REPORTS OF THE TIBOR T. POLGAR
FELLOWSHIP PROGRAM, 2013
David J. Yozzo, Sarah H. Fernald and Helena Andreyko
Editors
A Joint Program of The Hudson River Foundation and The New York State Department of Environmental Conservation
December 2015
ABSTRACT
Eight studies were conducted within the Hudson River Estuary under the auspices of the Tibor T. Polgar Fellowship Program during 2013. Major objectives of these studies included: (1) reconstruction of past climate events through analysis of sedimentary microfossils, (2) determining past and future ability of New York City salt marshes to accommodate sea level rise through vertical accretion, (3) analysis of the effects of nutrient pollution on greenhouse gas production in Hudson River marshes, (4) detection and identification of pathogens in aerosols and surface waters of Newtown Creek, (5) detection of amphetamine type stimulants at wastewater outflow sites in the Hudson River, (6) investigating establishment limitations of new populations of Oriental bittersweet in Schodack Island State Park, (7) assessing macroinvertebrate tolerance to hypoxia in the presence of water chestnut and submerged aquatic species, and (8) examining the distribution and feeding ecology of larval sea lamprey in the Hudson River basin.
iii
TABLE OF CONTENTS
Abstract ...... iii
Preface ...... vii
Fellowship Reports
Pelagic Tropical to Subtropical Foraminifera in the Hudson River: What is their Source? Kyle M. Monahan and Dallas Abbott ...... I-1
Sea Level Rise and Sediment: Recent Salt Marsh Accretion in the Hudson River Estuary Troy D. Hill and Shimon C. Anisfeld ...... II-1
Nutrient Pollution in Hudson River Marshes: Effects on Greenhouse Gas Production Angel Montero, Brian Brigham, and Gregory D. O’Mullan ...... III-1
Microbial Agents of Concern Detected in Water and Air at the Hudson River Estuary Waterfront Sherif Kamal and M. Elias Dueker ...... IV-1
Occurrence and Ecological Effects of Amphetamine Type Stimulants in Wastewater Effluent Alexis M. Paspalof, Daniel Snow, and Emma Rosi-Marshall ...... V-1
The Distribution of Invasive Celastrus Orbiculatus in an Anthropogenically Disturbed Riparian Ecosystem Shabana Hoosein and George Robinson ...... VI-1
Hypoxia Tolerance of the Invertebrates Associated with Water-Chestnut Beds (Trapa natans L.) in the Hudson River Mariana Carolina Teixeira and David L. Strayer ...... VII-1
The Distribution and Feeding Ecology of Larval Sea Lampreys in the Hudson River Basin Thomas M. Evans and Karin E. Limburg ...... VIII-1
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PREFACE
The Hudson River estuary stretches from its tidal limit at the Federal Dam at Troy, New
York, to its merger with the New York Bight, south of New York City. Within that reach,
the estuary displays a broad transition from tidal freshwater to marine conditions that are
reflected in its physical composition and the biota its supports. As such, it presents a major
opportunity and challenge to researchers to describe the makeup and workings of a complex
and dynamic ecosystem. The Tibor T. Polgar Fellowship Program provides funds for
students to study selected aspects of the physical, chemical, biological, and public policy
realms of the estuary.
The Polgar Fellowship Program was established in 1985 in memory of Dr. Tibor T.
Polgar, former Chairman of the Hudson River Foundation Science Panel. The 2013 program was jointly conducted by the Hudson River Foundation for Science and Environmental
Research and the New York State Department of Environmental Conservation and underwritten by the Hudson River Foundation. The fellowship program provides stipends and research funds for research projects within the Hudson drainage basin and is open to
graduate and undergraduate students.
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Prior to 1988, Polgar studies were conducted only within the four sites that comprise
the Hudson River National Estuarine Research Reserve, a part of the National Estuarine
Research Reserve System. The four Hudson River sites, Piermont Marsh, Iona Island, Tivoli
Bays, and Stockport Flats exceed 4,000 acres and include a wide variety of habitats spaced
over 100 miles of the Hudson estuary. Since 1988, the Polgar Program has supported
research carried out at any location within the Hudson estuary.
The work reported in this volume represents the eight research projects conducted by
Polgar Fellows during 2013. These studies meet the goals of the Tibor T. Polgar Fellowship
Program to generate new information on the nature of the Hudson estuary and to train students in estuarine science.
David J. Yozzo Henningson, Durham & Richardson Architecture and Engineering, P.C.
Sarah H. Fernald New York State Department of Environmental Conservation
Helena Andreyko Hudson River Foundation for Science and Environmental Research
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PELAGIC TROPICAL TO SUBTROPICAL FORAMINIFERA IN THE HUDSON RIVER: WHAT IS THEIR SOURCE?
A Final Report of the Tibor T. Polgar Fellowship Program
Kyle M. Monahan
Polgar Fellow
Institute for a Sustainable Environment Clarkson University Potsdam, NY 13676
Project Advisor:
Dallas Abbott Lamont–Doherty Earth Observatory Columbia University Palisades, NY 10964
Monahan, K. M. and D. H. Abbott. 2015. Pelagic Tropical to Subtropical Foraminifera in the Hudson River: What Is Their Source? Section I: 1-19 pp. In D.J. Yozzo, S.H. Fernald and H. Andreyko (eds.), Final Reports of the Polgar Fellowship Program, 2013. Hudson River Foundation.
I-1
ABSTRACT
Reconstruction of past climate events in the Hudson River basin is useful to
calibrate historical records, and increasingly important in light of recent hurricane
impacts on the area. The concentration of radiogenic nuclides such as Cs-137 and Pb-210
can be used as a sediment dating technique (Bopp and Simpson 1989), but do not provide historical climate data intrinsically. Microfossil identification was used, which provides
δ18O reconstructions and species specific tolerance ranges. Sediment samples from three
Hudson River sediment cores were sieved and picked for microfossils. Discrete layers of
marine pelagic foraminifera were found in one core, CD02-29A and replicated in CD02-
13. Tropical foraminifera species such as Globigerinoides ruber (pink) dominate the
assemblages from CD02-29A. Dating estimates of the cores were provided by Pb-210
and Cs-137 profiles. Pending carbon-14 dates from CD02-29A and CD02-13 will provide
more robust sedimentation rate assessments, and future hurricane event correlations.
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TABLE OF CONTENTS
Abstract ...... I-2
Table of Contacts ...... I-3
Lists of Figures and Tables ...... I-4
Introduction ...... I-5
Methods...... I-8
Results ...... I-10
Discussion ...... I-14
Acknowledgements ...... I-17
References ...... I-18
I-3
LIST OF FIGURES AND TABLES
Figure 1 – SEM photomicrographs of pelagic foraminifera ...... I-7
Figure 2 – Location of all sediment cores used in study...... I-9
Figure 3 – Photos of G. ruber (pink) from CD02-13 ...... I-10
Figure 4 – Photos of other marine planktonic foraminifera from CD02-13 ...... I-11
Figure 5 – CD02-29A updated core profile ...... I-11
Figure 6 – CD02-29A Cs-137 profile ...... I-12
Figure 7 – CD02-13 Cs-137 profile ...... I-13
Figure 8 – CD02-13 Total Pb-210 Activity ...... I-13
Table 1 – Species of foraminifera in Hudson River core CD2-29A ...... I-6
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INTRODUCTION
Reconstruction of past climate events is useful both for calibration of historical
records and providing a better baseline for comparing current atmospheric phenomena.
Identifying and developing sedimentary proxies can help ascertain the past intensity and
frequency of large drought and hurricane events, and help inform planning and adaption
to future events. Sediment dating techniques using the concentration of radiogenic
nuclides such as Cs-137 and byproducts of Pb-210 have been very successful in
providing sediment dates (Bopp and Simpson 1989), but do not provide historical climate
data. The use of microfossils in sediments for climate analysis is one method for
gathering climate data, through either δ18O measurement or through species specific
climatic tolerance ranges (Cronin et al. 2005).
In previous work, analysis of three sediment cores from the Hudson River from
off Manhattan Island to Piermont (shown in Fig. 2) determined that these cores contained
calcareous shells of marine organisms (Table 1).
When examining core CD02-29A, pelagic foraminifera were found in discrete layers. Globigerinoides ruber (pink) and other tropical species dominate the assemblages. None of these marine species could live in the Hudson River at normal salinity and temperature values (Weiss et al. 1978).
These tropical foraminifera were examined with a scanning electron microscope and have been shown to contain cocoliths cemented into their pore structure (Fig. 1).
Cocoliths are small carbonate plates produced by marine cocolithophore organisms
(Ramsey 1974).
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Table 1. Species of Foraminifera in Hudson River core CD2-29A Species Classification *Salinity *Temperature Range, ppt Range, °C G. ruber (pink) tropical 22-49 18-32 G. ruber (white) temperate to 22-49 14-32 subtropical G. sacculifer tropical 24-47 16-31
Orbulina universa cosmopolitan 23-46 10-30 Neogloboquadrina. tropical to 25-46 15-30 dutertrei subtropical Globorotalia inflata temperate 33-38* 5-27
Globorotalia tumida tropical to 34-38* 16-29 subtropical *Tolerance data from (Bijima et al. 1990) and (Tolderlund and Be 1971)
Cocolithophores are unilocular marine algae which live in marine environments and appear in the sediment record in the Upper Triassic (Bown et al. 2004). These plates suggest an open ocean source for these tropical foraminifera. The tropical and subtropical foraminifera were found in seven distinct layers in the sediment core, which suggests a unique transport mechanism.
The observed distinct layers of tropical marine foraminifera may have been deposited during a hurricane event which created a storm surge. A storm surge forms when a low pressure weather system with high winds piles up water near the coast, and can flood areas close by, especially if the storm occurs during high tide conditions.
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Figure 1. SEM photomicrographs of pelagic foraminifera from Hudson River cores. Top left: G. ruber (pink) from CD2-29A. Top right: Magnification of the black box in the figure to the left, showing coccoliths cemented onto the foraminiferal test. Bottom left: pelagic marine foraminifer from VM32-2. Bottom right: two coccoliths from area of foraminifer in box. Photo credits: Dee Breger, Micrographic Arts, Greenfield, NY.
Storm surges can form from both Nor’easters and hurricanes (tropical cyclones).
Warm core rings from the Gulf Stream travel to 40°N and contain tropical foraminifera
(Ruddiman 1969). Under most circumstances, warm core rings stay on the outer shelf.
The strong onshore winds and less intense rainfall produced by a hurricane whose eye has
its landfall south of the Hudson River (like Hurricane Sandy), could help to move warm
core rings from the outer shelf onto the inner shelf and subsequently into the Hudson. In
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this case, the foraminifera were most likely dead by the time they reached the Hudson
River.
Another possible transport mechanism is through drought conditions, which
would be evident in the movement northward of the “salt wedge” that is formed in an
estuary like the Hudson, where the dense salt water flows below the freshwater. This
would allow the foraminifera to live further up the river than the salinity tolerances of the
foraminifera species would originally suggest, and could result in such depositional
layers as the drought seasons come and go. This effect would be most likely to occur in
midsummer, a time when tropical to subtropical foraminifera would find it easier to
thrive in the offshore ocean, and may be exacerbated by strong offshore winds.
In this study, goals were to check the reproducibility of foraminifera bearing
layers in CD02-29A by re-sieving samples from the core, and to provide higher
resolution Cs-137 and Pb-210 dating of CD02-29A to further constrain the sedimentation
rate. Another objective was to process another similar core titled CD02-13 for Cs-137
concentration to assess sedimentation rate and sieve this core for microfossils.
Constraining the spatial distribution and absolute ages of the stratigraphic microfossil
layers is the first step to developing a useful paleoclimatic proxy.
METHODS
Core Sampling
Sediment cores CD02-29A and CD02-13 were kept in refrigerated storage at
Lamont Doherty Earth Observatory. Samples were taken from different depths in core in
2 cm3 sizes for analysis. Sampling depths in both cores were selected using lows in
I-8
acoustic impedance as a proxy for incursions of marine, clay-rich sediment. Some depths in core CD02-29 where foraminifera had been found previously were resampled to check on the reproducibility of the prior results. The sediment core CD02-29A was originally recovered during the “Can-Do” cruise on June 3rd, 2002 using a vibracore mechanism at
8.2 m water depth. Similarly, CD02-13 was taken on May 31st, 2002 at 11 m water depth, also a vibracore. Core locations are shown in Figure 2.
Figure 2. Map of the three sediment cores used in this study. VM32-2 is a piston core, CD02-13 and CD02-29A are vibracores.
Sample Collection
Sediment core samples were processed using wet sieving methods. Small subsamples were weighed and then dried in a sealed oven at 23° C for two days, and the weight was again recorded.
Next, the original sample was disaggregated in de-ionized water. The sample was sonified for 15 minutes, and allowed to settle overnight. This sonification step was repeated three times. In order to segregate different test size fractions, the samples were
I-9
washed through stacked 250μm, 125μm, 63μm, and 38μm sieve sizes, and rinsed with de- ionized water. Stacked sieves with samples were covered with a watch-glass and dried overnight. Sieves were cleaned through ultra-sonification in a water bath between each sample.
Fossil Picking
Fossil picking was accomplished with a Herrbrugg Wild MSA stereographic microscope for each size fraction at each depth sampled. The number and species of fossils were recorded and tests were stored in slides for reference and possible future isotopic analysis.
RESULTS
Multiple layers containing tropical planktonic foraminifera including a large number of G. ruber (pink) were recovered after sieving and fossil picking in CD02-13
(Fig 3, Fig 4). These assemblages were very similar in abundance and species distribution to CD02-29A.
Figure 3. SEM and visual light image of a G. ruber (pink) test from CD02-13, 100-102cm. Photo credits: Dee Breger, Micrographic Arts, Greenfield, NY.
I-10
Figure 4. Visual light images of other marine planktonic foraminifera tests from CD02-13.
Layers containing marine planktonic foraminifera were replicated in CD02-29A, and the core profile was updated to reflect the new values. Acoustic impedance was a good predictor of carbonate bearing layers (Fig. 5). New layers containing benthic foraminifera were also found in CD02-29A.
CD02-29A Updated Core Profile
Figure 5.
The acoustic impedance, number of pelagic forams, and the percent G. ruber (pink) is provided by depth in core.
I-11
Refined radiogenic isotope data for CD02-29A provided a higher resolution
dating profile for Cs-137. Peaks were resolved at 1971 for the Indian Point nuclear power
plant leak of radioactive material, in 1963 for the fallout maximum, and a final drop-off
in 1954 (Fig 6). For CD02-13, post-1954 Cs-137-bearing sediment was only detected in the first 2 cm, and mixed with post-1954 level sediment until 6cm. Below this line, Cs-
137 was inconclusive for CD02-13 (Fig 7).
Figure 6. The Cs-137 profile for CD02-29A shows a peak at 1971, during the leak at Indian Point, another fallout peak in 1963, and a final drop-off at 1954, all of which can be used as stratigraphic markers (Levinton and Waldman 2006).
I-12
CD02-13 Cs-137 Activity
Cs-137 (pCi/kg) -60 -40 -20 0 20 40 60 80 100 120 0
5 10 15 20
25 (cm) Core Depth 30
Figure 7. The Cs-137 profile for CD02-13 shows a slight increase at the beginning of the core of post-1954 Cs-137 and mixing down to 6cm, but lower values are within standard error variation.
Assuming a constant influx of sediment to the core location, an estimate of the sedimentation rate can be provided from the Cs-137 values. For CD02-29A, the estimated sedimentation rate is between 0.5 cm/year and 1.5 cm/year. Total Pb-210 activity from
CD02-13 was inconclusive.
Figure 8. The total Pb-210 data from CD02-13 with the production of refined lead in the Hudson Basin adjacent (Nelson, 2011).
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DISCUSSION
Previously suggested pelagic foraminifera-bearing layers were confirmed by repeat core sampling and sieving of CD02-29A. This result supports the previous layer identifications, and sampling at many depth levels corroborated the hypothesis that these are the main stratigraphic foram-bearing layers.
Similar assemblages of sub-tropical and tropical planktonic foraminifera were found in core CD02-13, located upriver from CD02-29A. This replicates the findings in CD02-
29A and suggests that deposits of pelagic foraminifera may occur throughout the lower
Hudson River. However, abundances of the fossils found were much lower than in
CD02-29A. Issues of possible contamination of the sieves by entrapped foraminiferal tests were addressed through sonification of sieves between each use, as in Green (2001).
In the future, methylene blue treatment of the sieves could provide a more robust protection against contamination, due to the low abundances present in CD02-13.
Radiogenic sediment dating for CD02-29A and CD02-13 provides valuable temporal
parameters for the former, but for the latter core the Cs-137 peaks were not resolved. This
is likely due to sedimentation rate changes or local dredging, as depositional patterns are
very spatially dynamic within the Hudson River Estuary (McHugh et al. 2004). Common
influences on the resolution of Cs-137 profiles within sediment cores include: reworking
of sediment (re-suspension and bioturbation), diffusional movement of Cs-137 through
pore waters, and particle size of sediments, which impacts particle association (Ritchie
and McHenry 1990). The most likely influence on the difference between resolved Cs-
137 horizons between CD02-29A and CD02-13 is changes in the sedimentation rate.
I-14
Sedimentation rates can vary seasonally, and commonly sedimentation in the Hudson
occurs both due to mechanisms of flocculation from an increase in salinity downstream
and kinetic settling of suspended matter in lower flow conditions (Woodruff et al. 2001).
High levels of rapid sediment deposition can occur in the estuarine turbidity maximum
(ETM) zone (Meade 1969). In this area, physical sediment trapping mechanisms such as
salinity-inducted flocculation and tidal effects cause a focusing and concentration of
suspended sediment (Nitsche et al. 2010).
The collection of these suspended particles acts as a sediment trap at the marine-
freshwater water interface, creating high sedimentation rates through rapid sheer velocity
changes, salinity gradients, and spatial stratification gradients (Ralston et al. 2013). The
depositional profiles at the ETM can change as the fluxes of freshwater from upriver flow
seasonally change (Woodruff et al. 2001). During times of high flow during snowmelt or
storm events, the increase in freshwater can push the ETM downstream, changing the
salinity gradients in the Hudson and allowing for higher rates of sediment deposition
downriver (Woodruff et al, 2001). Gathering bathymetric and seismic data can help to
address accumulation rates and depositional profiles (Nitsche et al. 2010), which may
help understand the dynamics of sediment deposition in the range of CD02-29A and
CD02-13.
A hurricane event could re-suspend sediment containing planktonic foraminifera tests
deposited from a tropical source (likely the Gulf Stream) which are then deposited once the tests reach the depositional area of the ETM within the Hudson River. This resuspended material could be deposited further upriver in the event of a storm surge.
One such storm surge event occurred in 1998 when high winds from the East produced
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water surface elevations into the Battery as far North as Albany, NY (Blumberg and
Hellweger 2006). The well-preserved nature of the tests suggests a rapid depositional
profile without post-depositional suspension, such as would be common in the ETM. This
assemblage of foraminifera from this paper is quite characteristic of the Gulf Stream, as
G. sacculifer has its highest abundances within the Gulf Steam, which was another
tropical foraminifera identified as a component of the tropical assemblage (Balsam and
Felssa et al. 1978). The high levels of fine-grained materials in the depositional layers
and correlation with acoustic impedance are consistent with high depositional rate event
sediment profiles in the Hudson, monitored by both sediment core and optical acoustic
measurements (Traykovski et al. 2004).
An external check for the climatic signal could be obtained through use of carbon-14
dating. Using this method in estuaries can be complex due to re-suspension and
incorporation of old carbon sources into shell material, however it has been used in the
past (Pekar et al. 2004) for this area in the Hudson. Pending carbon-14 dates from CD02-
29A and CD02-13 will provide more robust sedimentation rate assessments, and future
hurricane event correlations.
In the future, collection of another sediment vibracore in the depositional area near
CD02-29A would provide for replication of depositional layers in the sediment. Review
of other core candidates near CD02-13 which have a higher resolution Cs-137 profile is
suggested as well. Though greater levels of replication in identified stratigraphic layers are needed, the implications of current results still suggest a unique depositional mechanism and climatic signal, possibly due to storm surges.
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ACKNOWLEGMENTS
We would like to thank the Tibor T. Polgar Fellowship program and the Hudson
River Foundation for funding and support of this project. Thank you to Dee Breger for her SEM images and workup of the images in Figure 1. Thank you to Richard Bopp and
Miriam Katz at Rensselaer Polytechnic Institute for their support and continued advice, along with Bärbel Hönisch, Nichole Anest and Jon Stelling at LDEO.
I-17
REFERENCES
Bopp, R. F., and H. J. Simpson. 1989. Contamination of the Hudson River: The sediment record. Pages 401-416 in Contaminated Marine Sediments Assessment and Remediation. Marine Board, National Commission on Engineering and Technical Systems, National Research Council. National Academies Press, Washington, D.C. Bown, P. R., J. A. Lees, and J. R. Young. 2004. Calcareous nannoplankton evolution and diversity through time. Pages 481-508 in H. Thierstein and J. Young, (eds.), Coccolithophores: From Molecular Processes to Global Impact. Springer Berlin Heidelberg, New York. Bijima, J., W.W. Faber, and C. Hemleben. 1990. Temperature and salinity limits for growth and survival of some planktonic foraminifers in laboratory cultures: Journal of Foraminiferal Research 20:95-116. Blumberg, A.F., and F.L. Hellweger. 2006. Hydrodynamics of the Hudson River Estuary. Pages 1-19 in J. Waldman, K. Limburg, and D. Strayer, (eds.), Hudson River Fishes and Their Environment. American Fisheries Society Symposium 51. Cronin, T. M., R. Thunell, G. S. Dwyer, C. Saenger, M. E. Mann, C. Vann, and R. R. Seal. 2005. Multiproxy evidence of Holocene climate variability from estuarine sediments, eastern North America: Paleoceanography 20:1-21. Green, O.R. 2001. A manual of practical laboratory and field techniques in paleobiology: Klumer Academic Publishers. Boston. Levinton, J.S., and J.R. Waldman, eds. 2006. The Hudson River Estuary. Cambridge University Press. McHugh, C, S.F.Pekar, N Christie-Blick, B.F. Ryan, S. Carbotte, and R. Bell. 2004. Spatial variations in a condensed interval between estuarine and open-marine settings: Holocene Hudson River estuary and adjacent continental shelf: Geology 32.2:169-172. Meade, Robert H. 1969. Landward transport of bottom sediments in estuaries of the Atlantic coastal plain: Journal of Sedimentary Research 39:222-234. Nelson, T. 2011. Establishing a Platinum Group Element (PGE) Method of Sediment Chronology While Tracing New York Harbor’s Industrial Past: 1910-1982: Columbia College Senior Thesis. Pekar S.F., C.M. McHugh, N. Christie-Blick, M. Jones, S.M. Carbotte, and R.E. Bell. 2004. Estuarine processes and their stratigraphic record: paleosalinity and sedimentation changes in the Hudson Estuary: Marine Geology 209:113-129.
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Ralston, D.K., J.C. Warner, W.R. Geyer, and G.R. Wall. 2013. Sediment transport due to extreme events: The Hudson River estuary after tropical storms Irene and Lee: Geophysical Research Letters 40:5451-5455. Ramsey, A. T. S. 1974. Coccoliths: production, transportation and sedimentation: Marine Micropaleontology 1:65-79. Ritchie, J. C., and J.R. McHenry. 1990. Application of radioactive fallout cesium-137 for measuring soil erosion and sediment accumulation rates and patterns: a review: Journal of Environmental Quality. 19:215-233. Ruddiman, W. F. 1969. Historical stability of the Gulf Stream, foraminiferal evidence: Deep Sea Research and Oceanographic Abstracts 15:137-148. Tolderlund, D. S., and A.W. Be. 1971. Seasonal distribution of planktonic foraminifera in the western North Atlantic: Micropaleontology 17:297-329. Traykovski, P., R. Geyer, and C. Sommerfield. 2004. Rapid sediment deposition and fine‐scale strata formation in the Hudson estuary: Journal of Geophysical Research: Earth Surface (2003–2012) 109:1-20. Weiss, D., K. Geitzenauer, and F.C. Shaw, 1978, Foraminifera, diatom and bivalve distribution in recent sediments of the Hudson Estuary: Estuarine and Coastal Marine Science 7:393-400. Woodruff, J., R. Geyer, C., Sommerfield, and N. Driscoll. 2001. Seasonal variation of sediment deposition in the Hudson River estuary. Marine Geology 179:105-119.
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SEA LEVEL RISE AND SEDIMENT: RECENT SALT MARSH ACCRETION IN THE HUDSON RIVER ESTUARY
A Final Report of the Tibor T. Polgar Fellowship Program
Troy D. Hill
Polgar Fellow
School of Forestry and Environmental Studies Yale University New Haven, CT 06511
Project Advisor:
Shimon C. Anisfeld School of Forestry and Environmental Studies Yale University New Haven, CT 06511
Hill, T. D. and S. C. Anisfeld. 2015. Sea Level Rise and Sediment: Recent Salt Marsh Accretion in the Hudson River Estuary. Section II: 1-32 pp. In D.J. Yozzo, S.H. Fernald and H. Andreyko (eds.), Final Reports of the Tibor T. Polgar Fellowship Program, 2013. Hudson River Foundation.
II - 1
ABSTRACT
Salt marshes occupy a narrow elevation range within the intertidal zone. If they are to remain viable, marshes must rise at a rate commensurate with sea level. To understand the prospects for accommodating future sea level rise, this research asks whether vertical accretion in New York City salt marshes has kept pace with rising sea levels over the past two centuries, and how the resultant disparities have changed the flooding stress experienced by the marshes.
Sediment cores were retrieved from marshes on Staten Island, Jamaica Bay, and the Bronx, NY. Age-depth models were calibrated using multiple independent age markers; 210Pb, 137Cs, and total Hg content. Subsamples from depth increments were combusted to determine mineral and organic content. Water level loggers were deployed at each site and used to model historic tide data based on the 150-year tide data record for
New York City.
These data show that marsh surface accretion has been variable in time and space, but has overwhelmingly lagged behind sea level rise over the past two centuries. This de- coupling has led to a substantial reduction in the elevation of the marsh surface relative to sea level, and has dramatically increased the frequency and duration of flooding experienced by the marshes, although there is evidence of a rebound of marsh accretion in response to increased flooding. Contributions of mineral and organic sediment deposition to total accumulation rates are examined within the context of changing inundation regimes. These results suggest that an uncertain future awaits many coastal wetlands in the northeastern US.
II - 2
TABLE OF CONTENTS
Abstract ...... II-2
Table of Contents ...... II-3
List of figures and tables ...... II-4
Introduction ...... II-5
Methods...... II-7
Results ...... II-13
Discussion ...... II-25
Conclusions ...... II-27
Acknowledgements ...... II-29
References ...... II-30
II - 3
LIST OF FIGURES AND TABLES
Figure 1 – Annual mean sea level in New York City, 1856-2012 ...... II-6
Figure 2 – Map of coring locations ...... II-9
Figure 3 – Depth profiles of bulk density and organic matter ...... II-14
Figure 4 – Depth profiles of organic carbon and nitrogen ...... II-15
210 Figure 5 – Depth profiles of Pbxs activity ...... II-17
Figure 6 – Rates of vertical accretion and mass accumulation ...... II-19
Figure 7 – Fractional mineral accumulation rates ...... II-20
Figure 8 – Flooding frequency and duration curves ...... II-22
Figure 9 – Relative elevation of marsh surface, 1880-present ...... II-24
Table 1 – Quality control data for chemical analyses ...... II-16
Table 2 – Selected characteristics of sediment cores and 210Pb profiles...... II-17
Table 3 – Changes in hydroperiod between 1900 and 2012 ...... II-24
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INTRODUCTION
As urbanization proceeded in the New York City region, salt marshes suffered substantial losses due to dredging and filling (Niering and Bowers 1961). Salt marshes that survived the urban expansion of the early twentieth century have since been protected from direct destruction, in recognition of the ecological services they provide, including carbon storage, storm surge mitigation, and the provision of habitat for commercially important fauna (Boesch and Turner 1984; Chmura et al. 2003). Although dredging and filling activities have largely ceased, marshes are still affected indirectly by human activities.
Accelerated sea-level rise and altered sediment availability are two important and enduring human impacts on coastal wetlands. Sea level rise records for the Hudson River suggest average rates of 1.3 mm∙yr-1 from 4 kybp (thousand years before present) until the late 1800s (Engelhart and Horton 2012). During the last 150 years, the instrumental record shows that rates of sea level rise in New York City have averaged 2.8 mm∙yr-1
(Donnelly et al. 2004; PSMSL 2013). In the coming century, the rate of sea level rise in the New York City region may increase by an order of magnitude or more (Horton et al.
2010). It is unclear whether the region’s salt marshes will be able to accommodate dramatic increases in sea levels. Indeed, it is uncertain how marshes have responded to the moderately accelerated sea level rise observed over the past 150 years.
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Figure 1. Instrumental record of annual mean sea level in New York City, 1856-2012. Regression line slope is 2.8 mm∙yr-1 (Data source: PSMSL 2013).
To avoid flooding stress and maintain their spatial extent, marshes must rise vertically to accommodate rising sea levels. This is accomplished through the capture of inorganic tidal sediment and the production of belowground organic material (Redfield
1972; Reed 1995). Many empirical models of marsh accretion treat belowground production as a constant (e.g., Morris et al. 2002; Temmerman et al. 2004), emphasizing the importance of sediment supplied by tidal waters.
Sediment supply to the Atlantic coastal zone has varied dramatically over time
(Stevenson et al. 1985; Stevenson et al. 1988; Kirwan et al. 2011). Outside of the
Hudson River Valley, European colonization led to extensive landscape clearing, erosion, and increased sedimentation in many coastal areas. This trend eventually shifted and sediment supply declined as forests re-grew and rivers were dammed, severing the link
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between watershed erosion and coastal deposition, a shift to which salt marshes may still be responding (Kirwan et al. 2011).
However, Hudson River tidal wetlands show a sedimentation pattern that bears little resemblance to broader trends in sediment supply. Prior to European settlement, sedimentation rates in Hudson River marshes varied from 0.5 mm∙yr-1 in northern marshes to 3.0 mm∙yr-1 in the southern Piermont marsh (Pederson et al. 2005; Sritrairat et al. 2012). Sedimentation rates appear to have declined with colonization (Peteet et al.
2007), only to reach historic highs in the last century – accreting as much as 7.8 mm∙yr-1
(Sritrairat et al. 2012).
It remains an open question whether marshes on the seaward end of the estuary reflect similar trends, and there is reason to suspect they may not. In highly developed coastal environments such as NYC, sediment supplied by coastal erosion may be severely limited by seawalls and other forms of shoreline hardening.
The present research explores two research questions:
(1) Are coastal marshes in NYC keeping pace with local rates of sea level rise?
(2) How have deposition rates of inorganic material (IM) and organic material
(OM) changed over the twentieth century?
METHODS
Research sites
This research was conducted at three salt marshes in New York City: Saw Mill
Creek on Staten Island, Pelham Bay Park in the Bronx, and Spring Creek Park, part of the
Jamaica Bay wetlands complex in Queens (Fig. 2). In this report, references to Staten
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Island, the Bronx, and Jamaica Bay refer, respectively, to these research sites. All three sites are owned and managed by the New York City Department of Parks and Recreation.
The Hudson River is the dominant local source of natural freshwater to the region.
The Bronx marsh is separated from NY/NJ Harbor by Hell Gate and is near the
Hutchinson River. Jamaica Bay receives nearly all of its freshwater inputs from four major sewage treatment plants, which collectively discharge approximately 106 m3∙day-1 into the Bay (Beck et al. 2009). The Staten Island marsh is adjacent to the Arthur Kill waterway, an industrialized shipping lane and the endpoint for a number of small rivers, although wastewater discharges (approximately 105.7 m3∙day-1) also make up a large proportion of the freshwater inputs to Arthur Kill (Burger 1994).
Salinity at the sites averaged 27 psu at the Staten Island marsh, and 32-33 psu at the Bronx and Jamaica Bay sites. Tidal range varies from ~1.4 m at the Staten Island site to ~1.5 meters at Jamaica Bay and ~2.1 m at the Bronx marsh. Vegetation at the Staten
Island and Bronx sites is dominated by Spartina patens and Distichlis spicata, with a creekbank fringe of tall-form Spartina alterniflora. The Jamaica Bay site is dominated by tall-form S. alterniflora, with smaller patches of S. patens and D. spicata.
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Figure 2. Map of New York City region. Coring locations are indicated.
Sediment core retrieval
At each site, exploratory coring was conducted to understand subsurface features and locate an area of peat that was without obvious physical or hydrologic perturbations.
After finding suitable locations, sediment cores were collected from areas of high marsh
(D. spicata or S. patens) with a coring tool made from 15 cm diameter PVC pipe.
Sediment compression was quantified by measuring vertical displacement of the marsh surface following insertion of the coring tool.
Marsh surface elevations at coring locations were measured in National Geodetic
Vertical Datum of 1929 (NGVD29) using surveying equipment. RTK-GPS was used to establish local benchmarks in NGVD29, and a total station was used to measure relative
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elevations within sites. Vertical accuracy of RTK-GPS, used to determine the absolute elevation of a benchmark at each site, is conservatively estimated at 25 mm. The total station used to measure relative elevations of coring locations, water level loggers, and benchmarks, has 2 mm vertical accuracy.
Physical and chemical measurements
In the laboratory, cores were sectioned into 1.5-2 cm intervals for physical and geochemical analysis. Each core section was weighed before and after drying to 60oC, and was then pulverized in a Spex 8000 ball mill. Sub-samples from each core section were re-dried and weighed at 60oC, 105oC, followed by combustion at 500oC to determine organic content by loss on ignition. Mineral content was measured as the ash- free dry weight of a sample. Moisture content was determined by mass loss between field moisture and 105oC. Bulk density was measured as the moisture-free dry mass
(105oC) per unit volume for each core section.
Organic carbon (Corg) and total nitrogen content were measured on a
ThermoFinnigan CHN analyzer following digestion with 6 N HCl. Total Hg content was measured using a Direct Mercury Analyzer (DMA80). Replicates, blanks, and standard reference material were run with every 10 samples for Corg, N and Hg analysis.
Radionuclide activity
Activity of 210Pb, 214Pb, 137Cs, and 7Be were measured by gamma ray spectrometry using a low-background Ge detector. Activities were measured at energies of 46.5, 352.7, 661.7, and 477.2 keV, respectively, and corrected for detector efficiency
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and self-absorption. Dried, ground samples were sealed in 115 cm3 cans and equilibrated for 21 days before analysis. This equilibration period allows supported 210Pb (210Pb produced by in-situ decay of 226Ra, measured indirectly as 214Pb) to be distinguished from
210 210 210 unsupported or excess Pb ( Pbxs, the Pb activity due to atmospheric deposition of
210Pb).
210Pb has a half-life of 22.3 y, making it ideal for studying depositional processes over the past century. 210Pb has limited utility in dating sediment older than about five half-lives, where less than 3% of initial activity is presently detectable.
To estimate sediment accretion rates, constant rate of supply (CRS) models
210 (Oldfield and Appleby 1984) were applied to the Pbxs data. The sediment age, tx, for each depth, x, is calculated as:
1 Q0 푡푥 = ln (1) 휆 Qx
210 210 where λ is the Pb decay constant, Q0 is the Pbxs inventory for the entire core,
210 and Qx is the Pbxs inventory below depth x. Sediment accretion rates for a depth interval, i, are then calculated as:
xi − xi−1 푟푖 = (2) ti − ti−1
CRS models assume that atmospheric supply of 210Pb is constant, and as a
210 consequence, variations in Pbxs activity between adjacent sediment layers reflect differences in sedimentation rates.
137Cs is an anthropogenic radionuclide derived from atmospheric nuclear weapons testing. Peak deposition occurred in 1963 and provides an unambiguous horizon of
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known age in sediments, suitable as an independent check on 210Pb-derived sediment dates (Appleby 2008). The deepest sediment layer with detectable 137Cs activity is sometimes used to indicate the 1954 horizon, when atmospheric nuclear weapons testing began. The 1954 horizon was not used in the present study because of high uncertainty.
This uncertainty stems from the low initial 137Cs activity expected in a potential 1954 horizon combined with decay over the subsequent 60 years, which has reduced detectable activity to 25% of the initial level. As a result, 1954 137Cs horizons are likely to exhibit an upward bias.
Hg concentrations offer an additional independent check on sediment dates. Peak
Hg pollution occurred from approximately the 1950s-1970s in the NYC region
(Varekamp et al. 2003). Hg peaks are less reliable than 137Cs because, while 137Cs reflects a global-scale signal, local industry and pollution trends can impart an undetected bias to Hg depth profiles.
Sediment accretion rates were then used to estimate mass accumulation rates:
푀퐴푅푖 = 푟푖 × 휌푖 × [푐푖] (3)
-2 -1 where MARi is the mass accumulation rate (g ∙ cm ∙ yr ) for a depth interval i; ri
-1 is the corresponding sediment accretion rate (cm ∙ yr ), 휌i is sediment bulk density (g
-3 -1 sediment ∙ cm ), and [ci] is the concentration of a constituent of interest (e.g., g C ∙ g sediment).
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Hydrology
Water levels were measured by deploying pressure transducers (Solinst®
Levelogger LTC) at each of the three marshes. Transducers recorded pressure, temperature, and conductivity at 6 minute intervals for at least two months at each site.
Pressure measurements were corrected for atmospheric pressure to yield water depth measurements, and these were verified by manual field measurements. Surveying techniques were used to convert water depths to elevations in NGVD29.
RESULTS
Bulk density and organic matter profiles
Bulk density profiles (Fig. 3) show that these marshes generally have low density peat, consistent with field observations. The Bronx and Jamaica Bay cores have spikes in sediment density, although the depth of the peak varies from 9 cm in the Bronx to 34 cm in Jamaica Bay. Bulk density in the Staten Island core is homogenous.
Organic matter (OM) depth profiles (Fig. 3) were less variable than bulk density profiles. All three cores have slight subsurface increases in OM, followed by gradual declines with depth. In the Bronx and Staten Island cores, OM content subsequently rises again deep in the core. In the Bronx core, OM content near the base of the core is even higher than near the surface. These observed OM profiles are consistent with field observations of abundant OM throughout the cores, including fibrous rhizomes readily identifiable to species even at the base of the sediment cores.
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Figure 3. Depth profiles of bulk density (left side) and organic matter concentration (right side).
Elemental analysis
All of the cores exhibited two distinct peaks in organic carbon (Corg) content in the top 25 cm (Fig. 4). Peaks ranged from 25% - 40% Corg by mass. Corg content declined to 15-20% at depth in the Staten Island and Jamaica Bay cores, but in the Bronx core the
Corg concentrations deep in the core were comparable to levels observed in the near- surface peaks.
Nitrogen content was slightly elevated in the top ~10 cm of all cores. Staten
Island had the highest surface N concentration, nearly 3% N by mass. As with Corg, N declines with depth in the Staten Island and Jamaica Bay cores, but in the Bronx N content returned to levels seen at the marsh surface.
II - 14 Hg showed subsurface peaks of 0.5, 11.4, and 2.3 ppm in the Bronx, Staten
Island, and Jamaica Bay, respectively. Surface concentrations ranged from 0.1-2 ppm in the cores. Although background levels were not reached in the Jamaica Bay core, the other cores both declined to 0.02 ppm at depth.
Quality control data are presented in Table 1. Standard reference materials were consistently very close to expected values, and replicated samples showed close alignment. Replicates include samples duplicated within and among instrument runs, and reflect data quality on both scales.
Figure 4. Organic carbon (left three panels) and nitrogen (right panels) depth profiles.
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Standard reference material Coefficient of variation between
Element recovery replicates
C 99 ± 0.3% (63) 1 ± 0.1% (39)
N 100 ± 0.7% (63) 1 ± 0.1% (39)
Hg 102 ± 2% (128) 5 ± 1% (68)
Table 1. Quality control data for chemical analyses. Values reported are means ± 1SE (sample size in parentheses). Replicates include inter- and intra-run replicates.
210Pb, 137Cs, and sediment accretion rates
210 The Pbxs inventories for the Jamaica Bay and Bronx marshes were comparable to inventories from elsewhere in the region (Cochran et al. 1998) and those expected from local atmospheric deposition atmospheric 210Pb deposition (530 ±100 mBq∙cm-2;
210 Table 2; Turekian et al. 1983). Staten Island’s Pbxs inventory was 41% lower than the atmospheric inventory.
210 In all cores, Pbxs declined to undetectable levels within 30 cm of the surface
(Fig. 5). The Bronx core had a roughly log-linear decline whereas the other cores have significant departures from log-linearity (Fig. 5).
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Marsh surface Marsh surface Percent of 210 Pbxs inventory 210 elevation (m elevation (m rel. to expected Pbxs (mBq∙cm-2) Site NGVD29) mean high water) inventory
Bronx 1.598 0.106 439 82%
Staten Island 1.287 0.007 317 59%
Jamaica Bay 1.102 -0.017 563 106%
Table 2. Selected characteristics of sediment cores and 210Pb profiles.
210 Figure 5. Depth profiles of Pbxs activity. Note log scale x-axes.
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As an independent check on CRS accretion estimates, the 1963 137Cs peak was compared to the 210Pb-derived estimate for the 1963 layer. All cores showed excellent agreement between the two estimates. Peak Hg concentrations in the sediment profiles of the Bronx and Staten Island marshes also occur during the 1950s-1960s, within the expected date range (Varekamp et al. 2003). Based on the elevated Hg concentrations observed at the Staten Island and Jamaica Bay sites, local point sources of Hg are believed to be important at these sites and the chronology of Hg contamination may not reliably reflect the regional trend, despite the apparent alignment at Staten Island.
The CRS models show that accretion rates have varied spatially and temporally
(Fig. 6). All three marshes accreted sediment at a rate of 0.5-1 mm∙yr-1 from the late nineteenth century until approximately 1950, when Jamaica Bay began experiencing a sustained period of enhanced accretion. Accretion at Jamaica Bay peaked at 8.8 mm∙yr-1 around 1950, and subsequently remained between 2.5-5 mm∙yr-1. Sediment accretion at the Bronx marsh slowly increased to 1.7 mm∙yr-1 while the Staten Island marsh experienced a period of enhanced accretion from 1985-present.
Bulk sediment mass accumulation rates (Fig. 6) reflect the same patterns observed in sediment accretion rates. This similarity appears due to the relatively small variations in bulk density compared with those in sediment accretion rates.
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Figure 6. Rates of vertical accretion (top) and bulk sediment mass accumulation (bottom), over time. Mineral accumulation has varied over time in its contribution to total accumulation (Fig. 7). None of the marshes have experienced a simple linear trend with time; each has oscillated between mineral and organic dominance. In all cores, the past
II - 19 15-25 years have been a period of increasing mineral dominance, preceded by a 30-40 year period where mineral accumulation declined relative to organic material.
Figure 7. Mineral accumulation rates as a proportion of total accumulation rates.
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Hydrology and marsh accretion balance
Water level loggers captured 89 tidal cycles in Jamaica Bay, 120 in Staten Island, and 256 in the Bronx. These water level data were used to generate estimates of flooding duration and flooding frequency for each marsh, and to examine how those parameters vary with elevation. Water level data are used to estimate flooding of the marsh platform; the area of the marsh where sediment cores were taken and elevations measured.
The percent of time that a given elevation spends submerged is referred to as its flooding duration. This metric was calculated from the empirical cumulative distribution function (ECDF) for the full record of 6-minute water level measurements (Fig. 8).
Using these datasets, the three marshes surveyed in the Bronx, Staten Island, and Jamaica
Bay are submerged 8%, 11%, and 17% of the time, respectively.
Flooding frequency, the percent of high tides that flood a given elevation, was generated by isolating high tide heights from the period of water level logger deployment.
Local high tide heights were correlated with data from nearby harmonic NOAA stations; the Battery, NY was used for Staten Island and Jamaica Bay, and the Bridgeport, CT
NOAA station was used for the Bronx. The linear models relating local high tides to those at nearby NOAA stations were all highly significant (p < 10-16) and had strong explanatory power (R2 > 0.94). The resultant relationships were used to model local high tides over the entirety of 2012 (the most recent full calendar year of tidal data).
These year-long high tide datasets were used to calculate mean high water
(MHW) and flooding frequency as a function of elevation (Fig. 8). MHW was simply the average of all high tides in the dataset. The ECDF of each high tide dataset characterized the relationship between elevation and flooding frequency. These data show that in 2012,
II - 21 the three marshes in the Bronx, Staten Island, and Jamaica Bay were flooded by 32%,
50%, and 53% of high tides, respectively.
Figure 8. Relationship between vertical elevation and flooding frequency (black curves) and duration (red curves). Vertical elevations are relative to site-specific MHW in 2012. Black vertical lines show relative marsh surface elevations in 1900; blue vertical lines show relative elevations in 2012. Panel A: Bronx; panel B: Staten Island; panel C: Jamaica Bay.
The relative marsh surface elevation is defined as the difference between the elevation of the marsh surface and local mean high water (Table 2). The marsh platform in the Bronx had the highest relative elevation, sitting 11 cm above current MHW (Fig.
8). At Staten Island and Jamaica Bay, marshes were substantially lower, just 0.7 and -1.7 cm relative to current MHW.
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Annual mean sea level (MSL) data from NOAA’s tide station at the Battery, NYC
(1856-present) were used to characterize the rate of relative sea level rise. These trends include both eustatic sea level rise and region-wide subsidence following deglaciation.
The MSL trend over the entire period of record (2.8 mm∙yr-1) was used to estimate the elevation of MHW in the past. MHW data were not directly used to characterize rates of relative sea level rise because MHW data began being recorded in 1920 and do not cover the full period of 210Pb data. During periods where both MSL and MHW were recorded, the two datums increase at identical rates.
The elevation of MHW at some time in the past, MHWt2, was calculated as:
푀퐻푊푡2 = 푀퐻푊푡1 − (푟푠푙푟 ∙ (푡1 − 푡2)) (4)
Where MHWt1 is the mean high water elevation (m) at time point 1; rslr is the rate
-1 of relative sea level rise (assumed constant at 0.0028 m ∙ yr ); and t1 and t2 are time points of interest (units of years). Historic relative marsh surface elevations were calculated by comparing the vertical position of MHW (calculated from equation 4) with the vertical elevation of the dated core section corresponding to the time point of interest.
Although Jamaica Bay was the only marsh found to presently be below MHW, all sites have experienced declines in relative marsh surface elevation over time (Fig. 9). In
Jamaica Bay, the period of rapid accretion starting around 1950 (Fig. 6) occurred shortly after the marsh surface fell below MHW, and allowed the marsh to temporarily recover to an elevation just above MHW. Similarly, accretion at Staten Island began accelerating as the elevation of the marsh platform approached MHW in the 1980s. In the Bronx, the marsh surface elevation has not yet approached MHW, and accretion rates have remained low.
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The declines in relative marsh surface elevations have led to corresponding increases in flooding frequency and duration at all sites. Between 1900 and 2012, study sites in the Bronx and Staten Island experienced an approximate doubling of the frequency and duration of flooding. The Jamaica Bay site, already a relatively wet system in 1900, became ~25% wetter by 2012 (Table 3; Fig. 8).
Figure 9. Relative elevation of marsh surface, 1880-present.
Flooding frequency Flooding duration Absolute change
Site 1900 2012 1900 2012 Frequency Duration
Bronx 11% 32% 3% 8% 21% 5%
Staten Island 27% 50% 5% 11% 23% 6%
Jamaica Bay 43% 53% 13% 17% 10% 4%
Table 3. Changes in hydroperiod between 1900 and 2012.
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DISCUSSION
Research question (1) asks whether marshes in NYC are keeping up with sea level rise. At all sites, the relative marsh surface elevations have declined over the past century, indicative of accretion deficits. Integrating over the past century, these marshes have not risen at the same rate as sea levels. Losses in relative elevation have resulted in dramatic increases in the frequency and duration of tidal flooding.
But while the marshes have failed to maintain their initial relative elevations, there are some positive signs. Marshes in Staten Island and Jamaica Bay lost enough relative elevation to fall below MHW, but they subsequently began accreting at a rate that allowed them to remain at the same level as MHW. The marshes have not recovered or begun gaining ground against rising sea levels, but they may be in equilibrium. One caveat to this interpretation is related to the sea level rise estimates used. The 2.8 mm∙yr-
1 average derived from the 150 year tide record at the Battery may mask slightly higher rates of sea level rise in recent years, although shorter time series have much higher uncertainties. If sea level rise has recently accelerated above 2.8 mm∙yr-1, these marshes would have continued on a trajectory of accretion deficits and increased flooding.
The data presented in this study are generally consistent with those from other marshes in the region. Cochran et al. (1998) found that high marsh accretion rates in six western Long Island Sound marshes approximated or were significantly lower than the rate of sea level rise, and also observed recent accelerations in marsh accretion. A
Narragansett Bay high marsh also had an accretion rate that is slightly lower than the rate of sea level rise used in this study (Bricker-Urso et al. 1989). In Delaware, local rates of
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sea level rise also approximate or exceed marsh accretion rates (Church et al. 1981; Ward et al. 1998).
Some previous research has suggested that, generally, the northeastern US is an area where salt marsh accretion is safely in excess of rates of sea level rise (Stevenson et al. 1985; Reed 1990). This is inconsistent with the data reported in the present study and noted in the preceding paragraph. These incongruous results are likely due to different foci; the present study is focused on accretion of the marsh platform of non-riverine marshes. Low marsh accretion rates tend to be higher than on the marsh platform, although in the northeastern US the low marsh zones tend to be marginal and very limited in spatial extent (i.e., not representative of the system as a whole). Similarly, marshes along river margins tend to have higher accretion rates than marshes exposed to the more diffuse sediment supply of the coastal zone (Ward et al. 1998).
For marshes that are not keeping up with sea level rise, two extreme outcomes are possible. One possibility is that the process forms a negative feedback loop and self- regulates (Reed 1990; FitzGerald et al. 2008); greater flooding leads to increased sediment deposition and the marsh surface accretes rapidly enough to maintain or increase its relative elevation, resulting in a stable system. A second possibility is that marshes are unable to capture enough sediment to sufficiently increase their accretion rate. In this case, the high marsh community would transition to a Spartina alterniflora monoculture, which may continue to submerge until conditions become too stressful to support vegetation. These trajectories are complex, but the supply of tidal sediment to a marsh is an important determinant of a system’s resilience to sea level rise (Kirwan et al.
2010).
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Research question (2) asked how deposition of mineral material on the marsh surface has changed since 1900. Trends in mineral deposition can best be described as variable and bi-directional. The proportion of total accumulation attributable to mineral sediment is lower than it was in the middle of the twentieth century, but is also substantially higher than it was twenty years ago.
To an extent, especially at the lower, wetter sites (Jamaica Bay and Staten Island) the variations shown in Fig. 7 may be associated with greater tidal flooding. These variations may also reflect broader changes in sediment dynamics and nearby land uses.
Similarities between cores in the trends and the timing of inflection points (Fig. 7) suggest that landscape-scale factors are important drivers of mineral sediment availability. Determining specific regional-scale drivers of changes in sediment availability would be a useful extension of this work, and would raise the possibility of managing the watershed and harbor in ways that might help preserve coastal ecosystems.
CONCLUSION
Sediment cores from three marshes in NYC were used to create a detailed record of marsh response to sea level rise and changes in rates of mineral sediment deposition.
During the past century these marshes have all lost elevation relative to sea level. This has altered flooding regimes at the sites, and changes in hydroperiod were quantified.
Marshes that fell below MHW began accreting rapidly and were able to maintain their relative elevations as a result. This demonstrated ability to respond to sea level rise is encouraging, but is not decisive proof that these systems are entirely resilient.
Threshold rates of sea level rise may exist, and more work on sediment supply and
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ecosystem-scale vegetation and geomorphological changes will help discern the true resilience of these systems in the face of sea level rise.
The magnitude of sea level rise projections suggest that sea level rise is an existential threat to salt marshes. As projections of accelerated sea level rise become increasingly confident, coastal resource managers will benefit from an understanding of how marshes respond and what interventions might increase their viability. The present work contributes to this understanding and suggests that the fate of NYC salt marshes is not a foregone conclusion.
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ACKNOWLEDGMENTS
The author thanks Shimon Anisfeld for his constant mentoring and support. The author also thanks Andrew Kemp, Simon Engelhart, Ben Horton, Chris Vane, David
Yozzo, Ellen Hartig, and Jim Lynch, for helpful discussions and/or assistance with fieldwork. This research was funded by a 2013 Polgar Fellowship administered by the
Hudson River Foundation.
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Kirwan, M. L., A. B. Murray, J. P. Donnelly, and D. R. Corbett. 2011. Rapid wetland expansion during European settlement and its implication for marsh survival under modern sediment delivery rates. Geology 39:507-510.
Morris, J. T., P. V. Sundareshwar, C. T. Nietch, B. Kjerfve, and D. R. Cahoon. 2002. Responses of coastal wetlands to rising sea level. Ecology 83:2869-2877.
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NUTRIENT POLLUTION IN HUDSON RIVER MARSHES: EFFECTS ON GREENHOUSE GAS PRODUCTION
A Final Report of the Tibor T. Polgar Fellowship Program
Angel Montero
Polgar Fellow
School of Earth and Environmental Sciences Queens College, City University of New York Flushing, NY 11367
Project Advisors:
Brian Brigham and Gregory D. O’Mullan School of Earth and Environmental Sciences Queens College, City University of New York Flushing, NY 11367
Montero, A., B. Brigham, and G.D. O’Mullan. 2015. Nutrient Pollution in Hudson River Marshes: Effects on Greenhouse Gas Production. Section III: 1- 26 pp. In D.J. Yozzo, S.H. Fernald and H. Andreyko (eds.), Final Reports of the Tibor T. Polgar Fellowship Program, 2013. Hudson River Foundation.
III-1 ABSTRACT
The Hudson Valley has experienced rapid development and urbanization over the
last century. As a result, the release of untreated sewage from Combined Sewer
Overflows (CSOs) continues to be of concern as a nutrient loading vector to the environment. The majority of nutrient pollution in the Hudson River Estuary is transferred to the lower Hudson, including Iona Island Marsh, from New York City through tidal forcing. Wetland systems are primarily composed of anaerobic sediment that is regulated by different energy constraints than more well-studied terrestrial systems. A nutrient addition incubation experiment was performed with Iona Island wetland soil that measured the production of carbon dioxide and methane over a two- week period in response to different combinations of carbon and nitrogen additions. The
addition of carbon, in the form of acetate, to incubated wetland soils was found to cause
significantly increased production of carbon dioxide and methane, both potent
greenhouse gases. In contrast, nitrogen only additions, in the form of nitrate or ammonium, did not result in significantly increased greenhouse gas production compared
to the no nutrient addition control treatment. These results suggest that CSO releases into
the lower Hudson River Estuary are likely to stimulate increased pulses of both carbon
dioxide and methane from Hudson marshes and provide added rationale to more tightly
manage anthropogenic carbon release into the estuarine environment.
III-2
TABLE OF CONTENTS
Abstract ...... III – 2
Table of Contents ...... III – 3
Lists of Figures and Tables ...... III – 4
Introduction ...... III – 5
Methods...... III – 7
Results ...... III – 12
Discussion ...... III – 19
Conclusions ...... III – 21
Acknowledgements ...... III – 23
References ...... III – 24
III-3
LIST OF FIGURES AND TABLES
Figure 1 – Map of Sampling Site ...... III – 8
Figure 2 – Rate of CO2 production during incubation ...... III – 14
Figure 3 – Total CO2 production ...... III – 15
Figure 4 – Rate of CH4 production during incubation ...... III – 16
Figure 5 – Total CH4 production ...... III – 17
Figure 6 – Final Redox concentration ...... III – 18
Table 1 – Nutrient concentrations added per treatment ...... III – 10
III-4
INTRODUCTION
Wetlands are being subjected to increasing pollution from anthropogenic
phenomena (Bianchi 2011). The Hudson valley has seen rapid development and
urbanization over the last century, and it hosts one of the largest population densities in
the United States (US Census Bureau 2011). Restoration efforts following the Clean
Water Act of 1972 have done much to improve water quality in the Hudson (Brosnan et al. 2006); however, the input of nutrient pollution in the form of carbon (C) and nitrogen
(N) remains significant (Howarth et al. 2006). The majority of nutrient pollution to the lower Hudson, including Iona Island Marsh, originates in New York City and is transported through tidal forcing (Griffith and Raymond 2011; Yoon and Raymond
2012). In New York City, and many other local riverfront communities, street runoff and raw sewage are combined into a single sewer system. During periods of high precipitation, a portion of the water from sewers containing combined untreated sewage and street runoff are expelled directly into the Hudson as Combined Sewer Overflows
(CSOs), a matter of ongoing concern as a vector of nutrient loading in the environment
(Griffith and Raymond 2011).
Marshes are a major interface between terrestrial and aquatic habitats, providing
crucial ecosystem services. For example, marshes are believed to play a critical role in
filtering toxins from watersheds and providing a buffer to coastal flooding during major
storm events (Barbier et al. 2011). Estuaries and associated marshes are also important
fishery habitats and act as a nursery to many species of fish (Limburg et al. 2006).
Therefore, the wetland loss observed in recent decades has been a matter of concern in
coastal New York. For example, Jamaica Bay, NY, has experienced a steady increase in
III-5
salt marsh fragmentation and conversion to mud flats and pools since the 1950s (Hartig et
al. 2002). Small plot – scale experimental manipulations (Deegan et al. 2002; Turner et
al. 2009) and marsh-scale studies (Deegan et al. 2012) have shown that nutrient
enrichment can lead to changes in plant physiology leading to decreased formation of
below ground roots, consequently increasing erosion and wetland loss (Morris and
Bradley 1999; Turner et al. 2009); however, less work has been conducted to describe
the role microbial communities have in changing wetland dynamics following nutrient
enrichment (Bowen et al. 2009).
Wetland systems are primarily composed of anaerobic sediment that is regulated
by different energy constraints than more well-studied terrestrial systems (Reddy and
DeLaune 2008). In anaerobic sediments, microbial communities are forced to use alternative electron acceptors, and subsequently alternative metabolic pathways to
generate energy (Conrad 1996); however, these alternative acceptors have lower redox
potentials than oxygen and their reduction generates less energy per mol of organic
material degraded (Canfield et al. 2004; Thauer et al. 1977). The energy constraints of
oxygen-deprived microbial communities lead to the buildup of recalcitrant C pools in
anaerobic zones as organic C remains energetically unavailable (Bridgham et al. 2006).
Nutrient additions to anaerobic soil could trigger the activation of microorganisms that
promote increased greenhouse gas (GHG) emissions such as carbon dioxide (CO2) and
methane (CH4) and possibly facilitate soil C loss (Blagodatskaya and Kuzyakov 2008).
Therefore, elucidating the mechanistic interactions between nutrient additions, microbial
activity, and GHG production is crucial in order to understand the consequences of
sewage release into local waterways and more effectively manage estuaries.
III-6
The objective of this experiment was to quantify the microbial response to C and
N addition by measuring the production of CO2 and CH4 in anaerobic soil slurries
exposed to different C and N treatments. It was hypothesized that C additions to soil
slurries would increase the decomposition of organic matter measured by an increase in
both CO2 and CH4 rates, while N addition would not stimulate additional GHG
production.
METHODS
Overview. The experiment aimed to quantify soil anaerobic microbial response to C and N sources, and the subsequent production of the GHGs; CO2 and CH4.
Sediment cores and overlying water were collected from Iona Island Marsh, NY and
utilized to created soil slurries in one-liter Mason jars. A nutrient addition incubation
experiment was performed that measured the production of CO2 and CH4 in response to
different combinations and forms of C and N over a two-week period. Following the
incubation, units were broken down and extracted water and soil slurry samples were
collected for future biogeochemical and molecular analysis (beyond the scope of this fellowship report).
Site description. Iona Island marsh (Figure 1), located 60 km north of New York
City, was chosen for its proximity to Queens College, and its relatively low salinity
concentrations. In World War II it served as a U.S Navy ammunition depot. In 1965, it
was donated to the Palisades Interstate Park Commission. Today, it is one of four
wetlands composing the Hudson River National Estuarine Research Reserve. It is a
sanctuary for many bird species and marine and aquatic organisms like ducks and crabs,
which were observed during field excursions there. III-7
Figure 1a. (above) A map of Iona Island with
the sampling location circled in red.
Figure 1b. (left) A map of the lower Hudson River estuary marking the location of Iona Island.
III-8
Field sampling phase. Soil samples were collected from eight sites with an
average distance of 10 meters from each other. Overlying water (20 liters) was collected
in close proximity to collection sites. Sediment cores were collected from each site with
an AMS specialty wetland soil corer equipped with plastic core liners. The corer was
driven into the ground, rotated clockwise 90 degrees, and slowly pulled up to extract a
core. The cores varied from approximately 18 - 20 cm in length and were 5 cm in diameter. The core liners were then capped and stored in a collection bag at ambient temperature. Once all sites were visited, the soil cores were stored in an ice-cooled chest,
while the overlying water was stored in containers at ambient temperature.
Preparatory phase. Upon arrival in the laboratory, the material collected was
immediately processed and prepared for incubation. Soil cores were removed from liners
and placed at 4°C in Ziploc bags. Before soil slurry unit fabrication, individual cores
from each site were mixed together into homogenous slurry. The overlying water was
filtered (0.22 micron Sterivex filter), autoclaved, degassed, and refrigerated (4°C). To
prepare Mason jar units, approximately 105 g of wet soil was weighed. A small sample
of approximately 5 g was removed from this mass, weighed, and stored for soil C and N
analysis. The remainder (100 g) was weighed, and inserted into the Mason jar. The jar
was then filled with approximately 200 g of processed overlying water and shaken at low
rpm for 10 minutes. All the units were then flushed with nitrogen using a one hour
procedure to remove oxygen from both the headspace and water of the experimental
units. Seven replicate units from each site were constructed, for a total of 56 Mason jar
units. A single replicate from each site was randomly assigned to one of seven treatment
groups. The groups consisted of (Table 1): pre-treatment (PT); negative control (NT);
III-9
nitrate addition (NO3); ammonia addition (NH4); acetate addition (A); acetate + nitrate
(A+NO3); and acetate + ammonia addition (A + NH4).
Pre-incubation phase. After preparing Mason jar units, a two–week incubation
period was completed before nutrients were added. Units were stored in a temperature
controlled incubation chamber at 25°C for both the pre-incubation and the incubation period following nutrient addition. CO2 headspace concentrations were measured in the
pre-incubation units to monitor microbial activity. At the conclusion of the pre-
incubation, indicated by complete reduction in CO2 production rate, nutrients were added
as noted below (Table 1) to simulate expected nutrient loads following storm events
(Griffith and Raymond 2011). The amount added was approximately ten-fold higher than the daily rate modeled by Griffith and Raymond 2011 simulating expected nutrient loading during storm events (Yoon and Raymond 2012).
Immediately following the nutrient addition, all units were flushed to eliminate any accumulated GHGs before the incubation began.
Treatment Type Vol (mL) Nutrient Concentrations No Treatment (NT) 10 deionized water
Nitrate (NO3) 10 potassium nitrate to final concentration of 3.57 mmol N
Ammonium (NH4) 10 ammonium chloride to final concentration of 3.57 mmol N Acetate (A) 10 sodium acetate to final concentration of 16.67 mmol C potassium nitrate + sodium acetate for final concentrations of 3.57 mmol N/16.67
Acetate + Nitrate (A + NO3) 10 mmol C ammonium chloride + sodium acetate for final concentrations of 3.57 mmol 10 N/16.67 mmol C Acetate + Ammonium (A + NH4)
Table 1. Nutrients added to treatment groups with C and N concentrations per L of volume (wet soil sample + overlying water volume.)
Nutrient phase. During the main incubation phase, both CO2 and CH4
headspace concentration were measured in regular time intervals, every 24 hours and
every 48 hours, respectively. CO2 was measured with an EGM-4 (IRGA) environmental
III-10
gas monitor and CH4 was measured with a Hewlett-Packard 5890 Series plus II Gas
Chromatograph (GC) installed with a Flame Ionization Detector. A 60 ml plastic BD
syringe was used to extract gas samples from units. To extract a headspace sample, the
syringe was inserted through the unit’s septum, the headspace was then mixed by gently
pulling the syringe’s plunger up to 50 ml and back to 0 two times. The desired volume of
headspace sample was then extracted. During days when both CO2 and CH4 were
measured, 30 ml of headspace was extracted from the unit. From that 30 ml, 5 ml was
initially expelled, 10 ml was injected into the GC and analyzed for CH4, and the
remaining 15 ml inserted into the EGM-4 and analyzed for CO2. On days when only
CO2 was measured, only 15 ml of the unit’s headspace sample was extracted and
analyzed. The GC and EGM-4 were both flushed (air and N2 respectively) with gas after
each unit was sampled. The volume of gas extracted from the experimental unit’s headspace was replaced with nitrogen after sampling.
Breakdown. At the end of the pre-nutrient (pre-treatment units only) and nutrient phases the Mason jar units being sampled were broken down, and slurry-water mix tested for pH, redox, and salinity (Sensorex reference electrodes) with the Micro Observatory sensor system (Analytical Instrument Systems). Each unit was first shaken for ten minutes. Before sampling, the unit slurries were inverted three times to homogenize the contents. The units were then opened, and probes inserted into the slurry mix to be analyzed. After probe-measurements, soil and water samples were vacuum filtered, collected, and frozen for future biogeochemical measurements.
Statistics. Data from the end point of experimental incubations were analyzed
using software from the R project for statistical computing (www.r-project.org). Analysis
III-11
of Variance (ANOVA) was used to test for differences in the mean and if significant
differences were detected among groups, a post-hoc Tukey’s Range Test was used to adjust p-values for multiple comparisons to identify the pairs of experimental units with significantly different means.
RESULTS
Greenhouse gas production. CO2 production rates increased in carbon-treated
groups (A+NH4, A+NO3, A), most notably in days 8, 9, and 10, after which production
rates leveled off (Figure 2). A temporary decrease in CO2 accumulation was observed in
nd the A+NO3 group in the 2 day of the experiment. Total CO2 production differed significantly (p < 0.01) among treatments (Figure 3) with greater production found in C addition treatments. Only A (p<0.01), A+NO3 (p<0.01), and A+NH4 (p<0.01) treatment groups were significantly different from the control (NT) treatment. In contrast, N additions did not have a significant effect on CO2 production when compared to the no
addition control. In C–treated groups, an average of 460 µg C/ g of dry soil accumulated
as CO2 was produced. In contrast, an average of 229 µg C/ g of dry soil accumulated as
CO2 was produced in the NO3, NH4, and NT groups.
CH4 production was observed in C addition treatments but was not measurable in
N only and control treatments (Figure 4). Final CH4 differed among groups (ANOVA, p<0.01) (Figure 5); however, only A (p < 0.05) and A+NH4 (p < 0.01) treatments were
significantly different from the no addition control (NT) treatment. In C–treated groups,
an average of 206 µg C/ g of dry soil accumulated as CH4 was produced. In contrast, an
average of 3 µg C/ g of dry soil accumulated as CH4 in the NO3, NH4, and NT groups.
III-12
In total, 66% percent (µg C / g of dry soil) of gaseous C measured was in the
form of CO2, while CH4 made up 33% of gaseous C measured. C + N treatment groups
accumulated 3.3 times as much C compared to treatments with N only addition. CO2
comprised 51% of total C accumulation in acetate, 53% in A + NH4, 73% in A + NO3,
and >99% for all other treatments. CH4 comprised 49% of total C accumulation in
acetate, 47% in A + NH4, 27% in A + NO3, and < 1% for all other treatments.
Probe measurements Redox measurements conducted at the end of the incubation period demonstrate that the most significant reduction of experimental units occurred in C treated groups (Figure 6). Significant differences were detected among groups (ANOVA, p<0.01). The A (p<0.01), A+NH4 (p<0.01), A + NO3 (p<0.01), and
NO3 (p<0.01) treatments differed significantly from the no addition control treatment.
There were significant differences in pH among groups (p < 0.01). The measured
pH for the C–treated groups varied from 7 – 7.5. In contrast, the pH measured in N only
and control groups varied from 6.5 to 7.0. Only A (p < 0.01), A+NO3 (p < 0.01) and
A+NH4 (p<0.01) treatments significantly differed from the no addition control (NT).
The salinity levels measured in the field varied from 0.5 to 2.0 ppt.
III-13
Figure 2. CO2 production rate for the treatment groups through the two – week incubation period. The rate is measured as the amount µg C produced as CO2 since the last measurement point. The values were adjusted for slight variations in temperature and pressure, and normalized to the rate of C production per gram of dry soil weight present in the unit.
III-14
Figure 3. Box plot demonstrating the amount of C mineralized as CO2 at the end of the experiment. Concentrations were adjusted for slight temperature and pressure differences, and normalized to total µg C produced per gram of dry soil. Significant differences from the control are marked with an asterisk and groups that do not show significant differences among samples are designated with a letter.
III-15
Figure 4. CH4 production rate from the treatment groups through the two – week incubation period. The rate is measured as the amount µg C produced as CH4 since the last measurement point. The values were adjusted for slight variations in temperature and pressure, and normalized to the rate of C production per gram of dry soil weight present in the unit.
III-16
Figure 5. Box plot demonstrating amount of C mineralized as CH4 at the end of the experiment. Concentrations were adjusted for slight temperature and pressure differences, and normalized to total µg C produced per gram of dry soil. Note log scale of axis. Significant differences from the control are marked with an asterisk and groups that do not show significant differences among samples are designated with a letter.
III-17
Figure 6. Box plot showing redox condition at the end of experiment. Significant differences from the control are marked with an asterisk and groups that do not show significant differences among samples are designated with a letter.
III-18
DISCUSSION
Nutrient additions, similar to concentrations that would be expected in post-storm
nutrient pulses, were found to have variable influence on GHG production, depending on
the chemical composition of the addition. The addition of acetate, with or without
ammonium or nitrate, was found to cause significant increases in both CO2 and CH4
production in wetland soil slurries. Nutrient cycling in wetlands is generally thought to be controlled by the availability of C and N sources to produce energy for metabolism and growth (Conrad 1996). In the wetland system studied, anaerobic microbes use
alternative electron acceptors in lieu of oxygen; however, the energy available through
these compounds is very low in comparison to aerobic respiration (Thauer et al. 1977).
As a consequence, microbial growth is often sluggish in anaerobic environments as they are starved for energy and unable to utilize the recalcitrant C pool present. It appears that microbial communities in these soils were limited mainly by energy in the form of easily
degradable C. It is therefore not surprising that the redox potential in these treatments
was also found to decrease and that, under these reduced redox conditions, CH4 was produced in substantial quantities relative to the treatments lacking labile C additions and characterized by higher redox potential.
In contrast, the addition of nitrate or ammonium alone was not sufficient to increase GHG production from marshes as compared to the no addition control, suggesting that N limitation is not a major factor acting to suppress anaerobic metabolism, nor did it cause a reduction in redox potential of the experimental units by stimulating anaerobic metabolism. The addition of nitrate can also act as a favorable electron acceptor in anaerobic systems, and was observed to increase redox potential, as
III-19 expected, suggesting that denitrification might become a dominant energy producing pathway in this treatment. When adequate electron donors are available, microbial communities can be differentiated by their preference of electron acceptors that vary in metabolic efficiency. In the absence of aerobic respiration, due to the lack of oxygen, denitrification is the next most efficient catabolic process, followed by iron reduction, sulfate reduction, and lastly methanogenesis, which generally occurs only in highly reduced conditions (Conrad 1996). Microbial activity would be expected to be dominated by metabolic pathways providing the highest available energy yield, and as favorable electron acceptors become depleted, microbial activity would shift according to the redox condition. In saline environments, sulfate reduction has been found to produce ten times more energy than other metabolic activities (Howarth and Teal 1980) due to the high concentration of sulfate. Presence of sulfates has also been found to hamper methanogenic activities (Martens and Berner 1974; King and Wiebe 1980). Iona Island is typically fresh (0.5 - 2.0 ppt) suggesting that denitrification and methanogenesis would be expected to be important pathways utilizing available organic C; however, in this experiment the addition of N alone did not result in increased CO2 or methane production, indicating that the system was primarily limited by availability of electron donors, not electron acceptors. In addition, it is expected that the lower Hudson area already receives ample N from wastewater sources (Brosnan et al. 2006), which would make N widely available in marshes such as Iona Island.
These findings are important to interpret in the context of sewage and other nutrient pollution sources. Sewage contains high levels of labile C that, based on these findings, would be expected to increase GHG production from marshes in both the form
III-20 of CO2 and CH4, while also reducing the redox potential in impacted marsh soils, causing an increase in the relative utilization of C for anaerobic respiration. N can be released into the estuary either from sewage or other sources such as fertilizer usage. It appears that the addition of N may have little direct impact on GHG production in Iona Marsh sediment; however, the N pollution sources could still have indirect impacts on marsh soils by stimulating primary production in the estuary and thereby resulting in the addition of labile C to anaerobic marsh soils (Bianchi 2011). Management activities, such as reduction in CSOs and improved efficiency of wastewater treatment, that reduce
N pollution, and especially those that reduce C pollution, would be expected to reduce
GHG production from Iona Island marsh and similar wetland systems.
CONCLUSIONS
These results highlight the importance of labile C as a mediator of GHG production. This is importantly noted, as previous studies in northeastern U.S.A have largely focused on the role of N impacts on wetland habitats (Deegan et al. 2002). There has been little previous work on the combined effects of C and N on microbial communities in the Hudson River Estuary. The results from this experiment show that C has a larger effect on microbial activity in Iona Island marsh than N. The results show that C should be a more closely monitored element in water quality research and steps should be taken to minimize watershed exposure to C from anthropogenic sources.
Methanogenesis was shown to be strongly influenced by C addition. This is important to consider as CH4 is thought to have 25 times the warming potential of CO2 over a hundred year period (IPCC 2007). Studying the ways CH4 is produced and how
III-21 anthropogenic activities affect CH4 production is vital to help mitigate the effects of climate change. CSO releases would be expected to stimulate pulses of CH4 from
Hudson marsh environments. Coupled with the increasing vulnerability of this fragile system to climate change and increased storm events, a closer look at the intricate relationship between anthropogenic pollution and the Hudson’s health is warranted in order to better direct ecosystem conservation and restoration efforts.
III-22
ACKNOWLEDGEMENTS
We would like to thank the Hudson River Foundation for supporting this project and members of the Polgar Committee for their valuable feedback to improve this report.
In addition, we would like to thank Dr. Jeffrey Bird and members of the Bird laboratory at Queens College for their assistance in designing and conducting this project.
III-23
REFERENCES
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Bianchi, T.S. 2011. The role of terrestrially derived organic carbon in the coastal ocean: A changing paradigm and the priming effect. Proceedings of the National Academy of Sciences 49: 19473-19481.
Blagodatskaya, E., and Y. Kuzyakov. 2008. Mechanisms of real and apparent priming effects and their dependence on soil microbial biomass and community structure: critical review. Biology Fertility Soils 45:115-131.
Bowen, J.L., B.C. Crump, L.A. Deegan, and J.E. Hobbie. 2009. Salt marsh sediment bacteria: their distribution and response to external nutrient inputs. The ISME Journal 3:924-934.
Bridgham, S.D., J.P. Megonigal, J.K. Keller, N.B. Bliss, and C. Trettin. 2006. The carbon balance of North American wetlands. Wetlands 26:889-916.
Brosnan, T.M., A. Stoddard, and L.J. Hetling. 2006. Hudson river sewage inputs and impacts: Past and present. Pages 340–341 in: J.S. Levinton and J.R. Waldman, editors. The Hudson River Estuary. Cambridge University Press, New York.
Canfield, D. E., K.B. Sørensen, and A. Oren. 2004. Biogeochemistry of a gypsum- encrusted microbial ecosystem. Geobiology 2:133-150.
Conrad, R. 1996. Soil microorganisms as controllers of atmospheric trace gases (H2, CO2, CH4, OCS, N2O, and NO). Microbiological Reviews 60:609-640.
Deegan, L.A., J.E. Hughes, and R.A. Rountree. 2002. Salt marsh ecosystem support of marine transient species. Pages 333-365 in: M.P. Weinstein, and D.A. Kreeger, editors. Concepts and Controversies in Tidal Marsh Ecology. Springer Netherlands 333-365.
Deegan L.A., D.S. Johnson, R.S. Warren, B.J. Peterson, J.W. Fleeger, S. Fagherazzi, and W.M Wollheim. 2012. Coastal eutrophication as a driver of salt marsh loss. Nature 490: 388-392.
Griffith, D.R., and P.A. Raymond. 2011. Multiple-source heterotrophy fueled by aged organic carbon in an urbanized estuary. Marine Chemistry 124: 14-22.
Hartig, E.K, V. Gornitz, A. Kolker, F. Mushacke, and D. Fallon. 2002. Anthropogenic and climate-change impacts on salt marshes of Jamaica Bay, New York City. Wetlands 22:71-89
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Howarth, R.W., R. Marino, D.P. Swaney, and E.W. Boyer. 2006. Wastewater and watershed influences on primary productivity and oxygen dynamics in the lower Hudson River Estuary. Pages 121 - 139 in: J.S. Levinton and J.R. Waldman, editors. The Hudson River Estuary. Cambridge University Press, New York.
Howarth, R.W., and J.M. Teal. 1980. The role of reduced inorganic sulfur compounds. The American Naturalist 116:862-872.
IPCC. 2007. Changes in atmospheric constituents and in radiative forcing. Page 212 in: Solomon, S., D. Qin, M. Manning, Z. Chen, M. Marquis, K.B. Averyt, M. Tignor and H.L. Miller, editors. Climate Change 2007: The Physical Science Basis. Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, New York.
King, G. M., and W. J. Wiebe. 1980. Regulation of sulfate concentrations and methanogenesis in salt marsh soils. Estuarine and Coastal Marine Science 10: 215-224.
Limburg K.E., K.A. Hattala, A.W. Kahnle, and J.R. Waldman. 2006. Fisheries of the Hudson River Estuary. Pages 189 – 204 in: J.S. Levinton and J.R. Waldman, editors. The Hudson River Estuary. Cambridge University Press, New York.
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Reddy, K.R., and R.D. DeLaune. 2008. Biogeochemical Characteristics and Carbon. Pages 157-162 in: K.R. Reddy and R.D. DeLaune Biogeochemistry of Wetlands. CRC Press, Boca Raton.
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Yoon, B., and P. A. Raymond. 2012. Dissolved organic matter export from a forested watershed during Hurricane Irene. Geophysical Research Letters 39: L18402
III-26
MICROBIAL AGENTS OF CONCERN DETECTED IN WATER AND AIR AT THE HUDSON RIVER ESTUARY WATERFRONT
A Final Report of the Tibor T. Polgar Fellowship Program
Sherif Kamal
Polgar Fellow
Department of Environmental Studies Parsons The New School for Design 66 Fifth Avenue, New York, NY 10011
Project Advisor:
M. Elias Dueker School of Earth and Environmental Sciences Queens College, City University of New York Flushing, NY 11367
Kamal, S. and M. E. Dueker. 2015. Microbial Agents of Concern in Water and Air at the Hudson River Estuary Waterfront. Section IV: 1-27 pp. In D.J. Yozzo, S.H. Fernald and H. Andreyko (eds.), Final Reports of the Tibor T. Polgar Fellowship Program, 2013. Hudson River Foundation.
IV-1
ABSTRACT
Despite building evidence of the transfer of bacteria from contaminated water surfaces to coastal aerosols, the presence of antibiotic resistant bacteria and sewage indicating bacteria in the air in urban waterfronts has not been sufficiently studied. This represents an understudied route of pathogen exposure to human populations in proximity to sewage-contaminated surface waters. To bridge this gap, this study deployed new methodologies to detect and taxonomically identify these microbial agents of concern in the lower Hudson River Estuary (HRE). Aerosol and water surface samples were collected from the Newtown Creek Nature Walk, which is adjacent to a highly-urbanized tributary of the lower HRE, to assess whether microbial agents of concern were present and detectable. Antibiotic resistant bacteria and sewage indicators were detected in both water and air on each sampling day at this site. This occurred despite the fact that viable bacterial fallout was lower at this site than other previously studied waterfronts on the
HRE. The percentage of antibiotic resistance in bacterial aerosols (~20%) was significantly higher than in surface water bacteria (~8%). Molecular analysis of the 16S rRNA gene of viable bacteria sampled revealed a diverse bacterial community with a wide range of possible sources, including water, land, and sewage. Many of the resistant aerosols and water bacteria were members of genera known to contain human pathogens, including Massilia, Pseudomonas, and Roseomonas. These findings greatly expand the potential public health implications for sewage contamination in the urban coastal environment, adding aerosol exposure as a potential route for human contact with sewage-associated microbial agents of concern.
IV-2
TABLE OF CONTENTS
Abstract ...... IV-2
Table of Contents ...... IV-3
List of Figures and Tables...... IV-4
Introduction ...... IV-5
Methods...... IV-8
Results ...... IV-12
Discussion ...... IV-19
Conclusion ...... IV-22
Acknowledgments...... IV-23
References ...... IV-24
IV-3
LIST OF FIGURES AND TABLES
Figure 1: Microbial fallout at different HRE waterfront locations...... IV-14
Figure 2: Percentage of antibiotic resistance in water vs. air ...... IV-15
Figure 3: Microbial fallout vs. bacteria in surface waters ...... IV-16
Table 1: Meteorological context at Newtown Creek Nature Walk ...... IV-12
Table 2: Surface water quality at Newtown Creek Nature Walk...... IV-13
Table 3: Enterococcus faecalis in waterfront aerosols ...... IV-14
Table 4: Taxonomy of antibiotic-resistant bacteria found in water and air ...... IV-17
Table 5: Preliminary source analysis of antibiotic-resistant bacterial aerosols .. IV-18
IV-4
INTRODUCTION
Over 39% of the US population now lives near a coastline, and wastewater
inputs to rivers, estuaries and marine waters have been increasing (Niemi et al. 2004;
NOAA 2013). Raw sewage releases contain many pathogenic bacteria (Tourlousse et al.
2008; Korzeniewska et al. 2009) and antibiotic resistant bacteria (Kim et al. 2010;
Young et al. 2013), creating potential public health risks for human contact with these
waters. Direct contact with contaminated waters (ingestion, skin contact) is known to
lead to illnesses including gastroenteritis, conjunctivitis, and wound infections (US EPA
Office of Water 2012); however, contact through inhalation of aerosols created from
contaminated waters is an additional and understudied human exposure pathway.
Aerosols can be created from water surfaces by surface disruption such as wind-wave interactions, wave-shore interactions, and the recreational use of water (Monahan et al.
1983; Blanchard 1989; de Leeuw et al. 2000). This results in the movement of water surface materials to the air, which may include viable bacteria, as established in previous studies of coastal environments including the Hudson River Estuary (HRE) (Aller et al.
2005; Dueker et al. 2012a; Dueker et al. In Review).
This water-air movement of materials creates the opportunity for airborne human exposure to microbial agents of concern, including sewage pathogens and antibiotic-
resistant bacteria, when raw sewage is present in urban surface waters. Research
conducted at Louis Valentino Pier (LVP) in Red Hook, Brooklyn showed that the
number of culturable bacteria in the air correlated with the culturable bacteria residing
in surface waters (Dueker et al. In Review). Untreated sewage released to estuarine and
coastal environments gets restricted to the surface in a density-stratified layer.
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Therefore, any surface disruption of contaminated waters may cause the release of raw sewage materials into the atmosphere as indicated by previous research conducted at
Newtown Creek (Brooklyn, NY) (Dueker et al. 2012a).
Enterococci are routinely used as an indicator of human and animal fecal waste
in HRE surface waters and are associated with microbes that are of serious public
health concern (US EPA Office of Water 2012). Although Enterococcus presence in
HRE surface waters has been closely monitored, there is no current literature outlining
their presence in HRE waterfront air. Several studies have investigated the negative
public health prospects of aerosol exposure to aerosolized sewage microbes in relation
to sewage treatment plants (Woodcock 1955; Muscillo et al. 1997; Brandi et al. 2000;
Carducci et al. 2000; Radke 2005; Baertsch et al. 2007; Heinonen-Tanski et al. 2009;
Korzeniewska et al. 2009); however, the presence of these bacteria and their dominant
source in urban aerosols is a novel field of research.
Over the recent past the public has become increasingly alarmed by new
scientific findings of connections between the overuse of antibiotics in both medicine
and the agriculture–agrifood industry and the environmental emergence and spread of
antibiotic-resistant bacteria (Nikiforuk 1996; Feinmen 1998; Levy 1998;
Khachatourians 2013). Microbial resistance to antibiotics is on the rise. With fewer
new chemotherapeutic agents coming onto the market, the problem of microbial
resistance to drugs already in use has become a crisis in health care (Jungkind et al.
1995; US Office of Technology Assessment 1995). Previous studies conducted on
antibiotic resistant microbes in aerosols have been restricted to indoor hospital
environments (King et al. 2013; Muzslay et al. 2013) and agricultural environments
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(Liu et al. 2012; Schulz et al. 2012), where antibiotic resistant microbes are found to travel in viable state through the air.
Based on a recent study, high concentrations of Enterococci (indicating the presence of raw sewage) in HRE surface waters were directly related to concentrations of ampicillin and tetracycline-resistant bacteria in surface waters (Young et al. 2013).
Given the high frequency of raw sewage release to HRE water (Riverkeeper 2011) and the documented transfer of surface water materials to the air in coastal regions (Dueker et al. 2011; Dueker et al. 2012a; Dueker et al. 2012b) the potential exists for aerosolization of antibiotic resistant and sewage-associated bacteria when raw sewage is present. Despite these findings, little to no research has been conducted on the presence and potential sources for antibiotic resistant bacteria in outdoor air.
To address this gap, microbial agents of concern in the air and water were quantified and identified at a highly-urbanized lower HRE waterfront site. The goals of this study were to determine if microbial agents of concern were detectable in urban air and adjacent surface waters and to make a preliminary assessment of possible sources for these agents. The outcome of this study provides a unique dataset confirming the presence of microbial agents of concern both in water and air at the highly-urbanized lower HRE waterfront. These findings highlight the potential for unique pathways of exposure that may connect people to aquatic pollution, via aerosolized microbial agents of concern.
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METHODS
Study Site and Meteorology
Field sampling took place on five full days between 5 June 2013 and 4 July 2013 at the Newtown Creek Nature Walk (NCNW) (40.7368528 N, 73.9464472 W), a public park adjacent to the Newtown Creek Water Treatment Facility. Newtown Creek, once one of the busiest hubs of NYC industrial activity, is a 3.5 mile creek forming the northern and southern borders of Brooklyn and Queens has been heavily industrialized and traveled since the mid 1800’s (US EPA 2010). Newtown Creek was declared an EPA
Superfund site in September 2010 and is known to have frequent sewage contamination
(including sewage-indicators and antibiotic resistant bacteria) of surface waters
(Riverkeeper 2011; Young et al. 2013).
A portable Vantage Pro2 Plus Weather Station (Davis Instruments, Hayward,
CA) was used to measure meteorological conditions including air temperature, wind speed, and relative humidity. Total aerosol particle size and concentration was measured using a stationary Met One 9012 Ambient Aerosol Particulate Profiler (Met One
Instruments, Grants Pass, OR). The profiler was placed about 3 m above water level
(depending on tidal height). During each aerosol sampling event, water surface salinity and temperature were measured using a ThermoScientific Orion Star Portable
Multiparameter Meter (ThermoScientific, Waltham, MA).
Bacterial Aerosols
Culturable bacterial aerosol fallout was measured at the site during onshore wind conditions by exposing triplicate agar plates per media type to ambient aerosols on a
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platform oriented into the wind. This sampling method (Dueker et al. 2011; Dueker et al.
2012a; Dueker et al. 2012b) is not representative of the total concentration of bacteria in
the air, but has the advantage of enumerating bacteria that are viable and able to grow on
the provided media at the time of exposure. Four types of media (each deployed in
triplicate during each exposure event) were used in these exposures: Reasoner´s 2A agar
(R2A) media (Fisher Scientific), R2A amended with ampicillin (50 mg/L) and R2A
amended with tetracycline (10 mg/L) and MEI (Membrane Enterococcus Indoxyl-β-D-
Glucoside) Agar (manufacturer-prepared plates, Molecular Toxicology, Inc.). R2A has commonly been used in past aerosol studies to detect a broad spectrum of bacterial species (Lighthart and Shaffer 1995; Shaffer and Lighthart 1997) and MEI Agar selects for Enterococci, which are bacteria used as indicators of sewage presence as per the EPA
(US EPA Office of Water 2012).
Before each sampling day, 15 exposure plates were prepared for each media type using aseptic technique in a laminar flow hood. For each media type, three plates served as field controls, lab controls, 15-minute field exposures, 30-minute field exposures and
90 minute field exposures. After exposure in the field, all control and exposure plates were incubated for three days at 25 °C in the dark, then colonies growing on antibiotic- amended plates were transferred to fresh, unexposed antibiotic-amended media plates to ensure antibiotic resistance. After five days total incubation, CFUs (Colony-Forming
Unit) on all incubated plates (control, exposure, and transfer plates) were counted and then picked for future molecular analyses by transferring colony material to sterile
HyClone water (ThermoScientific, Waltham, MA). Bacterial aerosol fallout rate (CFU m−2 s−1) for each exposure event was calculated using plate counts, the surface area of the
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exposed petri dishes (0.0079 m2), and the duration of exposure.
Surface Water Bacteria
Surface water was sampled by collecting water in 50-ml centrifuge tubes that
were rinsed three times with sample water prior to collection. Immediately after sample
collection, tubes were kept in the dark and on ice until returned to the lab. To assess culturable bacterial concentrations in surface waters, a series dilution was created at a
1:10, 1:100 and 1:1000 dilution of sample water with autoclaved and filter-sterilized
HRE water. A 100 μL aliquot of each dilution was spread onto triplicate R2A,
R2A+Amp, and R2A+Tet plates in a laminar flow hood. Plates were incubated for five days and CFUs enumerated and picked for molecular analysis as outlined for aerosol plates above.
The presence of sewage-associated bacteria was inferred from the presence of the indicator organism Enterococcus in surface waters and aerosols as outlined by EPA regulations (US EPA Office of Water 2012). Colony growth on MEI selective media exposures was enumerated after aerosol exposure (exposed simultaneously with R2A plates above). After exposure, these plates were incubated in the dark at 41°C for 24 hours, after which CFUs were counted, picked for molecular analysis, and transferred to fresh MEI plates to further confirm media selectivity.
Water-surface samples were processed using the EPA-approved IDEXX
Enterolert system (IDEXX Laboratories, Westbrook, ME) (Riverkeeper 2011; Suter et al.
2011). Briefly, within 6 hours of collection, surface water was diluted 1:10 using autoclaved, filter-sterilized Newtown Creek surface water, and then added to liquid media before being sealed in a Quanti-tray (IDEXX) and being incubated in the dark at
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41°C for 24 hours. After incubation, growth of Enterococci was confirmed and quantified
by exposing the Quanti-tray to UV light and recording the number of wells fluorescing
blue. This resulted in the most probable number (MPN) of Enterococci per ml of sample
water.
Molecular Analyses for Taxonomic Identification of Resistant Bacteria
To gain a preliminary understanding of the types of antibiotic resistant bacteria
present in waterfront aerosols and surface waters, colonies were picked from at least one
antibiotic-amended, aerosol-exposed and water surface media plate per sample date and
suspended in 50 μL of HyClone sterile water. This material was then boiled for five
minutes to lyse cells and frozen until polymerase chain reaction (PCR) analysis was
performed.
The 16S rRNA gene from DNA of lysed cells was amplified using universal
bacterial primers 8F (5′-AGRGTTTGATCCTGGCTCAG-3′) and 1492R (5′-
CGGCTACCTTGTTACGACTT-3′) (Teske et al. 2002). Thermocycling conditions consisted of 35 cycles: 45s denaturation at 94°C, 45s of annealing at 55 ° C, and one min elongation at 72 ° C. After PCR, gel electrophoresis was performed on PCR products to ensure correct length of amplified fragments and that controls did not amplify.
Amplifications were then sent for single-pass Sanger sequencing to SeqWright Inc.
(Houston, TX). The sequence output files were edited using the Geneious software package (www.geneious.com), exported in FASTA format and uploaded to the
Ribosomal Database Project (RDP) webserver (http://rdp.cme.msu.edu/) for alignment and taxonomic classification to the level of genus with 80% confidence unless otherwise noted.
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A GenBank top-hits analysis (Dueker et al. 2012a) was performed on aerosol sequences to assess potential sources for these bacteria. Briefly, sequences were blasted against the GenBank database using Geneious’ Megablast function, and the sequence hit with the highest bit-score was designated the top hit for each sequence. The
environmental source reported for this top hit was then recorded.
RESULTS
During sampling days the mean velocity of onshore winds was very low, at 1.4 ±
0.3 m s-1, the average temperature was 23 ± 1 ºC and average relative humidity was 66 ±
4% (Table 1). Average water temperature was 22 ± 0.5 ºC, and average salinity was 16.9
± 0.4 ppt (Table 2). Enterococci were present in surface waters every sampling day, but
only exceeded EPA standards (EPA threshold for safe water contact = 104 cells/100 ml
for single-sample values) on 4 July 2013, when it exceeded the threshold by an order of
magnitude at 9,104 cells/100 ml (Table 2). These high sewage-indicator concentrations
followed a heavy rain event occurring late afternoon/evening of 3 July 2013 (Table 2).
Table 1. Meteorological context at Newtown Creek Nature Walk on sampling days. Mean of parameter during aerosol exposures, ± 1 standard deviation.
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Table 2. Surface water quality at Newtown Creek Nature Walk on sampling days. Mean of parameter ± 1 standard deviation. Italics, bolded numbers in Enterococcus column indicate that Enterococcus concentrations exceeded EPA standards for single-sample values (104/100 ml).
Microbial fallout on R2A-only (no antibiotics) plates was significantly lower at NCNW in comparison to the English Kills portion of Newtown Creek (Dueker et al. 2012a) and Louis Valentino Pier (Dueker et al. In Review) (Fig. 1). Antibiotic-resistant bacteria (Fig. 2) and sewage-indicating bacteria were detected in water (Table 2) and aerosol (Table 3) samples on each sampling day at this site. Colonies that initially grew on exposed ampicillin plates transferred to new, unexposed R2A+Amp plates with 100% success, confirming ampicillin resistance. The same was true for R2A+Tet plates that were exposed for 15 or 45 minutes, but not for 90 minute exposures. R2A+Tet plates exposed for 90 minutes had < 60% transfer success rates, suggesting that the sun was breaking down the tetracycline during 90 minute exposures. Data and colonies from 90-minute R2A+Tet exposure plates were therefore excluded from further analyses. All colonies growing transferred to new MEI plates with 100% regrowth.
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Figure 1. Microbial fallout on R2A plates (no antibiotics) at Newtown Creek Nature Walk compared to previous studies conducted at Louis Valentino Pier (Dueker et al., In Review) and Newtown Creek English Kills (Dueker et al 2012b)
Table 3. Enterococcus growing on exposed MEI agar plates, with transfer plate (selection confirmation) and molecular analysis results. ND = Not Determined (sequencing not yet performed).
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The percentage of bacterial fallout made up of antibiotic resistant bacteria was significantly higher than the percentage of antibiotic resistant bacteria in the water (p <
0.01 for both R2A+Amp and R2A+Tet) (Figure 2). With the exception of exposures conducted on 4 July 2013, bacterial fallout and surface water concentrations were linearly related for R2A (R2 = 0.75), R2A+Amp (R2 = 0.48), and R2A+Tet (R2 = 0.69) (Fig. 3).
Molecular analysis conducted on colonies grown on MEI plates confirmed that they were formed by Enterococcus faecalis (Table 3). After quality control, a total of 140 sequences from antibiotic resistant bacteria were used in analyses, 17 from aerosol exposures and 123 from surface waters. RDP classification of these sequences revealed a taxonomically diverse community of antibiotic-resistant bacteria in both water and air, with Aeromonas, Acinetobacter, and Acidovorax dominating surface water resistant
Figure 2. Percentage of antibiotic resistance in surface water bacteria and bacterial aerosols (amp = ampicillin, tet=tetracycline). Black bars indiate standard error of geometric mean, letters above bars indicate significant difference (p < 0.01).
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bacteria and Roseomonas and Bacillus dominating viable resistant bacterial aerosols
(Table 4). The GenBank top-hits analysis of the resistant aerosols resulted in a broad range of possible sources, including soil, water, and sewage (Table 5).
Figure 3. Microbial fallout vs. bacterial concentrations in surface waters. R2 and p-value of linear models (excluding 4 July 2013 data) noted.
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Table 4. RDP taxonomic classification (genus level) of antibiotic-resistant bacteria in air and water at the study site
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Table 5. Top-hits analysis of antibiotic-resistant bacterial aerosols sequenced in this study
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DISCUSSION
Microbial agents of concern at the HRE waterfront
This study confirmed the presence of microbial agents of concern both in water
and air at the highly-urbanized NCNW waterfront. Antibiotic resistant and sewage-
indicating bacteria were detected at NCNW in both air and water on each sampling
day. This occurred despite the fact that bacterial fallout at this site was significantly
lower than fallout at other previously-studied HRE waterfront locations (Dueker et al.
2012a; Dueker et al. In Review). This lower fallout was most likely due to differences
in environmental and aerosol production conditions between sites. The Newtown
Creek English Kills site is currently being remediated through underwater bubbling,
which has been shown to increase aerosol particle concentrations over the waterway
(Dueker et al. 2012a). Studies at Louis Valentino Pier (LVP) (Dueker et al. In Review)
were conducted under much higher mean wind speeds than at the NCNW where wind
speeds were < 2 m s-1. Also, LVP is on a very busy harbor, with much more boat
activity (and therefore aerosol production) than at NCNW. The presence of microbial
agents of concern in NCNW aerosols under low fallout conditions strongly suggests
the presence, and possibly higher abundance, of microbial agents of concern at HRE
waterfront sites with higher fallout rates and contaminated surface waters.
Many of the resistant bacterial genera in water represented include potential pathogens, including Aeromonas and Pseudomonas. Almost all of the resistant bacteria were also in genera representing potential pathogens, including Bacillus, Burkholderia,
Comamonas, Massilia, and Roseomonas. Both the taxonomic composition and percentages of antibiotic resistant bacteria in surface waters at this site agree with
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previous research conducted by Young et al. (2013). The presence of these genera in aerosols at the waterfront indicates potential for public health impacts.
Preliminary Assessment of Antibiotic Resistant Aerosol Sources
Previous research at the HRE waterfront and Newtown Creek specifically has shown that both terrestrial and aquatic sources influence the types of viable bacteria present in aerosols (Dueker et al. 2012a; Dueker et al. In Review). Similar source potentials for microbial agents of concern were found in these aerosols, with sewage contamination strongly implicated.
Interestingly, higher percentages of the bacterial aerosols were antibiotic-resistant
(~22%) compared to bacteria found in surface waters (~2-8%) at this site. This indicates different sources for antibiotic-resistant bacterial aerosols and/or selection processes occurring through the aerosolization process. Aerosolized bacterial cells experience extreme environmental stresses including desiccation and increased UV exposure.
Survival of these stresses, particularly under the relatively dry conditions (66% RH) experienced during sampling, may require traits that are selected for during the aerosolization process, favoring viability of certain bacteria over others. Although determination of this is beyond the current study, it is interesting to note that the resistant aerosols detected at NCNW include Bacillus, which can form endospores to survive stressful environments (Leggett et al. 2012), Pseudomonas, which is known to survive aerosolization in indoor environments (Walter et al. 1990), and a Roseomonas sequence that closely matched an organism sampled previously from the Kennedy Space Center clean room floor (accession: EU705114, Vaishampayan et al., unpublished).
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This study also found preliminary evidence of sewage-contaminated waters as a source for microbial agents of concern detected in the waterfront air. The presence of viable Enterococcus faecalis in aerosols is a strong indication of sewage influence on aerosols at the site. Given the site’s proximity to the Newtown Creek Waste Water
Treatment Plant, sources for these organisms include the treatment facility itself.
However, these indicators were also present in surface waters every sampling day. The
GenBank top hits analysis of resistant bacterial aerosols also confirmed that many of the resistant bacterial aerosols had been detected previously in aquatic and sewage- contaminated environments.
Rain is known to introduce untreated raw sewage into the water (Riverkeeper
2011) and as per previous studies there is a linear connection between sewage presence and antibiotic resistant bacteria (Young et al. 2013). Dueker et al. (in Review) found a linear relationship between culturable bacterial aerosols and culturable surface water bacterial concentrations at Louis Valentino Pier, indicating a connection between water and air quality at this site. A similar relationship was observed at NCNW, with the exception of sampling conducted on 4 July, with microbial fallout much lower than expected given the high surface water bacterial concentrations. The outlier of 4 July 2013 may be explained by the fact that it had rained for 4 consecutive days just prior to sampling (Table 2), which would cause spikes in bacterial concentrations of surface waters through subsequent CSO releases.
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CONCLUSION
The outcome of this study provides a unique dataset confirming the presence
of microbial agents of concern both in water and air at the highly-urbanized lower
HRE waterfront. Many of the antibiotic resistant bacteria in water and air at the
NCNW site were members of bacterial genera known to contain human pathogens, indicating potential public health implications for exposure to waterfront aerosols.
Sources for these bacterial aerosols included terrestrial and aquatic environments. The presence of viable Enterococcus faecalis and other aquatic and sewage-associated antibiotic-resistant bacteria in waterfront aerosols indicate that sewage pollution may play an important role in the presence of microbial agents of concern at this site.
Similarities between antibiotic-resistant bacteria found in surface waters at NCNW and Flushing Bay suggest that this study’s findings may be applicable to other sewage- contaminated HRE waterfront sites. Further research into sources for these aerosolized microbial agents of concern is merited.
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ACKNOWLEDGMENTS
This report was made possible by generous support from the Hudson River
Foundation. Special thanks to the Tibor T. Polgar committee for study guidance and feedback, Gregory O’Mullan for his participation in study design and use of his CUNY
Queens College lab, and to Roman Reichert and Angel Montero for laboratory analysis training and support.
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OCCURRENCE AND ECOLOGICAL EFFECTS OF AMPHETAMINE TYPE STIMULANTS IN WASTEWATER EFFLUENT
A Final Report of the Tibor T. Polgar Fellowship Program
Alexis M. Paspalof
Polgar Fellow
School of Natural Resources University of Nebraska-Lincoln Lincoln, NE 68508
Project Advisors:
Daniel Snow School of Natural Resources University of Nebraska-Lincoln Lincoln, NE 68508
Emma Rosi-Marshall Cary Institute of Ecosystem Studies Millbrook, NY 12545
Paspalof, A.M., D. Snow, and E. Rosi-Marshall. 2015. Occurrence and Ecological Effects of Amphetamine Type Stimulants in Wastewater Effluent. Section V: 1-24 pp. In D.J. Yozzo, S.H. Fernald and H. Andreyko (eds.), Final Reports of the Tibor T. Polgar Fellowship Program, 2013. Hudson River Foundation.
V-1 ABSTRACT
The presence of illicit drugs and their metabolites has become an increasingly
important topic worldwide, and their potential ecological effects are essentially unknown.
Some of the most interesting of these illicit compounds are the amphetamine type
stimulants (ATSs). Very little research has been conducted in the Hudson River Valley
(HRV) to determine the presence of various pharmaceuticals, let alone illicit drugs. This
report provides information from the summer of 2013 as to the presence of ATSs and
several other pharmaceutical compounds at six different sites in the HRV. Amphetamine
(AMPH) was detected at one site, the outflow of Kingston’s wastewater treatment plant
(WWTP). The detection at a combined sewer overflow (CSO) point indicates the release
of ATSs into the HRV, but lack of detects suggests these compounds are subjected to
high dilution and therefore persist at concentrations below current technologies limit of
detection. The ability to derivatize and quantify dopamine and other catecholamines in
algae is discussed in this report. This is in response to possible biological effects ATSs
have on the presence of catecholamines such as dopamine within algae. Detection by
high performance liquid chromatography (HPLC) coupled with fluorescence was first
evaluated, but determined to lack the sensitivity desired for this research. Currently a
method utilizing derivatization with N-Methyl-N-trimethylsily Trifluoroacetamide
(MSTFA) and detection using gas chromatography mass spectrometry (GC-MS) is being developed. This research lays the framework for artificial stream experiments that will be conducted in the spring of 2014 in which chronic exposure of biofilms will be monitored for potential biological effects.
V-2 TABLE OF CONTENTS
Abstract ...... V-2
Table of Contents ...... V-3
List of Figures and Tables...... V-4
Introduction ...... V-5
Methods...... V-8
Results ...... V-13
Discussion ...... V-17
Acknowledgements ...... V-20
References ...... V-21
V-3 LIST OF FIGURES AND TABLES
Figure 1- Structures of amphetamines ...... V-8
Figure 2 - Structures of catecholamines...... V-9
Figure 3 - Map of sampling sites ...... V-10
Figure 4 – GeopumpTM used during sampling ...... V-11
Figure 5 – Extent of pharmaceutical contamination at sample sites ...... V-14
Figure 6 – Kingston WWTP CSO outflow ...... V-15
Figure 7 – Ion chromatograms of silyl derivatives ...... V-16
Table 1 – Concentrations of pharmaceuticals at sampling sites ...... V-15
V-4 INTRODUCTION
The use of pharmaceuticals to combat against commonly known diseases has
been on a steady climb over the years. Until recently, there has been no sign that the use
of pharmaceuticals has had anything but a positive impact on the world; however, there is
mounting evidence indicating that pharmaceuticals excreted from human waste,
improperly disposed, or released through manufacturing processes are occurring in
combined sewer overflows (CSOs) and wastewater treatment plant (WWTP) effluent
(Ternes 1998; Zuccato et al. 2000; Heberer 2002; Castiglioni et al. 2006a; Zuccato et al.
2008). Though there has been extensive research indicating the presence or absence of
various pharmaceuticals present in aquatic systems, the effects of these compounds on
aquatic organisms such as algae and invertebrates are presently unknown (Pal et al.
2012).
Currently most research in wastewater effluent has focused on evaluating the use
and misuse of illicit amphetamines, such as methamphetamine (METH) and AMPH,
through the occurrence of these compounds in aquatic systems. A majority of these
studies have been conducted in various areas of Europe, though several rivers in the
United States have been investigated (Zuccato et al. 2008; Huerta-Fontela et al. 2008;
Castiglioni et al. 2006b; Jones-Lepp et al. 2004). Few studies in the Hudson River Valley
(HRV) have examined the presence of pharmaceuticals (Kolpin et al. 2002; Yu et al.
2006; Palmer et al. 2008; Cantwell et al. 2010; Li and Brownawell 2010; Reiner and
Kanna 2011; Liao et al. 2012 ). None have focused on the presence of illicit ATS.
Stimulant drugs affect dopamine, a chemical associated with pleasure, and other catecholamine levels in the brain of higher-level organisms. Other pharmacological
V-5 effects include loss of appetite as well as increased wakefulness and attentiveness. These
desired biological effects have led to the increased abuse of stimulant medications in the
treatment of diseases such as attention deficit hyperactivity disorder (ADHD) and
narcolepsy. Since these medications are comprised of similar chemicals as many illicit
substances, they are now on a list of abused substances worldwide (CDC 2005; Kessler et
al. 2006; Biederman and Faraone 2005). An example is AdderallTM; this ADHD
medication is a 3:1 mixture of d- and l-enantiomers of amphetamine salts (Cody et al.
2003). Based upon the increases in both medicinal and illicit uses, there is cause to
speculate that the metabolism and excretion of them and related medications, contribute
to the increased release of stimulants to various aquatic environments across the globe.
Previous work on illicit amphetamines has focused on the origin of these
contaminants in wastewater and their persistence in surface waters (Ternes 1998; Zuccato
et al. 2000; Heberer 2002; Jones-Lepp et al. 2004; Castiglioni et al. 2006a; Castiglioni et
al. 2006b; Huerta-Fontela et al. 2008; Zuccato et al. 2008). To date, there has been very
little research directed to the understanding of potential long-term effects that may occur
in organisms exposed to WWTP effluent and how much can be contributed to occurrence
of stimulants. A recent study demonstrated that effluent-influenced concentrations of a specific benzodiazepine anxiolytic drug alter the behavior of European Perch (Perca fluviatilis) (Brodin et al. 2013).
While the effects of stimulants on aquatic organisms are not currently understood
(Pal et al. 2012), the literature does provide background to support expected effects.
Catecholamines have been found in 44 plant families in which they normally serve a protective role against predators, injuries, and detoxification of nitrogen (Kuklin and
V-6 Conger 1995). Recent studies have revealed that naturally produced dopamine in the
marine green algae Ulvaria obscura, compromising 4.4% of the algae’s dry weight, was
related to decreased grazing by sea urchins (Strongylocentrotus droebachiensis) (Van
Alstyne et al. 2006). Dopamine is a monoamine neurotransmitter that many believe to be regulated by amine stimulants in higher organisms. If exposure to trace amounts of stimulants results in a change of natural dopamine production in aquatic organisms
(similar to what occurs in mammals) this suggests potential pathways toward altering microorganism metabolism and productivity, as well as ecosystem function. Exposure of algae to these dopamine-regulating substances may alter their natural production of dopamine, if it is present.
The first major goal of this study was to measure the concentrations of stimulants in the HRV. It is hypothesized that ATSs will be present in the watershed due to the high population density surrounding the sampling sites and large number of combined sewer overflows (CSOs) that occur in the area. The second goal was to collect and process algae samples to quantify dopamine and other catecholamine concentrations. These samples will test the hypothesis that chronic exposure to amphetamines increases the presence of dopamine and other catecholamines in algae. This fieldwork will lay the framework for later artificial stream experiments that will determine whether these stimulants influence algae and algal consumers (e.g., stream invertebrates).
V-7 METHODS
INVESTIGATED COMPOUNDS
Pharmaceuticals
The compounds researched in this study have been reported in previous research studies monitoring effluent at several sites across both North America and Europe (Jones-
Lepp et al. 2004; Castiglioni et al 2006a; Castiglioni et al 2006b; Huerta-Fontela et al.
2008; Kasprzyk-Hordern et al. 2008; Postigo et al. 2008; Bartelt-Hunt et al. 2009;
Bijlsma et al. 2009). This study focused on substituted amphetamines such as AMPH,
METH, 3,4-Methylenedioxymethamphetamine (MDMA), 3,4-
Methylenedioxyamphetamine (MDA) and others (Figure 1).
METH AMPH
MDMA MDA
Figure 1. Structures of amphetamine type stimulants being researched in this study.
Water samples were also processed for several compounds that are not within the amphetamine stimulant family. These compounds include: acetaminophen (a common pain reliever), caffeine, 1,7-dimethylxanthine (a metabolite of caffeine), carbamazepine
V-8 (used to treat seizures), cimetidine (used in the treatment of ulcers), cotinine (a metabolite
of nicotine), diphenhydramine (DPH) (Benadryl), morphine (a common pain reliever),
sulfadimethoxine (a sulfonamide antibiotic), sulfamethazine (a sulfonamide
antibacterial), sulfamethoxazole (a sulfonamide bacteriostatic antibiotic) and
thiabendazole (a fungicide and parasiticide).
Catecholamines
Three catecholamines were measured in algae: dopamine, epinephrine, and
norepinephrine (Figure 2). The expected levels of catecholamines in algae are virtually
unknown, but Van Alsytne et al. (2006) extracted dopamine from the green algae Ulvaria
obscura suggesting the potential of occurrence in other species. Catecholamine salts used for calibrations and standards were purchased from Sigma.
Epinephrine Norepinephrine Dopamine
Figure 2. Structures of catecholamines.
STUDY SITES
The main stem of the Hudson River was sampled at three sites June 4, 2013
(Figure 3). These sites included Albany, Castleton and Coxsackie. The goal was to assess
the levels of pharmacetucials in the Hudson where efflent occurs and collect algae
samples for dopamine analysis.
Originially three additional sites were to be tested (Hudson, Kingston,
Poughkeepsie), but due to weather and equipment issues, samples were only collected
V-9 from the three northern sites. All six sites and an additional two sites, Fort Montgomery
and Haverstraw Bay, are a part of a long term research project directed by the Cary
Institute of Ecosystem Studies. This group has water quality data dating back over twenty
years from a study that has focused on the impacts of the zebra muscle Dreissena
polymorpha (Strayer et al. 2004), making their sites perfect locations for the study of
pharmaceutical analysis.
Three other sites in surrounding tributaries of the Hudson River were sampled in
June 2013 (Figure 3 D-F). These sites were chosen based upon collaboration with the
Hudson River Keeper, and coordinated with
several sampling sites they visit on a
regular basis to monitor the prevalence of
the sewage indicating bacteria Enterococci.
One location on the Esopus, Catskill, and
Rondout Creeks was visited with the goal
of again determining the extent to which
these areas are polluted with amphetamines
and to collect algal samples.
SAMPLING Figure 3. Locations of the six sampling sites visited within the HRV: Study sites: Hudson River A - Albany, B – Castleton, C – Coxsackie, At each sampling site, a single D – Catskill Creek, E – Esopus Creek, water sample and six algae samples were F – Kingston WWTP collected. Water samples were filtered in
the field using a GeopumpTM (Figure 4). The water was collected directly off the boat and
V-10 then placed in an amber glass jar for analysis at a later date. At these sites, algae was sampled directly from the water column. In the case of the Coxsackie site, water and algae samples were collected directly from shore due to equipment issues concerning the boat that was being used. Algal samples were filtered from the water column onto
0.47μm glass fiber filter paper using the Geo-pump. Water was filtered directly from the river into a graduated cylinder to ensure accurate filtrate measurements.
Study sites: Tributaries
Water samples were collected from
the three tributary sites in a similar fashion
to that described above for the Coxsackie
site; however, at these three sites benthic
algae was collected instead of algae within
the water column. Six rocks at each site
Figure 4. Water sample collection were bagged and put on ice for transport to off the side of the Cary Institute boat at the the Cary Institute of Ecosystem Studies, Albany sampling site. where each rock was processed. Each rock was individually cleaned by rinsing with deionized water and carefully scrubbed in order to remove biofilms. The water used to rinse and wash the rocks was measured and recorded. A fixed amount of this water was filtered through 0.47μm glass fiber filter paper to determine chlorophyll a, ash free dry mass, and potentially other pigments.
Algae were then filtered onto three additional filters for later determination of potential catecholamine concentrations.
V-11 PROCESSING WATER SAMPLES
Collected water samples were put on ice and transported back to the Cary Institute where they were frozen upon arrival. To quantify ATSs and other pharmaceutical compounds solid phase extraction (SPE) was conducted using HLB 6cc cartridges. Each
cartridge was initially primed with deionized water and acetone. Cartridges with sample
extracts were then transported to the University of Nebraska-Lincoln, where they were
eluted with 6 ml of Optima grade acetone. Samples were then blown down to dryness under vacuum and constant stream of nitrogen gas. For each sample, 200 μl of
methanol/water (50:50) was added, as well as 100 ng of internal standard. Samples were
analyzed using liquid chromatography tandem mass spectrometry (LC/MS), a technique that has been proven to provide best analytical results for illicit drugs and metabolites
(Castiglioni et al. 2006a; Castiglioni et al. 2006b; Pal et al. 2012).
After initial analysis, it was determined that acetone alone did not provide quantitative recovery for some compounds. Cartridges were re-eluted with 6 ml of
Optima grade methanol. This was combined with the original extracts and the steps above were repeated for re-analysis. This additional step improved recovery of compounds in a method limit detection test.
MEASURING DOPAMINE
Algae samples were collected on glass fiber filters as stated above and then kept frozen until analysis. Dopamine and other catecholamine levels in algae were extracted using methods described by Van Alstyne et al (2006) combined with a derivatiztion method described by Suzuki et al. (2003). Approximately 0.1g of algae was frozen for 36 hours in 80% aqueous methanol (MEOH) at -16°C. Then, 2 ml of each extract was
V-12 filtered using a GF/A glass fiber filter followed by a 0.2 μm filter following Van Alstyne
et al. (2006). Extracts were then blown down to dryness under vacuum and a steady
stream of nitrogen gas. Dried samples had 50μl MSTFA with 1% TCMS and 50μl
pyridine added before being capped and placed on a vortex for approximately 10 seconds
(Suzuki et al. 2003). This mixture was then incubated at 40°C for 30 minutes and a 1 μl sample analyzed via GC-MS.
RESULTS
PRESENCE OF CONTAMINANTS
Initial analysis of water samples from each field site resulted in no detects of
ATSs; however, there were consistent detects of 1,7-dimethylxanthine, acetaminophen,
caffeine, and DPH in a majority of locations sampled. Highest detects were seen in
tributary sites, which is likely a result from sampling of smaller streams. Re-elution of the
HLB cartridges with methanol was able to improve the detection of amphetamine,
carbamazepine and cotinine.
Figure 5 (A and B) refers to the compounds that were detected at the Hudson
River sites and tributaries, respectively. Amphetamine was only detected at the Kingston
WWTP site (Figure 6). Fewer detects of ATSs potentially resulted from high dilution of
the contaminants being released within the HRV. A more complete overview of the
amounts of each compound detected is presented in Table 1.
V-13 A 1
g/L)
μ 0.1 Albany
Castleton 0.01 Coxsackie Concentration (
0.001
B 10
Esopus
g/L) 1
μ Creek Catskill 0.1 Creek Kingston WWTP 0.01 Concentration (
0.001
Figure 5. (A) Concentrations of pharmaceutical compounds detected at the three sampling sites on the main stem of the Hudson River. (B) Concentrations of pharmaceutical compounds detected at three tributary sites within the HRV.
V-14 Figure 6. Sample collection at the Kingston WWTP CSO that empties into Rondout Creek.
Esopus Catskill Kingston Albany Castleton Coxsackie Creek Creek WWTP 1,7- Dimethylxanthine 0.04 0.021 0.023 0.011 0.126 6.533 Acetaminophen 0.098 0.048 0.048 0.131 0.233 1.292 AMPH ** ** ** ** ** 0.018 Caffeine 0.291 0.101 0.153 0.077 0.833 14.563 Carbamazepine 0.003 0.002 0.003 0.006 ** 0.15 Cotinine 0.006 0.002 0.006 0.002 0.046 0.66 Diphenhydramine 0.084 0.014 0.011 0.033 0.094 1.552 Thiabendazole ** ** ** ** ** 0.006
Table 1. Concentrations of pharmaceutical compounds found at each sample site (μg L-). Spaces marked with (**) showed no detections at this site.
IDENTIFICATION OF CATECHOLAMINES
Catecholamines were identified after extraction from algae and derivatization.
Figure 7 shows the ion chromatograms of the three catecholamines being analyzed. 3,4-
Dihydroxybenzylamine hydrobromide was purchased from Arcos Organics to be used as
V-15
Figure 7. Ion chromatograms of silyl derivatives for (A) Dopamine, (B) Epinephrine and (C) Norepinephrine.
V-16 the internal standard. The m/z ratio for each compound is as follows: m/z 116
epinephrine, m/z 174 norepinephrine, m/z 174 dopaomine, and m/z 179 3,4-
Dihydroxybenzylamine. M/z of 355 and 147 for both norepinephrine and epinephrine
were also monitored since initial tests indicated multiple derivatives were produced for
these compounds.
DISCUSSION
The results of this study did not demonstrate large amounts of amphetamines within effluent and receiving water in the HRV; however, detection of amphetamine at
Kingston’s WWTP effluent outflow does indicate the potential for the occurrence of these stimulants. Limited detections of target compounds are most likely due to high dilution factors. Due to the high cost of analysis, few water samples could be processed.
This limited the number of sampling sites visited (Figure 3).
A majority of sampling was done after large rain events. The relevance of this is due to the use of CSOs along the Hudson River. CSOs are designed to collect rainwater, domestic sewage and industrial wastewater within the same pipe and route it all to various WWTPs. Several issues arise during or immediately following large rain events as pipes become overloaded and raw sewage bypasses WWTPs, to be released directly into streams and rivers (EPA 1994). Therefore, the pharmaceutical information presented in this paper is a good indication of what can be expected to be present after large rain events. Further research could potentially lead to detection of compounds not analyzed in this study, as well as detection of compounds that were not seen at these six sampling sites.
V-17 Of the compounds that were consistently detected, there is cause for some
concern regarding the persistence of DPH and carbamazepine (Table 1). DPH has been
known to cause behavioral changes in fathead minnows (Pimephales promelas)
(Berninger et al. 2011). Carbamazepine has been identified as a compound that alters feeding behavior (Nassef et al. 2010), impairs growth (Van den Brandhof and Monforts
2010), and induces stress responses in several different fish species (Li and Brownawell
2010). The detection of these two compounds as well as amphetamine, acetaminophen, and thiabendazole provides new information in addition to what has been previously identified (Kolpin et al. 2002; Yu et al. 2006; Palmer et al. 2008; Cantwell et al. 2010; Li
and Brownawell 2010; Reiner and Kanna 2011; Liao et al. 2012).
Determination of catecholamines in algae has not been completed at this point in
time. Initially, catecholamines were to be derivatizated with 6-aminoquinolyl-N- hydroxysuccinimidyl (AQC) and then measured with HPLC-fluorescence. Initial tests yielded broad peaks and relatively low sensitivity. To improve the derivatization of catecholamines, new AQC was synthesized following Cohen and Michaud (1993) and a new C-18 column was purchased from Thermo Scientific. When results did not improve, a borate buffer with ascorbate and EDTA was used to aid in the derivatization. Boughton et al. (2011) indicated that this small change should improve the recovery of target cartecholamines and slow the breakdown of derivatives; however, these changes did not result in improved peaks or sensitivity, so it was determined that HPLC was not sensitive enough.
It was decided that GC-MS analysis similar to methods in Suzuki et al. (2003) was a more practical way to measure dopamine and other compounds within algae. Initial
V-18 analysis of algae collected at Esopus Creek, Catskill Creek and the Kingston WWTP
showed no detects at the 0.25 ng/μl detection limit. Samples from Albany, Castleton and
Coxsackie have not been processed at this time due to the small algal samples collected from these sites. The lack of detection of catecholamines from the samples that were processed may have resulted from small sample sizes (~0.1g), or, alternatively, the specific algal species collected does not possess the ability to create these catecholamines. The algae processed by Van Alstyne et al. (2006) was Ulvaria obscura, a marine green algae. A pure culture experiment will be run in the future using a species similar to U. obscura.
The information presented in this study is a precursor to experiments that will be conducted within artificial streams. Stream experiments were originally planned for fall
2013, but since this work involves the use of controlled substances a special permit was needed before any work could begin. Though the permit has now been obtained, these experiments will be postponed until spring 2014. This will allow for determination of the best possible method to extract catecholamines from algae, as well as determine concentrations of amphetamine to spike artificial streams.
V-19 ACKNOWLEDGMENTS
The authors wish to thank the Hudson River Foundation and the Polgar
Fellowship Committee for their support in the efforts of this research. A special thanks also goes to the researchers at the Cary Institute of Ecosystem Studies and River Keeper, especially Captain John Lipscomb, for their assistance.
V-20 REFERENCES
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Castiglioni, S., E. Zuccator, E. Crisci, C. Chiabradno, R. Fanelli, and R. Bagnati. 2006b. Identification and Measurement of Illicit Drugs and Their Metabolites in Urban Wastewater by Liquid Chromatography−Tandem Mass Spectrometry. Analytical Chemistry 78:8421–8429.
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Suzuki, M., M. Mizoguchi, F. Yano, U. Hara, M. Yokoyama, and N. Watanabe. 2003. Changes in catecholamine levels in short day-induced cotyledons of Pharbitis nil. Journal of Biosciences 58:220-224.
Ternes, T.A. 1998. Occurrence of drugs in German sewage treatment plants and rivers. Water Resources 32:3245-3260.
V-23 Van Alstyne, K.L., A.V. Nelson, J.R. Vyvyan, and D.A. Cancilla. 2006. Dopamine functions as an antiherbivore defense in the temperate green alga Ulvaria obscura. Oecologia 148:304–311.
Van den Brandhof, E.J., and M. Montforts. 2010. Fish embryo toxicity of carbamazepine, diclorenac and metoprolol. Ecotoxicology and Environmental Safety 73:1862- 1866.
Yu, J.T., E.J. Bouwer, and M. Coelhan. 2006. Occurrence and biodegradability studies of select pharmaceuticals and personal care products in sewage effluent. Agriculture Water Management 86:72-80.
Zuccato, E., D. Calamari, M. Natangelo, and R. Fanelli. 2000. Presence of therapeutic drugs in the environment. The Lancet 355:1789-1790.
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V-24
THE DISTRIBUTION OF INVASIVE CELASTRUS ORBICULATUS IN AN ANTHROPOGENICALLY DISTURBED RIPARIAN ECOSYSTEM
A Final Report to the Tibor T. Polgar Fellowship Program
Shabana Hoosein
Polgar Fellow
Ecology and Evolutionary Biology Program State University of New York at Albany Albany, NY 12222
Project Advisor:
George Robinson Ecology and Evolutionary Biology Program State University of New York at Albany Albany, NY 12222
Hoosein, S. and G. Robinson. 2015. The Distribution of Invasive Celastrus orbiculatus in an Anthropogenically Disturbed Riparian Ecosystem. Section VI: 1-32 pp. In D.J. Yozzo, S.H. Fernald and H. Andreyko (eds.), Final Report of the Tibor T. Polgar Fellowship Program, 2013. Hudson River Foundation.
VI-1
ABSTRACT
The Hudson River Estuary has been colonized by numerous terrestrial invasive
plant species, due in part to its history of anthropogenic and natural disturbance riparian
dynamics. This study investigates the spatial patterns of a widespread invasion by
Oriental (or Asiatic) bittersweet (Celastrus orbiculatus Thunb) in Schodack Island State
Park, Rensselaer and Columbia Counties. The Park is home to rare species and
communities, several of which are threatened by the encroachment of bittersweet.
Bittersweet populations were mapped and surveyed on a fixed grid throughout the island,
to determine distribution patterns. Stem densities were approximately 50% higher in sites
with dredged material substrate. Local experimental tests were carried out to investigate the establishment limitations of new populations. Five site types were selected based on substrate properties (dredged material and forest floodplain) and local bittersweet densities. Ten greenhouse-grown seedlings (transplants) and 100 seeds were planted at each site, and then tracked for their survival, growth, and evidence of mycorrhizal inoculation. Transplants survived and grew at similar rates among the five experimental site types, but seed germination varied, with the highest rate in the dredged material
zones (p=0.05). No ectomycorrhizae were found on roots of experimental transplants,
seedlings, or wild-growing bittersweet, consistent with previous greenhouse studies by
other researchers. These results indicate that controlling bittersweet populations and
expansion should be focused on habitat management in dredged material-covered areas,
which represent a large fraction of the island’s upland area.
VI-2
TABLE OF CONTENTS
Abstract ...... VI-2
Table of Contents ...... VI-3
Lists of Figures and Tables ...... VI-4
Introduction ...... VI-5
Methods...... VI-9
Results ...... VI-17
Discussion ...... VI-24
Conclusions ...... VI-26
Acknowledgements ...... VI-28
References ...... VI-29
VI-3
LIST OF FIGURES AND TABLES
Figure 1- Map of bird conservation area, trails and survey grid lines ...... VI-11
Figure 2 - Oriental bittersweet time to germination ...... VI-13
Figure 3 - Map of habitat communities and experimental sites ...... VI-14
Figure 4 - Relationship between bittersweet stem diameter and age ...... VI-17
Figure 5 - Map of survey sites depicting stem densities and diameters ...... VI-18
Figure 6 - The effect of soil type and mean densities ...... VI-20
Figure 7 - Relationship between soil type and mean pH ...... VI-20
Figure 8a - The effect of soil type and cane-forming stem densities ...... VI-21
Figure 8b - The effect of soil type and liana-forming stem densities ...... VI-21
Figure 9 - The effect of bittersweet densities and
soil type on seed germination ...... VI-23
Table 1 - Matrix of experimental site categories ...... VI-14
Table 2 - Results of tests of soil samples ...... VI-22
Table 3 - ANOVA results from tests of seedling germination ...... VI-23
VI-4
INTRODUCTION
Understanding invasive species expansion is an important and difficult challenge in the Hudson River (HR) watershed. Riparian zones, which are naturally dynamic ecosystems, can be rapidly invaded by exotic plant species, particularly in sites where anthropogenic degradation has been layered over natural disturbance regimes (Stohlgren et al. 1998; Brown and Peet 2003). Disturbed ecosystems facing a series of exotic plant invasions are an ideal location to investigate the factors that may be promoting invasive species establishment and expansion.
Few studies have investigated the expansion of Oriental bittersweet after a disturbance (Ladwig 2009; Pavlovic et al. 2009; Pavlovic et al. 2010; Silveri et al. 2001).
Disturbances may disrupt natural cycles within the native ecosystem, through the alteration of soil qualities and communities, and could respond to and mediate invasions
(Kourtev et al. 2002; Ehrenfeld 2003; Reinhart and Callaway 2006). Anthropogenic modifications to soils may set the stage for invasion by altering the native community structure and enabling invasive plants to more easily establish sub-surface (mycorrhizal) interactions (Ferren and Schuyler 1980; Marler et al. 1999). Native communities may not be able to adjust to changes within the soil community quickly, increasing the opportunity for exotic species invasions.
A prime example is Schodack Island, once a series of natural islands and now a single landmass, created by repeated placement of dredged material over many decades
(C&S Engineers 1998; USACE 2000). As a disturbed ecosystem facing a series of exotic plant invasions, Schodack Island is an ideal location to investigate the factors that may be promoting invasive species establishment and expansion.
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Species characterization
One very aggressive and increasingly prominent invasive species on Schodack
Island is the exotic woody vine Oriental bittersweet (Celastrus orbiculatus Thunb.).
Documented as early as 1895, Oriental bittersweet was initially introduced in North
Carolina from Eastern Asia (McNab and Meeker1987). In recent years, Oriental
bittersweet densities have been steadily increasing in the northeastern region of North
America (McNab and Meeker1987; Hutchinson 1992; Virginia Native Plant Society
1995; Shepard 1996; TNC 2001; Pooler et al. 2002; Swearingen 2009). Its dispersal is
emphasized through various human and bird interactions. Since its introduction, Oriental
bittersweet has been cultivated for its bright colored fruits that are used in decorative
crafts (Dreyer 1994). Its distribution among riparian habitats is most likely the result of
various native and migratory bird species. Its strong dispersal potential is characterized
by high seed viability as well as being one of the few food sources for birds in the winter
(Browder 2011; Dreyer et al. 1987).
In forests and on forest edges, bittersweet is a superior competitor due to its fast
and opportunistic growth characteristics. Above ground, Oriental bittersweet can become
established in both the canopy and the understory due to its two growth forms: cane-
forming (understory layer) and liana-forming (canopy layer). Growing up to 3m per year,
bittersweet has a substantial height advantage, which inhibits the growth of understory
species (McNab and Loftis 2002; Patterson 1974; Silveri et al. 2001). Despite various
light conditions, bittersweet has the ability to crowd, overtop and eventually kill
established trees causing alterations in light gap dynamics within the understory (Albright
VI-6 et al. 2009). In addition, in its liana form, bittersweet can adjust growth to reduce intraspecific competition (Leicht-Young et al. 2011).
As with other invasives, bittersweet benefits from anthropogenic disturbances
(Silveri et al. 2001; Ladwig 2009; Pavlovic et al. 2009; Pavlovic et al. 2010). Bittersweet persists in a wide range of habitats including various types of forests, highly to slightly disturbed areas and gaps as well as low-light environments (Leicht-Young et al. 2009;
Leicht-Young et al. 2011; Browder 2011). Due to its high dispersive potential and morphological plasticity, it succeeds in a variety of open field, canopy, and sub-canopy environments to the detriment of the surrounding native plant communities (Browder
2011).
It is apparent to many people that Oriental bittersweet is becoming more prevalent throughout the northeast. On Schodack Island, bittersweet grows at high densities, often dominating canopies and some open areas. The invasion is of special concern for native and migratory bird species in the canopy layer throughout the island. Habitats are first altered by crowding, then by light gaps created when large trees are brought down by the weight of bittersweet. The impacted habitats include threatened and rare species, as well as natural communities of special interest on the island (NYS OPRHP 1998). This invasion has progressively intensified over the years, causing a serious problem for threatened and rare species inhabiting Schodack Island (Barbour and Kiviat 2001).
Mycorrhizal Associations
Below-ground competition from older bittersweet plants could also affect native vegetation (Leicht-Young et al. 2011). Like many woody invasive species, Oriental bittersweet tends to rely on generalist mycorrhizal associations that seem to be easily
VI-7
formed (Lett et al. 2011). The invasion of an exotic species can be magnified through the facilitation of various local mycorrhizal species, which may enhance competitive effects
in exotic species causing detrimental impacts on the native community structure (Marler
et al. 1999; Pringle et al. 2009; Richardson et al. 2000; Rejmanek 2000). These
associations may depend on the localized environmental conditions, such as soil
properties, which may determine the magnitude of ecosystem-level impacts (Ehrenfeld
2003). Such soil properties are specific to the land use history of an area, which has the
potential to exacerbate microbial competition.
General properties of freshwater dredged material varies considerably, but it has
the potential to exacerbate invasive species expansion (Zedler 2001; Ohimain 2004).
Freshwater dredged material deposition may alter the natural soil community dynamics
causing site specific vulnerabilities that encourage the establishment and overall success
of Oriental bittersweet. Little is known about the recovery of native ecosystems from
freshwater dredged material deposition; however, these disturbances have negatively
impacted native species diversity and caused significant range reductions within the
native species communities (Ferren and Schuyler 1980).
Several studies have shown that many invasive species have a demonstrated
ability to alter ecological processes to their advantage (Ehrenfeld 2003; Wolfe and
Kilronomos 2005). Various factors contribute to the success of Oriental bittersweet
invasion including a wide range of light and habitat tolerance (Dreyer et al. 1987; Leicht
and Silander 2006; Leicht et al. 2007). Few studies have documented Oriental bittersweet
soil ecosystem preferences (Leicht-Young et al. 2009; Lett et al. 2011); however, none
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have investigated the soil and microbial communities associated with its expansion and distribution in disturbed habitats.
The Objectives of this study were to determine the driving forces influencing the
distribution of Oriental bittersweet in a disturbed riparian habitat, with three specific
aims:
• To determine the approximate arrival time of Oriental bittersweet on Schodack
Island and its current distribution pattern, using field surveys.
• To test for site-specific conditions that may favor or limit the establishment and
growth of Oriental bittersweet individuals.
• To aid in the development of management strategies to restore key habitats and
limit further spread of the invasive bittersweet.
Null hypotheses were that (1) distribution patterns are random, with no focal zones of establishment; (2) sites with lower bittersweet densities are no less prone to invasion than sites with higher densities; (3) sites with native floodplain soils are no less prone to invasion than sites with soils derived from dredged material. Vulnerability to invasion was determined by a combination of mapping, field measurements, and experimental tests.
METHODS
General Site Information
Schodack Island State Park is a peninsula in the northern HR Estuary, bordered by
the main channel on the west and by Schodack Creek/Muitzes Kill to the east. Schodack
Island, once a series of natural islands and now a single landmass, created by repeated
VI-9
placement of dredged material over many decades (C&S Engineers 1998; USACE 2000).
Although surrounded by tidal wetlands, Schodack is dominated by dredged material and
floodplain forests. Ecological communities include successional old field, successional
shrubland, dredge “spoil” forest, locally rare freshwater intertidal mudflat, freshwater
tidal marsh, freshwater tidal swamp, and floodplain forest (NY NHP 2005). The park is a
designated Bird Conservation Area (NYS DEC 2012), Department of State Significant
Coastal Fish and Wildlife Habitat, and Significant Scenic Area (NYS DOS 2012). Since
at least 1965, the site has supported a breeding population of Cerulean Warbler
(Setophaga cerulea; NYS Species of Special Concern), with 13 singing males counted in
1997 (Barbour and Kiviat 2001), and their presence reconfirmed in 2012 by NY Audubon
volunteers. Bald Eagle (Haliaeetus leucocephalus) maintain two active nests, Osprey
(Pandion haliaetus) are commonly observed roosting and foraging, and a resident Great
Blue Heron (Ardea herodias) colony contains about 50 nests.
Schodack Island began as a set of alluvial islands, once used for agriculture by
Native Americans. Beginning in the 1920’s, dredged material was deposited up to 2 m above river level, creating a single land mass that is a mosaic of natural and artificial topography and soils (Miller et al. 2006). Schodack Island State Park, which includes all but the southern tip of the island (still an active USACE dredged material placement area), was established in 2002, and several rounds of planning identified the significance of its avian habitats as well as the threats they faced (NYS OPRHP 1998; C&S Engineers
1998; Barbour and Kiviat 2001). The most prominent invasive species is Oriental bittersweet.
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Experimental Design
Survey
In summer-fall 2012, operating with a research permit from NY State Parks, thirty transects were mapped out along the E-W axis of the island. Each transect was separated by 230 m on the N-S axis (which is totaled at 9.5 km) creating a grid within the park boundaries. Bittersweet populations were surveyed at 146 grid points (Figure 1).
Figure 1. Map of bird habitats, roads/trails and experimental survey grid on Schodack Island. Survey points are noted as the intersection between N-S and E-W transects.
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Within a 2 m radius of each point, UTM coordinates, stem densities of cane-
forming and climbing bittersweet, thickest stem diameters, and reproductive status were
recorded. All climbing bittersweet were cut at 0.5 m above ground level, to investigate
future regrowth patterns. Rapid assessments were made along roads and trails throughout
the park to locate larger bittersweet lianas (> 2 cm dia.), which were mapped, measured,
cut and aged.
Phytometer Experiment
In November 2012, approximately 4,000 bittersweet fruits (with 3-5 seeds each)
were collected from random locations across Schodack Island, and approximately 5,000
seeds were cleaned, batched, and stored under refrigeration (Young and Young 1992).
Preliminary tests indicate germination rates of about 70% in the greenhouse (Figure 2). In
March 2013, more than 1,000 seeds began germination in the greenhouse, planted in pots of sterile medium, watered and fertilized, in preparation for field transplantation, which occurred May 29-June 23, 2013. Transplants were grown in pots with 10 other seeds in the greenhouse. During installation, transplants were separated in to groups with 2-4 individuals, thinning out weaker seedlings to keep densities low.
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Figure 2. Scatter plot showing Oriental bittersweet time to germination in a greenhouse. Graph shows a logarithmic regression with a cluster of individuals germinating within 20 days and a second cluster of individuals germinating between 40-50 days.
Phytometer Bittersweet Densities
Twenty-five experimental sites were selected on the basis of five site categories
(Figure 3), with five replicates of each. High, low and absent densities of Oriental bittersweet were quantified into categories based on the field survey (Figure 5). High densities consisted of survey points that had more than 16 stems per 2m2. Low densities consisted of survey points that had between 1-10 stems per 2m2. Absent densities consisted of survey points that had 0 stems per 2m2. Soil classifications were roughly based on habitat community maps provided by The New York Natural Heritage Program
(Figure 3b).
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Table 1. Matrix of experimental site categories, designated on the basis of previous surveys and observations. Bittersweet has not been observed at high densities in wetter, low-lying areas on Schodack Island.
Site-specific Dredged Flood- bittersweet densities Material plain High X Low X X Absent X X
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Figure 3. Map showing habitat communities and phytometer experimental sites. The number correlated with each point identifies the waypoint for each site. Warm colored points (red, pink, purple) represent sites on the dredged material habitat. Cool colored points (blue, green) represent the sites on the native floodplain forest. H (high), L (low), A (absent) is a quantification of the adult bittersweet communities at each site.
Native vs. Non-native Soil Type
At each site, 10 transplant seedlings and 100 seeds were planted in mapped and marked locations. Transplants were grouped in clusters of 2-4 stems and planted within a one meter diameter of the data collection point. Seeds were spread evenly within a two meter diameter from the data collection point at about a half meter in width. Stem length was recorded during installation and monitoring which occurred three times throughout the summer and fall (50, 40 and 30 days apart). During monitoring, percent germination was also recorded. The monitoring period ended after 120 days in which evidence of herbivory was noted for each individual plant. In addition, soil samples were collected from each of the 25 sites and sent to Cornell University‘s Nutrient Analysis Laboratory for soil fertility (Morgan extractable P and NO3; K, Ca, Mg, Fe, Mn, Zn, and Al; pH;
buffer pH; % H20 and % organic matter).
Seedling Growth
Surviving seedlings were dried and weighed (separating roots and stems). Roots
of collected seedlings were separated from stems (excised immediately above the
uppermost root) and examined visually for evidence of ectomycorrhizae. Stems were
dried in paper bags for 24 hours at 60°C. Weights of paper bags before and after drying
were also taken into consideration for data collection.
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Statistical Analyses
Liana and cane densities and diameters were compared between dredged materials
and floodplain sites were made with two-tailed paired t-tests, using data from all survey
points (N = 85). The same test was applied to pH values for the subset of points with soil
samples (N = 26). The means and standard deviations for the phytometer experimental
tests based on soil type (forest floodplain vs. dredged material) and adult densities
(absent, low and high) were calculated. Experimental transplants were pooled into their original clusters (3-4 per site), using means of the 2-4 plants per cluster. This rule was applied because the transplants were moved and planted as groups (just as seeds are carried 2-4 per fruit by dispersers), and because they tended to share microenvironmental factors, such as substrate condition and herbivory. Survival (%) and growth increments
(above-ground) were compared among the five treatment levels using one-way Analysis
of Variance (ANOVA), following tests for normality and variance homogeneity. Mean
seedling germination rate (%) was counted as the maximum value per site over three
survey periods, and growth as the mean above-ground stem dry weight per site, again
using one-way ANOVAs. A power analysis indicates that with five replicates at five
levels, achieving a power of 0.8 (at alpha = .05) required a standard deviation < 1/2 the
mean difference among treatments). Thus, a 20% difference in means would have an
80% probability of achieving a significant effect with a standard deviation < 10%.
One low-density dredged material site was found to contain well over 100 new
seedlings during the final survey, well beyond the observed range of germination periods
for both greenhouse and field trials. Evidence from the surrounding area indicated a high
level of natural germination, confounding the results, so this site was ruled out in
VI-16 statistical comparisons of germination rates and seedling biomass. All analyses were performed using SYSTAT 13 (Systat Software, Inc. Chicago, IL).
RESULTS
Surveys
Oriental bittersweet showed a variety of age ranges, with the oldest stem observed being approximately 20 years old at 7 cm (Figure 4). The distribution of the largest stems seemed more proximate to the roads and trail edges. Larger stems in higher densities are generally seen in the dredged material habitat (Figure 5). Most of the fruit-bearing stems are lianas > 2.4 cm (7-10 yrs.), therefore younger populations may only be spreading through clonal expansion. Stem diameter was positively, but weakly correlated with density (R2 = .289).
Figure 4. Scatter plot showing the relationship between Oriental bittersweet stem diameter and age. Graph shows a logarithmic regression with the oldest bittersweet stem at approximately 20 years old.
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Figure 5. Map showing the habitat communities and survey points on Schodack Island. The figure on the left represents the northern part of the island, while the figure on the right represents the southern half. Survey points that are directly overlaid indicate higher densities than the single survey points.
The stem density of established Oriental bittersweet throughout the Schodack
Island was significantly correlated with soil type (Figure 6). On average, stem density per
2 m2 on the dredged material was greater than the native soil stem density (p=0.002).
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Oriental bittersweet stem diameter for the largest individual per survey site averages at
about 1.75 cm (or approx. 5-7 yrs.) on Schodack Island and is not strongly correlated
with stem densities (R2 = .088, N = 138 local populations). Sample sites were clustered
near roads and trails, which may explain similarities in pH (Figure 7), however additional soil analyses from phytometer sites also found no significant differences (Table 2). In contrast to density differences, stem diameters for both cane and liana forming bittersweet stems were similar among site types (Figure 8a, b).
Figure 6. Line graph showing relationship between soil type and mean stem density at each survey point.
Figure 7. Line graph showing relationship between soil type and pH.
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Figure 8. Line graphs showing (a) relationship between soil type and mean diameter of cane-forming bittersweet vines that mainly occupy the understory layer, and (b) relationship between soil type and mean diameter for liana-forming bittersweet vines that mainly occupy the canopy.
Phytometer Site Soil Nutrients
The native floodplain soils were much richer in organic matter, and in several major nutrients (Table 2). These differences may reflect differences in nutruent cycles as well as physical properties of the two substrates. Higher levels of Ca, Mg, and Mn may be attributable to greater clay content in native floodplains, in addition to a higher fraction of organic matter. Likewise, highly soluble nitrate may be bound less tightly in the sandy soils of dredged material origin.
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Table 2. Results of tests of soil samples taken from dredged material sites (D) and floodplain sites (F). LOI = loss on ignition; OM = organic matter. Adjusted p values represent Bonferroni corrections for multiple t-tests (ns = p > .05).
Test Site Adjusted Variable Type N Mean SD p-Value H2O (%) D 15 0.395 0.294 <0.0001 F 10 1.306 0.296 pH D 15 5.901 0.408 ns F 10 6.217 0.379 LOI (%) D 15 2.227 1.644 <0.0001 F 10 6.134 1.84 OM (%) D 15 1.327 1.15 0.001 F 10 4.063 1.289 P (mg/Kg) D 15 2.026 1.559 ns F 10 4.693 2.597 Al (mg/Kg) D 15 6.315 3.006 ns F 10 5.607 3.536 Ca (mg/Kg) D 15 750.943 533.024 <0.0001 F 10 2,448.75 411.737 Fe (mg/Kg) D 15 3.753 1.597 ns F 10 5.846 5.977 K (mg/Kg) D 15 52.136 23.051 ns F 10 77.284 25.011 Mg(mg/Kg) D 15 80.41 70.149 <0.0001 F 10 234.494 63.661 Mn (mg/Kg) D 15 5.336 4.197 <0.0001 F 10 14.617 2.725 Zn (mg/Kg) D 15 3.094 2.007 ns F 10 5.975 2.455
NO3 (mg/Kg) D 15 4.171 3.099 0.027 F 10 10.802 4.919
Phytometer Bittersweet Germination, Survuval, and Growth
Of the 2,500 seeds that were planted throughout the island, only 211 seeds germinated. Final seedling counts ranged from zero (in 12 sites) to 27, for a mean of 4.7
VI-22 per site, or <5% of the total introduced seeds. These values stand in contrast to the high greenhouse germination rates (above), but are not atypical of woody plants in natural systems (Harper 1977). Germination rates varied among treatments (Figure 9; Table 3).
Once germinated, 64.2% of the new seedlings survived the duration of the experiment,
Figure 9. Bar graph representing the differences in soil type and adult bittersweet densities with mean percent germination of planted seeds (ANOVA results in Table 3).
Table 3. ANOVA results from tests of seedling germination per site type.
Source Type III SS df Mean Sq F-Ratio p
Site type 358.358 4 89.590 2.872 0.05
Error 592.600 19 31.189
Eighty-five percent of the total transplanted seedlings survived for the first 120 days. For the period studied, high levels of variance ruled out significant treatment differences (mean 3.53 seedlings lost; sd 5.08; ANOVA F4,74 = 1.66, p = .167). Sizes of
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plants from field-germinated seeds were equivalent among treatments (ANOVA F4,74 =
2.17, p = .915).
Mean transplant growth was all fairly low over the 120 day period. In the most
extreme cases, averaging under 5 cm across all treatments, which had similar values
(mean 1.71 cm; sd 3.66; ANOVA F4,74 = 1.66, p = .167). Under optimal conditions,
established plants of Oriental bittersweet can grow up to 100 cm in 120 days (Patterson
1974; Leicht and Silander 2006). Ten of 250 transplants were clipped with an angular cut,
indicating possible herbivory by meadow voles (Microtus pennsylvanicus; Manson et al.
2001). An additional 42 (16.8%) were not found, and categorized as unexplained losses.
Phytometer Ectomycorrhizal Inoculation
Among the 214 seedlings collected (including the potential natural seedlings from one site), all roots examined had unmodified root caps and visible root hairs, with none of the swelling or bifurcation associated with ectomycorrrhizae. This finding is consistent
with results from Lett et al. (2011), whose experiments demonstrated that C. orbiculatus
could not be inoculated with generalist ectomycorrhizal fungi in a laboratory setting.
DISCUSSION
Survey Results
Ring counts of the largest bittersweet lianas indicate a maximum age of
approximately 20 yr. This is consistent with its relatively recent spread in the region, and
during the course of this project, similar-aged specimens were found on nearby
Papscanee Island, to the north. Given that bittersweet seeds are bird-dispersed and have
VI-24 high germination potential, the invader probably began arriving on Schodack Island less than 25 years ago. In contrast to null predictions, distributions of bittersweet were non- random – stem densities were greater in the dredged material-derived soils compared to the native soils. In addition, high-density bittersweet populations were not encountered in any native soil locations.
Native substrate was uniformly classified as floodplain soil because it was found in low-lying areas and on shorelines. Therefore, waterlogging could be a factor in limiting bittersweet in these areas. Confirmation would require a longer study in conjunction with manipulation experiments. Soil nutrients are an unlikely explanation for site-specific differences, because the native soils were richer in several key nutrients, and bittersweet is not known as a poor-soil specialist (Leicht and Silander 2006; Leicht et al. 2007). A further explanation for greater stem densities in dredged material soils may lie in differential patterns of bird dispersal, coupled with clonal propagation. Large stems were found in several locations across the island, but its expansion could potentially be traced back to a few major colonization points that have expanded. The positive (although weak) correlation between stem diameters and stem densities supports this latter possibility. Other research has found evidence for positive density dependence in bittersweet populations (Leicht-Young et al. 2011).
Phytometer Results
Germination rates of planted seeds differed among treatments, with lower rates in native soils and dredged material sites with bittersweet absent. Transplant survival and growth as well as growth of seedlings from planted seeds were equivocal among
VI-25 treatments. However, the duration studies was limited to a single season, and final results of planned concluding work may show clearer differences. In addition, the experimental design was sensitive to the levels of variance encountered, as indicated in the power analysis (see Methods). Differential herbivory may have played a role, but it was variable among sites and not directly investigated.
CONCLUSIONS
The two main results provide support for a difference between substrate type in the establishment and success of bittersweet. What remains to be demonstrated are clearer mechanisms for the apparent role of substrate. For example, although ectomycorrhizae seem unimportant, VAM inoculations can be quantified in bittersweet plants removed from the two different substrates. Mycorrhizae are sensitive to nutrient levels (Allen
1991), and invasive plants may owe a large part of their success to an ability to form mycorrhizal mutualisms in new settings (Pringle et al. 2009). Therefore, it is conceivable that properties of the native soils may limit bittersweet’s potential to establish and grow via indirect interactions mediated by soil fungi.
If further work confirms that the dredged material-derived soils are preferred growth zone, more targeted management strategies can be developed for the control of bittersweet in Schodack Island State Park. Sensitive habitats and other areas of conservation interest have been demarcated for the park (NYS OPRHP 1998), and the next step is to assess bittersweet populations in their vicinity. Targeted removal of the largest stems has already begun, but supplies of new recruits are plentiful on and off the island. Using the new information from the study in this report, it seems reasonable to
VI-26 focus broader control efforts on source populations growing on dredged material-derived soils. An additional strategy seems apparent. Much of the Hudson River (western) shoreline of the park is rimmed with bulwarks and other hardened features, a legacy of previous use that continues to disrupt tidal floodplain connections. Restoring those connections may provide a measure of resistance to further spread of bittersweet on
Schodack Island.
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ACKNOWLEDGEMENTS
I would thank the Tibor T. Polgar Fellowship program for a wonderful opportunity that has inspired, taught and furthered me in my research career. I am grateful for Dr. George Robinson who has helped me in my constant struggles and has provided great resources to help me excel with my project. New York State Office of
Parks and Recreation and Casey Holtzworth were very helpful in providing ideas and providing information about the island. I would also like to thank the State University of
New York Research Foundation and State University of New York Graduate Student
Office for supporting this research and its continued exposure, and The New York State
Natural Heritage Program for providing various maps and advice throughout the process.
And finally, I wouldn’t have gotten this far without the help of volunteers from The
Audubon Society (Capital District Chapter), Nick Osseni, Alyssa Resey and Kate Foley.
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Ohimain, E. 2004. Environmental impacts of dredging in the Niger Delta: Options for sediment relocation that will mitigate acidification and enhance natural mangrove restoration. Terra et Aqua 97: 9-19.
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VI-32
HYPOXIA TOLERANCE OF THE INVERTEBRATES ASSOCIATED WITH WATER-CHESTNUT BEDS (TRAPA NATANS L.) IN THE HUDSON RIVER
A Final Report of the Tibor T. Polgar Fellowship Program
Mariana Carolina Teixeira
Polgar Fellow
Freshwater Ecology Post-Graduation Program Universidade Estadual de Maringá Maringá, PR, Brazil
Project Advisor:
David L. Strayer Cary Institute of Ecosystem Studies Millbrook, NY 12545
Teixeira, M. C. and D. L. Strayer. 2015. Hypoxia Tolerance of the Invertebrates Associated with Water-chestnut Beds (Trapa natans L.) in the Hudson River. Section VII: 1-29 pp. In D. J. Yozzo, S.H. Fernald and H. Andreyko (eds.), Final Reports of the Tibor T. Polgar Fellowship Program, 2013. Hudson River Foundation.
VII-1
ABSTRACT
Aquatic macrophytes provide structures to be colonized and used as refuge, shelter and feeding sites by many organisms. As ecosystem engineers, they also have major effects on the physical and chemical features of water and sediment. Low oxygen concentrations are one of the recurrent impacts of dense beds of aquatic macrophytes, so the macroinvertebrates that inhabit macrophyte beds may have to cope with hypoxia.
This study tested the survival of six groups of macroinvertebrates associated with different species of macrophytes (Trapa natans or submerged species) under hypoxia.
The hypothesis tested was that the organisms that inhabit stands of the invasive T. natans are more resistant to low oxygen concentrations than those that inhabit submerged aquatic vegetation (SAV). Chironomids and amphipods associated with T. natans were more resistant than those associated with submerged species. Gastropods, ostracods and planarians from the two habitats were equally resistant to hypoxia, with high survival in both cases. Zygopterans were the most impaired group, with no survival at all. The results indicate the responses to hypoxia vary among groups, and so do the mechanisms through which they occur.
VII-2
TABLE OF CONTENTS
Abstract ...... VII-2
Table of Contents ...... VII-3
Lists of Figures and Tables ...... VII-4
Introduction ...... VII-5
Methods...... VII-10
Results ...... VII-14
Discussion ...... VII-17
Acknowledgements ...... VII-23
References ...... VII-24
VII-3
LIST OF FIGURES AND TABLES
Figure 1 - Submerged growth form of Vallisneria americana versus
the floating-leaved form of Trapa natans ...... VII-7
Figure 2 -Sampling sites in the Tivoli North and South Bays ...... VII-11
Figure 3 -Survival under hypoxia of organisms associated
with T. natans (TNA) and submerged aquatic vegetation (SAV)
for six and eighteen hours exposures ...... VII-16
Table 1 –Trapa natans vs. SAV-dwelling
macroinvertebrates’ survival to hypoxia ...... VII-15
VII-4
INTRODUCTION
Biological invasions have become one of the greatest threats to the integrity of
Earth’s ecosystems (Strayer 2010). Invasive species are usually exotic (but not always)
to the habitat they colonize and very successful in them, often to the detriment of the
former inhabitants. They bring in new traits that represent new pressures to the system,
frequently culminating in undesirable impacts to native communities and the environment. In freshwater habitats, which are especially sensitive to anthropogenic impacts (Dudgeon et al. 2006), invasions are recurrent and their effects have become obvious and a cause of concern (Strayer 2010).
Aquatic macrophytes create distinctive habitats within aquatic ecosystems. They provide structures to be colonized by an array of organisms, and are used as refuge, shelter and feeding sites (Agostinho et al. 2007; Mormul et al. 2010). They also have major effects on the physical and chemical features of water and sediment (Urban et al.
2009; Bini et al. 2010; Kleeberg et al. 2010; Tall et al. 2011; Rodríguez et al. 2012), and, because of that, are considered ecosystem engineers (Bouma et al. 2005; Caraco et al.
2006). When macrophytes are invasive, their “engineering” can become a problem if it causes changes that the original inhabitants of the invaded area cannot cope with.
Macrophytes have a great potential of becoming invasive because they adapt easily to a large range of habitats, have high rates of vegetative reproduction and fast growth
(Santamaría 2002).
A few macrophyte traits are often related to impacts in their environment and to other organisms when they become invasive. Biomass production, photosynthesis and decomposition, for example, are three functions through which macrophytes can alter
VII-5
habitat complexity, light penetration, pH and dissolved oxygen concentration (Schultz
and Dibble 2012). Low oxygen concentrations have been identified as one of the recurrent impacts of dense growth of aquatic macrophytes (Caraco et al. 2006; Bunch et al. 2010; Turner et al. 2010; Desmet et al. 2011). Overnight respiration in large
macrophyte beds usually creates a daily variation pattern in dissolved oxygen (DO)
levels, with lower values at sunrise and higher values when light is most available for
photosynthesis (Jones et al. 1996; Miranda et al. 2000; Thomaz et al. 2001; Colon-Gaud
et al. 2004). Besides plants’ metabolic activities, oxygen dynamics also depend on the
physical structure of the stands, as they regulate light incidence and water circulation
(Jones et al. 1996; Miranda and Hodges 2000; Bunch et al. 2010). In floating-leaved and
emergent species, for example, the diel pattern can be overcome by the constant shading
and reduced exchange at the water-air interface caused by the biomass that lies above the
surface. Furthermore, the oxygen produced during photosynthesis by the aerial biomass is
liberated to the atmosphere while the oxygen dissolved in the water is still consumed by
respiration and decomposition. This unbalanced uptake can lead to more prolonged
hypoxia or anoxia in dense beds of floating-leaved or emergent plants, which has been shown to occur in many macrophyte beds (Bunch et al. 2010; Turner et al. 2010; Desmet et al. 2011), among them water-chestnut beds in the Hudson River (Caraco and Cole
2002).
The water-chestnut Trapa natans L. is native to temperate and tropical Eurasia and Africa (Crow and Hellquist 2000), and was introduced in the United States in 1874
(Hummel and Kiviat 2004). In 1930, it reached the Hudson River (Hummel and Kiviat
2004), where it is now widespread, forming stands of hundreds of hectares (Nieder et al.
VII-6
2004). This macrophyte is rooted in the sediment, with a submerged stem and feather leaves, and a dense canopy of floating leaves (Hummel and Kiviat 2004). T. natans has already replaced part of the native aquatic vegetation, dominated by the submerged species Vallisneria americana Michx. (water-celery), in many parts of the Hudson River
(Strayer et al. 2003). The native V. americana is also rooted in the sediment, but its leaves are long and flat, grow from rhizomes, and the whole plant is submersed. Such different growth forms, shown in Figure 1, create distinct habitats for all macrophyte- associated organisms, as much in their physical structuring as in their influence on physical and chemical parameters of water and sediments.
Vallisneriaamericana Trapa natans
Figure 1. Comparison between the submerged growth form of Vallisneria americana (water-celery) and the floating-leaved form of Trapa natans (water-chestnut).
The case of T. natans in the Hudson River has been thoroughly investigated
(Caraco and Cole 2002; Hummel and Findlay 2006; Goodwin et al. 2008). Tidal
VII-7
influence has been shown to add to its engineering on DO dynamics. Caraco and Cole
(2002) showed that the DO levels in T. natans beds varied in a 12.5 h tidal cycle, with
high peaks (ca. 7.5 mg.L-1) during high tides as fresh, well-oxygenated water flows in
from the open water of the river, and lows of 0 to 2.5 mg.L-1 during low tides. In V.
americana stands however, DO never dropped below 5 mg.L-1 and was always higher
than in the open water surrounding the bed (Caraco and Cole 2002). Therefore, the hypoxic to anoxic events that take place in T. natans beds are temporally limited by the
tides, lasting approximately six hours with DO values from 0 to 2.0 mg.L-1 (Caraco and
Cole 2002). These low DO values could be maintained for much longer periods in large beds that are not subjected to periodic flushes of oxygen-rich water.
Such different habitats created by T. natans in relation to the former dominant
species V. americana could have led to changes in the macrophyte-dwelling
communities. Strayer et al. (2003) compared the communities of aquatic
macroinvertebrates in stands of T. natans and V. americana in the Hudson River. They
found that although some community attributes, such as species richness and diversity,
did not differ between the native and invasive stands, taxonomic composition differed
greatly. The ability of species to deal with the distinctive dissolved oxygen dynamics
with its periodic hypoxic events was thought to explain part of the differences between T.
natans- and V. americana-associated macroinvertebrate communities, along with
predator pressure and food availability, both which would also be influenced by T. natans
engineering (Strayer et al. 2003). Kornijów et al. (2010) attempted to unveil the role of
dissolved oxygen dynamics in determining the macroinvertebrate community in T. natans
beds. They found that species richness, diversity and density did not differ among
VII-8
microhabitats (beds’ edges versus center; leaves versus sediments) with different degrees
of oxygen depletion in T. natans beds, meaning that the macroinvertebrates inhabiting
these beds can and do cope with periodic short-term hypoxia.
Dissolved oxygen concentration in aquatic habitats has diverse effects on the physiology, behavior, life cycle, growth capacity, distribution, reproductive success and immunological system of animals (Ekau et al. 2010). Morphological, physiological and behavioral adaptations to low levels of dissolved oxygen have been reported for many aquatic organisms, including auxiliary respiratory appendages (Robinson et al. 1991;
Marziali et al. 2006), migration to oxic microhabitats (Apodaca and Chapman 2004;
Sesterhenn et al. 2013), use of atmospheric air (Hanley and Ultsch 1999; Seuffert and
Martin 2010; Penha-Lopes et al. 2010), induction of respiratory pigments (Spicer 1993;
Panis et al. 1996) and anaerobic metabolism (Frank 1983; Hamburger et al. 1995;
Hamburger et al. 2000).
The aim of this study was to investigate whether the organisms that inhabit stands of the invasive T. natans are more resistant to low oxygen concentrations than those that inhabit submerged aquatic vegetation (V. americana and Myriophyllum spicatum L.
(European milfoil)) stands. This study also aims to discuss which adaptations could provide T. natans-dwelling macroinvertebrates with a better capacity to cope with hypoxia.
VII-9
METHODS
Study area
The North and South Tivoli Bays, shown in Figure 2, encompass ca. 290 ha of freshwater tidal marshes, subtidal shallows and intertidal mudflats, along 3 km of the eastern shore of the Hudson River. The marshes are dominated by narrow-leaved cattail
(Typha angustifolia L.), the subtidal shallows are dominated by the submerged species V. americana and M. spicatum, and the South Bay mudflats are dominated by the invasive
T. natans (Yozzo et al. 2005). All these macrophytes have been shown to support dense and diverse macroinvertebrate communities (Strayer et al. 2003; Kornijów et al. 2010;
Yozzo and Osgood 2013). These areas are also used by many fishes, such as black bass
(Micropterus sp.), white perch (Morone americana) and common carp (Cyprinus carpio), waterfowl, many birds, and snapping turtles (Yozzo et al. 2005).
VII-10
North Tivoli Bay
South Tivoli Bay
Hudson River km 158-161 Dutchess County, NY, USA
Figure 2. Sampling sites in the Tivoli North and South Bays.
Experiments
Sampling occurred from mid August until late September 2013, a period that encompassed peak biomass for invasive T. natans (coded “TNA” throughout), and for
VII-11
submerged V. americana and M. spicatum (coded “SAV” throughout) (Caraco and Cole
2002). Animals were collected from a large TNA bed in Tivoli South Bay and from an
SAV bed in Tivoli North Bay by putting whole plants into a submerged bucket, gently
shaking them, pouring the water into a plankton net (118 µm mesh size) and collecting
the retained material, which often included mud and debris, in jars. In SAV beds, some
samples were collected by sweeping the net in the beds to avoid destroying the plants.
All collected material was kept on ice and immediately taken to the lab for sorting.
Organisms were sorted under a stereomicroscope, separated by groups and kept in
jars with river water. The most abundant groups in the samples were Chironomidae,
Gastropoda, Amphipoda, Ostracoda, Zygoptera and Planariidae. All jars were kept in a
large tray containing water and ice packs to keep temperature low. With the exception of
Amphipoda, no organisms were kept longer than 24 hours before the trials, and the
Amphipoda that were kept longer (up to five days) were in a large plastic tray (ca. 10 L)
with an air pump providing high oxygen levels, and live chironomids were added as food.
To determine the resistance of organisms to hypoxia, a series of laboratory assays were run by manipulating dissolved oxygen levels and measuring the survival of different invertebrates. River water was sparged with N2 gas until low concentrations of DO (0.5
-1 to 1.0 mgO2.L ) were achieved for the hypoxic treatment and aerated (when necessary)
-1 until saturation for the normoxic treatment (8 - 9 mgO2.L ). These concentrations are
within the range of variation found by Caraco and Cole (2002) in TNA (low DO) and
SAV (high DO) beds. The assays were run in 300 ml biological oxygen demand bottles.
Water temperature was not controlled and averaged 22.2±0.7°C (mean±SD). After all
bottles were filled with water, and oxygen and temperature were measured with a VII-12
handheld YSI ProODO meter, the organisms were added and bottles sealed. Each trial
consisted of one group being tested in a set of five replicates for each treatment, run according with organism’s availability. In a few cases, only three replicates were run, or
the five replicates were run in blocks of two and three replicates. Five individuals were
used in each bottle with the exception of ostracods, for which ten individuals were used
per bottle.
Two different exposure times were tested: six hours, which was based on the
length of hypoxic events observed in the Hudson River (Caraco and Cole 2002), and
eighteen hours, which meant to represent an environment without tidal influence and
consequently longer hypoxic periods. At the end of each trial, DO and temperature were
measured in each bottle and the organism’s survival was verified. Animals were
checked, under magnification when necessary, and considered alive when moving
spontaneously or in response to touching. The total number of organisms recovered,
which differed for some groups from the initial number probably due to predation and/or
difficulties in handling, was used to calculate the survival percentage. All organisms
were kept in 70% alcohol for later taxonomic identification.
Statistical analysis
Welch t-tests (Welch 1947) were used to test for differences between the survival
of organisms from different origins (TNA vs. SAV) within each taxonomic group
(Amphipoda, Odonata, Gastropoda, Chironomidae, Ostracoda and Planariidae) and
exposure time (six and eighteen hours). A paired Welch t-test was also run including data from all taxonomic groups to test for differences in survival between origins within
VII-13
each exposure time. In all analyses, the response variable was the percentage of survival in the hypoxic treatment divided by the percentage of survival in the normoxic treatment.
Tests were performed with R (R Core Team 2013).
RESULTS
DO initial and final concentrations in hypoxic treatments averaged 0.69±0.2
-1 -1 mgO2.L and 0.25±0.3 mgO2.L , while normoxic concentrations averaged 8.65±0.3
-1 -1 mgO2.L at the beginning of experiments and 7.29±1.5 mgO2.L at the end. The
normoxic controls of all trials (SAV and TNA, 6 and 18 hours, n=117) averaged 97±9%
of survival, which indicates that handling and experimental conditions did not by
themselves kill many animals; however, the survival values used in the analyses did take
into account the survival in normoxic treatments, as mentioned in methods.
The results of t-tests can be seen in Table 1 and in Figure 3. When all taxa were
considered, there were no significant differences in survival between organisms
associated with TNA or SAV beds for either exposure time. When each taxonomic group
was tested separately, significant differences were only found for Amphipoda in the six
hour trial and for Chironomidae in the eighteen hour trial. Both groups showed higher
survival of organisms associated with TNA rather than SAV, as predicted in the
hypothesis. Nevertheless, none of the remaining groups (Zygoptera, Ostracoda,
Gastropoda, Planariidae) differed in their survival according to organisms’ origin.
Zygoptera had no survival at all in low DO at either exposure time and regardless
of the origin of the organisms. On the other hand, Ostracoda had high survival in all
hypoxic treatments, independent of originating from TNA or SAV.
VII-14
Table 1. Comparison (Welch t-test) between Trapa natans- and submerged aquatic vegetation-dwelling macroinvertebrates’ survival to hypoxia. Significant differences (p<0.05) are in bold.
Groups Exposure t Df p All 6h 1.52 20 0.1442 18h -0.85 27 0.4048 Amphipoda 6h 5.72 5.54 0.0016 18h - - - Zygoptera 6h -1.44 4 0.2228 18h - - - Gastropoda 6h -1 2 0.4226 18h -2.25 5.74 0.0676 Chironomidae 6h 1.12 4.53 0.3181 18h 3.64 4 0.0220 Ostracoda 6h -1.05 4 0.351 18h 1.01 7.34 0.344 Planariidae 18h -1.04 5.68 0.342
When considering all groups, Amphipoda, Gastropoda and Chironomidae, survival was higher in the six hour trials than in the eighteen hour ones. Moreover, when survival between origins in different exposure times was compared, there seemed to be an interaction between these factors, with a smaller difference between origins in the six hour treatment than in the eighteen hour one.
VII-15
AMPHIPODA 1.0 ALL GROUPS 1.0
0.8 0.8
0.6 0.6 *
0.4 0.4 SURVIVAL SURVIVAL
0.2 0.2
0.0 0.0
TNA SAV TNA SAV ZYGOPTERA 1.0 CHIRONOMIDAE 1.0
0.8 0.8
0.6 0.6
0.4 0.4 SURVIVAL SURVIVAL
0.2 * 0.2
0.0 0.0
TNA SAV TNA SAV
1.0 1.0 1.0 PLANARIIDAE
0.8 0.8 0.8
0.6 0.6 0.6
0.4 0.4 0.4 SURVIVAL SURVIVAL SURVIVAL
0.2 0.2 0.2
0.0 0.0 0.0 GASTROPODA OSTRACODA TNA SAV TNA SAV TNA SAV
Figure 3. Survival under hypoxia (mean ± standard error) of organisms associated with T. natans (TNA) and submerged aquatic vegetation (SAV) in six (■) and eighteen (●) hours exposure. The * indicates significant difference between TNA and SAV (p<0.05).
VII-16
DISCUSSION
Mixed support was found for the hypothesis that macroinvertebrates associated with TNA beds are more resistant to hypoxia than those inhabiting SAV. Data for
Amphipoda and Chironomidae supported this hypothesis, while data for the other tested groups did not. Thus, the results indicate that the differences in community composition found by Strayer et al. (2003) between TNA and Vallisneria americana (which is one of the two species in SAV) stands can be attributed only in part to differential resistance to hypoxia.
The mechanisms involved in some groups’ resistance to hypoxia can only be inferred to a certain extent. Higher survival of Chironomidae associated with TNA could be related to the presence of hemoglobin in animals living in TNA beds and its absence or lower quantity in SAV organisms, since there is great inter-specific variation in hemoglobin content among chironomids (Panis et al. 1996). However, Strayer et al.
(2003) found more hemoglobin-bearing chironomid species in V. americana than in TNA beds, so this might not be the main mechanism here. Chironomids also possess morphological adaptations, such as thoracic horns and fringed anal lobes that tend to be more developed in species coping with low DO (Marziali et al. 2006). These structures, together with undulatory movements of the abdomen (Panis et al. 1996), can increase
ventilation, but the animals would still rely on DO. It has been shown though that some
chironomids can turn to an anaerobic metabolism that could give them resistance to short
(a few days) and even long-term (a few months) hypoxia (Frank 1983; Hamburger et al.
1995; Hamburger et al. 2000). Species differ in their ability to perform anaerobic
VII-17 metabolism (Frank 1983), and this could account for the higher survival of TNA- dwelling chironomids in relation to the SAV ones.
The other group that showed higher survival to hypoxia when collected from TNA beds was Amphipoda, in the six hour trials. In the eighteen hour trials, however, all animals died, revealing that their resistance is limited to short-term exposure to hypoxia.
Previous studies (Hoback and Barnhart 1996; Spicer et al. 2002) showed that the species
Gammarus pseudolimnaeus and Gammarus pulex were highly sensitive to low DO, with up to 100% mortality in 90 minutes of anoxic conditions for G. pulex. An attribute that could enhance survival of amphipods in low DO environments would be their ability to escape to more oxygenated areas due to their good swimming skills; however, in this experiment, this mechanism was impaired because the animals were enclosed in an arena with uniformly low DO. Moreover, in the field, this behavior might increase predation risk and thus mortality (Kolar and Rahel 1993). The presence of the respiratory pigment hemocyanin was demonstrated in many amphipods (Stucliffe 1984; Spicer 1993) and it might be an important physiological adaptation to hypoxia. On the other hand, a low DO- induced anaerobic metabolism was shown to be poor and of little importance for G. pulex
(Spicer et al. 2002), so the survival during short-term hypoxia of Amphipoda species in this experiment is more likely to be related to the presence of respiratory pigments.
Out of the groups tested, Zygoptera was the most strongly impaired by hypoxia, with no survival at all in low DO at either exposure time and regardless of the origin of the organisms. As tracheal gill breathers, they do tend to be more sensitive to hypoxia than other groups (Chapman et al. 2004). It was also observed that these animals were much more abundant in SAV (data not shown), meaning that DO might play a part in
VII-18
their distribution. They possess a morphological accessory, their caudal lamellae, that
can assist oxygen uptake among other functions (Apodaca and Chapman 2004;
Sesterhenn et al. 2013); however, even though most of the animals used had their
lamellae intact, they did not survive even the short-term hypoxia. This suggests they must
possess some kind of behavioral adaptation to escape it, since they were found in TNA
beds. In fact, this group is known to migrate towards the water’s surface to find higher
DO concentrations (Robinson et al. 1991; Apodaca and Chapman 2004; Sesterhenn et al.
2013). This might be an important mechanism for the Zygoptera in Hudson River
habitats, since Strayer et al. (2003) found them to occur on plants but not sediments, and
Kornijów et al. (2010) collected them from TNA leaves but not from other parts of the
plants, suggesting a dependence on more oxygenated areas of TNA beds.
At the other extreme, Ostracoda were highly resistant to hypoxia, even in the eighteen hour exposure treatment, regardless of origin, with an average survival of 80%.
These results agree with former works that showed high survival of freshwater ostracods under hypoxia (Hagerman 1969; Rossi et al. 2002), although others have stated Ostracoda
can be very sensitive to low DO levels (Rosseti et al. 2004; Ruiz et al. 2013). A migratory
behavior towards oxygen-rich microhabitats was described by many authors as an
adaptation of ostracods to hypoxia (Hagerman 1969; Corbari et al. 2004; Corbari et al.
2005), especially because they seem incapable of regulating their oxygen uptake
according to DO availability (Corbari et al. 2004; Corbari et al. 2005). Although no
evidence was found in the literature for anaerobic metabolism in ostracods (this study;
Rossi et al. 2002), some species contain hemoglobin (Fox 1957), which may assist their
survival under hypoxia.
VII-19
Gastropods did not differ in survival by habitat of origin and had relatively high
survival rates under hypoxia. Calow (1975) showed around 60% survival after five days
of anoxia for Ancylus fluviatilis and up to 80% after weeklong anoxia for Planorbis
contortus (values estimated from his figure 7). Animals in this group can be air- or
water-breathers or both, and although some species might rely on their access to the
atmosphere (Penha-Lopes et al. 2010; Seuffert and Martin 2010), some, even pulmonates,
do not (this study; Burky and Burky 1977; Hanley and Ultsch 1999). The presence of
hemoglobin was considered unimportant for Planorbella duryi and Helisoma anceps,
since they showed similar responses to hypoxia as non-hemoglobin-bearing species
(Hanley and Ultsch 1999). Also, pulmonate and prosobranch snails did not differ in their
responses to hypoxia (Hanley and Ultsch 1999). No papers were found demonstrating the
occurrence of anaerobic metabolism in Gastropoda, so their survival is more likely to be
related to their great capacity of regulating their oxygen consumption under varying DO
concentrations (Hanley and Ultsch 1999).
Planarians also showed high resistance to hypoxia and no effect of origin in these
experiments; however, Rivera and Perich (1994) identified DO as the most critical out of
six variables affecting survival and reproduction of four species of planarians (Dugesia
dorotocephala, D. tigrina, Cura foremanni and Dendrocelopsis vaginatus). Their
experiments consisted of a two week exposure to different levels of DO, and none of the
species survived the 0 and 2 mg.L-1 DO levels. In this shorter-term (18 hours) hypoxia
treatment though, planarians were very successful. Some flatworms contain hemoglobin
(Crompton and Smith 1963), but no studies were found reporting the occurrence of anaerobic metabolism in these animals.
VII-20
Macroinvertebrates inhabiting hypoxic macrophyte beds such as TNA beds in the
Hudson have two alternatives to cope with the hypoxic habitat within the beds: they can escape to more oxygenated micro-habitats, or possess adaptations to survive through hypoxia. The first alternative was not available for the animals in this experiment, so with the exception of Zygoptera, which must survive hypoxia by escaping to oxygenated zones, all other groups tested must have adaptations to survive hypoxia in place. Some groups in these experiments (Gastropoda, Ostracoda, Planariidae) displayed the same ability to survive hypoxia whether their habitat requires them to do so (TNA beds) or not
(SAV). On the other hand, Chironomidae and Amphipoda inhabiting TNA beds seem to have been selected by their better resistance to hypoxia than their relatives in SAV.
Because of their engineering activities on DO, macrophyte beds can impose challenges as well as benefits for the macroinvertebrates that live there. Whether through their regular dark respiration and daylight photosynthesis or growth form-related DO fluctuations, large macrophyte beds represent habitats with abiotic filters that may require specific adaptations of organisms. The results show that these challenges can be met in several different ways by macroinvertebrates: probable movement to nearby oxygenated habitats by zygopterans, short-term (6-18 hour) tolerance of hypoxia by amphipods and chironomids, and long-term (>18 hour) tolerance of hypoxia by gastropods, ostracods and planarians. The fact that different invertebrates meet the challenge of hypoxia in different ways means that different kinds of hypoxic macrophyte beds may support different assemblages of macroinvertebrates. Specifically, the spatial and temporal extent of hypoxia will be critical (whether well-oxygenated refuges are nearby, or whether the hypoxia lasts for longer than the invertebrate can tolerate). Likewise, shifts in
VII-21 macrophyte species composition caused by species invasions or other reasons may have differential effects on macroinvertebrates that are caused by different oxygen regimes.
VII-22
ACKNOWLEDGEMENTS
I would like to acknowledge the Hudson River Foundation and the Tibor T.
Polgar Fellowship Program for funding this research and Dr. David L. Strayer for the opportunity to work with him and his valuable advising. I am also deeply grateful to
Mary P. Budd for her help in sampling, sorting of organisms and especially for the great
time we had while working together. I am also grateful to David T. Fischer, Paul T.
Henne and all Cary Institute’s personnel, who were always welcoming and willing to
help, and my advisor in Brazil Dr. Sidinei M. Thomaz for always encouraging me to
move forward in science.
VII-23
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VII-29
THE DISTRIBUTION AND FEEDING ECOLOGY OF LARVAL SEA LAMPREYS IN THE HUDSON RIVER BASIN
A Final Report of the Tibor T. Polgar Fellowship Program
Thomas M. Evans
Polgar Fellow
Department of Environmental and Forest Biology State University of New York, School of Environmental Science and Forestry Syracuse, NY 13210
Project Advisor:
Karin E. Limburg Department of Environmental and Forest Biology State University of New York, School of Environmental Science and Forestry Syracuse, NY 13210
Evans, T. E. and K. E. Limburg. 2015. The Distribution and Feeding Ecology of Larval Sea Lampreys in the Hudson River Basin. Section VIII: 1-50 pp. In D. J. Yozzo, S. H. Fernald and H. Andreyko (eds.), Final Reports of the Tibor T. Polgar Fellowship Program, 2013. Hudson River Foundation.
VIII - 1
ABSTRACT
Anadromous fishes are important members of the Hudson River fish community,
but not all members have been well studied. Sea lampreys (Petromyzon marinus) have largely been ignored, and as a result even their distribution within the Hudson River is poorly understood. Sea lampreys are currently present in four tributaries of the Hudson
River: 1) Cedar Pond Brook, 2) Catskill Creek, 3) Roeliff Jansen Kill, and 4) Rondout
Creek. Catskill Creek and its tributary, Kaaterskill Creek, produce the majority of sea lamprey in the Hudson River. The Roeliff Jansen Kill is also important for sea lampreys in the Hudson River, as sea lampreys appear to grow faster here than at other tributaries.
Sea lampreys are likely to increase their range in the Hudson River in the near future as the removal of barriers to migration continues, but global climate change poses a serious long term threat to lamprey populations in the Hudson River.
Isotopic analysis demonstrated that site influenced the importance of terrestrially derived plant material (allochthonous) and aquatic plant sources (autochthonous) to sea lamprey larvae. For instance, larval lamprey from the Kaaterskill Creek depended on up to 60% on allochthonous sources, while those at Cedar Pond Brook only obtained ~1% of their nutrition from the same source. Changes in watershed land use were not associated with nutritional source dependence, and local sources may be more important than watershed level effects. Larval lamprey were isotopically distinct from all macroinvertebrates measured, suggesting that they exploit resources in different proportions from measured invertebrates and help more fully utilize organic material in streams. This study provides important data about the current distribution of sea lampreys in the Hudson River, and is the first to examine sea lampreys and the invertebrate community simultaneously with stable isotopes.
VIII - 2
TABLE OF CONTENTS
Abstract ...... VIII-2
Table of Contents ...... VIII-3
List of Figures and Tables ...... VIII-4
Introduction ...... VIII-5
Methods...... VIII-9
Results ...... VIII-16
Discussion ...... VIII-22
Conclusions ...... VIII-42
Acknowledgements ...... VIII-44
References ...... VIII-45
VIII - 3
LIST OF FIGURES AND TABLES
Figure 1: Tributaries to the Hudson River below the Troy Dam (A), in the
mid-Hudson River(B), and in the Lower Hudson (C) sampled ...... VIII-10
Figure 2: Logistic regression model predicting the presence or absence of sea
lamprey (Petromyzon marinus) in Hudson River ...... VIII-18
Figure 3: Average size of young of year lamprey from all streams ...... VIII-21
Figure 4: Stable isotope ratios of macroinvertebrates, young of year sea lamprey
larvae, and potential food sources ...... VIII-23
Figure 5: Median percent contributions of different nutritional resources to
macroinvertebrate and sea lamprey larvae ...... VIII-39
Table 1: Sites assessed for sea lamprey (Petromyzon marinus) in the Hudson
River ...... VIII-17
Table 2: Physical and chemical properties measured at the sample sites in the
Hudson River ...... VIII-19
Table 3: Macroinvertebrate data collected in the present study listed by site ..... VIII-20
Table 4: Sites for which sea lamprey (Petromyzon marinus) have been
reported in the Hudson River and the capture confirmation or explanation
of discordance ...... VIII-24
VIII - 4
INTRODUCTION
Sea Lamprey in the Hudson River
Anadromous fishes, or those species that spawn in freshwater but migrate to
marine ecosystems for the majority of their growth, are important members of many
coastal ecosystems. The historic abundance of anadromous fishes, combined with their
ease of capture during spawning migrations, has provided economic and social opportunities for numerous societies (Hardy 1999; Bolster 2006; Bottom et al. 2009).
Early European settlers, familiar with migratory fishes, were still impressed by the abundance and diversity of species that greeted them upon their arrival to North America
(Collette and Klein-Macphee 2002). However, after initially utilizing these vast resources, Europeans rapidly overexploited the fisheries, while simultaneously limiting access to and destroying the freshwater habitats upon which anadromous fishes depended
(Hardy 1999; Lichter et al. 2006; Limburg and Waldman 2009). Current estimates suggest that >90% of the historic biomass of these migratory fishes has already been lost
(Limburg and Waldman 2009).
The Hudson River is a biologically diverse ecosystem which currently contains ten anadromous fish species (Levinton and Waldman 2006) and has been the focus of extensive research (Jackson et al. 2005). The anadromous fish community in the Hudson
River has been, in general, well studied. However, research efforts have not been evenly divided between all species, and gaps in the current understanding of this important group remain (Waldman 2006). Sea lampreys (Petromyzon marinus) are the most poorly studied member of the Hudson River anadromous fish community. Sea lampreys were neglected because they were not considered economically or socially important in North America,
VIII - 5
are often difficult to sample effectively, and perceptions about this species have been
dominated by research efforts on invasive populations in the upper Great Lakes.
Sea lampreys belong to the most primitive group of vertebrate lineages (Order:
Cyclostomata), and exhibit an anadromous, semelparous (i.e., a species that spawns only once) life history. Unlike most other anadromous fishes, which spend a shorter time in fresh than in the marine waters, sea lampreys spend protracted periods in freshwater (up to 17 years) before a relatively brief time at sea (1-3 years; Beamish 1980). In addition, unlike many other anadromous fishes in the Northeast where at least some adults are iteroparous (i.e., they spawn more than once), sea lampreys are an obligatory semelparous species (Saunders et al. 2006). Because of the limited information available, there is no way to know historical abundance, or which streams were utilized by sea lampreys prior to large scale habitat alteration by European settlers. Even recent records of sea lampreys (i.e., those within the last 100 years) are difficult to find, and are often anecdotal.
Although sea lampreys have not often been considered an important member of the anadromous fish community, they may provide or have provided many of the same ecosystem services as other anadromous species (Holmlund and Hammer 1999). Sea lamprey larvae and carcasses of spent adults may have provided important food subsidies to freshwater aquatic and terrestrial systems (Nislow and Kynard 2009). In Maine, sea lamprey spawning coincides with first feeding of salmonid parr (juveniles), and spent adults, eggs, and embryos of sea lampreys may all provide important food sources to these young salmonids (Saunders et al. 2006). In addition, sea lampreys continue to spawn after other anadromous species during the spawning run (Saunders et al. 2006). A
VIII - 6
sea lamprey adult was observed at a spawning site as late as June 26th during the course of this study. Also, sea lamprey body shape and swimming style are unlike any other anadromous species, and they may prefer habitats not utilized by other anadromous species. In the Great Lakes (where sea lampreys are invasive) barriers are designed that specifically exploit their climbing style and exclude them from upstream reaches
(McLaughlin et al. 2006).
Stable Isotope Analysis of Lower Trophic Levels in Stream Food Webs
To better understand what role sea lampreys may play in freshwater food webs stable isotope ratio analysis was used to examine young-of-year (YOY) larval lamprey, different functional feeding groups of aquatic insects, and primary producers. Briefly, many elements consist of two or more stable isotopes that differ slightly in mass due to different numbers of neutrons in the nucleus. These small differences can be quantified with mass spectrometers, and the ratios can provide important insights into biophysical processes. All living organisms develop isotopic signatures based upon the source materials they utilize for growth and development (Michener and Lajtha 2007). The lighter isotope is preferentially incorporated into biological materials because it is energetically less expensive to fix (i.e., fractionation), but, when the organism produces waste, it favors compounds with the lighter isotope for expulsion. Ultimately, the organism becomes more enriched in the heavier isotope than the food source. These differences in fractionation develop natural tracers in food webs, which can be used to follow the flow of nutrients as they move through the ecosystem. The predictable
VIII - 7
fractionation of isotopes as they move through food webs also allows for the estimation
of the trophic level at which an organism is feeding (Peterson and Fry 1987).
Although extensive visual examination of larval lamprey gut content has
documented materials they consume (primarily detritus), the source(s) of the detrital
material is still largely unknown (Moore and Mallatt 1980, Sutton and Bowen 1994,
Mundahl et al. 2005). In contrast to visual identification methods, natural abundance
isotopic analysis represents potentially a more robust and quantitative approach for assessing the sources of organic matter (OM) supporting the somatic growth and
energetic maintenance of an organism or population (Cole and Solomon 2012; Roach et al. 2011). Naturally occurring isotopes of C, N, H, O and S in OM may be used to assess, both qualitatively and quantitatively, different OM sources supporting organism nutrition
and through food webs and even ecosystems, provided the isotopes of these elements can
be measured in both the potential food sources and the target organisms (Peterson and
Fry 1987; Michener and Lajtha 2007). Simultaneous use of multiple natural isotopes can
help resolve with greater accuracy the dietary and nutritional sources of a given species
than with a single isotope (Peterson and Fry 1987; Caraco et al. 2010). Natural abundance stable isotopes can also be used to help establish at which trophic level organisms are feeding using predictable fractionation values (Post 2002).
Understanding the nutritional sources supporting larval lamprey is important for their conservation and to understand better how they interact with the larger stream community. In addition, although stable isotopes have only been applied in a limited number of studies of larval lampreys (Hollett 1995; Shirakawa et al. 2009; Limm and
Power 2011), isotopes have been widely used to examine macroinvertebrates (Vander
VIII - 8
Zanden and Rasmussen 1999; Finlay 2001). Many macroinvertebrates have a similar feeding ecology to larval lamprey, but to date only a single study has simultaneously
examined larval lampreys and stream macroinvertebrates with stable isotopes together
(Bilby et al. 1996). All species of larval lampreys are filter feeders (Mallatt 1982), as are many aquatic invertebrates (Voshell 2002), but it is unclear if they utilize food sources in the same proportions. This study attempts to develop a more complete picture of the stream community by concurrently examining stable isotopes of both invertebrates and lamprey.
Therefore, the purpose of this study was twofold: 1) to identify the current distribution of sea lampreys in the Hudson River estuary below the Troy Dam, and 2) to determine the likely OM supporting YOY larval lampreys and selected macroinvertebrates utilizing stable isotope analysis. It was hypothesized that YOY sea lamprey nutrition will be predominantly composed of allochthonous materials, and that larval lamprey will be isotopically similar to macroinvertebrate filterers. It was also
hypothesized that the increased urbanization and agricultural land uses within the
watershed will increase the reliance on autochthonous sources, because more nutrients
and light will enter the stream. This will be reflected by more positive isotopic values in
the consumers (including sea lamprey larvae).
METHODS
Surveys for Sea Lamprey
In the present study, 23 tributaries of the Hudson River below the Troy Dam were
surveyed for sea lampreys in the summer of 2013 (Figure 1). At 19 of these sites a full
VIII - 9
A B C
No record and not captured Recorded but not captured Record and capture
Figure 1: Tributaries to the Hudson River below the Troy dam (A), in the mid-Hudson River (B), and in the lower Hudson (C) sampled. The length of each stream is the area below the first impassable barrier for sea lamprey (Petromyzon marinus). Numbers are tributaries and are as follows: 1-Poesten Kill, 2-Normans Kill, 3-Vloman Kill, 4-Vlockie Kill, 5-Muitzes Kill, 6-Hannacroix Creek, 7-Coxsackie Creek, 8-Stockport Creek (including Claverack and Kinderhook Creek), 9-Catskill Creek, 10-Kaaterskill Creek, 11-Roeliff Jansen Kill, 12-Saw Kill, 13-Rondout Creek, 14-Indian Kill, 15-Black Creek, 16-Quassaick Creek, 17-Moodna Creek, 18-Indian Brook, 19-Annsville Creek, 20-Cedar Pond Brook, 21-Minisceongo Creek, 22-Furnace Brook, 23-Croton River.
VIII - 10 workup was performed, while at the remaining four only sampling for sea lampreys was performed. A full workup included measuring local habitat characteristics, physical and chemical sampling of the stream, a macroinvertebrate sample, and a survey for sea lampreys at the site. In those sites where a full workup was not conducted, conditions were determined to be extremely unlikely to support lampreys, usually because of siltation or high water temperatures. However, to validate the expectation, sampling for lamprey was conducted in the same manner as at all other sites. Individuals who visited the site during sampling were also interviewed. Any pertinent information they provided to the study was noted and is reported.
To determine the composition of the surface area of the sampled stream reach, a diagonal transect across the stream from bankfull to bankfull was walked. Bankfull was defined as the point at which water would exit the river and enter the flood plain. At every step along the transect two points were randomly selected and the bottom composition was determined following the New York State Department of Environmental
Conservation (NYSDEC) bottom composition scale for at least 100 points (NYSDEC
2009). If a transect was too short to sample 100 points, a second transect was set and the process repeated.
Aquatic macroinvertebrate samples were collected by sampling the nearest riffle to the sample site, walking along a 5 m diagonal transect, and kicking into a D-frame net
(mesh size 500 µm) for five minutes. During two sampling events the 500 µm net was lost and a 600 µm sampler was used instead (at the Black Creek and Moodna Creek). The entire sample was immediately placed in 70% ethanol and stored at ambient temperature.
Upon return from the field invertebrates were picked from detrital material and classified
VIII - 11
to order. Cleaned invertebrate samples were then subsampled in a tray divided into 25
blocks (1.7 cm2 per block). Blocks were randomly selected and all macroinvertebrates
with heads present in the block were identified to genus. Macroinvertebrates which could not be identified to genus were identified to the next lowest level and then assigned to a genus, based upon the assignment of other positively identified members of the same group. Blocks were selected until at least100 individuals had been identified. Individuals belonging to Chironomidae and Simuliidae were never classified below the family level, and did not count towards the 100 classified macroinvertebrate individuals. The number
of individuals belonging to each group in the overall sample was calculated by
multiplying the measured proportions in the subsample by the number of individuals in
the overall sample. The estimated sample was then used to calculate the Shannon diversity index (SDI) and Hilsenhoff‘s Biological Index (HBI). The SDI was calculated as follows:
[1] = 𝑅𝑅 ′ 𝑆𝑆 − � 𝑝𝑝𝑖𝑖𝑙𝑙𝑙𝑙 𝑝𝑝𝑖𝑖 𝑖𝑖=1 Where pi is the proportion of individuals in the overall sample (excluding Chironomids)
belonging to the lowest taxonomic group identified. The HBI was calculated as follows:
[2] = 𝑅𝑅 N 𝑛𝑛𝑖𝑖𝑎𝑎𝑖𝑖 𝐻𝐻′ � 𝑖𝑖=1 Where ni is the number of animals in the taxonomic grouping, ai is the biological
tolerance value, and N is the total number of animals in the sample (excluding
Chironomids; Bode et al. 1996).
To determine if sea lamprey were present at a site, a backpack electrofisher
(Haltech, HT-2000) was used to electrofish for larval lampreys. A low shock (5 Hz, 150 VIII - 12
V, 2:2 pulse pattern) was used to cause larval lampreys to emerge from the substrate.
Animals that emerged from the substrate were captured before they could burrow. It was observed that the weak pulsed electrical output encouraged larval lampreys to emerge and attempt to burrow rapidly, in contrast to a continuous pattern which generally caused them to forego burrowing and attempt to escape the electric field or remain burrowed.
Electrofishing occurred for 15 minutes at each site, which allowed the sampling of all habitats that appeared likely to support sea lamprey larvae (a mixture of loosely packed sand and fine organic matter; Slade et al. 2003).
Collections for Stable Isotopes Analysis
For stable isotope analysis, YOY larval lampreys were collected following the procedure described above. The smallest size class at a site was assumed to be YOY
(Hardisty 1961). Larval lamprey were then stored at ~0°C in sand and stream water from where they were collected, until they could be returned to laboratory to be frozen at -
20°C. Animals were never held alive for more than 24 hrs. Sand was added to allow larval lamprey to rest; otherwise, it was observed that they would continue to attempt to burrow until they were too exhausted to move. Common aquatic invertebrates of the local macroinvertebrate community were collected by kick netting in nearby riffles and then hand-picking individuals. Predatory animals were separated from other groups and all were stored in stream water at 0-4°C for 36-48 hours to allow them to void their guts.
Potential primary food sources were also collected simultaneously by hand picking terrestrial leaves, soil, aquatic plants and algae, and undisturbed sediment from the areas in which ammocoetes were collected.
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Samples for stable isotope analysis were dried to constant mass at ~60°C and then
homogenized by grinding in a clean porcelain mortar. Samples were then stored in
desiccation chambers. Stable isotope values were calculated as follows:
[3] = 1000 1 𝑅𝑅𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠 𝛿𝛿𝛿𝛿 � 𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠 − � where Rsample is the𝑅𝑅 ratio of the heavy to light isotope in the sample and Rstandard is the ratio in an agreed upon standard. To eliminate leading zeroes and ease readability the sample was then multiplied by 1000. Values were reported as “per mil” (‰) and designated by the notation δ with a superscript designating which heavy isotope has been measured. Subsamples of every sample type were packed in clean tin capsules and analyzed for δ13C and δ15N at the University of California, Davis (PDZ Europa ANCA-
GSL (EA) attached to a PDZ Europa 20-20 isotope ratio mass spectrometer (IRMS)).
Standard deviations for replicate analyses of standards for the instrument were < 0.2‰
for δ13C and < 0.3‰ for δ15N.
Analysis of Distribution, Abundance, and Isotope Data
A logistic regression model was used to quantify the probability of encountering
sea lampreys at a site (i.e., presence or absence). Logistic regression calculates the
probability of an event occurring based upon categorical or quantitative variables. If sea
lamprey presence could be linked to macroinvertebrate data, extensive archived data on
the latter could be used to predict potential sea lamprey distribution. The final model
included only the percent Ephemeropteran (mayfly) abundance, as no other measured
values were important for the model.
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To determine if sea lamprey isotopic values were driven by land use changes, the percent land use in each watershed was calculated from the 2006 National Land Cover
Database (Fry et al. 2011) and the National Hydrography Dataset (USGS 2013). Land use was divided into four groups: 1) urban/developed, 2) agriculture/pasture, 3) forested/natural, and 4) open water. The percent land use of the watershed was correlated with the average isotopic values of YOY larval lampreys from that site.
Stable Isotope Mixing Modeling
The Bayesian stable isotope mixing model MixSIR (Moore and Semmens 2008)
was used to estimate the contributions of potential food sources to YOY larval lamprey nutrition. Bayesian statistics allow for backward reasoning (after observation of an event)
to predict the likely occurrences that led to that observation. MixSIR applies this
statistical analysis to stable isotope mixing models, allowing for the incorporation of the uncertainties around source isotopic values and fractionation estimates to better predict food source dependence and confidence of the importance of a given dietary item to the organism (Moore and Semmens 2008).
Each site was modeled separately because of apparent differences in isotopic values of primary producers and consumers. Potential food sources for lampreys and invertebrates at each site were divided into two groups (i.e., autochthonous and allochthonous). Autochthonous sources were considered to be all types of aquatic plants
collected at a site, including algae and macrophytes (e.g., Elodea sp., Potamogeton sp.,
and Myriophyllum sp.). Allochthonous sources were the terrestrial plants collected at a
site and included Acer spp., Vitis riparia, Quercus sp. and Rosa multiflora. Terrestrial
VIII - 15
surface soils and aquatic surface sediments were also considered for inclusion within
allochthonous and autochthonous sources in the model respectively, but they were
intermediate between these terrestrial and aquatic plants. Therefore, because these
sources were likely amalgamations of terrestrial and aquatic plants, which were already
accounted for in the model, they were not included explicitly in the final model. The final
models included all the measured terrestrial plants at a site as the allochthonous source,
and all of the measured aquatic plants, including algae, as the autochthonous source.
Isotopic values of larval lampreys, invertebrates, and potential nutritional sources
were determined from measurements of samples collected in the present study. On the
basis of prior work (Sutton and Bowen 1994), larval lampreys were assumed to be one
trophic level above primary producers and follow published fractionation values of
0.4±1.3‰ for δ13C and 3.4 ± 1.0‰ for δ15N (Post 2002). Invertebrates were also
considered to be one trophic level above primary producers and the same fractionation
values as for larval lampreys were utilized. All models were run with 1,000,000 iterations
(i.e., MixSIR attempted to find a solution to the model for each iteration). Results
conformed to the recommended guidelines for determining if the model output had
estimated true posterior distributions (Moore and Semmens 2008).
RESULTS
Sea lampreys were found in four tributaries to the Hudson River: the 1) Roeliff
Jansen Kill (Columbia County, confluence at river km 178), 2) Catskill Creek (Greene
County, river km 182), 3) Rondout Creek (Ulster County, river km 146), and 4) Cedar
Pond Brook (Rockland County, river km 63; Figure 1). Sea lampreys were not found in
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all Hudson River tributaries for which prior records of these animals existed (Table 1;
Figure 1).
Table 1: Sites assessed for sea lamprey (Petromyzon marinus) in the Hudson River. Lamprey River Lamprey Stream Name County Prior Source observed in km Reported this study Annsville Creek Westchester 71 N NA N Black Creek Ulster 135 Y Schmidt and Limburg 1989 N Catskill Creek Greene 182 Y Numerous Y Cedar Pond Brook Rockland 63 Y www.piplon.org (2002) Y Coxsackie Creek Greene 204 N NA N Croton River Westchester 55 N NA N Furnace Brook Westchester 63 N NA N Hannacroix Creek Albany 212 Y Greeley and Bishop 1933 N Indian Brook Putnam 85 N NA N Indian Kill Dutchess 137 N NA N Schmidt and Cooper 1996 Kaaterskill Creek Greene 182 Y Y Bryan et al. 2005 Moodna Creek Orange 92 N NA N Minisceongo Rockland 64 N NA N Creek Muitzes Kill Rensselaer 217 N NA N Normans Kill Albany 225 N NA N New York State Museum specimen Poesten Kill Rensselaer 241 Y N (2007) Quassaick Creek Orange 100 Y NYSDEC specimen (2005) N Roeliff Jansen Brussard et al. 1981 Columbia 178 Y Y Kill Waldman 2006
Greeley and Greene 1937 Rondout Creek Ulster 146 Y Y Hudson River Almanac 2007b Saw Kill Dutchess 158 Y Smith 1985 N Hudson River Almanac 2005 Stockport Creek Columbia 195 Y N Anonymous 2013 Vlockie Kill Rensselaer 219 N NA N Vloman Kill Albany 220 N NA N
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Sea lamprey presence was not correlated with percent bottom cover (i.e., percent
cover of rock, rubble, gravel, sand, or silt; Table 2), percentage of Trichopterans,
Plecopterans, Dipterans, or Chironomids in the invertebrate sample, species richness, the
SDI, or the HBI (Table 3). Although it was not significant (Chi-square, df = 1, X2 = 2.20, p-value = 0.14), sea lampreys were never found at sites that did not also have
Plecopterans. A logistic regression found that sea lamprey presence was likely (> 50%)
when the percent of Ephemeropterans exceeded 44% (coefficient estimate = 0.10,
standard error = 0.05, z-value = 2.05, p-value = 0.041; Figure 2). Yes Presense of Sea Lamprey
1 ( ) = 1 + e ( 4.60 + 0.104 ) 𝐿𝐿 𝑥𝑥 − − 𝑥𝑥 No Plecopterans not present Plecopterans present
0 20 40 60 80 100 Percent Ephemeroptera in Sample
Figure 2: Logistic regression model predicting the presence or absence of sea lamprey (Petromyzon marinus) in Hudson River tributaries as a function of percent Ephemeroptera in the macroinvertebrate sample. Plecoptera presence was not included in the model, but sea lamprey were always found at sites which also had Plecopterans.
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Table 2: Physical properties and pH measured at the sample sites in the Hudson River.
Width Bottom type (%) Stream Name County Embeddedness pH (m) Rock Rubble Gravel Sand Silt Annsville Creek Westchester 15.7 4 7 65 16 8 100 7.4 Black Creek Ulster 77.0 0 5 75 9 12 <20 8.0 Catskill Creek Greene 59.0 0 36 63 0 0 0 8.1 Cedar Pond Brook Rockland 17.5 4 37 46 12 2 <5 8.7 Claverack Creek Columbia 51.0 2 15 56 11 17 ND 8.5 Coxsackie Creek Greene 24.0 3 40 32 10 14 50 8.4 Croton River Westchester 140.0 ND ND ND ND ND 10 ND Furnace Brook Westchester 10.2 2 9 83 6 0 10 7.9 Hannacroix Creek Albany 16.5 22 36 33 5 4 ND 8.8 Indian Brook Putnam 6.0 6 22 68 2 2 <5 7.9 Indian Kill Dutchess 7.0 ND ND ND ND ND ND ND Kaaterskill Creek Greene 62.0 21 17 48 8 5 80 7.7 Kinderhook Creek Columbia 44.0 67 5 25 4 0 ND 8.2 Moodna Creek Orange 19.0 11 16 42 4 28 20 7.8 Minisceongo Creek Rockland 15.5 5 29 42 16 8 10 8.4 Muitzes Kill Rensselaer 19.0 7 35 36 10 12 60 8.5 Normans Kill Albany 45.0 83 5 1 5 6 0 8.0 Poesten Kill Rensselaer 14.5 35 23 14 24 4 10 8.0 Quassaick Creek Orange 24.4 ND ND ND ND ND ND ND Roeliff Jansen Kill Columbia 30.0 17 32 44 6 1 0 8.1 Rondout Creek Ulster 49.0 ND ND ND ND ND ND 7.8 Saw Kill Dutchess 18.0 35 19 40 0 5 80 8.0 Stockport Creek Columbia 335.0 0 10 67 12 11 66 8.3 Vlockie Kill Rensselaer 1.1 32 0 44 17 7 50 8.1 Vloman Kill Albany 125.0 ND ND ND ND ND 100 ND
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Table 3: Macroinvertebrate data collected in the present study listed by site.
Number of Percent Composition Stream Name County SDI HBI Species Ephemeroptera Plecoptera Trichoptera Coleoptera Diptera Chironomidae Non-insects Annsville Creek Westchester 16 1.68 4.75 10 1 9 51 14 13 14 Black Creek Ulster 18 2.27 3.70 10 3 38 7 39 38 0 Catskill Creek Greene 16 2.04 4.50 72 1 21 2 4 3 0 Cedar Pond Brook Rockland 8 1.57 5.37 52 <1 31 0 17 15 0 Claverack Creek Columbia 10 1.83 5.01 29 0 41 2 25 14 2 Coxsackie Creek Greene 3 0.13 5.98 1 0 0 1 31 31 67 Furnace Brook Westchester 8 2.61 4.51 6 0 53 30 10 6 1 Hannacroix Creek Albany 13 1.09 6.88 6 1 2 6 5 5 80 Indian Brook Putnam 15 2.16 3.07 18 22 48 1 10 9 1 Kaaterskill Creek Greene 13 1.90 4.65 51 <1 34 1 14 11 0 Kinderhook Creek Columbia 19 2.42 4.99 36 <1 27 24 12 4 0 Moodna Creek Orange 16 2.05 4.24 5 1 60 13 13 9 0 Minisceongo Creek Rockland 12 1.46 4.93 34 0 34 4 27 25 0 Muitzes Kill Rensselaer 14 1.36 4.60 7 0 16 15 60 52 2 Normans Kill Albany 19 2.48 5.73 40 1 2 4 48 46 4 Poesten Kill Rensselaer 21 2.33 4.72 40 2 15 3 40 36 0 Roeliff Jansen Kill Columbia 22 2.53 4.35 29 3 35 12 14 13 6 Rondout Creek Ulster 18 2.27 4.55 47 0 19 20 13 3 0 Saw Kill Dutchess 18 1.77 4.17 14 0 77 2 4 2 1 Stockport Creek Columbia 9 1.49 5.45 30 0 2 1 34 34 30 Vlockie Kill Rensselaer 14 1.99 4.35 8 0 30 53 9 6 0
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The average length of YOY larval lamprey was different between sites (ANOVA, df = 4, F-test = 11.53, p < 0.001). The Roeliff Jansen Kill YOY larval lamprey were larger (28.8±3.5 mm, SD) than larvae from all other sites (18.8±2.6 mm, SD; Figure 3). 40
b 30
a a a a Total Length (mm) 20 10 0
Kaaterskill Creek Catskill Creek Rondout Creek Cedar Pond Brook Roeliff Jansen Ki
Stream Name
Figure 3: Average size of young of year lamprey from all streams at which lamprey were captured. The black line is the median and boxes are the second and third quartiles. Sites that are different are marked by different letters.
Isotope Modeling Results
YOY larval lamprey isotopic values were different from measured macroinvertebrates at every site, even though many of the macroinvertebrate groups measured are also filterers (i.e., Hydropsychidae, and Isonychiidae; Figure 4).
Contributions of autochthonous and allochthonous sources to YOY larval lamprey varied
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by site from almost complete dependence on autochthonous sources (median contribution
98.7% at Cedar Pond Brook), to less than half of their nutritional needs (40.3% at
Kaaterskill Creek; Figure 5). Larval lamprey dependence on autochthonous or
allochthonous sources was not associated with urban land (df = 2, t-value = 1.6, p-value =
0.25), agricultural lands (df = 2, t-value = -1.17, p-value = 0.36), or forested lands (df = 2, t-value = 0.34, p-value = 0.77).
DISCUSSION
Sea lamprey presence/absence was associated with only a few of the metrics collected. These results are consistent with prior work that has found that sea lamprey distribution at the level of the stream difficult to predict from habitat data (Neeson et al.
2007). Sea lamprey larvae utilize marginal habitat in the hyporheic zone (i.e., area along the edge of the stream in which ground and surface waters mix), potentially limiting the feasibility of utilizing high level generalizations about streams to predict their presence.
Subsurface flows, ground water temperature, and oxygen content may all play important roles at the local level and are difficult to scale up. In addition, larvae may be experiencing conditions different from the surface waters and the macroinvertebrate communities that were examined during the course of this work.
Macroinvertebrate assemblages provide a tool to help other researchers in the
Hudson River select sites where they might expect to find sea lampreys. Plecopterans are associated with cool, unpolluted, highly oxygenated waters (Stewart and Stark, 2008), conditions which are highly favorable to sea lamprey growth and development (Morman
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Figure 4: Stable isotope ratios of macroinvertebrates, young of year sea lamprey larvae, and potential food sources. δ13C vs. δ15N for (a) Cedar Pond Brook, (b) Kaaterskill Creek, (c) Roeliff Jansen Kill, and (d) Rondout Creek
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Table 4: Sites for which sea lamprey (Petromyzon marinus) have been reported in the Hudson River and the capture confirmation or explanation of discordance.
Lamprey Lamprey Stream Name Reason for Disagreement With Historic Data Reported Observed
Black Creek Y N Single larvae captured during large sampling effort, animals may be extremely rare Catskill Creek Y Y
Cedar Pond Y Y Brook Hannacroix Y N Historic record (1934), possible lampreys currently spawn higher than sampling site Creek Kaaterskill Y Y Creek Poesten Kill Y N Single record of an adult, possible migrant looking for spawning site Quassaick Creek Y N Single record of a dead adult, possible migrant looking for spawning site Rondout Creek Y Y
Roeliff Jansen Y Y Kill Maybe currently extirpated, interview with an anonymous resident suggest this to be the case; Stockport Creek Y N no current observations Saw Kill Y N American Brook Lamprey record, sea lampreys may occasionally attempt to spawn here
et al. 1980; Jobling 1981). The association of sea lamprey with Ephemeropterans may also be because Ephemeropterans can be abundant and most diverse in rocky-bottomed, 2nd and 3rd order streams (Waltz and Burian 2008), often where adult sea lampreys spawn. In areas in which sea lampreys are no longer able to access historic grounds, the percentage of mayflies may provide a way to help establish where sea lamprey occurred historically and inform potential restoration efforts.
Assessed Tributary Narratives
A narrative of each sampled stream listed alphabetically by stream name, is reported below. Sites for which a record of sea lampreys existed are summarized in Table 4. This survey was conducted in the summer following Hurricane Sandy, and within two years of Hurricane
Irene and Tropical Storm Lee, all of which caused serious flooding in the Hudson River.
Flooding events can wash larval lampreys out of streams (Potter 1980), and these events likely
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depressed lamprey populations in the Hudson Valley. As a result, 2013 may have been an atypical year for sea lampreys in the Hudson River.
Annsville Creek (Westchester County, river km 71)
Annsville Creek enters the Hudson River in Peekskill, NY, but is surrounded by large stretches of forest. Residential communities are adjacent to much of the river and many small historic dams remain limiting habitat availability for sea lampreys. Sampling access below barriers to migration is also limited. Sea lampreys were not found at this site although the appropriate habitat for adult spawning and larval rearing appeared to exist. Plecopterans were present, but Ephemeropteran diversity (three genera) and percent composition (10%) was low. If efforts with local land owners were made to remove many of the small historical dams (which likely also present a public safety risk) sea lampreys and other migratory species might begin to exploit this stream.
Black Creek (Ulster County, river km 135)
Black Creek is a small, high quality stream south of Kingston, NY whose watershed is primarily forested. Black Creek still supports a large alewife run (Schmidt and Limburg 1989), and these workers also reported capturing a single sea lamprey larva. However, there are no reports of adults from this stream. There is abundant and extensive sandy habitat for larvae, and a series of riffles throughout the stream which would provide spawning habitat. In addition, the invertebrate community was very diverse, and included Plecopterans. However, it was composed primarily of Chironomids (38%), with Ephemeropterans making up only 10% of the invertebrate community (Table 3). Neither sea lamprey adults nor larvae were observed at this site, though
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the site was visited during the spawning period in June and then again in July. It is unclear why
they are not common within this stream, but it may be related to temperature. Upper reaches of
the Black Creek are dammed by many small (< 2 m) barriers, which potentially allow heating to
occur, even though large stretches are forested. Temperatures at or above 31°C are lethal to larval
lampreys in the laboratory (Jobling 1981). In natural systems even approaching thermal tolerance
values likely acts synergistically with other stressors to limit lamprey populations.
Catskill Creek (Greene County, river km 182)
Catskill Creek is known to support sea lampreys, and numerous authors have collected
sea lampreys or reported their spawning here (Greeley and Greene 1937, Smith 1985, Waldman
2006, NYSDEC 2012a). During interviews, multiple individuals reported seeing sea lampreys spawning there (Anonymous, personal communication). Catskill Creek splits into two tributaries outside of the city of Catskill, N Y, the southerly branch is the Kaaterskill Creek (discussed separately) and the northerly branch is Catskill Creek. Catskill Creek has a relatively short stretch of unobstructed stream that sea lampreys can access (~3 km) before they reach the first impassable barrier (a dam in Leeds, NY). Larval lampreys were easy to find (catch per unit effort
1.2 animals per minute) and a spawning adult was observed on June 26th. Catskill Creek appears
to be sediment starved (potentially as a result of dams further up river), and it may have supported higher numbers of sea lampreys in the past. The total number of lampreys utilizing this stream was hard to estimate because many areas of the stream were too deep (> 2 m) to sample
with a backpack electrofisher. Human activities upstream should take into account that sea
lampreys downstream may be vulnerable to degradation of the aquatic habitat, especially during
the May-June spawning period.
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Cedar Pond Brook (Rockland County, river km 63)
Cedar Pond Brook flows into the Hudson River at Stony Point, NY. The lowest portions
of this stream are heavily urbanized, but the headwaters are located in extensive state park lands.
Sea lampreys have been documented from Cedar Pond Brook recently (2002, www.piplon.org/fishes.htm), but their distribution and abundance within the river is unknown.
An impassable dam for sea lamprey is located ~ 4 km upstream of the mouth of this stream. Sea
lampreys utilized all portions of the stream that were above the tidally influenced areas (first two
km). The majority of larvae observed were YOY, and they were locally abundant (two
individuals/m2) but patchy. Sea lamprey larvae did not appear to utilize all appropriate habitats, and the majority of animals observed were from a single sand bar. In addition, animals older than
YOY were rare, constituting < 5% of the observed population. The apparent lack of age classes
suggests that sea lampreys are not surviving well in this stream. However, sea lampreys may still
be recovering from hurricanes that likely washed large numbers larvae out of the stream, and age
class structure may reappear with enough time.
To conserve sea lampreys in this stream, excessive salting of nearby areas during the
winter should be discouraged (salt is lethal to larval lamprey at low concentrations; Reis-Santos
et al. 2008). In addition, bank erosion, which is occurring at a number of locations within the
stream, should be addressed wherever feasible. Finally, removal or passage of dams may allow
sea lampreys to better exploit more upstream habitat.
Coxsackie Creek (Greene County, river km 204)
Coxsackie Creek, near Coxsackie, NY, is surrounded by agricultural lands, and is highly
influenced by their contribution. The invertebrate diversity was poor (three genera were present,
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besides Chironomids) and composed primarily of the crustacean Gammarus sp., which are omnivorous and tolerant (Table 3). Siltation was also clearly occurring, with 62% of bottom samples having 0.5 – 5 mm of deposition. Eutrophication is also likely occurring, as large mats of filamentous algae were present. No sea lampreys were observed while surveying, and in its current state this stream would not support them.
Croton River (Westchester County, river km 55)
The Croton River, near Croton-on- Hudson, NY, is fed by the New Croton Reservoir, the final receiving reservoir in the East-of-Hudson component of New York City’s water supply system. Eutrophication appeared to be an issue as water was turbid and visibility was limited to <
50 cm. Sea lampreys were not observed at this site, although abundant larval lamprey habitat appeared to be available. Only a short stretch would be available to sea lampreys and may preclude successful spawning.
Furnace Brook (Westchester County, river km 63)
Furnace Brook near Cortlandt, NY, is extensively dammed although there is a short stretch below the first complete dam and the head of tide sea lampreys could access. The invertebrate community was only moderately diverse (six genera, not including Dipterans), and included no Plecopterans and few Ephemeropterans (Table 3). Sea lampreys were not observed at this site, although it may have historically supported them. The extensive damming along this stream may be preventing natural processes and is likely degrading the quality of the limited habitat available. It is hard to recommend large scale efforts to remove or mitigate dams here,
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because of how numerous they are in the stream. Many of these dams are not functional, and
deserve to be inspected, as they may begin to pose a public safety hazard as they continue to age.
Hannacroix Creek (Albany County, river km 212)
Hannacroix Creek was historically a tributary utilized by sea lampreys (Greeley and
Bishop 1933) and is currently surrounded by forest and agricultural lands. Reaches lightly
influenced by tidal inputs were sampled, although these areas are not ideal for sea lampreys,
because access to higher reaches was difficult. Invertebrates were generally dominated by non-
insect groups (~80%, i.e., Crustacea, primarily Gammarus sp.), but Plecopterans (Agnetina sp.)
were present, suggestive of high quality habitat potentially upstream (Table 3). No larval
lampreys were observed though American eel (Anguilla rostrata) were found occasionally. Sea
lampreys may be able to exploit upper reaches, and further efforts should be made to determine if
they still utilize Hannacroix Creek.
Indian Brook (Putnam County, river km 85)
Indian Brook enters the Hudson River by flowing through Constitution Marsh, a large
wetland. The stream has a short stretch (~100 m) beyond head of tide of moderate gradient with
potential lamprey habitat, before a series of falls that would prevent migration further upstream.
The water quality at this site is likely extremely high as evidenced by the macroinvertebrate
community, which included three Plecopteran genera (Pteronarcys sp., Acroneuria sp., Leuctra
sp.) and seven genera which are listed as sensitive to very sensitive (e.g., Dolophilodes sp.,
Lepidostoma sp.; Table 3). During the spring a fyke net, operated by the Audubon Society, is
deployed across the mouth of the stream to capture in-migrating American eels (glass eels).
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Audubon staff reported that they have not seen or captured any lampreys. Water quality does not prevent lampreys from utilizing this stream, but the limited habitat available may prevent successful exploitation. Other tributaries to Constitution Marsh provide similar habitat, and lampreys may intermittently spawn in them, but they are unlikely to succeed to any serious
degree.
Indian Kill (Dutchess County, river km 137)
The Indian Kill is a small stream that passes through Margaret Lewis Norrie State Park
before meeting the Hudson River. No lampreys were seen during the survey at this stream
although American eels were common. Schmidt and Cooper (1996) said that this stream had high
water quality, but reported that DO was < 100% saturation. The stretch has excellent potential larval lamprey habitat as extensive loosely packed sand is present. However, the stream is very shallow and the survey did not observe any riffles that would support spawning adult lampreys.
Kaaterskill Creek (Greene County, river km 182)
Kaaterskill Creek is a tributary of Catskill Creek, but is more turbid and much sandier than the Catskill Creek. The additional sand provides excellent and extensive larval lamprey habitat. Interestingly, a private land owner reported, during an interview, that he had observed
“…thousands of eels below the falls [at the junction of Cauterskill Rd and Cauterskill Ave] when kayaking on the river” (Anonymous, personal communication). He had not determined if they were sea lampreys (which are often called lamprey eels) or American eel, but it appears likely they were sea lampreys. Sea lampreys can be visible in the day during the spawning run, and would congregate below a natural barrier if it slowed their migration upstream. In contrast,
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American eel generally are nocturnal and congregate in sanctuaries together but not in the open during daylight. Above the falls, larvae of multiple sizes were easily collected, suggesting multiple successful spawning events over the course of numerous years. The New York State
Museum also reported in their collection a lamprey more than 9 km from the falls. Therefore,
Kaaterskill Creek is likely the most important stream for sea lamprey in the Hudson River. Based on the unobstructed length of the Kaaterskill (~17 km) and the observation that habitat quality generally appeared to be high along its entire length, the majority of sea lampreys in the Hudson
River may originate in this river. Efforts should be made to continue “business as usual” in this stream as it appears critical for sea lamprey success in the Hudson River.
Minisceongo Creek (Rockland County, river km 64)
Minisceongo Creek, in West Haverstraw, NY, is a heavily urbanized stream that has extensive heavy industrial activity near its mouth. However, the water was cool and clear when sampling occurred. In addition, numerous trout (likely Brown, Salmo trutta or Brook trout,
Salvelinus fontinalis) were observed at the sample site. Sea lampreys were not captured at this site, although there appeared to be appropriate habitat for adults and larval lamprey. Large amounts of trash and some signs of sedimentation and pollution were also observed at the site. In addition, this stream is immediately adjacent to a river which does contain sea lamprey. Local conditions may be limiting sea lamprey usage of this stream, and cleanup efforts may be able to establish a sea lamprey run here.
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Moodna Creek (Orange County, river km 92)
It was not possible to sample the main body of Moodna Creek and the first tributary on the southern side was sampled instead. Moodna Creek was highly modified by hurricane Irene and large portions of both banks were eroded. Construction activities occurred throughout the summer to stabilize the banks and rebuild washed out bridges/roads which precluded access to the river. Sea lampreys were not captured in the tributary, but American eels were not observed either, suggesting a barrier further downstream. It is possible that the tributary has a blockage not visible from aerial photographs or roads that is precluding migratory fish access. Plecopterans were present at the sampled site, although mayflies were extremely rare (<5% of invertebrate community). Without being able to sample the main stem, the presence of sea lampreys in this tributary cannot be conclusively ruled out.
Normans Kill (Albany County, river km 225)
The Normans Kill is south of Albany, NY, and urban sprawl is encroaching on a number of banks, likely increasing erosion and sedimentation. The water of the main body was always turbid near Albany and further upstream, though it was visited during both low and high flow events. Upstream habitat is protected by forested areas in some regions, and some tributaries appear to be transporting large amounts of loosely packed sand, which would provide excellent habitat for sea lamprey larvae. There are natural rapids near the mouth of the river, but these likely do not prevent sea lamprey access to the river. The Normans Kill is noted as a high quality warm water stream (NYSDEC 2007), and as a result likely does not support lamprey. There are no reports of sea lampreys in the Normans Kill though the NYSDEC has surveyed the stream
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multiple times (NYSDEC 2007). If sea lampreys utilize the Normans Kill they may use small
tributaries that are cool enough to support larvae year round.
Poesten Kill (Rensselaer County, river km 241)
The Poesten Kill is a highly urbanized stream that passes through Troy, NY, and the
stream was filled with trash and refuse when sampled. However, the macroinvertebrate community was rich, and two species found at this site (i.e., Siphlonurus sp., Haploperla sp.) were found nowhere else in the Hudson River during the course of this study (Table 3).
American eels were extremely abundant, and both large and small eels were encountered frequently. The site appeared to offer excellent conditions for sea lampreys from a community perspective. However, no sea lampreys were found although a sea lamprey adult has been collected in recent years within the creek (Hudson River Almanac 2007a). The adult may have been brought in attached to another fish or may have been searching for spawning sites. It is also possible that adults do spawn in this stream but that conditions are inappropriate for larval development. American eel are known to feed on larval lamprey (Perlmutter 1951; Denoncourt and Stauffer 1993), and the proximity to the urban environment may result in high levels of salt runoff during winter.
Quassaick Creek (Orange County, river km 100)
Quassaick Creek is a small, highly urbanized stream, in Newburgh, NY with barriers to sea lamprey migration ~1 km from its mouth. Trash and filamentous algae were present throughout the stream when sampled. However, water clarity was high, and the bottom was visible at all depths (~1 m). No lamprey larvae were found, but American eels were common. A
VIII - 33 recent report (June 27 2005) of a deceased sea lamprey adult in this stream (Hudson River
Almanac 2005) suggest that adults do occasionally enter the creek. Currently, however conditions appear wholly unsuitable for larval growth and development. Besides a variety of pollutants, salting of the local roads may also pose a threat to sea lamprey larvae here.
Roeliff Jansen Kill (Columbia County, river km 178)
The Roeliff Jansen Kill (locally known as the “Roe-Jan”) is a high quality, cool water stream that has a long stretch (> 6 km) before the first barrier (NYSDEC 2012b). Land use in downstream reaches is primarily forest with small amounts of residential, while the upstream is primarily agricultural. Sea lampreys have been captured here before (Brussard et al. 1981;
Waldman 2006), and numerous individuals who were interviewed along the Roe-Jan had seen sea lampreys in the stream. One individual reported observing them spawning yearly at the Mill dam immediately below the Mill Road bridge (Anonymous, personal communication). At this location there is an old dam on top of a natural rapids which likely limits further upstream movement. High rains and the closure of the Mill Road precluded visual confirmation of adult migrations within the stream during the spawning run. Prior to construction of the dam, this may still have been the upper limit of sea lamprey migration. To confirm if the barrier was completely impassable, the survey sampled above the barrier where habitat conditions also appeared suitable for lampreys, but did not find any individuals. Likely lamprey redds were observed further downstream in June.
Larvae were abundant at the local level, but their distribution was patchy and often areas that appeared appropriate by visual inspection did not contain any larvae. YOY larvae captured here were larger (29 ± 4 mm, SD) than at all other sites (19 ± 3 mm, SD; Figure 3), suggesting
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growing conditions for lamprey in the Roe-Jan are the most favorable in the Hudson River.
Farmers in this watershed should be encouraged to maintain buffer strips and local residents
should be educated not to develop highly mowed and treated lawns (wherever possible) to
maintain this habitat. Any work that is conducted within the stream or immediately adjacent
should take into account that the habitat is highly unique and important for sea lampreys.
Rondout Creek (Ulster County, river km 146)
Rondout Creek is a large tributary of the Hudson that passes through Kingston, NY
before emptying into the Hudson River. The stream is highly urbanized around the Kingston
metropolitan area, but generally forested further upstream. Agriculture is also an important land
use in the watershed, but is primarily concentrated around the main body of the stream. A historic
record (Greeley and Greene 1937) and recent reports (Hudson River Almanac 2007b), have
documented sea lamprey utilization of this river. Larval lampreys were found immediately downstream from High Falls, NY, and this is likely the upper limit for sea lampreys. Appropriate habitat is available upstream but the steep waterfalls and high flows likely preclude adults reaching these areas. Larvae were common in sandbars along the shoreline at High Falls.
However, the depth of the stream (>2 m) made it difficult to estimate how many larvae might be present. There may be numerous larval lampreys utilizing this river in deeper areas. There appear
to be no immediate threats to sea lamprey usage of this habitat, unless access is limited by the
currently passable Eddyville Dam further downstream.
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Saw Kill (Dutchess County, river km 158)
The Saw Kill has a short stretch of stream (~100m) before the first large natural falls and above the head of tide. Invertebrate diversity was high (17 families were collected not including
Chironomids), and there were numerous groups (five genera) that are listed as sensitive (Table
3). The Saw Kill is unusual in that it is the only stream in the Hudson River, north of New York
City, for which a record of American brook lamprey (Lethenteron appendix) has been reported
(Smith 1985). When surveyed no lampreys of any kind were found, although adult sea lampreys
have been observed at the falls (E. Kiviat, personal communication). The habitat available for
larval lampreys below the falls is extremely limited and likely not a successful rearing area for
sea lampreys. American brook lampreys are small (usually < 200 mm) non-parasitic lampreys
that likely migrate only short distances to spawn. This animal probably did not migrate from
more southerly populations near New York City to the Saw Kill, but was possibly washed
downstream from the headwaters of the Saw Kill. There may be, or was, a population of
American brook lampreys further upstream that attracts sea lampreys occasionally to the stream
(Fine et al. 2004). In the present survey, sampling for American brook lamprey was attempted at
upstream locations, but suitable access points were unavailable. Researchers working with
permission from land owners should examine the headwaters of the Saw Kill to determine if an
undocumented population of American brook lampreys is present.
Stockport Creek (Columbia County, river km 195)
Stockport Creek, located north of Hudson, NY, has a long, tidally influenced delta region
that is surrounded by private residences and forest. There is a small waterfall (approximately three m in height) under the Route 9 bridge that separates the upper river from the tidal portion,
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but likely does not limit sea lamprey migration. Upstream of the falls Stockport Creek eventually splits into Claverack Creek and Kinderhook Creek. The Claverack Creek has a dam within 500 m of its mouth, but the dam has been severely damaged and would not prevent migration.
However, only eight insect genera where present (excluding Dipterans; Table 3) and water temperatures approached sea lamprey lethal limits (30°C in stream, lethal limit is 31°C).
Although extensive appropriate habitat appeared to be present no lampreys were found.
Kinderhook Creek appeared to have high water quality and provide potential habitat. The invertebrate community was very diverse (18 genera excluding Dipterans; Table 3) and abundant. Upstream of the mouth are multiple dams currently used for hydropower that are built on top of a series of natural falls. Water temperatures in the stream were also close to the lethal limit for sea lamprey (30°C in stream, the lethal limit is 31°C). Therefore, lower reaches are unlikely to support larval lampreys, and adults likely need to penetrate further upstream to successfully reproduce.
An interviewed individual currently working at one of the hydropower facilities reported regularly seeing small sea lampreys damaged or killed in the turbines when removing dead, impinged, American eels (Anonymous, personal communication). Their size and direction of capture (i.e., as they moved downstream) suggest these animals were migratory transformers heading out to sea to feed. He also reported not having seen any for “…a couple of years…”
Surveys for lampreys above the dams in Valatie, NY were conducted, but no lampreys or eels were found. Habitat upstream looked to be of high quality and it is likely if sea lampreys could access the habitat, they would utilize it. The long, protracted larval period may allow rare adult penetrations into upper reaches to produce successful migrants for long periods. This may result
VIII - 37 in a net sink for the lamprey population as migrating adults are attracted to the scent of larval lamprey and congregate below successful spawning areas, wasting their reproductive output.
Vlockie Kill (Rensselaer County, river km 219)
The Vlockie Kill, near Castleton-on-Hudson, is a small stream that has short stretch
(~0.5-1 km) available to migratory sea lamprey. Although this site was surrounded by forest there was a large amount of siltation present (possibly from tidal influence), human refuse, and few fish were observed. The Vlockie Kill may be suffering from modifications upstream that have allowed the mobilization of silt/fines. No sea lampreys were found and they will not exploit this habitat unless siltation is significantly reduced.
Vloman Kill (Albany County, river km 220)
The Vloman Kill is south of Albany, NY and has a very limited stretch available for migratory fishes (<1 km) below the first barrier. The delta floodplain was surveyed for sea lampreys during low tide but none were found. The water was very turbid and the bottom was primarily fine clay particles which were anoxic at a few mm in depth. It was therefore not suitable for sea lamprey larvae. Two barriers to upstream movement are present, and these are likely able to restrict all anadromous fish movement into higher reaches. Individuals nearby reported that they had never seen sea lampreys at the Vloman Kill, but had observed them frequently and in large numbers in the Catskill Creek.
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Stable Isotope Analysis
At Cedar Pond Brook, lampreys were primarily supported by autochthonous sources
(Figure 5), because aquatic plants had isotopic values high enough to explain the high isotopic values of lampreys (both in respect to δ13C and δ15N). Although the model explained larval lamprey isotopic values with autochthonous sources, lampreys at this site may actually be receiving N and C contributions from anthropogenic waste sources, such as sewage or leaky septic systems, however, these sources was not sampled. Anthropogenic waste is more positive with respect to both δ13C and δ15N than natural sources (McClelland et al. 1997; Cravotta 2002),
Figure 5: Median percent contributions of different nutritional resources (i.e., autochthonous and allochthonous) to macroinvertebrates and sea lamprey larvae from different streams calculated by the Bayesian model MixSIR. Lower and upper error bars correspond to the 5% and 95% posterior proportional contributions respectively.
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and larval values (range of δ13C = -27 to -23, and of δ15N = 6 to 9) would be easily explained by
this source (δ13C = -25 to -20, and of δ15N = -2 to 13; Cravotta 2002). The stream was filled with large amounts of trash, and is immediately downstream of a county park. Larval lampreys may
also have access to liquefied wastes traveling in subsurface waters, and may utilize this source if it exists.
The macroinvertebrates measured at the site did not appear to be relying on the same materials and were well explained by sources. Therefore, larval lamprey (which spend time burrowed in sediments and rarely emerge) may be relying on sources not often consumed by the invertebrates measured. Interestingly, one larval lamprey had lower isotopic values (δ13C = -27.1,
δ15N = 5.7) than the other two lampreys at the site (δ13C = -23.3 and -24.0, δ15N = 9.2 and 8.6;
Figure 4). The first lamprey was larger (24 mm) than the other two lampreys (18 and 19 mm),
suggesting it may have been spawned earlier in the year. If it was spawned earlier it was likely
further upstream and was washed downstream to the site where it was collected later. It may still
have been coming to isotopic equilibrium with its sources at the collection site (Tieszen et al.
1983), and its isotopic signature may have reflected sources further upstream. Larval lampreys
may be very sensitive to differences in local site conditions, and may offer a way to detect
anthropogenic waste in streams.
Predicted contributions of sources from modeling to different groups at Kaaterskill Creek
were similar, and uncertainty was large for all estimates (Figure 5). A single aquatic algae sample
had a δ13C signature similar to terrestrial plants, which resulted in a large standard deviation for
autochthonous sources in the model. Kaaterskill Creek appears to be more dependent on
allochthonous sources as all groups had more elevated values than at any other site. The water at
the site was moderately turbid, and the bottom of the stream was not readily visible at all depths.
VIII - 40
The turbidity appeared to be the result of suspended particulate matter, which may have been the
result of runoff from local fields transporting soil material. If this is the case the system may be more dependent on allochthonous carbon, because particulates shade algal growth but also provide a source of nutrition (Rounick et al. 1982; Bartels et al. 2012).
At the Roeliff Jansen Kill (where growth of lampreys was fastest; Figure 3), larval lampreys were dependent approximately equally on autochthonous and allochthonous sources
(Figure 5). Larval lamprey growth may be dependent on the quality of food available, but is also known to depend on the density of animals (Morman 1987). Further work could be done to determine if lamprey in the Roeliff Jansen Kill are growing faster because of better nutrition or for another reason. Source contributions to larval lamprey at this site were more similar to those of Heptageniid mayflies, which are scrapers, than the collecting-filtering Isonychiids mayflies.
Larval lamprey filtering/collecting is assumed to be similar to other invertebrates, but may not be similar to other groups. Lampreys are unlike all other filter feeders, and have no good proxy among any extant group (Mallatt 1982). Part of the success of lamprey as a group maybe their ability to exploit a habitat and feeding style that no other group has adapted to fill.
At Rondout Creek, autochthonous sources were more important than allochthonous sources for all groups. Again lampreys were more similar to Heptageniid mayflies (scrapers) and
Hydropsychiid caddisflies (sedentary collector-filterers), than the more mobile collector-filtering
Isonychiids. The Rondout is wide (> 10 m) and slow moving above the barrier to sea lamprey
migration, and likely produces high amounts of algal growth. Large black fly (Simuliidae)
larvae, which are also filter feeders, were also very common at the site, often completely
covering rocks in areas of high current. It seems likely that algal sources upstream provide large
VIII - 41
amounts of productivity to communities immediately downstream where these animals were
captured. Lamprey throughout the Rondout may not be as dependent on autochthonous sources.
Larval lampreys appear capable of exploiting a variety of sources to meet nutritional requirements. In addition, larval lampreys may be sensitive to some types of human pollution, which could be detected with isotopic analysis. Further work is needed to determine if they could be useful as bio-monitoring tools. Their limited distribution and the difficulty of sampling for larval lamprey may prevent wide scale use, but at sites in which they do occur, they could offer unique information about their environment. Larval lampreys were also isotopically unique from the macroinvertebrate community in which they occur, and appeared to rely on a different proportion of autochthonous and allochthonous sources.
CONCLUSIONS
The current distribution of larval lampreys in the Hudson River is restricted to only a few tributaries (i.e., Cedar Pond Brook, Catskill Creek, Roeliff Jansen Kill, Rondout Creek).
However, densities of larval lampreys at the local scale can be high (~2 individual/ m2). Sea lampreys were found in fewer locations than previously reported in the Hudson River. However,
many of the reports are for adult animals, which were likely seeking suitable spawning habitat
and are not as sensitive as larvae. In at least one case (i.e., Stockport Creek), there appears to be a
true reduction in their range in recent years. Sea lampreys may be able to make rare, successful
spawning events that produce larvae for extended periods, but attract easily visible adults for
long periods by means of odor trails (Fine et al. 2004). The barriers to migration at Stockport
Creek would be difficult for sea lamprey to pass except during exceptionally high flow years.
Animals unable to penetrate above the falls and dams likely do not produce larvae, but any adult
VIII - 42
that made it upstream would have excellent habitat available for spawning and the subsequent
larvae. In at least one location (i.e., Saw Kill), undocumented populations of American brook lampreys may be present based on a prior capture report (Smith 1985) as well as reports of sea lampreys entering the stream. Sea lampreys are attracted to the larvae of any lamprey species
(Fine et al. 2004), but adult sea lampreys would not be able to successfully reproduce in the
mouth of the Saw Kill.
Currently, Catskill Creek is the most important producer of sea lampreys, and the
Kaaterskill Creek (which is a tributary to the Catskill Creek) is its most important tributary. The
Kaaterskill Creek has a long stretch (> 10 km) of unobstructed habitat appropriate for both sea
lamprey larvae and adults, making this unique to the Hudson River. The Roeliff Jansen Kill is
also an important tributary for sea lamprey in the Hudson River. Sea lampreys in the Roe-Jan
appear to grow more quickly than at any of the other tributaries where they were captured. More
work is needed to determine if growth rates are actually higher throughout the Roe-Jan, or if
larval growth is more dependent on local conditions.
Sea lampreys prefer cool water, and as a result climate change is likely to restrict and
reduce sea lamprey numbers in the Hudson River. Although climate change is a threat to the
future of sea lamprey in the Hudson River, past and present anthropogenic activities have also likely limited populations through the construction of dams and pollution of stream waters.
However, sea lampreys in the Hudson River were likely never as abundant as other members of
the diadromous fish community. The erosion of decaying dams and the removal of barriers to
migration may allow sea lampreys to exploit more habitats, potentially providing a buffer against
climate change impacts. Sea lampreys are an unusual member of the Hudson River that have
been little studied, but their uniqueness and importance as members of the anadromous fish
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community make them deserving of more directed research efforts. Additionally, their presence
indicates high stream quality and thus restoration efforts might be able to use their presence as a
criterion for achievement of some restoration goals.
ACKNOWLEDGEMENTS
I thank the Tibor T. Polgar Fellowship Program of the Hudson River Foundation and the Edna S.
Bailey Sussman Foundation for funding and support of this project. I also thank Kathy Hattala and Robert Adams for advice and guidance on selecting locations and choosing streams to survey. In addition, thanks to all members of the NYSDEC who provided help and reports of sea lamprey, and especially to those working on the Hudson River Almanac. Thanks also to David
Yozzo who allowed me to come field sampling with him and for providing information. Thanks to the University of California Davis Stable Isotope Facility as well as Joy Matthews for their excellent work and rapid turnaround of our samples. Special thanks to John Waldman, Robert
Schmidt, and Erik Kiviat for their help this summer. Thanks to Karin Limburg for her help and support throughout my work on this project, and for reviewing and editing drafts of the final manuscript. Thanks also to reviewers at the Hudson River Foundation whose comments improved the manuscript before its completion. Finally, thanks to Caitlin Eger, my wife, for all her help and for assisting in almost all of the field sampling for this project.
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Pictured (left to right): Shabana Hoosein, Thomas Evans, Alexis Paspalof, Kyle Monahan, Sherif Kamal, Angel Montero, Troy Hill, and Mariana Teixeira
Special thanks to Melissa Wei for assistance in formatting manuscripts.