are related, in large 

part, to water depth (Coops and others, 1996). Cattail 

productivity in the Ellison Park wetland, as measured by cattail 

height, density, and biomass, also could be explained in terms 

of general soil wetness (Bernard and Seischab, 1991; 1997). 

Results of the precontrol (1991 and 1996) surveys indicate that 

cattail productivity was low in dry areas along streambanks and 

in deep-water areas where water depth exceeded about 1.3 ft 

(40 cm), and high in areas where the water level was about 

1.0 ft (30 cm) above the top of the cattail mat. Data from the 

three consecutive surveys from 1999 through 2001 indicate an 

alternative interpretation, however. Cattail productivity during 

2001, when water levels were intermediate between the dry 

conditions of 1999 and the wet conditions of 2000, was lower 

than in either of the previous 2 years; no direct correlation 

between water level and productivity within a given year was 

discernible. The degree of wetness during a given year can 

directly affect the growth of rhizomes during that year, 

however, and might indirectly affect productivity during the 

subsequent year (Noll and others, 2002). Thus, the dry 

conditions of 1999 might have been conducive to 

rhizome growth and the late-summer sprouting of 

numerous shoots, which grew during the 2000 growing 

season, when wet conditions might have retarded 

rhizome growth, thus reducing the number of shoots that 

would be available for growth during the 2001 growing 

season. This interpretation of the multi-year relation 

between water level and cattail productivity suggests 

that wetland studies of 1-yr, and possibly 2-yr, duration 

might be too short, and lack the necessary data for 

accurate identification and explanation of this relation.

Chemical Composition of Cattail Tissues

Results of the chemical analyses of cattail tissue 

are presented in table 10; the values represent mean 

concentrations of major nutrients and trace elements in 

above- and below-ground tissues collected along the 

three permanent transects in the wetland during 1991, 

1996, 1999, and 2001. Most constituents, except 

nitrogen, were found at higher concentrations in belowground tissues than in above-ground tissues; nitrogen 

concentrations were 2 to 3 times greater in the aboveground tissues. Concentrations of barium, manganese, 

phosphorus, and potassium were inconsistent from year 

to year. Aluminum concentrations in below-ground 

tissue decreased through the study period.

Several constituents showed large differences from 

year to year, especially iron. The high iron 

concentrations in below-ground tissues in 1991 and 

1996 probably reflected the increased availability of 

iron, which was released from the sediments under 

inundated (reduced) conditions in the rhizosphere, and 

the bioconcentration of iron when close to the thin, 

oxidized zone surrounding individual cattail rhizomes. 

The sharp (4-orders-of-magnitude) decrease in iron 

concentrations from the 1996 to 1999 samples probably 

reflects the low-water oxidized conditions that would 

have prevailed in the wetland in 1999; iron would have 

remained in insoluble forms in the sediment and would 

not have been available for bioconcentration on the 

cattail rhizomes. Intermediate water levels in 2001 

apparently created less extreme oxidation-reduction 

conditions and resulted in intermediate iron 

concentrations in below-ground tissues. Manganese 

Figure 10. Cattail (Typha glauca) biomass measured during five surveys along 

three permanent transects in Ellison Park wetland, Monroe County, N.Y., 

1991-2001
Effects of flow modification 31

Table 10. Mean concentrations of selected elements in above- and below-ground cattail tissues collected along three 

permanent transects in Ellison Park wetland, Monroe County, N.Y., 1991-2001 

[Values are in milligrams per kilogram. FCS, flow-control structure; --, not measured; <, less than; bdl, below analytical detection

limit. Dates of sample collection are the same as those for biomass samples listed in table 8. No tissue samples were collected from 

the Cove site in 1991. Locations of transects are shown in fig. 7.]

Element

Downstream from FCS Upstream from FCS Cove site

1991 1996 1999 2001 1991 1996 1999 2001 1996 1999 2001

Number of 

samples 26 5 10 9 31 5 9 9 5 3 3

A. Above-ground Cattail Tissues

Aluminum 15.9 5.0 126 56.0 19.6 7.1 86.7 34.9 7.7 86.7 16.2

Barium -- -- 9.30 14.2 -- -- 7.28 8.63 -- 16.4 14.9

Boron 7.3 6.1 -- 63.5 15.4 7.9 -- 247 5.4 -- bdl

Cadmium bdl bdl bdl bdl bdl 0.16 bdl bdl bdl bdl bdl

Calcium 10,000 11,000 12,300 14,700 10,000 6,000 13,700 16,600 10,000 12,800 10,200

Chromium 1.2 0.06 0.97 3.76 a 1.5 0.18 0.76 bdl 0.06 0.73 2.72

Cobalt < 0.1 0.06 0.33 5.35 0.2 0.08 0.42 4.12 a 0.06 bdl 6.11

Copper 4.5 4.1 7.33 9.58 5.3 3.8 10.7 8.00 a 3.3 5.63 20.4

Iron 58.8 56.9 50.0 35.0 72.3 56.4 46.7 31.6 62.8 40.0 12.6

Lead 0.5 -- 0.7a

 a Estimated value; three or more concentrations were below the analytical detection limit.

6.36 0.7 -- 0.7 a bdl -- 0.6 a 6.46

Magnesium 2,000 2,000 2,180 2,510 1,800 1,600 1,800 1,740 2,000 2,770 1,720

Manganese 478 284 289 247 607 188 244 360 368 378 245

Molybdenum < 0.1 0.3 0.40 7.01 0.6 0.8 0.66 4.94 a -- 0.26 23.7

Nickel 1.6 0.8 0.83 4.66 a 1.6 1.3 0.69 bdl 1.2 0.58 11.6

Nitrogen 22,500 22,000 17,400 11,900 26,000 24,000 19,300 12,400 23,000 13,400 11,200

Phosphorus 2,000 2,700 2,890 9,670 2,400 2,900 2,830 8,120 2,500 2,300 7,680

Potassium 11,000 11,000 12,600 1,420 14,000 23,000 6,580 1,050 13,000 3,090 824

Sodium 4,000 -- 5,180 15,500 3,490 -- 4,760 11,300 -- 3,520 7,860

Tin -- -- 248 11.8 -- -- 224 14.4 -- 240 9.71

Zinc 17.5 13.2 15.8 18.9 15.0 13.9 13.0 16.9 10.4 9.13 12.1

B. Below-ground Cattail Tissues

Aluminum 1,720 1,200 351 94.8 2,030 1,200 351 138 1,320 290 38.7

Barium -- -- 8.28 9.10 -- -- 13.2 9.83 -- 11.7 14.3

Boron 53.3 43.2 -- -- 212 49.8 -- 244 43.5 -- 91.6

Cadmium 0.2 bdl 0.61 bdl bdl bdl 0.54 0.78 a 0.1 bdl bdl

Calcium 13,000 15,000 8,470 10,700 14,000 13,000 8,790 9,370 12,000 7,960 8,080

Chromium 6.9 3.6 1.12 3.86 a 5.2 6.8 1.12 bdl 4.2 0.89 bdl

Cobalt 0.2 0.5 0.52 6.59 1.0 0.7 0.33 5.78 a -- bdl 6.75

Copper 15.5 12.7 13.5 16.4 15.7 15.2 8.70 7.39 11.6 18.2 13.7

Iron 20,300 13,300 1.39 207 15,700 12,400 3.35 1,050 13,900 0.13 65.2

Lead 14.7 -- 1.82 9.16 a 23.1 -- 1.2 a 16.1 a -- 1.87 bdl

Magnesium 5,000 5,000 3,320 4,510 5,000 6,000 6,430 3,870 7,000 4,770 3,440

Manganese 407 420 86.5 100 508 397 80.9 159 582 104 151

Molybdenum 0.2 0.5 -- 15.9 1.0 0.5 -- 5.48a 0.2 -- 20.8

Nickel 10.9 10.9 0.88 5.93 a 8.4 12.6 0.80 bdl 13.1 0.78 13.6

Nitrogen 7,600 7,000 7,500 4,700 9,000 9,000 6,000 5,500 8,000 4,400 6,100

Phosphorus 2,000 2,000 3,160 8,840 2,200 2,000 3,240 10,100 3,000 2,440 8,470

Potassium 5,000 6,000 8,360 1,100 2,000 7,000 7,120 926 9,000 5,880 993

Sodium 10,700 -- 10,700 1,680 9,330 -- 11,200 13,600 -- 7,480 9,170

Tin -- -- 281 3.23 -- -- 311 8.84 -- 269 4.10

Zinc 46.1 49.7 22.2 93.8 73.6 47.3 16.8 50.9 42.5 12.6 42.0
32 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

concentrations, although lower than iron concentrations, 

followed a similar pattern through the surveyed years. 

Above-ground tissues assimilated an average of 1.12 to 

2.6 percent of their dry weight in nitrogen; below-ground 

tissues assimilated an average of 0.44 to 0.9 percent. Aboveand below-ground tissues assimilated similar quantities of 

phosphorus—an average of 0.2 to 1.0 percent of their dry 

weight. Generally, upward or downward trends in the 

concentration data of given constituents from the southern 

transect were also found in the data from the northern transect. 

Additional sampling would be needed to detect or confirm any 

real change in the chemical composition of cattail tissues since 

the installation of the FCS, however.

Fish

Four characteristics of the fish community upstream and 

downstream from the FCS were evaluated to discern 

differences between the two areas; these were (1) species 

richness and distribution within the wetland; (2) use of the 

wetland by nonresident species and their distribution; (3) 

growth rates of selected species; and (4) trophic status of the 

fish community. Additionally, contaminant-related stress 

indicators in brown bullhead were identified. Neither 

electrofishing in 1996 nor video monitoring of fish passage 

through the FCS in 2001 yielded any results that contributed to 

the evaluation of the effects of the FCS on the fish community.

Species Richness and Distribution

The Ellison Park wetland is used by a diverse fish 

community; 44 species were found during the 1991-2001 study 

period (table 11). Species richness did not change greatly 

during the study period (Crabtree and Ringler, 2002); each of 

four surveys identified from 25 to 29 species. The dominant 

species usually included bluegill and pumpkinseed but also 

could have included several other species, including alewife, 

black crappie, brook stickleback, gizzard shad, brown bullhead, 

and yellow perch. Dominance varied from year to year, or, in 

some years, from season to season and by wetland area. The 

dominant species often greatly outnumbered the combined 

number of all other individuals caught and resulted in an uneven 

fish community, especially during four periods—May 1992, 

when 1,066 of 1,833 individuals (58 percent) were alewife; 

October 1999, when 5,119 of 6,922 individuals (74 percent) 

were pumpkinseed; July 2001, when 4,070 of 6,446 individuals 

(63 percent) were yellow perch; and July 1999, when 23,100 of 

24,600 individuals (94 percent) were gizzard shad (table 12). 

Fourteen common species were found during at least one season 

of each survey year—black crappie, bluegill, bowfin, brown 

bullhead, common carp, gizzard shad, golden shiner, johnny 

darter, largemouth bass, northern pike, pumpkinseed, spottail 

shiner, white perch, and white sucker. Eight rare species were 

identified by the capture of only one individual of a species 

during the entire study period; these were American eel (1996), 

central stoneroller (1991), coho salmon (1996), fathead minnow 

(1991), freshwater drum (2001), green sunfish (1992), mimic 

shiner (1992), and rosyface shiner (1992). Brook stickleback 

and logperch were found during only 1 year’s survey—520 

brook stickleback were captured in May 1996, and 5 logperch 

were captured during 2001. 

The wetland was used for spawning and nursery habitat, as 

evidenced by the presence or capture of young-of-the-year 

(individuals less than 1 year old) and juvenile individuals, adult 

fishes in breeding colors or with nuptial tubercles, and adult 

fishes with easily released milt or eggs (Crabtree and Ringler, 

2002). In 1991-92, at least 16 species that were confirmed to use 

the wetland for spawning and(or) rearing habitat, and 6 species 

that probably did, were identified (Miller and Ringler, 1992). In 

2001, 15 species were found as either young-of-the-year, 

juveniles, or spawning adults. Of these, 10 species were found 

upstream and downstream from the FCS. The presence of 

migratory salmonids (brown trout and rainbow trout) indicated 

that the wetland provides a pathway between Irondequoit Bay 

and Lake Ontario and spawning areas in Irondequoit Creek 

upstream of the wetland.

Comparison of the fish-catch size and species richness 

upstream from the FCS with the same characteristics 

downstream was possible for 11 of the 12 survey seasons (3 

seasons per survey year); extremely low water levels in the 

upstream wetland area during September 2001 prevented fall 

sampling that year. Generally, larger catches per unit effort 

were made in the area downstream from the FCS than in the 

area upstream (table 12), especially during those seasons when 

the catch size was affected by a large influx of fish from 

Irondequoit Bay (May 1992, July 1999, and July 2001). 

Upstream catches exceeded downstream catches during July 

and September 1991, September 1996, and June 1999. Fewer 

species were found upstream from the FCS than downstream 

during 5 of the 11 surveyed seasons; more species, or an equal 

number of species, were found upstream than downstream 

during the other six seasons (table 12). The FCS did not appear 

to affect the relations between upstream and downstream fishcatch size or species richness.

Resident- and Nonresident-Fish

Of the 44 species found in the wetland, 25 were considered 

resident species—those that are likely to spend all of their lives 

within a marsh setting (table 11) according to the classification 

of Jude and Pappas (1992). The remaining 19 species were 

considered nonresident species—those that migrate into a 

marsh environment for only part of their life cycle. The resident 

species outnumbered the nonresident species in all survey 

years. The number of resident species found in the wetland 

ranged from 15 to 17, whereas the number of nonresident 

species ranged from 14 in 1991 to 10 in 1999 (table 11). The 

number of nonresident species found in the area upstream from 

the FCS during the precontrol period ranged from 0 to 4 less 

than the number found in the area downstream from the FCS 
Effects of flow modification 33

Table 11. Fish species found in Ellison Park wetland, Monroe County, N.Y., 1991-2001 

Species Wetland

resident

species

July and 

September

1991

May

Common name Scientific name 1992 1996 1999 2001

Longnose gar Lepisosteus osseus yes XXX

Bowfin Amia calva yes X XXXX

American eel Anguilla rostrata no X

Alewife Alosa pseudoharengus no X XX X

Gizzard shad Dorosoma cepedianum yes X XXXX

Goldfish Carassius auratus yes XXX

Common carp Cyprinus carpio no X XXXX

Central stoneroller Campostoma anomalum yes X

Spotfin shiner Cyprinella spiloptera yes XX X

Common shiner Luxilus cornutus yes X X

Emerald shiner Notropis atherinoides no X XX

Spottail shiner Notropis hudsonius no X XXXX

Rosyface shiner Notropis rubellus yes X

Mimic shiner Notropis volucellus no X

Bluntnose minnow Pimephales notatus yes X X XX

Fathead minnow Pimephales promelas yes X

Golden shiner Notemigonus crysoleucas yes X XXXX

Rudd Scardinius erythrophthalmus yes X XX X

Quillback Carpiodes cyprinus yes X X

White sucker Catostomus commersoni no X XXXX

Shorthead redhorse Moxostoma macrolepidotum no X X

Brown bullhead Ameiurus nebulosus yes X XXXX

Channel catfish Ictalurus punctatus no X X

Northern pike Esox lucius no X XXXX

Central mudminnow Umbra limi yes X XXX

Coho salmon Oncorhynchus kisutch no X

Rainbow trout (steelhead) Oncorhynchus mykiss no XX X

Brown trout Salmo trutta no X X

Trout perch Percopsis omiscomaycus no X X

Brook silverside Labidesthes sicculus no X X

Banded killifish Fundulus diaphanus yes X XX

Brook stickleback Culaea inconstans yes X

White perch Morone americana yes X XXXX

Rock bass Ambloplites rupestris yes X XXX

Green sunfish Lepomis cyanellus yes X

Pumpkinseed Lepomis gibbosus yes X XXXX

Bluegill Lepomis macrochirus yes X XXXX

Smallmouth bass Micropterus dolomieu no X X

Largemouth bass Micropterus salmoides yes X XXXX

Black crappie Pomoxis nigromaculatus no X XXXX

Johnny darter Etheostoma nigrum no X XXXX

Yellow perch Perca flavescens no X XXX

Logperch Percina caprodes yes X

Freshwater drum Aplodinotus grunniens yes X

Total number of species 29 29 26 25 29

Number of resident species / number of nonresident species 15 / 14 17 / 12 15 / 11 15 / 10 17 / 12
34 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Table 12. Results of fish surveys in Ellison Park wetland, Monroe County, N.Y., 1991-2001

[FCS, flow-control structure.]

Fish-community 

characteristic

1991-92 1996

May 1992a

 a May 1992 data are listed before 1991 data to maintain seasonal uniformity among the surveys. 

July 1991 Sept. 1991 May July September

A. PRECONTROL PERIOD 

Number of individuals caught

Area downstream from FCS 1,439 344 217 954 398 94

Area upstream from FCS 394 417 253 115 138 120

Number of species identified: resident / nonresident

Area downstream from FCS 14 / 9 11 / 9 8 / 10 12 / 6 7 / 7 5 / 6

Area upstream from FCS 14 / 9 11 / 9 7 / 7 6 / 2 6 / 5 6 / 6

Most abundant species

Area downstream from FCS Alewife Bluegill Pumpkinseed Brook stickleback Alewife Pumpkinseed

Area upstream from FCS Alewife Black crappie Bluegill Brook stickleback Pumpkinseed Bluegill

Other abundant species (more than 20 individuals caught), in order of abundance

Area downstream from FCS White perch

Gizzard shad

Pumpkinseed

Brown bullhead

Spottail shiner

Brown bullhead

Alewife

Common carp

Pumpkinseed

Gizzard shad

Gizzard shad

Brook silverside

Pumpkinseed

Bluegill

Black crappie

Pumpkinseed

Bluegill

Area upstream from FCS Pumpkinseed

Gizzard shad

White perch

Bluegill

Brown bullhead

Gizzard shad

Bluegill

Largemouth bass

Bluntnose 

minnow

Pumpkinseed

1999 2001

June July October May July September

B. POSTCONTROL PERIOD

Number of individuals caught

Area downstream from FCS 106 23,692 4,348 347 6,209 3,119

Area upstream from FCS 265 965 2,574 217 237 ---b

 b Extremely low water levels prevented use of trap nets. 

Number of species identified: resident / nonresident

Area downstream from FCS 7 / 5 10 / 7 11 / 6 9 / 7 13 / 8 13 / 6

Area upstream from FCS 8 / 5 9 / 4 10 / 8 10 / 8 5 / 6 ---

Most abundant species

Area downstream from FCS Pumpkinseed Gizzard shad Pumpkinseed Brown bullhead Yellow perch Pumpkinseed

Area upstream from FCS Pumpkinseed Gizzard shad Pumpkinseed Pumpkinseed Pumpkinseed ---

Other abundant species (more than 20 individuals caught), in order of abundance

Area downstream from FCS Pumkinseed

White sucker

Emerald shiner

Common carp

Yellow perch

Largemouth bass

Spottail shiner

Bluegill

Golden shiner

Bluegill

Goldfish

Largemouth bass

Yellow perch

Common carp

White sucker

Golden shiner

Pumpkinseed

Emerald shiner

Brown bullhead

Pumpkinseed

Golden shiner

White sucker

Bluntnose 

minnow

Common carp

Bluegill

Yellow perch

Bluegill

Golden shiner

Gizzard shad

Area upstream from FCS Emerald shiner

Bluegill

Gizzard shad

Golden shiner

Pumpkinseed

Common carp

Largemouth bass

Bluegill

White sucker

Bluegill

Gizzard shad

Common carp

Yellow perch

Largemouth bass

Emerald shiner

Brown bullhead

Bluegill

Golden shiner

Yellow perch
Effects of flow modification 35

(table 12). This relation was found during 3 of the 5 sampling 

seasons during the postcontrol period, but the number of 

nonresident species in the upstream area exceeded that in the 

downstream area during October 1999 and May 2001. 

The wetland functioned as two different systems—the 

southern wetland area (upstream from the FCS) was dominated 

by resident species, whereas the northern area contained a 

mixture of residents and nonresidents, with a pulse of 

nonresidents (either spawners or young-of-the-year) during one 

season of each surveyed year (Crabtree and Ringler, 2002). The 

dominant nonresident species was alewife in May 1992 and 

July 1996; and yellow perch in 1999 and 2001 (the number of 

yellow perch caught in 2001 was 10 times that in 1999). These 

differences, as well as the fluctuations in total catch among 

years and seasons, probably reflected natural or background 

population dynamics throughout the Irondequoit Bay system, 

and possibly the greater southern Lake Ontario near-shore 

system, rather than the effects of the FCS (Crabtree and Ringler, 

2002).

Growth Rates of Selected Species

Growth rates of five species—black crappie, bluegill, 

largemouth bass, northern pike, and pumpkinseed—were 

calculated for each of the survey years (table 13). The most 

rapid growth for all species occurred during the first or second 

year. All of these species showed large year-to-year variations 

in growth rate, especially northern pike, bluegill, and 

pumpkinseed. Largemouth bass showed the most consistency in 

growth rates among the surveyed years. Statistically significant 

differences in growth rates were detected between certain years 

for some species (Thompson and others, 2000), but these 

differences could not be attributed to the FCS (Crabtree and 

Ringler, 2002). The fluctuations in growth rates probably result 

Table 13. Growth rates of selected fish species in Ellison Park wetland, Monroe County, N.Y., 1991-2001

[Values are in grams per year. DS (US), downstream (upstream) from flow-control structure.]

Species

Year

collected

Age, in years

0-1 1-2 2-3 3-4 4-5 5-6 6-7 7-8 8-9

Black crappie 1991 1.96 0.81 0.57 0.42 0.39 0.37 0.14 0.17

1996 2.52 1.05 .57 .27

1999 - DS 1.28 1.43 .60

1999 - US 1.46

2001 2.43 .86 .71 .33 - .12 - .11 .52 .52 .23

Bluegill 1991 .68 1.56 .76 .44 .33 .29

1996 .69 1.32 .72 .57 .39 .30 .17 .14 .10

1999 - DS 3.83 1.71 .88 .46 .38 .61 .24

1999 - US - .78 .03 .016 .007 .006 .004 .002

2001 - DS 1.43 1.36 .95 .34 .15 .07 .72

2001 - US 2.77 .65 .50 .24 .17 - .01

Largemouth bass 1991 1.79 1.04 .69 .45 .35 .43 .19 .30

1996 1.45 1.16

1999 - DS 1.21 .90

1999 - US 1.29 1.17 .62 .44

2001 Growth rates not computed

Northern pike 1991 4.06 2.60 1.22 .87

1996 4.33 1.87 1.23 .50

1999 - DS 8.61 .043 .043 .027 .008 .007 .013

1999 - US 7.10 .008 .004 .003 .002

2001 Growth rates not computed

Pumpkinseed 1991 .98 1.14 .57 .94 .52 .30

1996 .88 1.14 .84 .56 .48 .37 .39

1999 - DS .32 1.61 .88 .36 .30 .18 .28 .15

1999 - US .38 1.86 .80 .44 .43 .19 .16

2001 - DS 2.77 1.07 .66 .18 .30

2001 - US 1.72 .97 .60 .38 .50
36 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

from natural changes in annual temperature and prey or 

predator densities. 

Trophic Status of Fish Community

The trophic status of the Ellison Park wetland community 

changed seasonally and from year to year. Insectivores 

dominated the wetland during most seasons and represented 

88 percent of the fish found in the northern wetland area during 

May 1996 (table 14). Invertivores dominated the northern 

wetland area during May 1992 and July 1996 because of an 

unusually large influx of nonresident alewife. Omnivores, 

which usually represented less than 30 percent of the fish 

community, dominated the northern and southern wetland areas 

during July 1999, when gizzard shad were found in extremely 

large numbers. Piscivores, such as walleye and trout, were 

relatively rare (usually less than 15 percent of the fish 

community), probably because these are cool-water fish that do 

not tolerate the warm water temperatures of the wetland for 

prolonged periods (Thompson and others, 2000), but also 

because piscivores typically represent a small percentage of a 

fish assemblage (D.B. Chambers, U.S. Geological Survey, 

written commun., 2002). This trophic class dominated the 

northern wetland area during July 2001 with the large influx of 

yellow perch; these fish were young-of-the-year, however, and 

less piscivorous than adult yellow perch (D.B. Chambers, U.S. 

Geological Survey, written commun., 2002). Use of generalized 

trophic classes such as used in these fish studies could 

oversimplify, and possibly misrepresent, the Ellison Park 

wetland community (D.L. Crabtree, State University of N.Y., 

College of Environmental Science and Forestry, written 

commun., 2002). 

Apparent shifts in the trophic status of the wetland were 

strongly related to the presence of a dominant fish species or a 

combination of nondominant species that were found in fairly 

equal numbers in the wetland. The fish community in the 

southern wetland area (upstream from the FCS) during May 

1992 included many species in fairly equal numbers; the 

trophic-group percentages clustered between 14 and 34 percent 

(table 14). Conversely, the fish community in the northern 

wetland area during the same season was dominated by a single 

species (alewife), and the resulting trophic status of that 

wetland area reflected the trophic class of alewife, that is, 

invertivores. Similarly, brook stickleback greatly affected the 

trophic status of the wetland during May 1996, as did gizzard 

shad during July 1999.

Apparent changes in trophic composition of the fish 

community either from year to year, seasonally, or within a 

given area of the wetland, were probably related to the effect of 

a dominant species, the availability of prey, or regional trends 

in fish-population sizes. The data from the four fish surveys did 

not identify any shift in the trophic status of the fish community 

in the southern wetland area that was not seen in the northern 

wetland area or that could be attributed directly to the FCS, or 

indirectly to changes in sediment- and nutrient-accumulation 

rates. 

Contaminant-Related Stress Indicators in 

Brown Bullhead

Contaminant-related stress indicators in brown bullheads 

were noted during the July 1991, May 1992, and 1999 sampling 

periods. Stress indicators in bottom-feeding fish result from 

sediment contamination by organic compounds and metals 

(Hickey and others, 1990; Hirethota and Ringler, 1992). Of the 

164 brown bullheads captured during the 1991-92 survey, 128 

were examined for melanoma, tumors, lesions, and barbel 

deformities (shortened, bent, burnt, bumpy, or missing). Of 

these, about 75 percent showed contaminant-related stress 

indicators (table 15). All 38 individuals captured during the 

1999 survey showed melanotic pigmentation, lesions, oral 

Table 14. Trophic status of fish community in Ellison Park

wetland, Monroe County, N.Y., 1991-2001

[Values are percentage of total catch in each trophic group. DS (US),

downstream (upstream) from flow-control structure. --, no data.]

Season 

Sampling

area

Trophic Group

InsectivoreInvertivoreOmnivorePiscivore

July 1991 DS 34 36 17 13

US 26 36 29 9

September 1991 DS 58 0 23 19

US 74 2 18 6

May 1992 DS 11 74 7 8

US 34 27 25 14

May 1996 DS 88 5 5 2

US 82 3 12 3

July 1996 DS 29 64 4 3

US 72 5 9 14

September 1996 DS 79 2 13 6

US 77 2 10 11

June 1999 DS 75 1 12 11

US 63 6 24 6

July 1999 DS 4 0 94 2

US 26 1 67 7

October 1999 DS 86 0 7 7

US 84 0 8 7

May 2001 DS 78 9 6 7

US 61 20 9 10

July 2001 DS 30 1 3 66

US 57 5 24 14

September 2001 DS 56 20 4 20

US -- -- -- --
Effects of flow modification 37

ulcers, bent barbels and other indicators of stress (table 15). No 

note of contaminant-related stress indicators was made on 

brown bullheads that were captured during the 1996 and 2001 

surveys. 

The most common indicator noted in 1991 and 1992 was 

burnt barbel, which was found on 44 to 42 percent of the 128 

fish examined, respectively. This indicator was not recorded 

during the 1999 survey, but other barbel deformities were found 

on at least 26 percent of the fish captured that year. During 

1999, the most prevalent indicator was melanoma (skin 

pigmentation), which was noted on 22 (58 percent) of the 

captured fish. The incidence of contaminant-related stress 

indicators appeared to be related to fish size and location in the 

wetland; large fish in the northern part of the wetland (close to 

Irondequoit Bay) showed more signs of stress than small fish in 

that area and more signs than fish in the southern part of the 

wetland. Some compounds that could be the cause of these 

indicators in brown bullhead, such as polycyclic aromatic 

hydrocarbons, have been found in high concentrations in the 

wetland sediments. (See “Sediment Quality” section, further 

on.) The actual source and extent of the contaminant-related 

stress indicators in brown bullheads are obscured by the 

proximity of Irondequoit Bay, into and from which brown 

bullhead can freely migrate; however, the FCS probably had no 

direct relation to the prevalence of these stress indicators in the 

wetland (Thompson and others, 2000). 

Turtles

The capture of fish often resulted in the capture of turtles 

as well, mostly in the trapnets (table 16). The only two species 

found during all sampling seasons were painted and snapping 

turtles; two spiny softshell turtles were caught in 1999. As many 

as 208 painted turtles (1999) and 105 snapping turtles (1996) 

were caught in a year. Generally, more turtles of both species 

were caught in the southern wetland area (upstream from the 

FCS) than in the northern area. During 1991, 115 turtles were 

individually marked with notches filed in their scutes. During 

May 1992, 12 of the 91 marked painted turtles from the 

southern wetland area were recaptured; this allowed the 

painted-turtle population of the southern wetland area to be 

estimated at 478 ± 227. No turtles were recaptured in the 

northern wetland area. The spiny softshell turtle (Trionyx 

spiniferus) is listed as a New York State species of concern and 

was not known to inhabit the Ellison Park wetland (Thompson 

and others, 2000). 

Benthic Macroinvertebrates

No benthic macroinvertebrate survey was conducted after 

installation of the FCS; therefore, the effect of the FCS on the 

benthic-macroinvertebrate community could not be assessed. 

Nonetheless, the macroinvertebrate data are presented to 

characterize, in combination with other survey results, the 

precontrol ecological condition of the wetland.

During 1991-92, the benthic macroinvertebrate 

community of the wetland was characterized through an 

analysis of the stomach contents of 99 individual fish of 9 

species (Miller and Ringler, 1992). Thirteen taxa of 

macroinvertebrates were identified, generally to the lowest 

taxonomic level possible. The most abundant taxa was 

Cladocera (crustaceans—Daphnia, Bosmina), which 

represented 87.4 percent of all macroinvertebrates identified. 

The predominant insect family was Chironomidae (midges), 

which represented 82.8 percent of all insects, and 9.6 percent of 

all macroinvertebrates identified by this method. Other insect 

taxa included Trichoptera (caddisflies), Plecoptera (stoneflies), 

Zygoptera (damselflies), Veliidae (water striders), 

Table 15. Indicators of contaminant-related stress found in

brown bullheads in Ellison Park wetland, Monroe Country,

N.Y., 1991-2001

[Stress indicators not recorded during 1996 or 2001 surveys.]

Indicator

Frequency of occurrence 

(percentage of fish examined)

July

1991

May

1992 1999

Number of fish examined 23 105 38

Bent barbel 21.7 6.7 26.3

Bumpy barbel 0.0 10.5 5.3

Burnt barbel 43.5 41.9 --

Missing barbel 8.7 7.6 7.9

Shortened barbel 4.3 1.0 10.5

Lesion or ulcer 26.0 25.7 31.6

Melanoma 26.0 14.3 57.9

Tumor or raised-skin growth 13.0 7.6 21.0

No symptoms found 8.7 29.5 0.0

Table 16. Turtle species found in Ellison Park wetland, Monroe County, N.Y., 1991-2001

[US (DS), upstream (downstream) from flow-control structure.]

Species July

1991

May

1992

1996 1999 2001

Common name Scientific name US DS US DS US DS

Painted turtle Chrysemys picta 129 76 187 6 129 79 81 32

Snapping turtle Chelydra serpentina 46 16 90 15 43 52 52 6

Spiny softshell turtle Trionyx spiniferus 00 00 20 00
38 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Hydrachnidia (water mites), and Circulionidae (weevils). Other 

taxa identified were Copepoda (copepods), Amphipoda (scuds), 

Oligochaeta (segmented worms), and Astracidae (crayfish). 

This sampling method provided only general information on the 

macroinvertebrate community; the results cannot be considered 

comprehensive. The actual macroinvertebrate composition 

would be more diverse and also would change seasonally, 

depending on the availability of particular macroinvertebrates 

as food sources and on the selectivity of feeding fish (Miller and 

Ringler, 1992). 

During September and October 1996, macroinvertebrate 

samples were collected from three habitats—channel bottom, 

vegetation, and pelagic zone—at one creek-channel site and one 

wetland (backwater channel) site in each of the southern and 

northern parts of the wetland (Haynes and McNamara, 1998). 

Only 254 individuals from 12 species were found on the 

multiplate samplers in the pelagic (open-water) habitats; 

therefore, the data from this habitat are not included in the 

following analyses. 

Amphipods dominated the macroinvertebrate community 

at all four locations sampled and represented 74 percent of all 

macroinvertebrates collected (table 17). The only other 

abundant macroinvertebrate group found in the wetland was 

Oligochaeta, which represented 16 percent of all 

macroinvertebrates collected. Other groups found included 

nonamphipod Crustacea (2.3 percent), Chironomidae 

(2.6 percent), nonchironomid Insecta (3.0 percent), and 

Gastropoda (mollusks, 1.7 percent). Physical characteristics 

measured at the two creek sites and two wetland sites to assess 

comparability of the macroinvertebrate data (table 18) indicated 

that some factors, such as water temperature, specific 

conductance, dissolved-oxygen concentration, pH, percent 

canopy cover, and flow velocity, were similar, whereas others, 

such as water depth, amount of vegetation, and substrate 

composition, differed. In general, no major differences were 

noted; thus, comparison of the creek and wetland habitats in the 

southern part of the wetland with those in the northern part was 

considered valid. 

Community Indicators of Water Quality

The benthic-macroinvertebrate taxa found in the Ellison 

Park wetland during the 1996 survey were typical of slowmoving, organically rich, oxygen-poor aquatic habitats with 

low taxonomic diversity and large populations of species that 

Table 17. Benthic macroinvertebrates found in Ellison Park wetland, Monroe County, N.Y., September-October 1996

[Data from Haynes and McNamara (1998). P., Phylum; C., Class; O., Order; F., Family.]

Taxon

Creek sites 

(sediment and vegetation)

Wetland sites 

(sediment and vegetation) Openwater

sites

Number

in

taxon

Percentage

of total

in taxon

Southern

area

Northern

area

Southern

area

Northern

area

P. Platyhelminthes 14 3 7 0 4 28 0.07

P. Nematomorpha 12 12 0 6 .01

P. Annelida

C. Oligochaeta 740 2,720 2,228 1,243 6,931 16.06

Other Annelida 0 0 10 2 12 .03

P. Arthropoda

C. Crustacea

O. Amphipoda 10,256 5,239 10,839 5,369 160 31,863 73.85

Other Crustacea 118 153 527 143 53 994 2.30

C. Insecta 

O. Diptera 

F. Chironomidae 201 376 296 237 1 1,111 2.58

Other Insecta 366 205 584 124 8 1,287 2.98

C. Arachnida 27 23 3 3 0 56 .13

P. Mollusca

C. Gastropoda 258 198 133 113 28 730 1.69

Other Mollusca 2 31 88 10 0 131 .30

Subtotal, by area 11,983 8,950 14,716 7,246 254 43,149

Percentage of total, by area 27.77 20.74 34.11 16.79 0.59 100.00
Effects of flow modification 39

1 Simpson s diversity index combines species richness and community balance (evenness) to quantify ecological health of a habitat.

0 = highly diverse, 1.0 = nondiverse (single-species dominance)

2 PMA, a measure of community-composition similarity to an unaffected community. Values > 55 indicate little adverse effect.

3 DOM-3, a measure of community balance evenness of distribution of individuals among the species. DOM-3 is the combined percent 

contribution of the three most numerous species. High values (approaching 100) indicate unbalanced communities.

4 HBI, a combined measure of species abundance and tolerance. Scale is 1 to 10; low values indicate clean-water conditions.

5

 Qualitative terms, such as severe, moderate, slight, and none, can be used to describe the level of ecological impact that is quantitatively 

defined by the biotic indices.

Table 18. Physical characteristics of sampling sites and benthic-macroinvertebrate measures of ecological health in southern 

and northern areas of Ellison Park wetland, Monroe County, N.Y., September-October 1996

[Data from Haynes and McNamara (1998). μS/cm at 25  ̊C, microsiemens per centimeter at 25 degrees Celsius.]

Site characteristics and 

measures of ecological health

Creek sites Wetland (backwater-channel) sites

Southern area Northern area Southern area Northern area

Physical Characteristics of Sites

Dominant substrate material Sand Mud Mud Mud

Water depth (feet) 1.80 3.08 1.94 2.00

Channel width (feet) 60 60 26 46

Water temperature (degrees Celsius) 14.9 10.5 16.3 15.7

Specific conductance (μS/cm at 25  ̊C) 600 555 540 600

Dissolved oxygen (milligrams per liter) 6.9 12.6 6.1 5.0

pH 8.1 7.8 7.7 8.0

Percentage of area covered by submerged vegetation 80 60 80 0

Flow velocity (feet per second) 0.95 0.59 0.10 0.20

Percent canopy cover 20 0 0

Measures of Ecological Health

Substrate habitat 

Number of organisms collected 879 2,856 1,732 1,851

Species richness (no. of species or taxa in a sample) 

number of species 18 19 17 20

species with more than 1 individual 12 14 13 15

species with more than 4 individuals 9 10 9 9

Simpson s diversity index 1 0.566 0.399 0.418 0.629

Percent model affinity (PMA) 2 46 42 39 44

Dominance-3 model (DOM-3) 3 92 97 96 91

Hilsenhoff biotic index (HBI) 4 8.13 9.41 8.59 7.21

Qualitative impact assessment 5 moderate moderate to severe moderate to severe moderate

Vegetation habitat

Number of organisms collected 11,104 6,094 12,984 5,395

Species richness

number of species 50 38 50 43

species with more than 1 individual 36 27 37 25

species with more than 4 individuals 17 15 27 13

Simpson s diversity index 1 0.414 0.431 0.530 0.451
40 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

can survive under these conditions (Haynes and McNamara, 

1998). Oligochaetes (segmented worms, mainly Limnodrilus 

hoffmeisteri) were most abundant in the muddy sediment 

samples, whereas crustaceans (mainly amphipods, Gammarus 

spp.) dominated submerged-vegetation habitats. Several 

measures of ecological health, such as species richness, and 

biotic indices such as Simpson’s diversity index, percent model 

affinity, dominance, and Hilsenhoff’s biotic index (Bode and 

others, 1996), were calculated (table 18). Species richness is the 

total number of species or taxa found in a community; high 

values (greater than about 20) are mostly associated with cleanwater conditions. Species richness in the channel-bottom 

(substrate) habitats ranged from 17 to 20, but these values were 

halved when only those species represented by more than four 

individuals were counted. Species richness was greater in the 

submerged-vegetation habitats than the substrate habitats; 

values ranged from 38 to 50 but decreased to 13 to 27 if only 

species represented by more than four individuals were 

counted.

A species-diversity index, such as Simpson’s index, 

combines species richness and community balance (evenness) 

to quantify the ecological quality of a habitat. Simpson’s 

diversity index, which is the probability that two individuals, 

chosen at random and independently from the population, will 

be found to belong to the same group, can range from 0 for a 

highly diverse and balanced community to 1.0 for a nondiverse 

community that is dominated by one or only a few species 

(Washington, 1984). Simpson’s diversity index ranged from 

0.399 to 0.629 in the substrate habitats and from 0.414 to 0.530 

in the vegetation habitats. Percent model affinity (PMA) is a 

measure of similarity to a nonimpaired model community, as 

described by the percent abundance of seven major taxonomic 

groups; values greater than 55 percent indicate similarity and 

good ecological quality. Wetland substrate PMA values ranged 

from 39 to 46. Dominance is a measure of community balance, 

or evenness of the distribution of individuals among the species; 

high values that approach 100 indicate unbalanced communities 

that are strongly dominated by one or only a few species. 

Dominance-3 model (DOM-3) is the combined percent 

contribution of the three most numerous species; DOM-3 

values ranged from 91 to 97 in the substrate habitats. The 

Hilsenhoff biotic index (HBI) is a combined measure of species 

abundance and tolerance to pollution; on a scale of 0 to 10, low 

values are indicative of clean-water conditions. Wetland 

substrate HBI values ranged from 7.21 to 9.41. These biotic 

index values reflect the macroinvertebrate-community’s 

response to many environmental factors, including habitat and 

sediment quality, as well as water quality, and indicate that the 

ecological health of the Ellison Park wetland is moderately to 

severely impaired (table 18; Haynes and McNamara, 1998). 

Benthic-macroinvertebrate community differences 

between the southern and northern wetland areas were 

summarized by Haynes and McNamara (1998) as follows:

1. The sand substrate of the southern creek site contained fewer 

species and organisms per sample than the mud substrate of 

the northern creek site, but the southern creek site had better 

balance among species and, thus, better ecological quality 

than the northern creek site as indicated by biotic-index 

values. 

2. Conversely, the vegetation-habitat samples from the southern 

creek site showed greater species richness and higher 

organism counts than those from the northern creek site, but 

the evenness component of the diversity index indicated 

slightly better ecological quality in the northern-creek 

vegetation habitat. 

3. Of the two wetland backwater sites, which both had mud 

substrates, the southern site had lower species richness and 

lower numbers of organisms than the northern site, and much 

lower biotic-index values than the northern site. 

4. Of the two backwater vegetation-habitat sites, the southern 

site was far more densely vegetated than the northern site, 

and had greater species richness, higher organism counts, 

and a greater diversity value. 

5. Differences between the macroinvertebrate communities in 

the southern creek and wetland sites and those in the 

northern sites were probably due to habitat differences, such 

as substrate composition, water depth, and density of 

submerged aquatic vegetation.

Species Indicators of Water Quality

Among the taxa of benthic macroinvertebrates that could 

be used as indicators of current water-quality conditions and of 

future water-quality changes in the Ellison Park wetland are 

Gammarus spp. (scuds or sideswimmers) and chironomids 

(Haynes and McNamara, 1998). Pollution-tolerance criteria for 

macroinvertebrate taxa range from 0 for highly intolerant of 

pollution to 10 for highly tolerant (Bode and others, 1996). A 

high ratio of Gammarus pseudolimnaeus, which has a tolerance 

value of 4, to G. fasciatus, which has a tolerance value of 6, is 

indicative of good water quality. Both species are abundant in 

the Ellison Park wetland; G. pseudolimnaeus (27,040 

individuals) represented 85 percent of the genus, and G. 

fasciatus (4,823 individuals) represented the other 15 percent. 

The fairly high ratio of 5.67 for these two species indicates good 

water quality in the wetland, but an increase in numbers of G. 

fasciatus might indicate deterioration in water quality.

Another indicator of environmental pollution is deformed 

mouthparts in chironomids (Warwick, 1990). Among 

chironomid taxa, Procladius is considered to be more tolerant 

of pollution than Chironomus (Warwick, 1989). Of 51 

Procladius individuals collected in the wetland, only one was 

deformed, and of 392 Chironomus individuals, 30 (7.6 percent) 

had mouthpart deformities. Other studies have concluded that 

deformity rates of less than 14 percent for chironomids 

(Warwick, 1990), and less than 12 percent for Chironomus

(Lenat and Barbour, 1994) are typical in organically enriched 

sediments characterized by low dissolved-oxygen 
Effects of flow modification 41

concentration and high biological oxygen demand, such as in 

the Ellison Park wetland. A deformity rate greater than 

20 percent would indicate exposure to toxic chemical pollution 

(Lenat and Barbour, 1994). Less than 4 percent of the 

chironomids collected, and 7.6 percent of Chironomus, had 

mouthpart deformities; therefore, the observed incidence of 

deformities is attributed to exposure to organic enrichment of 

the sediment rather than to toxic chemicals (Haynes and 

McNamara, 1998).

Fish-Stomach-Contents Data

Benthic macroinvertebrates found in the stomach contents 

of 40 fish—20 bluegill (Lepomis macrochirus) and 20 

pumpkinseed (L. gibbosus)—that were captured by 

electrofishing during September and October 1996, were 

compared with those identified in the stomach contents of 7 

bluegill and 8 pumpkinseed captured during July or September 

1991 or May 1992. The diets of these species differed sharply 

between the two sampling years—1991-92 stomach contents 

were dominated by chironomids (midges), and even within that 

survey period, bluegill preferred fish eggs over any 

macroinvertebrate during May 1992; whereas 1996 stomach 

contents were dominated by amphipods (scuds). In addition to 

the discrepancies in sample sizes, fish-capture methods, and 

sampling seasons, these differences were probably a result of 

natural variability in the dominant food species for these fish. 

The large percentage of amphipods found in the 1996 stomach 

contents—92 percent of all macroinvertebrates identified 

—reflected the abundance of amphipods—74 percent of all 

macroinvertebrates identified—throughout the wetland. 

Amphipods made up less than 3 percent of the total diet of 

bluegill and pumpkinseed, and less than 1 percent of the 

stomach contents of all fish analyzed in 1991-92, however. In 

contrast, the preferred food source of bluegill and pumpkinseed 

in 1991-92 (chironomids) represented less than 10 percent of 

the macroinvertebrates identified in stomach contents of all fish 

examined during that survey, and did not reflect the wetlandwide dominance of Cladocera, which was the preferred food 

source of black crappie. Large numbers of fish eggs consumed 

during May 1992 simply reflected the abundance of that food 

source during spawning season.

Birds

Three surveys were conducted in the wetland (1991-92, 

1996, and 2002) to identify the presence and absence of bird 

species during the pre- and postcontrol periods. An earlier 

(1980-85) statewide survey by the New York State Department 

of Environmental Conservation and the Federation of New 

York State Bird Clubs to identify bird species and map the 

extent of their breeding ranges in New York provided an initial 

baseline listing of birds that were likely to be found in or around 

the wetland during the early 1980’s (Andrle and Carroll, 1988). 

The data collected during that study were compiled by survey 

blocks, each of which covered about 10.5 mi2 in the Ellison 

Park area. The wetland, which occupies less than 1 mi2, lies 

within the boundaries of two of these blocks; therefore, the 

1980-85 data refer to bird species identified over a two-block 

(21 mi2) area that contains a second major wetland, the 

Thousand Acre Swamp, about 3.5 mi east of the Ellison Park 

wetland. The 1980-85 survey reported 95 species of birds and 

classified them as possible, probable, or confirmed breeders. 

Those species that are known to breed in or near a wetland 

environment, and those that were identified in the 1991-92, 

1996, or 2002 surveys, are listed in table 19. Of the 25 wetland 

species identified in the 1980-85 survey, most were probably 

sighted in the Ellison Park wetland, although a few species, 

such as the great blue heron, would be considered visitors to the 

Ellison Park wetland and breeders in the Thousand Acre 

Swamp (McKinney, 1996). 

The 1991-92 survey (Miller and Ringler, 1992) identified 

30 species of birds by sight or call but did not note the breeding 

status of these species (table 19). Of these species, 23 were 

identified during the 1980-85 statewide breeding survey. Three 

species sighted in 1991-92 (pied-billed grebe, herring gull, and 

osprey) were not identified during any of the other bird surveys.

The 1996 survey (McKinney, 1996) identified 60 species 

by sight or call (table 19); twice as many as during the 1991-92 

survey. Several common species were notably absent (such as 

sora, American bittern, and least bittern), but the high water 

levels that year might have hindered breeding activities by these 

species (McKinney, 1996). Two wetland species that had been 

found in the past—black tern and blue-winged teal—have not 

been seen in the area for many years. Two other 

species—American black duck and sharp-shinned hawk—were 

not identified during any of the other surveys.

The 2002 survey (Spahn, 2002) was similar in scope and 

detail to that of McKinney (1996) and reported 91 species 

(table 19). Thirty-five of these species, of which 19 are 

potential breeding species, were not detected during the 1996 

survey, and 20 were not identified during any other survey. The 

newly reported species included mute swan (the only confirmed 

breeder), yellow-headed blackbird (locally very rare), and many 

warblers. Of the wetland ground-nesting species, no evidence 

of American bittern, American coot, common moorhen, or 

pied-billed grebe—which had been reported during at least one 

of the earlier surveys—was found; least bittern and sora were 

reported, however. Evidence of cerulean warbler was expected, 

but this species’ absence is general to the area and not 

connected to any changes in the wetland (Spahn, 2002).

The composition of the breeding-bird population during 

1996 (precontrol period) differed from that in 2002 (postcontrol 

period) in that 15 additional breeders were identified in 2002; 

all breeding species reported in 1996 were also identified in 

2002. Several ground-nesting species—ring-necked pheasant, 

common snipe, blue-winged teal, and American 

woodcock—were not reported during any of the study-period 

surveys, but only during the 1980-85 survey, which covered an 
42 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Table 19. Bird species observed in or near Ellison Park wetland, Monroe County, N.Y., 1980-2002, and designation of breeding 

status in wetland (possible, probable, or confirmed) 

[Based on available data and criteria in Andrle and Carroll (1988). PO, possible breeder; PR, probable breeder; CO, confirmed breeder. 

 X, bird sighted, breeding status unknown. Z, bird sighted but not considered an Ellison Park wetland breeder.]

Common name Scientific name

Year of survey Groundnesting

species e 1980-85 a 1991-92 b 1996 c 2002 d

Grebe, pied-billed Podilymbus podiceps X Yes

Heron, black-crowned night Nycticorax nycticorax Z

 great blue Ardea herodias PO X Z Z

 green-backed (green) Butorides striatus PR X Z PO

Bittern, American Botaurus lentiginosus PO X Yes

Bittern, least Ixobrychus exilis PR PO Yes

Swan, mute Cygnus olor CO Yes

Goose, Canada Branta canadensis X PO CO Yes

Mallard Anas platyrhynchos CO X CO CO Yes

Teal, blue-winged Anas discors CO Yes

Duck, American black Anas rubripes Z

Duck, wood Aix sponsa CO CO CO

Rail, Virginia Rallus limicola CO PR PO Yes

Sora Porzano carolina PR PO Yes

Moorhen, common Gallinula chloropus CO CO Yes

Coot, American Fulica americana X Z Yes

Killdeer Charadrius vociferus CO X CO Yes

Sandpiper, solitary Tringa solitaria Z

 spotted Actitis macularia PR X PO PO Yes

Snipe, common Gallinago gallinago CO Yes

Woodcock, American Scolopax minor CO Yes

Gull, ring-billed Larus delawarensis X Z

 herring Larus argentatus X

Tern, caspian Sterna caspia Z

Vulture, turkey Cathartes aura Z Z

Hawk, red-tailed Buteo jamaicensis PO X PO CO

 sharp-shinned Accipiter striatus Z

Osprey Pandion haliaetus X

Pheasant, ring-necked Phasianus colchicus CO Yes

Turkey, wild Meleagris gallopavo PO

Dove, mourning Zenaida macroura CO PR CO

 rock Columba livia CO CO PO

Owl, great-horned Bubo virginianus CO PO

 eastern screech Otus asio CO

Swift, chimney Chaetura pelagica PO Z

Hummingbird, ruby-throated Archilochus colubris PR PR PO

Kingfisher, belted Ceryle alcyon CO X PR CO

Flicker, northern Colaptes auratus CO PR PO

Woodpecker, downy Picoides pubescens CO CO CO

 hairy Picoides villosus CO CO CO

 pileated Dryocopus pileatus CO PO

 red-bellied Melanerpes carolinus CO X CO PR

Kingbird, eastern Tyrannus tyrannus CO X CO PO

Flycatcher, alder Empidonax alnorum PO

 great crested Myiarchus crinitus CO PR CO

 least Empidonax minimus PO PO

 willow Empidonax traillii CO PO PR

Pewee, eastern wood- Contopus virens CO PR CO

Phoebe, eastern Sayornis phoebe CO CO CO

Swallow, bank Riparia riparia CO CO PO

 barn Hirundo rustica CO X CO PO

 cliff Hirundo pyrrhonota Z

a

 Includes species identified during 1990-2002 surveys, as well as species identified during New York State Department of Environmental 

Conservation and Federation of New York State Bird Clubs Breeding Bird Atlas Project (Andrle and Carroll, 1988) that are likely to use a 

wetland environment for breeding purposes (Robert Spahn, Genesee Ornithological Association, written commun., 1995). 

b

 Miller and Ringler (1992) c McKinney (1996) d Spahn (2002) e

 Wetland species that nest on, near, or above water surface or on the ground.
Effects of flow modification 43

Swallow, northern rough-winged Stelgidopteryx serripennis CO PO

 tree Tachycineta bicolor CO X PR PO

Jay, blue Cyanocitta cristata CO X PO CO

Crow, American Corvus brachyrhynchos CO CO PR

Titmouse, tufted Parus bicolor CO CO CO

Chickadee, black-capped Parus atricapillus CO X CO CO

Nuthatch, white-breasted Sitta carolinensis CO PR PO

Wren, Carolina Thryothorus ludovicianus CO

 house Troglodytes aedon CO CO PR

 marsh Cistothorus palustrus CO X CO CO

Gnatcatcher, blue-gray Polioptila caerulea CO CO

Bluebird, eastern Sialia sialis PR PO

Thrush, wood Hylocichla mustelina CO PR PR

Veery Catharus fuscescens CO PR PR Yes

Robin, American Turdus migratorius CO X CO CO

Catbird, gray Dumetella carolinensis CO X PR PR

Mockingbird, northern Mimus polyglottos X PR

Waxwing, cedar Bombycilla cedrorum CO X PO PO

Starling, European Sturnus vulgaris CO PO CO

Vireo, red-eyed Vireo olivaceus CO PO PR

 warbling Vireo gilvus CO PR PR

 yellow-throated Vireo flavifrons PR PR PR

Warbler, bay-breasted Dendroica castanea Z

 blackburnian Dendroica fusca Z

 black-throated blue Dendroica caerulescens Z

 black-throated green Dendroica virens Z

 blue-winged Vermivora pinus PR PO

 Cape May Dendroica tigrina Z

 cerulean Dendroica cerulea CO

 chestnut-sided Dendroica pensylvanica Z

 magnolia Dendroica magnolia Z

 Nashville Vermivora ruficapilla Z

 Tennessee Vermivora peregrina Z

 yellow Dendroica petechia CO PR CO

 yellow-rumped Dendroica coronata Z

Parula, northern Parula americana PO Z

Yellowthroat, common Geothlypis trichas CO PR PR

Redstart, American Setophaga ruticilla CO PR

Grosbeak, rose-breasted Pheucticus ludovicianus CO PR CO

Cardinal, northern Cardinalis cardinalis CO X CO CO

Bunting, indigo Passerina cyanea PR PO

Sparrow, chipping Spizella passerina CO CO

 house Passer domesticus CO CO CO

 song Melospiza melodia CO X CO CO Yes

 swamp Melospiza georgiana CO PR PR Yes

 white-crowned Zonotrichia leucophrys Z

 white-throated Zonotrichia albicollis Z

Blackbird, red-winged Agelaius phoeniceus CO X CO CO Yes

 yellow-headed Xanthocephalus xanthocephalus Z

Cowbird, brown-headed Molothrus ater CO CO PR

Grackle, common Quiscalus quiscula CO CO CO

Oriole, northern (Baltimore) Icterus galbula CO X CO CO

Tanager, scarlet Piranga olivacea CO PR PR

Goldfinch, American Carduelis tristis CO X PO PR

Finch, house Carpodacus mexicanus CO CO PR

Total number of species 77 30 60 91 20

Table 19. Bird species observed in or near Ellison Park wetland, Monroe County, N.Y., 1980-2002, and designation of breeding 

status in wetland (possible, probable, or confirmed)—continued

[Based on available data and criteria in Andrle and Carroll (1988). PO, possible breeder; PR, probable breeder; CO, confirmed breeder. 

 X, bird sighted, breeding status unknown. Z, bird sighted but not considered an Ellison Park wetland breeder.]

Common name Scientific name

Year of survey Groundnesting

species e 1980-85 a 1991-92 b 1996 c 2002 d
44 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

area much larger than just the Ellison Park wetland. The 

absence or scarcity of certain marsh species that were present 

before or during the early 1980’s, such as American bittern, 

pied-billed grebe, and blue-winged teal, appeared to be regional 

(Robert Spahn, Genesee Ornithological Association, written 

commun., 2002), and not associated with the FCS. 

Sediment Characteristics

The composition of channel-bottom material within the 

Ellison Park wetland varies according to flow velocity within a 

given area. The predominant bed material in the main channel 

(Irondequoit Creek), in which flow is rapid, was sand; from 82 

to 98 percent of the bed material sampled was sand sized or 

larger (Coon, 1997). The predominant bed materials in the 

open, backwater areas, where velocities are slow, were silt and 

clay; from 60 to 96 percent of the bed material was within these 

particle-size classes (table 20; Coon, 1997). The sediments in 

these backwater areas contained only small amounts of organic 

carbon (2 to 7 percent by weight), probably reflecting the 

relatively slow accumulation of biomass from small 

submergent aquatic macrophytes (such as Myriophyllum sp.

and Ceratophyllum demersum). Substrate samples from the 

cattail-covered areas contained much more organic matter (7 to 

79 percent by weight), as a result of the high biomassaccumulation rates in these areas (Hunter and others, 2000; Noll 

and others, 2002).

The FCS was expected to increase the accumulation of 

sediment and the concentrations of associated metals and 

organic compounds upstream from the FCS as a result of 

increased dispersal and detention of stormflows. Sedimentation 

rates were monitored, and chemical analyses of sediment 

samples were performed, to assess the effects of the FCS on 

sediment deposited in the wetland areas upstream and 

downstream of the FCS. 

Sedimentation Rates

Plastic sedimentation disks were anchored to the cattail 

mat in the wetland at randomly selected locations (fig. 8). 

Sediment thickness above each disk was measured at least 

annually, and sedimentation rates were calculated. Sediment 

was found to be deposited and removed from the disks in a 

pattern that would be expected for a system in a state of 

dynamic equilibrium—episodic deposition of a large quantity 

of sediment in one year, followed in subsequent years by partial 

resuspension and removal. The sediment included 

decomposing organic matter, primarily cattail shoots and 

duckweed; therefore, measurements were adjusted to reflect the 

inorganic fraction. Errors probably resulted from this 

adjustment, but the same method was used upstream and 

downstream from the FCS throughout the study to ensure 

comparability of data. Therefore, the calculated differences 

between pre- and postcontrol sedimentation rates are assumed 

to be valid. 

Net sediment accumulation was measured at all sites; 11 of 

the 12 sites (except SR13; fig. 8) showed a decrease in sediment 

thickness for at least one annual measurement (table 21). The 

results from the six long-term measurement sites (SR1 through 

SR6; fig. 8) indicated that the mean precontrol sedimentation 

rate for three sites upstream from the FCS was 3.0 mm/yr, and 

the mean for three sites downstream from the FCS was 2.7 mm/

yr. These rates agreed closely with (1) the historic rate of 3 mm/

yr computed by Young (1992) from radiocarbon dating of 

organic matter extracted from two sediment cores from the 

outlet of Irondequoit Bay and one from the vicinity of 

Irondequoit Creek near Browncroft Boulevard; (2) the glacial 

isostatic-rebound rate of 1 ft per century (3 mm/yr) that has 

been estimated for the south-central shore of Lake Ontario 

(Clark and Persoage, 1970; Larsen, 1985); and (3) the nearrecent rate of 2.8 to 3.7 mm/yr calculated by Schroeder (1985) 

from lead-210 and cesium-137 radioisotopic measurements of 

four bottom-sediment cores taken from Irondequoit Bay in 

1980. Mean postcontrol sedimentation rates, based on data from 

the six long-term measurement sites, increased in both wetland 

areas—to 7.3 mm/yr upstream from the FCS, and 4.6 mm/yr 

below it (table 21). 

Results from the six short-term measurement sites (SR7 

through SR10, SR12, and SR13; fig. 8) indicated an 

inconsistent pattern of postcontrol sediment deposition in the 

wetland; sedimentation rates at these sites ranged from 3.0 to 

66.8 mm/yr. The highest postcontrol sedimentation rate 

(66.8 mm/yr), which was recorded at SR13, probably reflected 

(1) this site’s location just downstream from where the Millrace 

channel enters the wetland proper, and the sharp decrease in 

streamflow velocity as the channelized flows move into this 

low-gradient, cattail-covered area; and (2) increased sediment 

loads as a result of the 1997-98 channel modifications that 

increased the diversion of bankfull stormflows from the creek 

to the Millrace and caused erosion of the Millrace as the channel 

adjusted to the increased flows (as identified by comparison of 

1992 and 2003 cross-sectional data). No site-to-site trend in 

sedimentation rates was indicated downstream from SR13; 

although the rate at the next site downstream (SR12) was 

8.0 mm/yr, the next highest sedimentation rate (27.0 mm/yr) 

was recorded at site SR10 (fig. 8), and rates in the intervening 

area ranged from 3.4 to 10.0 mm/yr (table 21). The high 

sedimentation rate at SR10 could reflect the site’s proximity to 

the FCS, which was about 1,000 ft downstream, but the 

sedimentation rate at SR3, which was between SR10 and the 

FCS, was 10.6 mm/yr, less than at SR10. 

The sedimentation-rate data indicate that the FCS strongly 

affected sediment deposition in the wetland, as expected, but 

the annual fluctuations in sediment thickness at any given site 

were puzzling. The sedimentation rates at upstream and 

downstream sites are plotted in figure 11 with peak flows in 

Irondequoit Creek, before and after the installation of the FCS. 

Note that (1) data from site SR13 (upstream from the FCS) are 

excluded to make the scale of the two plots and the detail of the 

data comparable; and (2) four of the nine peak flows that were 

recorded during the postcontrol period exceeded all peak flows 
Table 20. Size distribution of sediment samples collected for chemical analyses from Ellison Park wetland, Monroe County, N.Y., 1994-2001, 

particles finer than index-particle size[SQ, sediment-quality site. US (DS), upstream (downstream) from flow-control structure. Locations are shown in fig. 8.] 

Sampling

site and

wetland

area

Date of 

sample 

collection

Index particle size, in millimeters Perc

2 1 0.5 0.25 0.125 0.062 0.031 0.016 0.008 0.004 0.002 0.001 falls in 

Percentage of particles finer than particle size given above San

SQ1 - US 10/1994 100 99.99 99.96 99.84 99.28 92.39 89.04 62.58 37.11 21.62 11.55 7.6 

10/1997 100 99.8 89.0 69.39 43.59 29.80 19.57 11.56 6.23 11.0 

11/1999 100 98.4 94.5 65.97 43.57 29.96 19.00 9.92 6.99 5.5 

10/2001 100 99.8 99.2 94.9 78.8 55.8 38.7 35.1 27.0 5.1 

SQ2 - US 10/1994 100 99.65 99.16 97.35 88.69 76.46 54.03 34.23 24.76 16.46 11.3 

10/1997 100 98.7 93.6 82.4 74.18 57.70 43.38 29.50 18.22 12.15 17.6 

11/1999 100 98.2 94.8 87.0 81.9 59.8 30.3 21.1 13.3 7.3 7.1 18.1 

10/2001 100 98.9 95.5 82.6 66.0 41.5 27.0 23.6 4.5 

SQ3 - DS 10/1994 100 99.93 99.77 99.38 97.87 82.35 35.47 32.41 19.93 14.80 10.91 17.6 

10/1997 100 99.4 79.2 57.39 36.02 26.52 18.21 11.87 8.31 20.8 

11/1999 100 93.50 89.3 59.5 36.50 24.97 15.64 8.98 4.58 3.03 40.5 

10/2001 100 95.2 67.7 35.2 22.4 17.1 15.1 12.0 32.3 

Sampling

site and 

wetland

area

Annual thickness of accumulated sediment, in millimeters

Precontrol period 

(1991-96)

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002

Net

sediment

thickness

(mm)

Period

of data

collection

(years)

Net 

sedimentation 

rate

(mm/yr)

Ne

sedim

thick

nes

(mm

Long-term sites

SR1 - US 0 2 16 9 6 6 30 48 33 37 43 47 6 5 1.2 17 SR2 - US 0 4 15 -- 6 8 -- -- -- 20 40 -- 8 5 1.6 32a 

SR3 - US 0 17 16 12 15 31 50 63 66 86 91 103 31 5 6.2 53 SR4 - DS 0 20 11 15 -- 11 79 99 97 106 91 99 11 5 2.2 20 SR5 - DS 0 30 -- 9 -- -- 21 35 30 39 9 3 3.0 30a SR6 - DS 0 9 -- -- -- 38 49 33 24 28 9a

 a Sediment thickness estimated.

3 3.0 19a 

Short-term sites

SR7 - US 0 2 4 38 36 44 44 -- 40 SR8 - DS 0 26 29 23 30 36 36 SR9 - DS 0 7 -- 4 27 15 15 

SR10 - US 0 18 14 13 98 135 135 SR12 - US 0 0 2.5 28 27 40 40 SR13 - US 0 105 140 256 278 334 334 

Table 21. Sediment thickness and net sedimentation rates in Ellison Park wetland, Monroe County, N.Y., 1991-2001[SR, sedimentation-rate measurement site. mm/yr, millimeter per year; --, no data; US (DS), upstream (downstream) from flow-control structure. Locations are 
46 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

4

1990 1992 1994 1996 1998 2000 2002

0

140

20

40

60

80

100

120

THICKNESS OF ACCUMULATED SEDIMENT, IN MILLIMETERS

0

3,500

500

1,000

1,500

2,000

2,500

3,000

DISCHARGE, IN CUBIC FEET PER SECOND

Peak discharge at Blossom 

Road (wetland inflow)

SR1

SR2

SR3

SR7

SR10

SR12

1990 1992 1994 1996 1998 2000 2002

0

140

20

40

60

80

100

120

THICKNESS OF ACCUMULATED SEDIMENT, IN MILLIMETERS

0

3,500

500

1,000

1,500

2,000

2,500

3,000

DISCHARGE, IN CUBIC FEET PER SECOND

Peak discharge at Blossom 

Road (wetland inflow)

SR4

SR5

SR6

SR8

SR9

Flow-control structure

installed during

late February 1997

Flow-control structure

installed during

late February 1997

A. UPSTREAM FROM FLOW-CONTROL STRUCTURE

B. DOWNSTREAM FROM FLOW-CONTROL STRUCTURE

Figure 11. Thickness of accumulated sediment at sedimentation-rate-measurement sites in Ellison Park wetland, Monroe County, 

N.Y., 1991-2002: A. Upstream from flow control structure. B. Downstream from flow-control structure.
Effects of flow modification 47

that occurred during the precontrol study period (1991-96) and 

the previous 10-year period (1981-90). Sediment thickness at 

all long-term sites (SR1 through 6) increased during 1996-98; 

the FCS was installed during 1997. The plotted data indicate 

that (1) most sites showed an increase in sediment accumulation 

that began in 1995 or 1996, before installation of the FCS; (2) 

the greatest increase occurred at SR4, downstream from the 

FCS; (3) many of the sites indicated a leveling off of sediment 

accumulation both upstream and downstream from the FCS 

after 1998; and (4) despite the increased frequency of high 

flows during the postcontrol period, no relation between the 

magnitude of peak flows and the annual change in sediment 

thickness was found. 

High peakflows did not necessarily result in greater 

sediment deposition in the wetland than low peak flows. Peak 

flows that coincided with seasonally high water levels in the 

wetland (fig. 4) might explain more of the variation in sediment 

accumulation than peak-flow magnitude alone. At most sites, 

sediment accumulation during high-water periods was greater 

than during low-water periods, regardless of the peak-flow 

magnitude, although even this relation was inconsistent. 

Therefore, although the FCS undoubtedly had an effect, the 

apparent increase in sedimentation rates after FCS installation 

could have been partly the result of other factors, such as the 

magnitude, frequency, and the timing (season) of peak flows.

Sediment Quality

Sediment samples were collected periodically from three 

sites in the wetland (fig. 8) and analyzed for major and trace 

elements (table 22A), polycyclic aromatic hydrocarbons 

(PAHs; table 22B), and organochlorine and organophosphate 

compounds (table 22C). The percentage of silt-plus-clay in the 

sediment can affect the concentrations of metals (Hem, 1985), 

and the percentage of organic matter can affect the 

concentrations of organic compounds (Smith and others, 1988). 

The percentage of fine-grained sediment in a sample, or the 

concentration of aluminum or titanium, can be used to 

normalize trace-element concentrations for identification of 

trends and adjustment for manmade effects, and to make data 

comparable among sites and among years (Horowitz, 1991); 

similarly, total organic carbon concentrations can be used to 

normalize concentrations of organic compounds (Daskalakis 

and O’Connor, 1995). Normalization of the Ellison Park 

wetland samples was unnecessary because (excluding the 1999 

sample from site SQ3, which was anomalous as explained 

below) (1) more than 68 percent of each sample was silt and 

clay size (table 20); (2) aluminum concentrations varied by 

only 0.7 percent among the 11 samples, and titanium 

concentrations by only 0.03 percent; and (3) total organic 

carbon concentrations varied by only 3.6 percent among the 11 

samples (table 22A).

The 1999 sample from site SQ3 was anomalous in that it 

contained many trace elements in unusually high concentrations 

and many major elements in unusually low concentrations in 

relation to the concentrations in samples from other sites and 

other years. This sample also had the smallest amount of silt and 

clay and the lowest organic-carbon concentration of all samples 

collected, but no other distinction was identified that could 

explain the discrepancies in element concentrations. The results 

from this sample are included in table 22A but were not used in 

the element analyses. 

The high concentrations of some major elements, 

including barium, manganese, strontium, and zinc, and some 

trace elements, including chromium, copper, lead, and 

vanadium (table 22A), were within the range expected for 

sediment derived from an urbanized basin (Coon and others, 

2000). The concentrations of most PAHs (coal-tar compounds) 

were below the analytical detection limit and are not included in 

table 22B. PAHs that consistently had measurable 

concentrations were benzo[a] anthracene, benzo[a] pyrene, 

benzo[b] and benzo[k] fluoranthene, benzo[g,h,i] perylene, 

bis(2-ethylhexyl) phthalate, chrysene, fluoranthene, 

phenanthrene, and pyrene (table 22B). Organochlorine 

compounds, including pesticides and polychlorinated 

biphenyls, were found at measurable levels, whereas 

organophosphate compounds (pesticides) generally were not 

(table 22C). The concentrations of many of the organochlorine 

compounds were comparable to or lower than those measured 

in 14 studies in the United States during the 1970’s and 1980’s, 

as reported by Smith and others (1988). Many of the organic 

compounds found in the sediment samples are (1) highly 

resistant to chemical or biological transformation and, 

therefore, extremely persistent in the environment; and (2) have 

low aqueous solubility and strongly partition from solution to 

biota, particulate material, and dissolved organic matter (Smith 

and others, 1988). Therefore, the presence and concentrations 

of these compounds were fairly typical of a depositional 

environment in a highly urbanized basin. 

Sediment-quality guidelines (SQGs) have been developed 

to assess sediment quality and to set targets for sediment quality 

that will sustain ecosystem health in aquatic environments 

(Canadian Council of Ministers of the Environment, 1995; 

Ingersoll and others, 2000). SQGs identify threshold 

concentrations of selected chemicals that, if exceeded in 

freshwater or marine sediments, are likely to cause adverse 

effects in aquatic life. Three freshwater SQGs for 7 metals, 11 

PAHs, and 9 organochlorine compounds are presented in 

table 22. The interim sediment-quality guideline (ISQG) is the 

concentration below which adverse effects to aquatic life are 

not expected to occur; the probable effect level (PEL) is the 

concentration above which adverse effects to aquatic life are 

expected to frequently occur (Canadian Council of Ministers of 

the Environment, 1995). The consensus-based probable effect 

concentration (PEC), which is based on agreement among 

various published SQGs (MacDonald and others, 2000), are 

alternative threshold concentrations above which adverse 

effects are expected to frequently occur (Ingersoll and others, 

2000). PEC values are generally greater than PEL values. 

Comparison of the measured concentrations of selected 

elements and compounds in the wetland sediment samples and 
Table 22. Chemical concentrations in sediment samples from Ellison Park wetland, Monroe County, N.Y., 1994-2001, and sediment-quality guidelin

[SQ, sediment-quality site. m, presence verified, but not quantified; e, estimated; <, less than. Site locations are shown in fig. 8.] Element Freshwater sediment quality guideline Southern wetland area (upstream from flow-control structure) North

(downstream fr

Site SQ1 Site SQ2 

ISQG1 PEL2 PEC3 1994 1997 1999 2001 1994 1997 1999 2001 1994 199

A. Major and trace elements, values represent concentrations in clay fraction (particle size < 63 microns), in micrograms per gram (parts per million) unless indicate

of sample by weight Aluminum, % 5 5 4.9 4.3 4.9 5 4.7 4.4 4.8 4

Antimony 0.4 0.4 0.4 0.5 0.2 0.4 0.4 0.7 0.6 0Arsenic 5.9 17.0 33.0 5 3.4 4.2 6 5 3.7 4.6 4.3 4 3Barium 460 430 410 410 450 430 400 400 470 450 Beryllium 1 <1 1.2 1.1 1 1 1.1 1.3 1 1 Bismuth <10 <10 <1 <1 <10 <10 <1 <1 <10 <10 Cadmium 0.6 3.5 4.98 0.8 0.6 0.8 0.8 0.9 0.7 0.8 0.8 3.8 2Calcium, % 6.5 6.5 6.4 7 6.7 6.5 8.1 8.9 4.7 4Cerium 44 49 46 41 45 48 46 44 39 44 Chromium 37.3 90.0 111 43 38 43 53 46 44 43 52 59 45 Cobalt 10 11 7 8 12 7 8 8 10 6 Copper 35.7 197 149 40 32 41 41 49 37 41 43 72 43 Europium <2 <2 1 <1 <2 <2 1 <1 <2 <2 Gallium 12 16 10 9 12 14 10 10 11 13 Gold <8 <8 <1 <1 <8 <8 <1 <1 <8 <8 Holmium <4 <4 <1 <1 <4 <4 <1 <1 <4 <4 Iron, % 2.7 2.3 2.5 2.5 2.8 2.7 2.6 2.5 2.4 2Lanthanum 25 25 24 19 25 27 24 21 23 24 Lead 35.0 91.3 128 45 42 49 52 62 48 49 50 89 53 Lithium 30 20 30 20 30 30 30 30 20 20 Magnesium, % 1.4 1.2 1.3 1.2 1.2 1 1.1 1.2 1.3 1Manganese 880 630 520 590 720 640 630 670 530 590 1 Interim freshwater sediment quality guideline; concentration below which adverse effects to aquatic life are not expected to occur (Canadian Council of Minis

Environment, 1999).

2 Probable effect level; concentration above which adverse effects to aquatic life are expected to frequently occur (Canadian Council of Ministers of the Environ3 Consensus-based probable effect concentration; concentration above which adverse effects to aquatic life are expected to frequently occur (Ingersoll and other
Mercury 0.17 0.486 1.06 0.1 0.12 0.14 0.16 0.14 0.1 0.13 0.16 0.33 Molybdenum <2 <2 0.5 0.6 <2 <2 0.5 0.6 <2 <Neodymium 26 25 22 22 24 25 22 22 23 2Nickel -- 36 48.6 18 15 20 18 20 19 21 19 20 1Niobium 12 9 8 8 11 10 7 8 11 Phosphorus, % 0.13 0.12 0.13 0.13 0.15 0.13 0.13 0.14 0.12 Potassium, % 1.7 1.7 1.6 1.4 1.6 1.6 1.6 1.5 1.6 Scandium 8788 8 888 7 Selenium 0.6 0.5 0.6 0.8 0.8 0.6 0.6 0.8 0.8 Silver 0.5 0.4 0.4 0.3 0.7 0.6 0.4 0.7 2 Sodium, % 1.2 1.2 1.1 0.84 1.1 0.95 0.99 0.79 1.3 Strontium 420 390 360 450 440 440 470 550 340 32Sulfur, % 0.34 0.63 0.64 0.81 0.6 0.54 0.49 0.68 0.35 Tantalum <40 <40 <1 <1 <40 <40 <1 <1 <40 <4Thallium <1 <1 <1 <1 Thorium 6555 6 555 6 Tin <5 <5 4 4 <5 <5 4 4 <5 <Titanium, % 0.26 0.25 0.26 0.24 0.25 0.27 0.25 0.24 0.27 Uranium 1.6 1.5 1.4 1.2 1.6 1.7 1.3 1.2 1.7 Vanadium 50 46 47 47 50 51 48 48 45 4Ytterbium 2222 2 222 2 Yttrium 20 22 22 20 19 22 21 19 19 2Zinc 123 315 459 220 170 200 240 230 190 200 250 250 20Total carbon, % 5.4 6.6 6.7 8.9 6.9 7.3 6.5 7.6 5.1 Inorganic 

carbon, % 1.9 1.8 1.7 2.1 1.9 1.8 2.1 2.5 1.4 

Organic 

carbon, % 3.5 4.8 5 6.9 5 5.5 4.4 5.2 3.7 

Table 22. Chemical concentrations in sediment samples from Ellison Park wetland, Monroe County, N.Y., 1994-2001, and sediment-quality guid

chemicals—continued

Element

Freshwater sediment

quality guideline

Southern wetland area 

(upstream from flow-control structure)

Nor

(downstream 

Site SQ1 Site SQ2 

ISQG1 PEL2 PEC3 1994 1997 1999 2001 1994 1997 1999 2001 1994 19
B. Polycyclic aromatic hydrocarbons, values represent concentrations in silt-clay fraction (particle size < 2 mm), in micrograms per kilogra1,2,5,6-Dibenz- anthracene <400 <400 m -- <400 <400 m -- <400 <40Acenaphthene 6.71 88.9 -- <200 <200 m e10 <200 -- m <150 <200 Acenaphthyl- ene 5.87 128 -- <200 <200 m e20 <200 -- m e30 <200 Anthracene 46.9 245 845 <200 -- m e100 <200 -- m e110 <200 Benzo[a] anthracene 31.7 385 1,050 500 -- e200 500 <400 -- e200 560 1,200 50Benzo[a]pyrene 31.9 782 1,450 600 -- e300 700 500 -- e300 750 1,300 60Benzo[b] fluoranthene 1,000 -- 600 1,000 800 500 700 1,100 1,300 1,00Benzo[k] fluoranthene 1,000 -- e200 900 700 500 e200 820 1,400 60Benzo[g,h,i] perylene <400 <400 e200 470 <400 -- e200 540 800 Bis(2- ethylhexyl) phthalate 500 -- e200 -- 500 -- e200 -- 800 30Chrysene 57.1 862 1,290 900 -- 500 1,000 700 -- 500 1,000 1,700 70Fluoranthene 111 2,355 2,230 1,800 300 700 2,000 1,100 500 700 1,900 3,300 1,60Fluorene 21.2 144 536 <200 <200 m -- <200 -- m -- <200 Indeno(1,2, 3-cd) pyrene <400 <400 e200 580 <400 -- e200 700 700 Isophorone <200 <200 <200 <150 <200 -- m <150 <200 <20Naphthalene 34.6 391 561 <200 <200 m e20 <200 -- m e20 <200 n-Butylbenzyl- phthalate <200 <200 m -- <200 <200 m -- <200 <20Phenanthrene 41.9 515 1,170 700 200 590 300 e200 460 1,700 50Phenol <200 <200 <200 e80 <200 <200 <200 e90 <200 Pyrene 53.0 875 1,520 1,400 300 600 1,500 800 400 600 1,500 2,900 1,20

Table 22. Chemical concentrations in sediment samples from Ellison Park wetland, Monroe County, N.Y., 1994-2001, and sediment-quality guid

chemicals—continued

Element

Freshwater sediment

quality guideline

Southern wetland area 

(upstream from flow-control structure)

Nor

(downstream 

Site SQ1 Site SQ2 

ISQG1 PEL2 PEC3 1994 1997 1999 2001 1994 1997 1999 2001 1994 19
C. Organochlorine and organophosphate compounds, values represent concentrations in silt-clay fraction (particle size < 2 mm), in microgram

billion)Organochlorine compoundsAldrin <0.1 0.5 <0.2 <2 <0.2 <0.5 <0.8 <3 <0.1 <0.3 

Chlordane 4.50 8.87 17.6 20 43 24 -- 18 40 38 -- 26 39 Dieldrin 2.85 6.67 61.8 2.0 2.2 1.4 m 1.2 1.5 3 <3 3.5 4.2 Endosulfan <0.1 <0.2 <0.2 <2 <0.2 <0.2 <0.2 <3 <0.1 <0.2 Endrin 2.67 62.4 207 <0.8 <1.2 <0.2 <4 <1.6 <1.1 <0.2 <6 <0.8 <0.5 Heptachlor <0.1 <0.2 <0.2 <2 <0.2 <0.2 <0.2 <3 <0.1 <0.2 Heptachlor  epoxide 0.60 2.74 16.0 <0.8 0.4 <0.4 <2 <1.6 0.4 <0.5 <3 <0.8 <0.2 Lindane 0.94 1.38 4.99 <0.1 <0.2 <0.2 <2 <0.2 <0.2 <0.2 <3 <0.1 <0.2 Methoxychlor <29 <2.5 <2.5 <10 <24 <2.5 <2.5 <15 <38 <2.5 Mirex <0.1 <0.2 <0.2 <2 <0.2 <0.2 <0.2 <3 <0.1 <0.2 p,p’-DDD 3.54 8.51 28.0 3.4 5.8 e6.7 3 4.5 4.6 e13 e3 6.5 7.p,p’-DDE 1.42 6.75 31.3 6.6 13 12 7 8.4 11 12 9 7.4 11 p,p’-DDT 1.19 4.77 62.9 1 1.4 <2.7 <4 0.6 1.2 <5.0 <6 0.8 1.PCB 34.1 277 676 26 46 39 <100 43 40 75 <150 52 98 PCN <1.0 -- -- -- <2.0 -- -- -- <1.0 -- Toxaphene 0.1 -- -- <10 <50 <50 -- <20 <50 <50 -- <10 <50 Organophosphate compoundsDiazinon -- 1.8 <0.4 <0.2 -- 1.7 0.5 <0.2 -- 3 Ethion -- <0.2 <0.4 <0.2 -- <0.2 <0.2 <0.2 -- <0.2 Malathion -- <0.2 <0.4 <0.2 -- <0.2 <0.2 <0.2 -- <0.2 Methyl 

parathion -- <0.2 <0.4 <0.2 -- <0.2 <0.2 <0.2 -- <0.2 

Parathion -- <0.2 <0.4 <0.2 -- <0.2 <0.2 <0.2 -- <0.2 Trithion -- -- <0.2 <0.2 -- -- <0.2 <0.2 -- -- 

Table 22. Chemical concentrations in sediment samples from Ellison Park wetland, Monroe County, N.Y., 1994-2001, and sediment-quality guidel

chemicals—continued

Element

Freshwater sediment

quality guideline

Southern wetland area 

(upstream from flow-control structure)

North

(downstream f

Site SQ1 Site SQ2 

ISQG1 PEL2 PEC3 1994 1997 1999 2001 1994 1997 1999 2001 1994 199
52 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

the SQGs indicates that (1) concentrations of metals exceeded 

the ISQG values but generally were lower than the PEL and 

PEC values; this implies that the wetland sediment quality is 

slightly impaired; (2) concentrations of PAHs greatly exceeded 

ISGQ values and often exceeded the PEL values, and exceeded 

the PEC values at Site SQ3 during 1994, which implies severe 

impairment; and (3) concentrations of several organochlorine 

compounds were less than the ISQG values, although the 

concentrations of DDD, DDE, DDT, and PCB indicated 

moderate impairment, and chlordane concentrations indicated 

severe impairment. These results generally indicate that the 

ecological health of the wetland is moderately to severely 

impaired, and agree with the evaluation of ecological health 

based on the 1996 macroinvertebrate survey. Except for the 

elevated concentrations of PAHs at site SQ3 downstream from 

the FCS in 1994, the data indicated a similar degree of 

impairment in each part of the wetland during a given year. 

Most of the elements and compounds showed no 

substantial fluctuation in concentrations, nor a consistent 

upward or downward trend, among the years sampled. The only 

elements or compounds whose concentrations indicated a clear 

upward trend from 1994 to 2001 upstream from the FCS were 

antimony, calcium, chromium, and possibly mercury and 

strontium (table 22A); the only one of these to show an upward 

trend downstream from the FCS was chromium. Downward 

trends were indicated for barium and uranium. Concentrations 

of many PAHs appeared to decrease in 1997 and 1999, but the 

2001 concentrations were close to those measured in 1994 

(table 22B). Conversely, concentrations of the few 

organochlorine compounds that had detectable concentrations 

in all four years appeared to increase in 1997 and 1999, but like 

many PAHs, their 2001 concentrations were comparable to 

those measured in 1994 (table 22C). Therefore, these results do 

not indicate that the FCS had any definable effect on sediment 

quality. 

Constituent Loads and Removal Efficiencies

Inflow and outflow loads of selected constituents, and the 

wetland’s removal efficiency for each constituent, were 

computed to evaluate the effect of the FCS on constituent loads. 

Bias in the concentrations of some constituents collected by the 

automatic samplers at the inflow and outflow water-quality 

monitoring sites might affect these results, as discussed below.

Bias in Water-Quality Data

Significant differences in the means of the paired EWIand Auto-sample data, and indications of systematic bias in the 

data collected by the automatic sampler at the inflow site 

(Blossom Road), were noted in the concentrations of total 

phosphorus, ammonia, and total suspended solids during the 

pre- and postcontrol periods, and in the concentrations of 

volatile suspended solids during the postcontrol period. Bias in 

the automatically collected samples from the outflow site 

(Empire Boulevard) was noted in the concentrations of 

ammonia during the pre- and postcontrol periods, and nitrateplus-nitrite nitrogen during the postcontrol period (table 23). 

The concentrations of automatically collected samples were not 

adjusted for these biases before calculation of loads for the 

following five reasons. 

1. The QA/QC analyses were performed on discrete samples 

that were associated with instantaneous discharges, whereas 

the automatic-sample analyses were performed on samples 

composited over periods ranging from several hours to 

several days and were associated with the mean discharge 

over that period. 

2. The majority of QA/QC samples were collected during lowflow periods, mostly when streamflow was less than 

300 ft3/s, and none were collected during annual peak-flow 

periods. 

3. No consistent relation between the errors in concentration 

and discharge was observed. The variation in error was 

generally greatest among low-flow samples and decreased 

with increasing discharge; therefore, any error in the 

computed loads presumably was greatest during low-flow 

periods, when the loads of a constituent in the creek were 

small, and was smallest during periods when the loads in the 

creek were large. 

4. For a given constituent, the range in concentration of the 

QA/QC samples collected during either the pre- or 

postcontrol period was much less than the range 

encompassed by all of the automatically collected samples 

that were collected during that period (see table 7). 

5. Biases, when identified for both monitoring sites during a 

given period, such as for ammonia concentrations, had the 

same sign (+ or -) at both sites. 

These observations indicate that error in the computed 

loads of the five constituents named above was unavoidable, 

but the relation between inflow and outflow loads, and the 

relative change in wetland removal efficiency, was valid. The 

possible effects of using unadjusted concentration data for 

estimation of chemical loads are discussed in the sections on 

specific constituents. 

Criteria for Evaluation of Flow-Control Structure

The effect of the FCS on constituent concentrations and 

loads passing through the wetland was assessed through a 

comparison of pre- and postcontrol data from the wetland’s 

inflow and outflow (1) chemical concentrations (table 24, 

fig. 12); (2) average annual loads (table 25); and (3) monthly 

loads (including postcontrol outflow loads that were predicted 

by the precontrol ESTIMATOR-regression equation) and flowweighted concentrations (figs. 13-21; appendix 1). These data 

were used to calculate annual and monthly pre- and postcontrol 

removal-efficiency values (figs. 13-21; appendix 1). Statistical 

analyses were performed to objectively identify significant 
Table 23. Mean difference (bias) between concentrations of selected constituents in water samples collected by automatic sampler (Auto) and e

method (EWI), and comparative statistics for concentrations of constituents in quality-assurance/quality-control samples (QA/QC) and all automaples at the inflow (Blossom Road) and outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, N.Y., 1990-2001

[Values are in milligrams per liter. Bias is mean difference between concentrations in Auto and EWI samples for those constituents for which a statistically sigbetween concentrations was found. A positive (negative) bias results when Auto concentrations are consistently greater than (less than) EWI concentrations. n, number of samples; Min, minimum value; Max, maximum value; <, less. than. Locations are shown in fig. 2.]

Constituent

Inflow (Blossom Road) Outflow (Empire Bouleva

QA/QC data All data QA/QC data 

n Bias Median Min Max n Median Min Max n Bias Median Min Max n M

Precontrol period (October 1990 through February 1997) Total phosphorus 19 0.035 0.10 0.02 0.37 1081 0.10 0.009 2.15

Ammonia 18 -.005 .01 < .01 .07 1045 .01 < .01 .32 14 -0.015 0.03 0.01 0.09 925 

Total suspended solids 15 32.7 56 6 188 324 139 3 1,510Postcontrol period (March 1997 through September 2001)Total phosphorus 17 0.018 0.055 .02 .20 636 0.075 .015 6.75

Ammonia 15 -.01 0.02 < .01 .04 657 < .01 < .01 .45 17 -0.016 0.06 .01 .09 694 Nitrate plus nitrite 17 .034 0.90 .51 1.4 633 Total suspended solids 15 9.34 13 7.7 142 195 83 4 5,800Volatile suspended solids 11 1.61 3 1.6 14 193 12 2 293

. 
Table 24. Concentrations and summary statistics of selected constituents, Ellison Park wetland, Monroe County, N.Y., 1990-2001 [Inflow site is Blossom Road; outflow site is Empire Boulevard. Concentrations are in milligrams per liter except those for cadmium, copper, lead, and zinc, which are in micrograms per liter. n, number of samples; <, less than. Locations are shown in fig. 2.] Constituent Precontrol period (October 1990 through February 1997) Postcontrol per

(March 1997 through Sep

Inflow concentration Outflow concentration Inflow concentration 

n Percentile n Percentile n Percentile n 

75th 50th 25th 75th 50th 25th 75th 50th 25th 

Total phosphorus 1081 0.17 0.10 0.06 1004 0.16 0.10 0.06 636 0.14 0.08 0.04 606 

Orthophosphate 1082 .016 .011 .007 947 .023 .016 .01 663 .02 .013 .009 704 Ammonia-plus- organic nitrogen 1076 .99 .77 .58 1003 .98 0.78 .61 633 .65 .46 .33 608 Ammonia nitrogen 1045 .02 .01 < .01 925 .05 .03 .02 657 .015 < .01 < .01 694 Nitrate plus nitrite 1066 1.2 .99 .81 961 1.2 .9 .7 623 1.2 1.0 .8 633 Total suspended solids 324 244 139 95 255 152 102 73 195 171 83 39 162 Volatile suspended solids 320 30.5 21 15 242 21 16 12 193 21 12 7 162 Chloride 1087 140 110 95 945 140 120 100 626 150 130 110 637 Sulfate 1076 180 140 110 934 170 140 110 627 180 150 110 637 Total organic carbon 22 6.7 6.0 5.0 23 6.3 5.6 4.4 141 8.2 6.1 5.2 147 Biochemical oxygen demand 50 3.0 < 2.0 < 2.0 53 2.2 < 2.0 < 2.0 139 3.0 2.1 < 2.0 145 Chemical oxygen demand 24 40 30 18 27 30 22 14 147 40 30 20 153 Cadmium 76 <1 < 1 < 1 78 < 1 < 1 < 1 153 1 1 < 1 159 Copper 70 70 40 20 70 70 40 30 153 60 40 30 159 Lead 74 13 7 < 5 76 8 5 < 5 153 19 11 6 159 Zinc 775 70 40 < 40 652 60 40 < 30 303 40 25 15 304 
Effects of flow modification 55

differences in loads and flow-weighted concentrations between 

inflow and outflow sites and between the pre- and postcontrol 

periods (table 26). Included in these analyses was the 

adjustment of postcontrol loads and flow-weighted 

concentrations by the mean differences between precontrol 

inflow and outflow loads and flow-weighted concentrations, 

respectively, to statistically remove the load-altering effect of 

the wetland (“wetland effect”) and theoretically isolate the 

effect of the FCS. (See “Statistical Analyses of Load Data” in 

the “Methods” section.)

Total Phosphorus

Inflow concentrations of total phosphorus (TP) generally 

were lower during the postcontrol period than the precontrol 

period (table 24; fig. 12), but the average annual inflow load 

increased 13 percent (table 25). This increase reflected the 

disproportionate runoff volume that occurred during 2 of the 5 

postcontrol years. Postcontrol outflow concentrations of TP 

were slightly lower than precontrol outflow concentrations, and 

average annual outflow loads decreased 14 percent. Outflow 

loads exceeded inflow loads during the dry summers of 1999 

and 2001 but generally were lower than the loads predicted by 

the precontrol regression equation during these periods; they 

also were much lower during the postcontrol high-flow months 

of January and July 1998 and May 2000 than during the 

precontrol high-flow months of July 1992 and April 1993 

(fig. 13). Outflow flow-weighted concentrations generally were 

lower than inflow flow-weighted concentrations throughout the 

study period. The wetland’s removal efficiency for TP 

increased from 28 percent during the precontrol period to 

45 percent during the postcontrol period (table 25). 

The statistical analyses indicate that outflow loads were 

significantly lower than inflow loads during the precontrol 

period (p < 0.0001); however, the extremely large differences 

between inflow and outflow loads that resulted from the floods 

of January and July 1998 and May 2000 increased the variance 

in the data such that no statistically significant difference 

between inflow and outflow loads was detected for the 

postcontrol period (p = 0.0932; table 26). By removing the 

effect of the hydrologic dissimilarity between pre- and 

postcontrol periods from the load data through an analysis of the 

flow-weighted concentrations, the outflow concentrations were 

significantly lower than inflow concentrations for both periods 

(p < 0.0001, p = 0.0078). Additionally, adjusting the 

postcontrol data by removing the precontrol “wetland effect” 

resulted in a significant improvement in TP load reduction that 

could be attributed to the FCS; this adjustment decreased the 

p value for the postcontrol differences in flow-weighted 

concentrations from 0.0078 to 0.0012. 

TP concentrations in the automatically collected samples 

from the inflow site showed a positive bias during the pre- and 

postcontrol periods (table 23); that is, the TP concentrations in 

the Auto (automatically collected) samples generally were 

higher than the mean concentration of EWI (cross-sectional 

equal-width-increment) samples. No bias was found in the 

outflow concentrations. The magnitude of the inflow bias 

varied greatly in low-flow samples; the difference between 

Auto and EWI concentrations ranged from 0 to almost 

80 percent and averaged about 30 percent. The variation 

decreased with increasing discharge—the difference between 

the Auto and EWI concentrations (0.19 and 0.18 mg/L, 

respectively) for the highest sampled flow (800 ft3/s) was 

within the analytical error of measurement. Therefore, the 

greatest error in the computed loads presumably results from 

low-flow periods, when the TP load was small, and the smallest 

error results from high-flow periods, when the TP load was 

large. If the error was large, the computed inflow load (and the 

associated removal efficiency) for the precontrol period would 

have been inordinately large. The average TP bias (0.018 mg/L) 

during the postcontrol period, was only half as large as the 

precontrol bias (0.035 mg/L); thus, the error in the computed 

load and removal efficiency would have been smaller. 

Therefore, the relation between inflow and outflow loads, and 

the relative change in removal efficiency, are considered valid, 

even though the actual loads and removal efficiencies contain 

some unknown amount of error. These results indicate that the 

apparent increase in the wetland’s TP removal efficiency from 

28 percent during the precontrol period to 45 percent during the 

postcontrol period can be considered, at least, a minimum 

(17 percent) improvement in wetland removal of TP.

Orthophosphate

Postcontrol orthophosphate (PO4) concentrations at the 

inflow and outflow sites generally were greater than the 

precontrol concentrations, and outflow concentrations were 

greater than inflow concentrations during both periods 

(table 24; fig. 12). Average annual postcontrol inflow and 

outflow loads were 12 and 7 percent, respectively, larger than 

precontrol inflow and outflow loads (table 25). The wetland 

generated and exported PO4 in quantities greater than those that 

entered the wetland throughout the study period. The PO4 

outflow loads represented about 12 and 15 percent of total 

phosphorus outflow loads during the pre- and postcontrol 

periods, respectively. The exportation of PO4 from the wetland 

appeared to decrease slightly; the precontrol removal efficiency 

of -38 percent increased to a postcontrol value of -31 percent. 

The average annual PO4 load generated by the wetland was 

about 0.6 tons during both periods, however (table 25), and 

indicated that the concurrent increases in PO4 loads at the 

inflow and outflow sites distorted the interpretation of the 

removal-efficiency values. 

Monthly outflow loads exceeded monthly inflow loads, as 

did the monthly flow-weighted concentrations, throughout the 

study period (fig. 14). Flow-weighted concentrations at both 

sites showed an annual pattern of peak concentrations during 

the summer, and a slight increase during late fall to early winter. 

The peak summer concentrations reflected an increase in 

microbial activity and decomposition of organic matter; the 
56 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

0

2.5

0.5

1.0

1.5

2.0

CONCENTRATION, IN MILLIGRAMS PER LITER

TOTAL PHOSPHORUS, AS P

1 extreme outlier

not plotted

EXPLANATION

25th percentile

Median

75th percentile

Data value less than or equal to 1.5 times the

interquartile range outside the quartile

Outlier data value less than or equal to 3

and more than 1.5 times the

interquartile range outside the quartile

Outlier data value more than 3 times the

interquartile range outside the quartile

Detection limit

0.20

0.02

0.04

0.06

0.08

0.10

0.12

0.14

0.16

0.18

0.20 ORTHOPHOSPHATE, AS P

0

5

1

2

3

4

CONCENTRATION, IN MILLIGRAMS PER LITER

AMMONIA PLUS ORGANIC NITROGEN, AS N

Inflow Outflow Inflow Outflow

Precontrol Postcontrol

0

0.7

0.1

0.2

0.3

0.4

0.5

0.6

AMMONIA NITROGEN, AS N

Inflow Outflow Inflow Outflow

Precontrol Postcontrol

Figure 12. Concentrations of 16 constituents in and discharges associated with surface-water samples collected during pre- and 

postcontrol periods from inflow (Blossom Road) and outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, N.Y., 1990-

2001
Effects of flow modification 57

0

3.0

0.5

1.0

1.5

2.0

2.5

CONCENTRATION, IN MILLIGRAMS PER LITER

NITRATE PLUS NITRITE NITROGEN, AS N 

0

1,600

200

400

600

800

1,000

1,200

1,400

TOTAL SUSPENDED SOLIDS

2 extreme 

outliers

not plotted

Inflow Outflow Inflow Outflow

Precontrol Postcontrol

0

300

50

100

150

200

250

CONCENTRATION, IN MILLIGRAMS PER LITER

VOLATILE SUSPENDED SOLIDS

0

600

100

200

300

400

500

CHLORIDE, DISSOLVED

Inflow Outflow

Precontrol Postcontrol

Inflow Outflow

EXPLANATION

25th percentile

Median

75th percentile

Data value less than or equal to 1.5 times the

interquartile range outside the quartile

Outlier data value less than or equal to 3

and more than 1.5 times the

interquartile range outside the quartile

Outlier data value more than 3 times the

interquartile range outside the quartile

Figure 12 (continued). Concentrations of 16 constituents in and discharges associated with surface-water samples collected during 

pre- and postcontrol periods from inflow (Blossom Road) and outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, 

N.Y., 1990-2001
58 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

0

300

50

100

150

200

250

CONCENTRATION, IN MILLIGRAMS PER LITER

SULFATE, DISSOLVED

EXPLANATION

25th percentile

Median

75th percentile

Data value less than or equal to 1.5 times the

interquartile range outside the quartile

Outlier data value less than or equal to 3

and more than 1.5 times the

interquartile range outside the quartile

Outlier data value more than 3 times the

interquartile range outside the quartile

0

25

5

10

15

20

TOTAL ORGANIC CARBON

0

16

2

4

6

8

10

12

14

CONCENTRATION, IN MILLIGRAMS PER LITER

BIOCHEMICAL OXYGEN DEMAND

Inflow Outflow

Precontrol Postcontrol

Inflow Outflow

0

300

50

100

150

200

250

CHEMICAL OXYGEN DEMAND

1 extreme 

outlier

not plotted

Inflow Outflow

Precontrol Postcontrol

Inflow Outflow

Figure 12 (continued). Concentrations of 16 constituents in and discharges associated with surface-water samples collected during 

pre- and postcontrol periods from inflow (Blossom Road) and outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, 

N.Y., 1990-2001
Effects of flow modification 59

0

14

2

4

6

8

10

12

CONCENTRATION, IN MICROGRAMS PER LITER

TOTAL RECOVERABLE CADMIUM

0

350

50

100

150

200

250

300

TOTAL RECOVERABLE COPPER

0

100

20

40

60

80

CONCENTRATION, IN MICROGRAMS PER LITER

TOTAL RECOVERABLE LEAD

2 extreme

outliers

not plotted

0

400

50

100

150

200

250

300

350

TOTAL RECOVERABLE ZINC

1 extreme

outlier

not plotted

EXPLANATION

25th percentile

Median

75th percentile

Data value less than or equal to 1.5 times the

interquartile range outside the quartile

Outlier data value less than or equal to 3

and more than 1.5 times the

interquartile range outside the quartile

Outlier data value more than 3 times the

interquartile range outside the quartile

Detection limit

Inflow Outflow

Precontrol Postcontrol

Inflow Outflow Inflow Outflow

Precontrol Postcontrol

Inflow Outflow

Figure 12 (continued). Concentrations of 16 constituents in and discharges associated with surface-water samples collected during 

pre- and postcontrol periods from inflow (Blossom Road) and outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, 

N.Y., 1990-2001
60 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

0

3,500

500

1,000

1,500

2,000

2,500

3,000

DISCHARGE, IN CUBIC FEET PER SECOND

DISCHARGE ASSOCIATED 

WITH WATER SAMPLES

EXPLANATION

25th percentile

Median

75th percentile

Data value less than or equal to 1.5 times the

interquartile range outside the quartile

Outlier data value less than or equal to 3

and more than 1.5 times the

interquartile range outside the quartile

Outlier data value more than 3 times the

interquartile range outside the quartile

Inflow Outflow

Precontrol Postcontrol

Inflow Outflow

fall-to-winter increase could reflect leaching and 

decomposition of leaf litter. The large decrease in removal 

efficiency that coincided with installation of the FCS in 

February 1997 probably resulted from disturbance and erosion 

of the channel in the vicinity of the FCS. The percentage of PO4

exported from the wetland decreased from 1998 through 2001 

in a trend similar to that which appeared to be occurring at the 

end of the precontrol period (late 1994-96; fig. 14). Statistical 

analyses confirmed that outflow loads were significantly 

greater than inflow loads, as were the flow-weighted 

concentrations (p < 0.0001; table 26). Adjustment of 

postcontrol loads and flow-weighted concentrations for the 

precontrol “wetland effect” revealed no significant difference in 

PO4 loads or concentrations that could be attributed to the FCS. 

Ammonia-plus-Organic Nitrogen

Postcontrol concentrations of ammonia-plus-organic 

nitrogen (TKN) generally were less than precontrol 

concentrations at both sites, and the outflow concentrations 

exceeded inflow concentrations during the postcontrol period, 

but were similar to inflow concentrations during the precontrol 

period, (table 24; fig. 12). Postcontrol average annual loads at 

the inflow site were 28 percent lower than the precontrol loads, 

and, at the outflow site, were 24 percent lower (table 25). The 

wetland’s overall removal efficiency decreased from 

6.1 percent during the precontrol period to 0.7 percent during 

the postcontrol period. Postcontrol outflow loads exceeded 

inflow loads during low-flow periods and also exceeded the 

loads predicted by the precontrol regression equation, but 

generally were lower than inflow loads during high-flow 

months (fig. 15). Monthly removal efficiencies appear to be 

more negative than positive during the postcontrol period (fig. 

15), but the small loads associated with the negative removal 

efficiencies are offset by the large loads associated with the 

positive values; hence, a low, but positive, removal efficiency is 

indicated. Precontrol monthly inflow loads appear similar to 

monthly outflow loads (fig. 15), and no significant difference is 

evident (p = 0.0512; table 26). Statistically significant 

differences were noted between postcontrol inflow and outflow 

loads and flow-weighted concentrations, however (p = 0.0004 

and 0.0026, respectively). Adjustment for the precontrol 

“wetland effect” decreased these p values to less than 0.0001 

and 0.0023, respectively; this indicates that the FCS caused a 

significant, although slight, decrease in TKN loads. 

Ammonia

Ammonia (NH3) concentrations at the outflow site 

exceeded those at the inflow site during both periods. The 

postcontrol inflow concentrations were lower than the 

precontrol inflow concentrations, whereas postcontrol outflow 

concentrations were similar to precontrol outflow 

concentrations (table 24; fig. 12). Average annual NH3 loads 

Figure 12 (continued). Concentrations of 16 constituents in and discharges associated with surface-water samples collected during 

pre- and postcontrol periods from inflow (Blossom Road) and outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, 

N.Y., 1990-2001
Effects of flow modification 61

1 Water year extends from October 1 of one year through September 30 of the following year.

Table 25. Average annual inflow and outflow loads of selected constituents, and wetland’s removal efficiency for each 

constituent during pre- and postcontrol periods, Ellison Park wetland, Monroe County, N.Y., 1991-2001

[Inflow site is Blossom Road; outflow site is Empire Boulevard. Loads are in tons. Removal efficiency, in percent, is computed from total 

(not average annual) loads for the period. Locations are shown in fig. 2.]

Precontrol period (water years 1991-96)1 Postcontrol period (water years 1997-2001)1

Constituent Inflow load Outflow load

Removal 

efficiency Inflow load Outflow load

Removal 

efficiency

Total phosphorus 26.2 19.0 27.6 29.6 16.4 44.6

Orthophosphate 1.69 2.33 - 38.1 1.90 2.50 -31.4

Total nitrogen 275 263 4.3 230 230 -0.1

Ammonia-plus-organic nitrogen 126 118 6.1 90.6 90.0 0.7

Ammonia nitrogen 2.59 4.76 - 84.0 2.07 5.25 -153

Nitrate plus nitrite 149 145 2.7 139 140 -0.6

Total suspended solids 28,500 15,200 46.7 24,500 11,900 51.5

Volatile suspended solids 3,540 2,080 41.0 2,560 1,580 38.1

Chloride 16,800 17,500 - 3.9 18,500 20,200 -8.7

Sulfate 16,000 15,700 2.0 15,100 15,400 -1.8

decreased 20 percent at the inflow site during the postcontrol 

period, and increased 10 percent at the outflow site (table 25). 

Monthly outflow loads were consistently greater than 

inflow loads throughout the study period, but were substantially 

lower than predicted by the precontrol regression equation for 

the postcontrol period (fig. 16). The outflow loads represented 

1.8 and 2.3 percent of total nitrogen loads for the pre- and 

postcontrol periods, respectively. Monthly flow-weighted 

concentrations in outflow also exceeded those in inflow, and 

showed a semiannual pattern of peak NH3 generation during 

June and July and during December and January of each year. 

The peak summer concentrations coincided with warm 

temperatures and high rates of microbial activity, which 

facilitated NH3 generation through mineralization of organic 

matter. Seasonally high water levels also might have 

contributed to NH3 loads by promoting diffusion of NH3 from 

anaerobic soils. The NH3 peaks in flow-weighted 

concentrations occurred at the inflow site—the winter peaks 

generally were higher than summer peaks—and indicated that 

NH3 generation was a basinwide phenomenon and not limited 

to the wetland. The reason for the midwinter peaks in flowweighted concentrations is uncertain but could reflect the 

flushing of NH3 released from the decomposition of organic 

matter. Removal efficiencies showed a downward trend 

(increasing generation and exportation of NH3) during the 

precontrol period and an annual pattern of fluctuation within a 

wide, but fairly stable, range of negative values during the 

postcontrol period. The statistical analyses of the loads and 

flow-weighted concentrations confirmed that outflow values 

were significantly greater than inflow values (p < 0.0001), and 

adjusting the loads and concentrations for the precontrol 

“wetland effect” did not indicate that the FCS had a significant 

effect on these variables (table 26). 

NH3 concentrations in the Auto samples showed a 

negative bias during both periods and at both sites. The mean 

bias for the precontrol period at the inflow site (-0.005 mg/L) 

was half as large as that for the postcontrol period (-0.01) and a 

third as large as the mean bias computed for both periods at the 

outflow site (-0.015; table 23). The variation in the Auto-toEWI differences in inflow samples was large for low-flow 

samples and decreased for high-flow samples; this variation in 

the outflow samples was fairly constant regardless of flow rate. 

All loads were underestimated to some unknown degree—the 

outflow loads more severely than the inflow loads—because 

Auto concentrations generally were lower than EWI 

concentrations. Therefore, (1) the actual differences between 

inflow loads and outflow loads are larger than the calculated 

differences, especially those during the precontrol period; and 

(2) the calculated load differences and removal efficiencies can 

be considered minimum values. This analysis of the biases in 

NH3 concentrations did not change the conclusions drawn from 

the previous analysis of the estimated load data, 

however—outflow loads would still exceed inflow loads, 

postcontrol outflow loads would still exceed precontrol outflow 

loads, and the removal efficiency would still decrease during 

the postcontrol period (table 25). Also, the NH3 load would still 

represent only a small percentage of the total nitrogen load.

Nitrate-plus-Nitrite Nitrogen

Nitrate-plus-nitrite nitrogen (NOx) concentrations during 

the postcontrol period were similar to those of the precontrol 

period at both sites, but the inflow concentrations generally 

exceeded the outflow concentrations (table 24; fig. 12). 

Average postcontrol inflow and outflow loads were less than 

precontrol loads (6.7 and 2.6 percent, respectively), and the 

removal efficiencies for both periods were close to zero 

(table 25). Monthly outflow loads were similar to monthly 

inflow loads throughout the study period, and postcontrol 
Table 26. Statistics (p values) related to tests of significant difference between the monthly constituent loads and flow-weighted concentrations

 computed for inflow (Blossom Road) and outflow (Empire Boulevard) of Ellison Park wetland, Monroe County, N.Y., 1990-2001 [Statistics are based on the differences between monthly paired data; that is, inflow (Blossom Road) value minus outflow (Empire Boulevard) value. A p val

suggests a significant difference between inflow and outflow values; the smaller the p value, the stronger the inference of a significant difference. <, less tha

Constituent Precontrol period (October 1990 through February 1997) Postcontrol period (March 1997 through September 2001) Wilcoxon Signed Ranks test Greater load (based on monthly flow- weighted concentrations) Wilcoxon Signed Ranks test 

Differences in monthly loads Differences in monthly flow- weighted concentrations Differences in monthly loads Differences in monthly flow- weighted concentrations Load differences adjusted by precontrol mean difference Flow-weighted

concentration differences adjus

by precontrol mean differenc

Total phosphorus < 0.0001 < 0.0001 inflow 0.0932 0.0078 < 0.0001 0.0012 Orthophosphate < .0001 < .0001 outflow < .0001 < .0001 .4841 .2403 Total nitrogen .0017 < .0001 inflow .0437 < .0001 < .0001 .0489 

Ammonia-plus-organic nitrogen .0512 .0120 inflow .0004 .0026 < .0001 .0023 Ammonia nitrogen < .0001 < .0001 outflow < .0001 < .0001 .1180 .0064 Nitrate plus nitrite .0178 < .0001 inflow .7408 < .0001 .1102 .0267 Total suspended solids < .0001 < .0001 inflow < .0001 < .0001 < .0001 .0003 Volatile suspended solids < .0001 < .0001 inflow < .0001 < .0001 < .0001 < .0001 

Chloride < .0001 < .0001 outflow < .0001 < .0001 .0014 .3389 Sulfate .0004 < .0001 inflow < .0001 < .0001 < .0001 .1414 
Effects of flow modification 63

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

1.0

0.2

0.4

0.6

0.8

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-60

100

-40

-20

0

20

40

60

80

REMOVAL EFFICIENCY, IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

20

5

10

15

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

TOTAL PHOSPHORUS, AS P

MONTHLY FLOW-WEIGHTED CONCENTRATION

Peak load

30 tons

Figure 13. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for total phosphorus, Ellison Park 

wetland, Monroe County, N.Y., 1990-2001
64 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

0.06

0.01

0.02

0.03

0.04

0.05

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-100

20

-80

-60

-40

-20

0

REMOVAL EFFICIENCY, IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

1.2

0.2

0.4

0.6

0.8

1.0

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

ORTHOPHOSPHATE, AS P

MONTHLY FLOW-WEIGHTED CONCENTRATION

Figure 14. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for orthophosphate, Ellison Park 

wetland, Monroe County, N.Y., 1990-2001
Effects of flow modification 65

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0.3

1.5

0.6

0.9

1.2

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-60

40

-40

-20

0

20

REMOVAL EFFICIENCY,

IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

50

10

20

30

40

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

AMMONIA PLUS ORGANIC NITROGEN, AS N

MONTHLY FLOW-WEIGHTED CONCENTRATION

Figure 15. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for ammonia-plus-organic nitrogen, 

Ellison Park wetland, Monroe County, N.Y., 1990-2001
66 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

0.08

0.02

0.04

0.06

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-600

200

-400

-200

0

REMOVAL EFFICIENCY, IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

2.5

0.5

1.0

1.5

2.0

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

AMMONIA NITROGEN, AS N

MONTHLY FLOW-WEIGHTED CONCENTRATION

Figure 16. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for ammonia, Ellison Park wetland, 

Monroe County, N.Y., 1990-2001
Effects of flow modification 67

outflow loads were substantially lower than those predicted by 

the precontrol regression equation (fig. 17). Monthly flowweighted concentrations at both sites showed an annual pattern 

of low concentrations during late summer and fall, and high 

concentrations during the winter. A significant difference 

(p < 0.0001) between monthly flow-weighted concentrations 

for the two sites was obtained for both periods; but monthly 

loads for the two sites differed significantly (p = 0.0178) only 

during the precontrol period (table 26). A decrease in maximum 

flow-weighted concentrations occurred during the postcontrol 

period at both sites and, thus, could not be attributed to the FCS. 

Removal-efficiency values reflected an annual pattern as well; 

with an increase from negative or near-zero in winter to positive 

values between 10 to 20 percent during the summer. 

Adjustment of load and flow-weighted-concentration 

differences by their respective precontrol “wetland effects” 

indicated that the FCS had little or no effect on NOx loads 

(table 26).

A statistically significant bias in NOx concentrations was 

noted in outflow samples collected during the postcontrol 

period only—the concentrations in Auto samples exceeded 

those measured in EWI samples by 0.034 mg/L on average 

(table 23). Adjusting the data for this small bias would have 

slightly decreased the computed postcontrol outflow loads and 

possibly resulted in a positive removal efficiency during this 

period, but the change would have been minimal and would not 

have affected the conclusion that the FCS had little effect on 

NOx loads.

Total and Volatile Suspended Solids

Total suspended solids (TSS) outflow concentrations were 

lower than inflow concentrations during both periods, and 

postcontrol inflow and outflow concentrations were generally 

lower than precontrol concentrations (table 24; fig. 12). The 

average annual outflow load was significantly lower than 

inflow loads during both periods (table 25). The average annual 

inflow load decreased 14 percent between the two periods; 

whereas average annual outflow load decreased 22 percent. 

Monthly outflow loads were lower than inflow loads during 

both periods, and the outflow loads during the postcontrol 

period were smaller than those predicted by the precontrol 

regression equation (fig. 18). The overall removal efficiency for 

TSS increased slightly from 47 percent before FCS installation 

to 52 percent thereafter. The monthly outflow loads and flowweighted concentrations were significantly less than the inflow 

loads and flow-weighted concentrations during both periods 

(p < 0.0001; table 26). Adjusting the postcontrol load and flowweighted-concentration differences by the respective 

precontrol mean differences yielded mixed results; the 

significance of the adjusted differences was no greater than 

those of the unadjusted differences. Therefore, the FCS cannot 

be said to have contributed to the decrease in TSS loads. 

As in the TP analysis, a positive bias was detected in TSS 

concentrations in the Auto samples from the inflow site during 

both periods (table 23); therefore, the computed inflow loads 

and removal efficiencies were erroneously high. The bias varied 

greatly among low-flow samples—from an 8-percent 

difference to about a 75-percent difference between Auto and 

EWI concentrations—and averaged about 54 percent. The bias 

decreased with increasing discharge, with only an 8-percent 

difference between Auto (158 mg/L) and EWI (146 mg/L) 

concentrations for the highest sampled flow of 800 ft3/s. 

Therefore, the error in loads presumably was greatest during 

low flows, when the total load of TSS was low, and least during 

high flows, when the total load was greatest. The postcontrol 

bias (9 mg/L) was less than the precontrol bias (33 mg/L). 

Therefore, the estimated inflow loads exceeded the actual loads 

by some unknown quantity; the percent error in the postcontrol 

loads was less than in the precontrol loads; and the actual 

removal efficiencies were less than those computed from the 

estimated load data; nevertheless, the indicated 5-percent 

increase in overall removal efficiency after FCS installation can 

be considered an improvement in TSS removal. 

The decrease in inflow TSS concentrations and loads 

during the postcontrol period coincided with and might be at 

least partly attributable to erosion-control projects that were 

completed on Irondequoit Creek in Linear Park during 1999 and 

on Allen Creek during 2000 (fig. 1)—both sites are about 3 mi 

upstream from the wetland. As much as half of the total 

sediment loading to the Ellison Park wetland has been estimated 

to have originated from erosion of the sand-and-silt bluffs in 

Linear Park (Young and Burton, 1993). Stabilization of the 

streambanks in these reaches had not resulted in a sharp 

decrease in the total suspended solids loads at the wetland 

inflow site, presumably because within channel and overbank 

storage of previously eroded sediment could continue to be 

sources of sediment loads for several years. Nevertheless, 

stabilization of these streambanks is expected to result in 

measurable decreases in future TSS loads. 

Volatile suspended solids (VSS) concentrations and loads 

mimicked TSS patterns during the study period (fig. 19). 

Postcontrol concentrations and loads were smaller than 

precontrol values (tables 24 and 25), and outflow 

concentrations and loads were significantly lower than inflow 

values (p < 0.0001; table 26). Removal efficiency for VSS 

decreased slightly after FCS installation—from 41 to 

38 percent during the pre- and postcontrol periods, respectively 

(table 25). Adjustment of the loads and flow-weighted 

concentrations for the precontrol “wetland effect” indicated 

significant differences between inflow and outflow values 

(p < 0.0001), but none that could be attributed to the FCS. A 

positive bias in the Auto samples was identified for the inflow 

site during the precontrol period (table 23). Adjustment of VSS 

concentrations for this bias would have decreased the inflow 

load and removal efficiency during the precontrol period by 

some unknown, but small, amount; thus, the apparent 

improvement in VSS removal by the wetland would be slightly 

less than indicated by the estimated loads. 
68 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0.6

1.8

0.8

1.0

1.2

1.4

1.6

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-40

40

-20

0

20

REMOVAL EFFICIENCY, IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

80

20

40

60

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

NITRATE PLUS NITRITE NITROGEN, AS N

MONTHLY FLOW-WEIGHTED CONCENTRATION

Figure 17. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for nitrate-plus-nitrite nitrogen, 

Ellison Park wetland, Monroe County, N.Y., 1990-2001
Effects of flow modification 69

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

800

200

400

600

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-40

100

-20

0

20

40

60

80

REMOVAL EFFICIENCY, IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

25,000

5,000

10,000

15,000

20,000

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

TOTAL SUSPENDED SOLIDS

MONTHLY FLOW-WEIGHTED CONCENTRATION

Figure 18. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for total suspended solids, Ellison Park 

wetland, Monroe County, N.Y., 1990-2001
70 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Chloride and Sulfate

Chloride (Cl) concentrations at the inflow and outflow 

sites during a given period were comparable, but the postcontrol 

concentrations slightly exceeded the precontrol concentrations 

(table 24; fig. 12). Average annual loads at the inflow and 

outflow sites during the postcontrol period exceeded the 

precontrol loads by 10 and 15 percent, respectively (table 25). 

outflow loads and flow-weighted concentrations were 

significantly greater than inflow loads (p < 0.0001; table 26; 

fig. 20). The average annual contribution of chloride from the 

area between the two monitoring sites increased from 

3.9 percent during the precontrol period to 8.7 percent during 

the postcontrol period (table 25). Adjustment of postcontrol 

flow-weighted-concentration differences by the precontrol 

mean difference indicated that the FCS did not have a 

significant effect on chloride loads.

Sulfate concentrations and monthly loads at the inflow 

were comparable to those at the outflow, and precontrol values 

were similar to postcontrol values. Monthly inflow flowweighted concentrations appeared to be greater than the 

corresponding outflow values, however (table 24; figs. 12 and 

21). Average annual loads at the inflow and outflow sites during 

the precontrol period decreased by 5.6 and 1.9 percent, 

respectively, during the postcontrol period (table 25). Average 

annual outflow loads were smaller than inflow loads during the 

precontrol period, but exceeded them during the postcontrol 

period, and caused the overall removal efficiency to switch 

from positive (2.0 percent) to negative (-1.8 percent) after 1997 

(table 25). Monthly flow-weighted concentrations showed a 

more consistent pattern during the study period than did the 

loads data; the monthly outflow flow-weighted concentrations 

were slightly, though significantly (p < 0.0001; table 26), lower 

than inflow flow-weighted concentrations during both periods. 

Adjustment of postcontrol flow-weighted-concentration 

differences by the precontrol mean difference indicated that the 

FCS did not have a significant effect on sulfate loads. 

Concentrations of Other Constituents

Storm-runoff samples were analyzed for several 

constituents, including total organic carbon (TOC), 

biochemical oxygen demand (BOD), chemical oxygen demand 

(COD), and metals (cadmium, copper, lead, and zinc). Far 

fewer samples were analyzed for these constituents (except 

zinc) during the precontrol period than during the postcontrol 

period (table 24). Analytical procedures were refined during the 

study period, and lower detection limits were used during the 

postcontrol period; thus, fewer concentrations of some 

constituents were censored during the postcontrol period than 

the precontrol period.

TOC inflow concentrations slightly exceeded outflow 

concentrations during the precontrol period; median values 

were 6.0 and 5.6 mg/L, respectively (table 24), but were 

comparable to the outflow concentrations during the 

postcontrol period, when median values were 6.1 and 6.2 mg/L, 

respectively. TOC concentrations were greater at both sites 

during the postcontrol period than during the precontrol period 

(table 24; fig. 12).

BOD concentrations were an order of magnitude smaller 

than COD concentrations, but their concentrations showed 

similar patterns during the study period. Outflow 

concentrations generally were lower than inflow concentrations 

during the precontrol period but were similar to inflow 

concentrations during the post-control period (table 24; fig. 12). 

Median BOD concentrations at both sites were below the 

detection limit during the precontrol period but increased to 2.1 

and 2.2 mg/L at the inflow and outflow sites, respectively, 

during the postcontrol period. Median COD concentrations at 

the inflow and outflow sites were 30 and 22 mg/L, respectively, 

during the precontrol period, but were 30 mg/L at both sites 

during the postcontrol period. 

Cadmium concentrations at both sites were below the 

analytical detection limit during most of the study period. 

Outflow concentrations, when detected, generally were lower 

than inflow concentrations.

Copper concentrations in outflow samples were similar to 

those in inflow samples in both periods (table 24). Median 

postcontrol concentrations were the same as median precontrol 

concentrations at both sites—40 mg/L. The 75th-percentile 

concentrations at both sites were lower during the postcontrol 

period (60 mg/L) than during the precontrol period (70 mg/L). 

High outlier values were more frequently detected at both sites 

during the postcontrol period than during the precontrol period, 

however (fig. 12). 

Lead concentrations in outflow samples generally were 

lower than in inflow samples during both periods. Median 

concentrations at the inflow and outflow sites increased from 7 

and 5 mg/L, respectively, during the precontrol period to 

11 mg/L at both sites during the postcontrol period (table 24). 

More high outlier values were measured at both sites during the 

postcontrol period than during the precontrol period (fig. 12). 

Zinc concentrations in outflow samples generally were 

lower than in inflow samples from both periods (fig. 12). 

Postcontrol concentrations were lower than precontrol 

concentrations at both sites. Precontrol median concentrations 

in the inflow samples decreased from 40 to 25 mg/L and, in the 

outflow samples, from 40 to 20 mg/L (table 24). 

The relatively high concentrations of most metals in 

Irondequoit Creek water samples probably result from 

urbanization within the basin. Additionally, zinc could be 

derived from the dissociation of zinc sulfide from dolomite, 

which underlies part of the basin and is also within the 

overlying glacial deposits and local soils (Kappel and others, 

1986). Depending on the element, decreases in metal 

concentrations within the Ellison Park wetland could result 

from several processes, including adsorption on mineral and 

organic particles, formation of insoluble compounds, 

sedimentation, ion exchange, bioaccumulation in plant tissues, 

and biomagnification in aquatic biota (Hem, 1989; Kadlec and 

Knight, 1996). Concentrations of cadmium, copper, and zinc 
Effects of flow modification 71

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

80

20

40

60

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-40

80

-20

0

20

40

60

REMOVAL EFFICIENCY, IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

2,000

500

1,000

1,500

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

VOLATILE SUSPENDED SOLIDS

MONTHLY FLOW-WEIGHTED CONCENTRATION

Figure 19. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for volatile suspended solids, Ellison 

Park wetland, Monroe County, N.Y., 1990-2001
72 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

50

300

100

150

200

250

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-30

30

-20

-10

0

10

20

REMOVAL EFFICIENCY, IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

6,000

1,000

2,000

3,000

4,000

5,000

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

CHLORIDE, DISSOLVED

MONTHLY FLOW-WEIGHTED CONCENTRATION

Figure 20. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for chloride, Ellison Park wetland, 

Monroe County, N.Y., 1990-2001
Effects of flow modification 73

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

50

250

100

150

200

MONTHLY FLOW-WEIGHTED CONCENTRATION,

IN MILLIGRAMS PER LITER

Inflow concentration

Outflow concentration

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

-20

20

-10

0

10

REMOVAL EFFICIENCY, IN PERCENT OF INPUT LOAD

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0

3,500

500

1,000

1,500

2,000

2,500

3,000

MONTHLY LOAD, IN TONS

Input load (Blossom Road)

Output load (Empire Boulevard)

Output load predicted by pre-control model

MONTHLY LOAD

REMOVAL EFFICIENCY

SULFATE, DISSOLVED

MONTHLY FLOW-WEIGHTED CONCENTRATION

Figure 21. Monthly loads and flow-weighted concentrations, and wetland’s removal efficiency for sulfate, Ellison Park wetland, 

Monroe County, N.Y., 1990-2001
74 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

generally were lower than those established by Environment 

Canada (2003) and New York State Department of Health 

(2003) as maximum acceptable concentrations in drinking 

water—5 μg/L for cadmium; 1,000 to 1,300 μg/L for copper; 

and 5,000 μg/L for zinc—but exceeded the maximum 

concentrations for the protection of aquatic life—0.017 μg/L 

for cadmium; 2 to 4 μg/L for copper; and 30 μg/L for 

zinc—below which negative effects to the most sensitive 

species of plants and animals would be unexpected 

(Environment Canada, 2003). Lead concentrations exceeded 

Canadian and New York State drinking-water (10 to 15 μg/L) 

and protection-of-aquatic-life guidelines (1 to 7 μg/L) some of 

the time.

Water-Quality Improvement of Irondequoit Bay

Improvement of the ecological and recreational status of 

Irondequoit Bay by decreasing the nutrient loads entering the 

bay has been a management objective of Monroe County since 

the 1970’s. The MCEHL has been monitoring progress toward 

this objective by estimating the annual total-phosphorus load 

entering the bay, and evaluating the bay’s trophic status through 

two limnological measures of lake fertility—concentrations of 

phytoplankton chlorophyll and biologically available 

phosphorus.

Annual Total Phosphorus Load

O’Brien and Gere (1983) estimated the annual load of total 

phosphorus (TP) entering Irondequoit Bay during 1980-81 to be 

15.7 tons, whereas Kappel and others (1986) estimated the 

annual TP load to be 20.2 tons; this latter value represented a 

10-percent decrease from the load estimated at Blossom Road 

as a result of passage through the wetland. Both analyses 

estimated that installation of a flow-control structure at the 

Narrows could achieve a total decrease in TP loads of 24 to 

25 percent, or a 15-percent improvement in the wetland’s 

removal efficiency. The current study found that the average 

annual TP load entering the bay during the precontrol period 

was 19.0 tons, which represented a wetland removal efficiency 

of 28 percent—almost three times the 10-percent effectiveness 

estimated by Kappel and others (1986). During the postcontrol 

period, the average annual TP load decreased to 16.4 tons, and 

the wetland’s removal efficiency increased to 45 percent—a 

17-percent improvement in response to the FCS. Although the 

1980-81 estimate of the wetland’s removal efficiency (10-

percent) was lower than the 1990-96 measured value (28-

percent), the measured improvement in removal efficiency 

(17 percent) that could be attributed to the FCS approximates 

the improvement estimated by the earlier researchers (15-

percent). Therefore, the FCS has met Monroe County’s 

management objective by this standard.

Trophic Status of the Bay

Chlorophyll_a concentrations have been measured in 

Irondequoit Bay since 1971 (Monroe County Environmental 

Health Laboratory, written commun., 2002). Water samples are 

collected biweekly at the deepest point in the bay from May 1 

through October 31 at about 3-ft intervals through a 20-ft depth 

in the epilimnion. Concentrations of chlorophyll_a within the 

water column are averaged for a given day, and the mean of 

these daily values is used as the summer mean concentration. 

The biologically available (potential) phosphorus concentration 

in the bay is computed from the annual total-phosphorus load, 

the surface-water discharge into the bay, and the bay’s mean 

depth and surface area (Vollenweider, 1976). A plot of the 

chlorophyll_a and potential phosphorus concentrations in the 

bay (fig. 22) depicts the bay’s trophic status from 1971 through 

2001. Eutrophic conditions prevailed during the 1970’s, when 

effluent from sewage-treatment plants was discharged into 

Irondequoit Creek. These conditions improved after the 

cessation of effluent discharges in 1977 and in response to 

erosion- and stormwater-control measures implemented by 

Monroe County since 1980 to decrease chemical loads from 

nonpoint sources in the basin. 

The trophic status of Irondequoit Bay has improved during 

the last 20 years and has entered the mesotrophic target range of 

the Irondequoit Bay water-quality management plan (Center for 

Governmental Research, Inc., 1985) three times during the 

study period—once before FCS installation (1995) and twice 

since FCS installation (1999 and 2001) (fig. 22). These 

apparent annual improvements in trophic status might result 

from two conditions—lower water levels in the wetland (fig. 4), 

and fewer and lower-magnitude storms (fig. 9); both would 

have resulted in lower TP loadings to the bay during these years 

than during the other study years. The FCS, which had a 

measurable effect on decreasing TP loads to the bay, 

presumably still made a contribution toward these 

improvements during the postcontrol years. 

SUGGESTIONS FOR FUTURE MONITORING

The results of this study suggest that future assessments of 

the effects of the FCS and the other flow modifications could 

focus on measurements of (1) sedimentation rates, through the 

annual measurement of sediment thickness at the 

sedimentation-rate-measurement sites; and (2) the wetland’s 

removal efficiency of total phosphorus, through continued 

monitoring of streamflow and water quality at the inflow and 

outflow sites. These two components of the study documented 

the most apparent concern (sedimentation upstream from the 

FCS) and most supportive argument (decreased loads of TP to 

Irondequoit Bay) for the continued operation of the FCS. Other 

components of the study contributed little to the assessment of 

adverse effects of the FCS. 
Summary and conclusions 75

71

77

78

8180 79

82

83

85

86

88

89

90 91 92

93

94

95

96

97

98

99

00

01

10 20 30 40 50 70 100 200 300 400 500 700 1,000

POTENTIAL PHOSPHORUS CONCENTRATION, IN MILLIGRAMS PER CUBIC METER

2

100

2.5

3

4

5

6

7

8

9

10

15

20

25

30

40

50

60

70

80

90

SUMMER MEAN CHLOROPHYLL_A CONCENTRATION,

IN MILLIGRAMS PER CUBIC METER

Data point and

year of measurement

TARGET RANGE OF

EUTROPHIC 

(Nutrient rich)

MESOTROPHIC 

(Nutrient balanced)

OLIGOTROPHIC 

(Nutrient poor)

WATER-QUALITYMANAGEMENT PLAN

88

EXPLANATION

Figure 22. Trophic status of Irondequoit Bay, 1971-2001, as function of measured summer mean concentrations of chlorophyll_a and

computed concentrations of potential phosphorus

Continued monitoring of the FCS’s effects on water levels 

are probably unnecessary because its flood-causing potential 

has been adequately evaluated. The ecological monitoring 

program documented the variable nature of the wetland’s 

biological systems and sediment quality, and the data indicated 

that identification of any adverse effect of the FCS in the midst 

of this natural variability would be difficult on the basis of 

surveys conducted on a greater-than-annual frequency. A 

monitoring program conducted at a yearly or seasonal 

frequency might be economically infeasible, however. 

Therefore, the value of the biological and sediment-quality 

surveys, which are required by State and Federal regulatory 

agencies, would be to provide an overview of the general 

ecological health of the wetland over time, and to detect any 

egregious change in wetland habitat. A decrease in the areal 

extent of cattails or open-water areas, or elevated 

concentrations of chemicals that might be harmful to wetland 

biota are likely to be detected long before irremediable effects 

occur, even with a greater-than-annual survey frequency. Once 

identified, these adverse effects might warrant repetition of 

surveys at a greater frequency to confirm their existence and 

extent, and(or) possible mitigation of the effect by removal of 

the FCS. 

SUMMARY AND CONCLUSIONS

Nutrients in Irondequoit Creek, which drains 151 mi2 of 

mostly residential and urban land in Monroe County, N.Y., 

contribute to eutrophic conditions in Irondequoit Bay on Lake 

Ontario. The Irondequoit Creek basin was studied in the 1979-
76 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

81 Nationwide Urban Runoff Program of the U.S. 

Environmental Protection Agency; one recommendation of that 

study was to install a flow-control structure (FCS) midway 

through the Ellison Park wetland, a 423-acre, predominantly 

cattail (Typha glauca) wetland at the mouth of Irondequoit 

Creek. An 11-year (1990-2001) study was conducted by the 

U.S. Geological Survey in cooperation with the Monroe County 

Department of Health to document the effects that flow 

modifications, which included installation of a FCS in 1997 and 

increased diversion of stormflows to the backwater areas of the 

wetland, would have on the wetland’s ability to decrease 

chemical loads transported by Irondequoit Creek into 

Irondequoit Bay. The FCS was intended to raise the watersurface elevation and thereby increase the dispersal and 

detention of stormflows in the southern (upstream) half of the 

wetland and, by increasing sedimentation and time for 

microbial utilization, decrease the loads of certain constituents, 

primarily phosphorus, that would otherwise be carried into 

Irondequoit Bay. Installation of the FCS was permitted by State 

and Federal wetland-regulatory agencies with the requirement 

that a comprehensive monitoring program be implemented to 

assess the effects of the FCS on the wetland’s water levels, 

biota, sedimentation rates, and chemical quality of water and 

sediment. Periodic surveys of the plant, fish, benthicmacroinvertebrate, and bird communities were conducted. 

Sedimentation rates were measured, and sediment quality was 

analyzed. Chemical loads entering and leaving the wetland 

were estimated and were used to calculate removal-efficiency 

values before and after FCS installation.

High-flow measurements of the water surface along 

Irondequoit Creek and discharge at the wetland inflow 

monitoring site above Blossom Road indicated that as much as 

1.5 ft of backwater was possible immediately upstream from 

the FCS, but that this backwater would not extend as far 

upstream as Blossom Road (about 13,000 ft). Rather, the 

backwater observed at Blossom Road was due to constriction of 

flow by the Browncroft Boulevard bridges, about 5,000 ft 

upstream from the FCS, or by local debris jams.

The plant surveys, which documented species richness, 

concentrations of nutrients and metals in cattail tissues, and 

cattail productivity (above- and below-ground biomass), 

indicated that observed differences among surveyed years and 

between the areas upstream and downstream from the FCS were 

due to seasonal water-level fluctuations that reflected the watersurface elevation of Lake Ontario, rather than water-level 

control by the FCS. Productivity during a given year apparently 

was affected by depth of inundation either during the current 

year, or during the end of the previous year’s growing season, 

when rhizome growth and sprouting of new shoots would be 

affected by water depth at that time, rather than by the water 

depth at the beginning of the next growing season. Results 

indicated no adverse effects to the cattails from either the 

naturally high water levels that prevail annually during the 

spring and summer in the wetland, nor from the short-duration 

elevated water levels that result from FCS operation. 

The fish surveys documented the use of the wetland by 44 

species, of which 25 to 29 species were found in any given year. 

Community composition was relatively consistent during the 

study, but seasonal and year-to-year variations in dominant 

resident and nonresident species were noted; these variations 

probably reflected natural or regional population patterns in 

Lake Ontario and Irondequoit Bay. The FCS did not pose any 

substantial obstruction to the passage of fish, and did not have 

other adverse effects on the fish community.

The bird surveys documented the use of the wetland by 

more than 90 species for breeding, feeding, and migration. The 

composition of the breeding-bird population during 1996 

(precontrol period) differed from that in 2002 (postcontrol 

period) in that 15 additional breeders were identified in 2002; 

all breeding species reported in 1996 were also identified in 

2002, however. Of wetland ground-nesting species, no 

evidence of American bittern, American coot, common 

moorhen, or pied-billed grebe—which had been reported 

during at least one of the earlier surveys—was found in 2002; 

least bittern and sora were reported, however. Several other 

ground-nesting species—ring-necked pheasant, common snipe, 

blue-winged teal, and American woodcock—were not reported 

during any of the study-period surveys, but only during the 

1980-85 survey, which covered an area much larger than the 

Ellison Park wetland. Absence or scarcity of specific species 

was not FCS-related, but was either a regional phenomenon or 

a result of seasonally high water levels that limited the extent of 

breeding habitat for some species, and the exposure of mud flats 

that attracted other species. 

The benthic-macroinvertebrate data were of minimal use 

for evaluating the effect of the FCS because no surveys were 

conducted after FCS installation; the precontrol results allowed 

assessment of the ecological quality of the wetland on the basis 

of biotic indices, however. The benthic-macroinvertebrate taxa 

in the wetland were typical of low-velocity, organically rich, 

oxygen-poor aquatic habitats with low taxonomic diversity, and 

large populations of those species that can survive. 

Macroinvertebrate-community measures of water quality 

indicated moderately to severely impaired conditions, whereas 

species indicators of water quality (such as the high ratio of 

Gammarus pseudolimnaeus to G. fasciatus, and low rates of 

chironomid mouthpart deformities) indicated “clean-water” 

conditions. Differences between the macroinvertebrate 

communities in the southern part of the wetland and those in the 

northern part are attributed to habitat differences, such as 

substrate composition, water depth, and density of submerged 

aquatic vegetation.

Sedimentation rates in the area upstream from the FCS 

increased from 3.0 mm/yr during the precontrol period to 

7.3 mm/yr during the postcontrol period, but a similar increase 

of 2.7 to 4.6 mm/yr was documented in the area downstream 

from the FCS. The concurrent increase in sediment 

accumulation downstream from the FCS, and the dynamic 

pattern of deposition and scour that was documented during the 

study period indicated that although the FCS undoubtedly 

played a major role in the apparent postcontrol increase in 
References Cited 77

sedimentation rates upstream from the FCS, other factors, such 

as the magnitude, frequency, and the timing (season) of peak 

flows, might also have contributed to this result.

Periodic sediment-quality analyses were used to monitor 

the concentrations of major and trace elements, polycyclic 

aromatic hydrocarbons, and organochlorine and 

organophosphate compounds in sediment samples from three 

long-term depositional sites in the wetland. The presence of 

certain constituents, and their concentrations, were fairly 

typical of a depositional environment in a highly urbanized 

basin. The concentrations of most constituents showed no 

substantial fluctuation or consistent upward or downward trend 

during the years of sampling. The only constituents whose 

concentrations indicated an upward trend from 1994 through 

2001 upstream from the FCS were antimony, calcium, 

chromium, and possibly mercury and strontium; of these, only 

chromium showed an upward trend downstream from the FCS. 

Downward trends were indicated for barium and uranium. 

Concentrations of many PAHs appeared to decrease in the 1997 

and 1999 samples, but the 2001 concentrations were close to 

those measured in 1994. Conversely, concentrations of the few 

organochlorine compounds that had detectable concentrations 

in all four sampling years appeared to increase in the 1997 and 

1999 samples, but like many PAHs, had concentrations in 2001 

that were comparable to those measured in 1994. Therefore, 

there was no definable effect of the FCS on sediment quality. 

Sediment-quality guidelines that are used to assess the 

ecological quality of substrate environments indicated that the 

wetland was moderately to severely impaired; this agrees with 

the benthic-macroinvertebrate biotic indices. 

During the precontrol period (1990-96), the wetland was a 

sink for total phosphorus and total suspended solids (removal 

efficiencies were 28 and 47 percent, respectively), but had little 

effect on conservative constituents (chloride and sulfate). The 

wetland was a source of orthophosphate and ammonia (removal 

efficiencies were -38 and -84 percent, respectively). The greater 

loads at the outflow than at the inflow were mainly the result of 

natural wetland processes, but buried wastes and composting 

operations in a landfill adjacent to the wetland could also have 

contributed to these increases.

During the postcontrol period (1997-2001), the wetland 

continued to be a sink for particulate constituents (removal 

efficiencies for total phosphorus and total suspended solids 

were 45 and 52 percent, respectively). The exportation of 

orthophosphate decreased (the removal efficiency increased to 

-31 percent), whereas the exportation of ammonia increased 

(the removal efficiency decreased to -153 percent). The outflow 

loads of orthophosphate and ammonia represented about 15 and 

2.3 percent of total phosphorus and total nitrogen loads, 

respectively. Changes in the loads of conservative constituents 

were negligible, and the overall removal efficiencies for other 

constituents during the precontrol period differed from those of 

the postcontrol period by no more than 5.4 percent. 

Statistical analyses of monthly inflow and outflow loads 

indicated significant differences between inflow and outflow 

loads of most constituents during both periods. Adjustment of 

the load data to remove the effect of hydrologic differences 

between the precontrol period and the postcontrol period 

(through calculation of flow-weighted concentrations) 

indicated similar results. Further adjustment of these data to 

remove the wetland effect—as defined by the precontrol mean 

differences between inflow and outflow loads and flowweighted concentrations—indicated that the FCS contributed 

significantly to the decrease in total phosphorus loads and 

slightly to a decrease in ammonia-plus-organic nitrogen loads, 

but had little or no significant effect on loads of other 

constituents. 

Annual measurements of summer chlorophyll_a and 

calculations of the inflow loads of biologically available 

phosphorus to monitor the trophic status of Irondequoit Bay 

indicated that eutrophic conditions improved to nearmesotrophic conditions three times during the study 

period—once before FCS installation (1995) and twice since 

FCS installation (1999 and 2001). These apparent annual 

improvements in trophic status might result in part from lower 

water levels in the wetland, and fewer and smaller storms 

during these years than during the other study years; still, the 

FCS presumably made a contribution toward these 

improvements during the postcontrol years by retaining 

phosphorus.

In conclusion, the FCS had no identifiable effect on the 

wetland biota—plants, fish, or birds—nor on flooding in the 

vicinity of Blossom Road. Localized increases in sediment 

accumulation upstream of the FCS were measured, but 

sediment quality did not appear to be adversely affected by the 

FCS. The FCS caused a statistically significant decrease in total 

phosphorus loading to the bay. 

References Cited

A list of reports on related biological studies produced as a 

result of the biological-monitoring program in the Ellison Park 

wetland and in fulfillment of contractual agreements with the 

U.S. Geological Survey or Monroe County Department of 

Health is given on page 81.

Adamus, P.R., and Stockwell, L.T., 1983, A method for 

wetland functional assessment—v. I, critical review and 

evaluation concepts: U.S. Department of Transportation, 

Federal Highway Administration Report FHWA-IP-82-23, 

176 p.

American Public Health Association, American Water Works 

Association, Water Environment Federation, 1995, Standard 

methods for examination of water and wastewater (19th ed.): 

Washington, D.C., American Public Health Association, 

paginated by section.

Andrews, N.J., and Pratt, D.C., 1978, The potential of cattails 

(Typha spp.) as an energy source—Productivity in managed 

stands: Journal of the Minnesota Academy of Science, v. 44, 

p. 5-8.
78 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Andrle, R.F., and Carroll, J.R., 1988, The atlas of breeding birds 

in New York State: Ithaca, N.Y., Cornell University Press, 

551 p.

Bagenal, T.B., and Tesch, F.W., 1978, Age and growth, in

Bagenal, T.B., (ed.), Methods for the assessment of fish 

production in freshwaters: Oxford, England, Blackwell 

Scientific Publications, p. 101-136.

Bernard, J.M., and Fitz, M.L., 1979, Seasonal changes in above 

ground primary production and nutrient contents in a central 

New York Typha glauca ecosystem: Bulletin of the Torrey 

Botanical Club, v. 106, p. 37-40.

Bernard, J.M., and Gorham, Eville, 1978, Life history aspects 

of primary production in sedge wetlands, in Good, R.E., 

Whigham, D.F., and Simpson, R.L., (eds.), Freshwater 

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82 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

A. Phosphorus Compounds 

Monthyear

Total phosphorus, as P Orthophosphate, as P

Inflow Outflow

RE

Inflow Outflow

Load Error Load Error Load Error Load Error RE

Oct-96 6.54 2.80 2.30 0.46 64.8 0.46 0.11 0.45 0.06 2.1

Nov-96 2.76 0.84 1.37 0.21 50.3 0.31 0.05 0.33 0.03 -4.6

Dec-96 3.71 1.10 1.63 0.24 56.0 0.45 0.07 0.43 0.04 5.2

Jan-97 0.67 0.12 0.48 0.05 27.7 0.12 0.01 0.13 0.01 -7.5

Feb-97 1.21 0.31 0.73 0.10 39.5 0.13 0.02 0.13 0.01 -4.7

Mar-97 2.10 0.33 1.21 0.13 42.7 0.10 0.01 0.14 0.01 -50.6

Apr-97 1.15 0.19 0.86 0.10 25.8 0.06 0.00 0.09 0.01 -54.2

May-97 0.93 0.19 0.74 0.09 20.5 0.06 0.01 0.09 0.01 -50.9

Jun-97 0.90 0.15 0.80 0.09 10.9 0.09 0.01 0.14 0.01 -57.2

Jul-97 0.45 0.07 0.45 0.05 0.3 0.07 0.01 0.11 0.01 -55.3

Aug-97 0.60 0.13 0.57 0.08 5.2 0.08 0.01 0.13 0.01 -64.0

Sep-97 1.69 0.67 0.85 0.15 49.8 0.10 0.01 0.15 0.02 -48.1

WY97 total 22.71 6.91 11.99 1.74 47.2 2.03 0.32 2.33 0.23 -14.7

Oct-97 0.54 0.10 0.53 0.06 1.0 0.06 0.01 0.11 0.01 -73.2

Nov-97 1.01 0.18 0.78 0.09 22.3 0.11 0.01 0.18 0.02 -60.8

Dec-97 0.78 0.11 0.73 0.07 6.6 0.13 0.01 0.22 0.02 -69.7

Jan-98 30.02 13.35 4.18 1.03 86.1 0.54 0.08 0.78 0.11 -44.4

Feb-98 1.95 0.39 1.08 0.13 44.8 0.15 0.01 0.18 0.02 -26.1

Mar-98 3.66 0.61 2.28 0.24 37.8 0.17 0.01 0.24 0.02 -40.9

Apr-98 1.04 0.16 1.04 0.11 0.1 0.07 0.01 0.10 0.01 -43.1

May-98 0.79 0.15 0.82 0.10 -3.3 0.07 0.01 0.10 0.01 -37.2

Jun-98 1.64 0.47 1.34 0.21 18.2 0.16 0.02 0.23 0.02 -43.9

Jul-98 13.09 5.31 5.33 1.13 59.3 0.58 0.07 0.85 0.11 -45.9

Aug-98 1.95 0.60 1.42 0.21 27.1 0.21 0.02 0.28 0.03 -34.0

Sep-98 0.96 0.16 0.96 0.11 -0.7 0.13 0.01 0.18 0.01 -39.5

WY98 total 57.44 21.60 20.49 3.49 64.3 2.38 0.27 3.45 0.37 -44.8

Oct-98 0.59 0.11 0.62 0.07 -3.8 0.08 0.01 0.11 0.01 -41.9

Nov-98 0.26 0.03 0.31 0.03 -21.9 0.05 0.00 0.08 0.01 -45.8

Dec-98 0.22 0.03 0.28 0.03 -28.1 0.06 0.00 0.09 0.01 -49.6

Jan-99 4.24 1.98 1.30 0.27 69.2 0.20 0.03 0.27 0.03 -30.7

Feb-99 0.87 0.14 0.89 0.10 -1.9 0.12 0.01 0.17 0.01 -39.8

Mar-99 4.51 0.90 3.39 0.41 24.9 0.21 0.02 0.31 0.03 -47.5

Apr-99 1.32 0.20 1.66 0.17 -25.6 0.10 0.01 0.15 0.01 -47.5

May-99 0.88 0.13 1.24 0.13 -42.1 0.10 0.01 0.14 0.01 -43.9

Jun-99 0.56 0.08 0.84 0.09 -50.9 0.10 0.01 0.14 0.01 -43.1

Jul-99 0.49 0.06 0.73 0.07 -49.2 0.11 0.01 0.16 0.01 -40.4

Aug-99 0.70 0.11 0.94 0.10 -33.6 0.14 0.01 0.19 0.01 -39.6

Sep-99 0.80 0.14 1.06 0.13 -32.9 0.12 0.01 0.18 0.02 -49.9

WY99 total 15.44 3.93 13.26 1.61 14.1 1.40 0.13 1.99 0.17 -42.4

Appendix. Monthly and annual inflow and outflow loads, associated errors, and removal efficiencies for 

selected constituents, Ellison Park wetland, Monroe County, N.Y., 1996-2001

[Inflow and outflow locations refer to Irondequoit Creek above Blossom Road and at Empire Boulevard, respectively. 

Load and error values are in tons. Error is standard error of prediction, which, when multiplied by 1.96 and added to 

and subtracted from the estimated load, provides 95-percent confidence limits of the load estimate. RE, removal 

efficiency, in percent, is computed from load values prior to rounding to two decimal places. Water year REs are based 

on total annual loads, not mean monthly values. Positive (negative) RE value indicates net retention in (export from) 

wetland, respectively. WY, water year. See Coon and others (2000) for 1990-96 data.]
Appendix 83

Oct-99 0.60 0.12 0.72 0.09 -19.8 0.09 0.01 0.13 0.01 -40.9

Nov-99 0.35 0.07 0.43 0.06 -22.1 0.07 0.01 0.10 0.01 -42.4

Dec-99 0.53 0.10 0.62 0.08 -16.4 0.12 0.01 0.17 0.01 -38.6

Jan-00 0.44 0.08 0.55 0.07 -25.6 0.11 0.01 0.15 0.01 -41.4

Feb-00 2.09 0.56 1.75 0.30 16.4 0.14 0.02 0.21 0.02 -49.9

Mar-00 1.30 0.18 1.49 0.15 -15.0 0.12 0.01 0.16 0.01 -30.0

Apr-00 4.65 0.97 3.92 0.49 15.7 0.20 0.02 0.28 0.02 -38.5

May-00 18.84 8.95 5.47 1.05 71.0 0.38 0.05 0.44 0.05 -14.8

Jun-00 4.55 1.22 3.57 0.55 21.6 0.38 0.04 0.48 0.05 -26.3

Jul-00 1.34 0.28 1.60 0.20 -19.6 0.24 0.02 0.31 0.03 -30.5

Aug-00 1.71 0.34 1.92 0.23 -12.7 0.27 0.02 0.36 0.03 -31.6

Sep-00 1.06 0.23 0.99 0.11 7.2 0.15 0.01 0.17 0.01 -13.5

WY00 total 37.46 13.10 23.02 3.38 38.5 2.27 0.22 2.95 0.27 -29.8

Oct-00 0.66 0.11 0.66 0.07 -0.2 0.10 0.01 0.11 0.01 -10.3

Nov-00 0.45 0.08 0.45 0.05 -0.2 0.09 0.01 0.10 0.01 -16.5

Dec-00 0.73 0.19 0.80 0.14 -9.0 0.14 0.01 0.19 0.02 -42.4

Jan-01 0.38 0.08 0.44 0.06 -15.1 0.09 0.01 0.11 0.01 -22.6

Feb-01 1.34 0.24 1.28 0.15 4.8 0.16 0.01 0.21 0.02 -25.6

Mar-01 5.70 1.27 3.52 0.45 38.3 0.24 0.02 0.29 0.02 -22.3

Apr-01 2.27 0.37 2.18 0.25 3.9 0.14 0.01 0.17 0.01 -24.9

May-01 0.72 0.11 0.97 0.10 -33.5 0.09 0.01 0.10 0.01 -21.7

Jun-01 0.50 0.07 0.76 0.08 -50.8 0.09 0.01 0.12 0.01 -32.3

Jul-01 0.39 0.06 0.56 0.06 -43.0 0.09 0.01 0.12 0.01 -27.0

Aug-01 0.52 0.09 0.62 0.07 -19.7 0.10 0.01 0.12 0.01 -20.8

Sep-01 1.11 0.41 0.85 0.15 23.6 0.11 0.01 0.13 0.02 -19.0

WY01 total 14.78 3.08 13.07 1.64 11.5 1.42 0.13 1.77 0.16 -24.2

Period of

record

WY97-01 147.83 48.61 81.84 11.86 44.6 9.50 1.06 12.48 1.20 -31.4

A. Phosphorus Compounds (continued)

Monthyear

Total phosphorus, as P Orthophosphate, as P

Inflow Outflow

RE

Inflow Outflow

Load Error Load Error Load Error Load Error RE

Appendix. Monthly and annual inflow and outflow loads, associated errors, and removal efficiencies for 

selected constituents, Ellison Park wetland, Monroe County, N.Y., 1996-2001 (continued)
84 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York 

rogen Nitrate plus nitrite Ammonia-plus-organic nitrogen Ammonia nitrogen

utflow

RE

Inflow Outflow

RE

Inflow Outflow

RE

Inflow Outflow

Load Load Error Load Error Load Error Load Error Load Error Load Error RE

22.92 6.7 12.49 0.88 12.46 0.72 0.2 12.06 1.91 10.45 1.09 13.4 0.12 0.05 0.53 0.12 -350

24.40 -0.2 14.36 0.77 14.55 0.68 -1.3 10.00 1.14 9.85 0.81 1.5 0.11 0.03 0.64 0.12 -485

37.81 5.9 25.18 1.29 24.54 1.09 2.5 14.99 1.63 13.27 1.06 11.5 0.21 0.05 1.14 0.20 -455

21.13 -4.3 14.91 0.72 15.45 0.65 -3.6 5.36 0.48 5.69 0.40 -6.2 0.11 0.02 0.73 0.12 -56326.26 1.1 19.61 1.09 19.53 0.93 0.4 6.95 0.76 6.73 0.54 3.2 0.15 0.04 0.75 0.14 -384

37.37 1.6 26.07 1.02 25.77 0.94 1.1 11.91 1.21 11.60 0.98 2.6 0.35 0.09 0.68 0.08 -92

21.24 0.7 14.54 0.59 13.61 0.52 6.4 6.84 0.74 7.63 0.68 -11.5 0.19 0.05 0.45 0.06 -139

13.31 6.1 9.16 0.37 7.59 0.29 17.2 5.02 0.61 5.72 0.56 -14.0 0.17 0.05 0.41 0.05 -145

11.35 3.9 7.50 0.30 6.21 0.24 17.2 4.31 0.50 5.14 0.49 -19.3 0.17 0.05 0.42 0.05 -152

5.52 5.9 4.11 0.16 3.25 0.12 20.8 1.76 0.19 2.27 0.21 -29.2 0.07 0.02 0.22 0.03 -228

5.75 -3.2 3.84 0.16 3.49 0.14 9.2 1.73 0.24 2.26 0.24 -30.7 0.04 0.01 0.18 0.02 -310

8.13 3.6 4.91 0.25 4.70 0.22 4.3 3.52 0.77 3.43 0.48 2.6 0.06 0.03 0.21 0.04 -260

35.20 2.5 156.68 7.60 151.16 6.53 3.5 84.45 10.17 84.05 7.53 0.5 1.75 0.49 6.37 1.04 -265

7.82 -19.2 4.32 0.17 4.77 0.18 -10.3 2.24 0.27 3.05 0.29 -36.5 0.04 0.01 0.23 0.03 -447

15.99 -10.5 8.82 0.36 9.70 0.37 -10.0 5.65 0.65 6.29 0.60 -11.3 0.15 0.04 0.51 0.07 -252

22.58 -17.3 12.77 0.48 14.97 0.53 -17.2 6.47 0.65 7.61 0.64 -17.7 0.25 0.06 0.73 0.08 -198

71.45 15.2 39.06 2.08 42.28 2.22 -8.2 45.19 9.54 29.18 4.68 35.4 1.71 0.82 1.82 0.34 -631.37 8.3 23.30 0.93 22.42 0.84 3.8 10.92 1.30 8.95 0.87 18.0 0.35 0.10 0.59 0.08 -67

45.19 0.2 28.77 1.07 29.84 1.03 -3.7 16.49 1.74 15.36 1.33 6.9 0.41 0.10 0.76 0.09 -87

18.77 -1.8 12.35 0.46 11.80 0.41 4.5 6.10 0.63 6.97 0.59 -14.4 0.14 0.04 0.38 0.05 -168

11.32 4.8 7.64 0.29 6.41 0.24 16.0 4.25 0.51 4.91 0.47 -15.5 0.12 0.03 0.32 0.04 -175

11.87 4.2 7.04 0.32 6.09 0.26 13.5 5.35 0.88 5.78 0.68 -8.0 0.17 0.07 0.40 0.06 -134

28.45 12.9 14.29 0.72 12.80 0.63 10.4 18.36 3.63 15.65 2.39 14.8 0.53 0.24 0.84 0.15 -59

9.59 7.6 6.18 0.27 5.43 0.22 12.2 4.19 0.68 4.16 0.46 0.7 0.08 0.03 0.26 0.04 -216

7.77 -2.2 4.84 0.19 4.59 0.17 5.1 2.76 0.31 3.18 0.29 -15.1 0.04 0.01 0.20 0.02 -356

82.18 5.1 169.39 7.34 171.08 7.08 -1.0 127.97 20.80 111.10 13.28 13.2 3.99 1.55 7.05 1.04 -77

nual inflow and outflow loads, associated errors, and removal efficiencies for selected constituents, Ellison Park 

wetland, Monroe County, 

refer to Irondequoit Creek above Blossom Road and at Empire Boulevard, respectively. Load and error values are in tons. Error is standard error of 

plied by 1.96 and added to and subtracted from the estimated load, provides 95-percent confidence limits of the load estimate. RE, removal efficiency, 

oad values prior to rounding to two decimal places. Water year REs are based on total annual loads, not mean monthly values. Positive (negative) RE 

n (export from) wetland, respectively. WY, water year. See Coon and others (2000) for 1990-96 data.]
Appendix 85

7.01 -7.7 4.19 0.16 4.21 0.15 -0.5 2.32 0.27 2.80 0.25 -20.6 0.04 0.01 0.19 0.02 -438

6.28 -12.1 3.87 0.14 4.02 0.14 -3.7 1.73 0.17 2.27 0.19 -30.8 0.04 0.01 0.21 0.03 -427

8.59 -14.1 5.47 0.21 5.93 0.21 -8.3 2.06 0.22 2.66 0.23 -29.5 0.07 0.02 0.28 0.03 -310

27.44 4.2 16.30 0.95 18.22 0.94 -11.8 12.34 2.90 9.23 1.38 25.2 0.39 0.21 0.62 0.11 -5624.24 -7.8 16.03 0.64 17.44 0.66 -8.8 6.44 0.72 6.80 0.64 -5.5 0.19 0.05 0.45 0.06 -135

49.77 -9.6 27.57 1.08 31.35 1.15 -13.7 17.85 2.05 18.42 1.71 -3.2 0.36 0.10 0.77 0.10 -114

23.19 -11.6 13.38 0.50 13.94 0.48 -4.2 7.41 0.74 9.25 0.78 -24.9 0.14 0.04 0.43 0.05 -198

14.19 -7.2 8.33 0.31 7.75 0.27 7.0 4.91 0.48 6.45 0.55 -31.3 0.11 0.03 0.37 0.04 -220

7.94 -4.0 4.95 0.18 4.26 0.15 13.9 2.69 0.26 3.68 0.31 -36.9 0.08 0.02 0.27 0.03 -251

5.83 -2.5 3.85 0.14 3.27 0.11 15.0 1.84 0.17 2.56 0.21 -39.0 0.05 0.01 0.20 0.02 -303

6.44 -4.4 4.09 0.15 3.74 0.13 8.6 2.07 0.22 2.70 0.23 -30.1 0.04 0.01 0.17 0.02 -363

7.58 -16.9 4.12 0.16 4.36 0.17 -5.8 2.36 0.28 3.22 0.32 -36.2 0.03 0.01 0.17 0.02 -476

88.53 -7.0 112.16 4.63 118.48 4.56 -5.6 64.03 8.49 70.04 6.79 -9.4 1.54 0.51 4.13 0.54 -168

7.25 -13.6 4.02 0.16 4.22 0.16 -4.9 2.36 0.30 3.03 0.30 -28.3 0.03 0.01 0.18 0.02 -498

7.20 -14.0 4.06 0.18 4.37 0.19 -7.5 2.25 0.32 2.83 0.33 -25.6 0.04 0.02 0.21 0.03 -384

14.57 -10.7 8.69 0.34 9.41 0.36 -8.3 4.47 0.52 5.16 0.51 -15.4 0.12 0.03 0.42 0.05 -252

16.33 -13.3 10.33 0.40 11.47 0.43 -11.0 4.07 0.49 4.86 0.48 -19.2 0.12 0.04 0.39 0.05 -21529.99 -16.0 16.53 0.80 19.71 0.95 -19.3 9.33 1.50 10.28 1.36 -10.2 0.19 0.07 0.47 0.08 -139

28.50 -6.8 18.26 0.67 19.13 0.66 -4.8 8.42 0.82 9.37 0.78 -11.2 0.15 0.04 0.42 0.05 -176

41.44 -6.7 21.18 0.86 22.32 0.88 -5.4 17.65 2.13 19.12 1.90 -8.3 0.27 0.08 0.66 0.09 -142

38.80 18.9 19.41 0.96 16.97 0.77 12.6 28.41 6.01 21.83 2.99 23.2 0.49 0.24 0.80 0.13 -63

24.71 7.4 12.98 0.56 11.18 0.46 13.9 13.69 2.01 13.53 1.58 1.2 0.30 0.10 0.65 0.09 -117

11.10 0.7 6.67 0.26 5.75 0.22 13.7 4.51 0.57 5.34 0.53 -18.5 0.09 0.03 0.31 0.04 -228

11.98 -2.9 6.93 0.27 6.47 0.24 6.7 4.72 0.58 5.52 0.52 -16.8 0.07 0.02 0.28 0.04 -286

7.63 5.3 4.88 0.20 4.34 0.16 11.2 3.18 0.42 3.30 0.31 -3.7 0.03 0.01 0.16 0.02 -390

39.49 -1.0 133.95 5.66 135.34 5.46 -1.0 103.06 15.68 104.16 11.60 -1.1 1.93 0.68 4.95 0.70 -157

(continued)

rogen Nitrate plus nitrite Ammonia-plus-organic nitrogen Ammonia nitrogen

utflow

RE

Inflow Outflow

RE

Inflow Outflow

RE

Inflow Outflow

Load Load Error Load Error Load Error Load Error Load Error Load Error RE

nual inflow and outflow loads, associated errors, and removal efficiencies for selected constituents, Ellison Park wetland, Monroe County, 

)
86 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York 

7.32 4.2 4.76 0.18 4.24 0.15 10.9 2.88 0.31 3.08 0.26 -6.9 0.03 0.01 0.17 0.02 -449 

8.35 -0.2 5.22 0.22 5.02 0.20 4.0 3.11 0.41 3.33 0.33 -7.2 0.05 0.02 0.23 0.03 -347 

17.56 -14.9 9.63 0.41 10.76 0.48 -11.7 5.65 0.82 6.81 0.89 -20.4 0.12 0.04 0.45 0.07 -262 

14.53 -4.6 10.16 0.42 10.29 0.41 -1.3 3.73 0.47 4.24 0.46 -13.7 0.09 0.03 0.30 0.04 -227 32.77 -6.5 20.92 0.84 22.48 0.85 -7.5 9.85 1.11 10.29 0.99 -4.5 0.20 0.05 0.50 0.07 -154 

53.06 -0.9 30.31 1.25 31.80 1.22 -4.9 22.29 2.77 21.26 2.10 4.6 0.30 0.09 0.68 0.09 -124 

31.53 -4.4 18.04 0.73 17.93 0.68 0.6 12.16 1.35 13.60 1.26 -11.9 0.16 0.04 0.46 0.06 -188 

12.80 -1.5 7.93 0.30 6.76 0.24 14.9 4.68 0.49 6.04 0.53 -29.2 0.08 0.02 0.28 0.03 -266 

8.44 -6.3 5.11 0.20 4.33 0.16 15.1 2.83 0.30 4.10 0.36 -44.8 0.06 0.01 0.23 0.03 -320 

5.45 -1.8 3.65 0.14 2.97 0.11 18.4 1.71 0.17 2.48 0.21 -45.0 0.03 0.01 0.16 0.02 -390 

5.35 -0.1 3.58 0.15 3.04 0.12 15.2 1.76 0.21 2.31 0.22 -31.2 0.02 0.01 0.12 0.02 -459 

7.25 0.3 4.22 0.22 3.90 0.20 7.5 3.05 0.62 3.34 0.46 -9.7 0.03 0.01 0.14 0.02 -432 

04.41 -3.6 123.53 5.07 123.51 4.81 0.0 73.70 9.03 80.90 8.06 -9.8 1.17 0.34 3.73 0.50 -218

49.81 -0.1 695.71 30.30 699.57 28.45 -0.6 453.22 64.17 450.24 47.27 0.7 10.37 3.57 26.23 3.80 -153

(continued)

rogen Nitrate plus nitrite Ammonia-plus-organic nitrogen Ammonia nitrogen

utflow

RE

Inflow Outflow

RE

Inflow Outflow

RE

Inflow Outflow

Load Load Error Load Error Load Error Load Error Load Error Load Error RE

nual inflow and outflow loads, associated errors, and removal efficiencies for selected constituents, Ellison Park 

wetland, Monroe County, 

)
Appendix 87

C. Suspended Solids 

Monthyear

Total suspended solids Volatile suspended solids

Inflow Outflow

RE

Inflow Outflow

Load Error Load Error Load Error Load Error RE

Oct-96 5,018 2,616 2,074 641 58.7 540 219 248 53 54.0

Nov-96 2,841 1,109 1,520 383 46.5 359 110 211 38 41.4

Dec-96 4,027 1,472 2,058 493 48.9 488 142 274 48 43.9

Jan-97 978 282 747 164 23.6 144 36 116 20 19.2

Feb-97 1,774 641 1,177 299 33.6 219 65 150 28 31.8

Mar-97 2,723 967 968 327 64.4 290 88 140 44 51.7

Apr-97 1,361 513 598 204 56.1 156 49 87 27 44.0

May-97 1,013 465 462 165 54.4 120 43 67 21 44.0

Jun-97 958 367 491 159 48.7 119 37 73 21 39.2

Jul-97 393 148 262 84 33.3 57 17 43 12 24.8

Aug-97 664 355 404 142 39.2 88 35 65 20 26.2

Sep-97 2,471 2,003 796 350 67.8 255 156 119 43 53.3

WY97 total 24,221 10,938 11,557 3,410 52.3 2,835 998 1,593 378 43.8

Oct-97 704 295 492 147 30.1 96 32 84 22 12.8

Nov-97 1,537 608 838 245 45.5 189 62 135 35 28.7

Dec-97 1,096 343 763 193 30.4 142 38 123 29 12.9

Jan-98 21,707 15,314 4,725 2,322 78.2 1,633 914 546 222 66.6

Feb-98 2,218 848 975 272 56.0 240 73 134 32 44.3

Mar-98 3,846 1,216 1,897 466 50.7 399 103 237 51 40.6

Apr-98 1,000 314 706 171 29.3 122 32 95 20 22.6

May-98 683 283 488 141 28.5 87 28 66 16 23.5

Jun-98 1,633 967 875 319 46.4 179 82 110 33 38.7

Jul-98 11,934 8,295 4,116 1,907 65.5 1,034 550 439 163 57.5

Aug-98 2,251 1,449 1,100 395 51.1 254 121 150 44 40.9

Sep-98 1,071 403 775 212 27.6 144 44 117 28 18.6

WY98 total 49,679 30,335 17,750 6,789 64.3 4,521 2,078 2,237 696 50.5

Oct-98 651 279 523 148 19.7 92 31 85 21 7.5

Nov-98 227 69 255 65 -12.3 37 10 46 11 -22.4

Dec-98 196 72 233 62 -18.7 31 9 41 10 -30.8

Jan-99 3,434 2,750 1,259 560 63.3 311 194 158 58 49.4

Feb-99 867 288 736 191 15.1 109 30 103 23 5.4

Mar-99 3,874 1,397 2,622 682 32.3 401 116 305 68 24.1

Apr-99 1,094 337 1,055 248 3.5 136 34 133 28 2.2

May-99 648 196 690 170 -6.5 88 22 91 20 -3.0

Jun-99 361 113 425 109 -17.5 54 14 60 14 -11.0

Jul-99 290 83 362 90 -24.6 47 11 54 12 -17.0

Aug-99 548 206 562 151 -2.6 81 24 85 20 -4.3

Sep-99 747 309 766 238 -2.5 106 34 115 30 -9.1

WY99 total 12,938 6,102 9,487 2,715 26.7 1,493 530 1,274 314 14.6

Appendix. Monthly and annual inflow and outflow loads, associated errors, and removal efficiencies for selected 

constituents, Ellison Park wetland, Monroe County, N.Y., 1996-2001 (continued)

[Inflow and outflow locations refer to Irondequoit Creek above Blossom Road and at Empire Boulevard, respectively. Load 

and error values are in tons. Error is standard error of prediction, which, when multiplied by 1.96 and added to and subtracted from the estimated load, provides 95-percent confidence limits of the load estimate. RE, removal efficiency, in percent, is computed from load values prior to rounding. Water year REs are based on total annual loads, not mean monthly 

values. Positive (negative) RE value indicates net retention in (export from) wetland, respectively. WY, water year. See Coon 

and others (2000) for 1990-96 data.]
88 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Oct-99 578 281 551 173 4.6 83 30 89 24 -6.9

Nov-99 328 159 341 121 -3.9 48 18 56 17 -17.8

Dec-99 527 207 514 157 2.4 74 23 82 21 -10.8

Jan-00 400 168 425 128 -6.3 56 18 67 17 -19.3

Feb-00 1,665 834 1,285 478 22.8 177 71 158 49 11.0

Mar-00 1,044 285 937 208 10.3 133 31 127 25 4.3

Apr-00 3,418 1,306 2,355 645 31.1 363 110 274 64 24.4

May-00 9,720 7,467 3,088 1,274 68.2 824 480 334 111 59.5

Jun-00 3,502 1,768 1,910 663 45.4 375 145 230 65 38.9

Jul-00 1,033 473 801 244 22.4 135 47 113 29 16.6

Aug-00 1,499 643 1,090 302 27.3 195 64 159 37 18.5

Sep-00 993 472 591 163 40.5 134 48 97 23 27.4

WY00 total 24,706 14,062 13,888 4,556 43.8 2,597 1,087 1,785 483 31.3

Oct-00 596 210 411 105 31.0 89 25 73 17 17.4

Nov-00 405 171 292 84 28.0 59 20 54 13 9.9

Dec-00 690 360 578 233 16.2 90 36 93 30 -3.0

Jan-01 297 138 258 88 13.1 43 15 45 12 -4.9

Feb-01 1,109 386 768 210 30.8 138 39 116 27 15.4

Mar-01 3,901 1,559 1,890 534 51.5 406 130 249 60 38.6

Apr-01 1,630 546 1,024 269 37.2 192 54 143 33 25.9

May-01 443 149 354 95 19.9 63 18 56 13 11.7

Jun-01 272 95 256 72 5.9 42 12 43 11 -1.5

Jul-01 192 63 182 51 5.0 32 9 34 9 -3.9

Aug-01 335 145 235 73 29.8 52 18 44 12 14.5

Sep-01 1,012 797 420 193 58.5 124 73 74 28 39.9

WY01 total 10,881 4,620 6,668 2,007 38.7 1,330 450 1,024 266 23.1

Period of

record

WY97-01 122,425 66,058 59,351 19,477 51.5 12,776 5,144 7,912 2,137 38.1

C. Suspended Solids (continued)

Monthyear

Total suspended solids Volatile suspended solids

Inflow Outflow

RE

Inflow Outflow

Load Error Load Error Load Error Load Error RE

Appendix. Monthly and annual inflow and outflow loads, associated errors, and removal efficiencies for selected 

constituents, Ellison Park wetland, Monroe County, N.Y., 1996-2001 (continued)
Appendix 89

D. Selected Anions (chloride and sulfate) 

Monthyear

Chloride Sulfate

Inflow Outflow

RE

Inflow Outflow

Load Error Load Error Load Error Load Error RE

Oct-96 1,321 71 1,395 72 -5.6 1,493 61 1,484 55 0.6

Nov-96 1,655 73 1,751 78 -5.8 1,597 59 1,572 55 1.6

Dec-96 2,741 117 2,847 124 -3.8 1,959 72 1,896 66 3.2

Jan-97 1,971 81 2,229 94 -13.1 1,298 47 1,316 45 -1.4

Feb-97 2,381 109 2,648 122 -11.2 1,370 54 1,376 50 -0.4

Mar-97 3,105 138 3,312 143 -6.7 1,926 66 1,896 61 1.6

Apr-97 1,699 79 1,820 83 -7.2 1,413 48 1,402 45 0.8

May-97 986 45 1,026 46 -4.1 1,137 37 1,102 34 3.1

Jun-97 752 35 806 37 -7.1 1,005 33 996 31 0.9

Jul-97 442 20 468 21 -6.0 753 25 727 22 3.4

Aug-97 442 21 494 23 -11.6 766 25 776 24 -1.4

Sep-97 588 33 627 33 -6.7 906 32 908 29 -0.3

WY97 total 18,083 825 19,423 876 -7.4 15,623 559 15,453 516 1.1

Oct-97 585 27 693 31 -18.5 895 28 961 28 -7.3

Nov-97 1,153 55 1,302 61 -12.9 1,229 39 1,284 38 -4.4

Dec-97 1,718 78 2,050 90 -19.3 1,376 42 1,488 42 -8.2

Jan-98 4,616 270 4,882 283 -5.8 2,501 100 2,394 87 4.3

Feb-98 3,141 149 3,127 145 0.5 1,693 55 1,614 49 4.7

Mar-98 3,727 165 3,972 171 -6.6 2,168 68 2,196 64 -1.3

Apr-98 1,625 72 1,739 75 -7.0 1,354 42 1,388 40 -2.5

May-98 935 43 962 43 -2.9 1,071 33 1,056 30 1.4

Jun-98 757 39 820 40 -8.2 988 32 1,013 31 -2.5

Jul-98 1,362 76 1,422 78 -4.4 1,544 54 1,559 50 -1.0

Aug-98 727 36 745 35 -2.5 1,031 33 1,037 31 -0.6

Sep-98 647 30 680 30 -5.0 964 30 988 29 -2.5

WY98 total 20,994 1,041 22,396 1,081 -6.7 16,813 554 16,977 517 -1.0

Oct-98 622 28 673 29 -8.1 920 28 950 27 -3.2

Nov-98 622 28 696 30 -11.9 812 25 836 24 -2.9

Dec-98 894 40 1,038 45 -16.1 870 27 912 26 -4.8

Jan-99 2,305 146 2,595 148 -12.6 1,355 52 1,409 48 -4.0

Feb-99 2,458 115 2,707 124 -10.2 1,372 44 1,433 43 -4.4

Mar-99 3,845 174 4,354 192 -13.3 2,173 70 2,336 71 -7.5

Apr-99 1,872 81 2,111 89 -12.7 1,470 46 1,590 46 -8.1

May-99 1,086 47 1,194 50 -10.0 1,165 35 1,225 35 -5.2

Jun-99 619 27 679 29 -9.6 861 27 891 26 -3.4

Jul-99 492 21 537 22 -9.2 780 24 799 23 -2.5

Aug-99 549 24 598 25 -8.8 850 26 885 25 -4.1

Sep-99 599 28 687 31 -14.6 895 28 973 29 -8.8

WY99 total 15,964 759 17,869 815 -11.9 13,523 433 14,238 422 -5.3

Appendix. Monthly and annual inflow and outflow loads, associated errors, and removal efficiencies for selected 

constituents, Ellison Park wetland, Monroe County, N.Y., 1996-2001

[Inflow and outflow locations refer to Irondequoit Creek above Blossom Road and at Empire Boulevard, respectively. Load and 

error values are in tons. Error is standard error of prediction, which, when multiplied by 1.96 and added to and subtracted from the 

estimated load, provides 95-percent confidence limits of the load estimate. RE, removal efficiency, in percent, is computed from 

load values prior to rounding. Water year REs are based on total annual loads, not mean monthly values. Positive (negative) RE 

value indicates net retention in (export from) wetland, respectively. WY, water year. See Coon and others (2000) for 1990-96 data.]
90 Effects of Flow Modification on a Cattail Wetland at the Mouth of Irondequoit Creek near Rochester, New York

Oct-99 636 29 707 31 -11.1 911 28 959 28 -5.2

Nov-99 670 34 756 38 -12.8 809 26 839 26 -3.7

Dec-99 1,399 64 1,560 70 -11.5 1,141 35 1,185 34 -3.9

Jan-00 1,727 79 1,983 89 -14.8 1,135 35 1,200 35 -5.8

Feb-00 2,576 141 3,023 166 -17.4 1,389 49 1,497 51 -7.7

Mar-00 2,867 126 3,111 134 -8.5 1,672 51 1,742 50 -4.2

Apr-00 2,871 137 3,180 150 -10.7 1,956 64 2,061 64 -5.4

May-00 2,259 123 2,266 116 -0.3 1,925 68 1,872 59 2.7

Jun-00 1,458 72 1,510 72 -3.6 1,483 48 1,500 45 -1.1

Jul-00 815 38 872 39 -7.0 1,038 32 1,068 31 -2.9

Aug-00 890 41 966 43 -8.5 1,125 34 1,184 34 -5.2

Sep-00 720 34 724 32 -0.6 978 31 973 28 0.6

WY00 total 18,889 919 20,658 982 -9.4 15,562 502 16,079 484 -3.3

Oct-00 765 34 756 32 1.2 1,000 30 969 27 3.1

Nov-00 865 43 904 43 -4.5 929 29 917 27 1.3

Dec-00 1,559 77 1,781 90 -14.2 1,183 37 1,224 37 -3.5

Jan-01 1,756 85 1,916 91 -9.1 1,094 34 1,095 32 -0.1

Feb-01 3,343 159 3,690 172 -10.4 1,619 54 1,667 52 -2.9

Mar-01 4,445 214 4,835 224 -8.8 2,285 78 2,324 74 -1.7

Apr-01 2,595 125 2,830 132 -9.1 1,710 57 1,758 55 -2.8

May-01 1,100 49 1,177 51 -7.0 1,112 36 1,115 33 -0.2

Jun-01 674 31 757 33 -12.3 856 28 873 27 -2.0

Jul-01 494 22 548 24 -10.9 738 25 737 23 0.2

Aug-01 512 24 553 25 -8.1 764 27 757 24 0.8

Sep-01 624 35 665 36 -6.6 848 32 837 29 1.2

WY01 total 18,732 899 20,413 953 -9.0 14,139 468 14,273 441 -0.9

Period of

record

WY97-01 92,662 4,443 100,759 4,707 -8.7 75,659 2,515 77,020 2,380 -1.8

D. Selected Anions (chloride and sulfate) (continued)

Monthyear

Chloride Sulfate

Inflow Outflow

RE

Inflow Outflow

Load Error Load Error Load Error Load Error RE

Appendix. Monthly and annual inflow and outflow loads, associated errors, and removal efficiencies for selected 

constituents, Ellison Park wetland, Monroe County, N.Y., 1996-2001 (continued)