Assessment of lower food web in Hamilton

Assessment of lower food web in Hamilton
ii
Assessment of lower food web in Hamilton
Harbour, Lake Ontario, 2002 - 2004.
R. Dermott, O. Johannsson, M. Munawar, R. Bonnell, K. Bowen,
M. Burley, M. Fitzpatrick, J. Gerlofsma, and H. Niblock
Great Lakes Laboratory for Fisheries and Aquatic Sciences
867 Lakeshore Road
Burlington, Ontario L7R 4A6
2007
Canadian Technical Report of Fisheries and
Aquatic Sciences 2729
Fisheries and Oceans
Canada
Pêches et Océans
Canada
iii
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i
Canadian Technical Report of
Fisheries and Aquatic Sciences 2729
2007
Assessment of lower food web in Hamilton Harbour, Lake Ontario,
2002 - 2004.
by
R. Dermott, O. Johannsson, M. Munawar, R. Bonnell, K. Bowen,
M. Burley, M. Fitzpatrick, J. Gerlofsma, and H. Niblock
Great Lakes Laboratory for Fisheries and Aquatic Sciences
Bayfield Institute
Fisheries and Oceans Canada
867 Lakeshore Road
Burlington, Ontario
L7R 4A6
ii
© Her Majesty the Queen in Right of Canada, 2007.
Cat. No. Fs 97-6/2729E ISSN 0706-6457
Correct citation for this publication:
Dermott, R., Johannsson, O., Munawar, M., Bonnell, R., Bowen, K., Burley, M.,
Fitzpatrick, M., Gerlofsma, J., and Niblock, H. 2007. Assessment of
lower food web in Hamilton Harbour, Lake Ontario, 2002 - 2004.
Can. Tech. Rep. Fish. Aquat. Sci. 2729: 120 p.
iii
TABLE OF CONTENTS
ABSTRACT ........................................................................................................
iv
RÉSUMÉ
v
........................................................................................................
Assessment of Hamilton Harbour; Background
.
...............................................................................
R. Dermott
1
M. Burley
9
M. Munawar, M. Fitzpatrick
43
Water Quality and Phytoplankton Photosynthesis
...............................................................................
An Integrated Assessment of Microbial and
Planktonic Communities of Hamilton Harbour
..................................................
Zooplankton in Hamilton Harbour: 2002-2004
.................................
J. Gerlofsma, K. Bowen, O. Johannsson
65
Benthic Fauna in Hamilton Harbour: 2002 - 2003
...........................................................
R. Dermott, R. Bonnell
91
iv
ABSTRACT
Dermott, R., Johannsson, O., Munawar, M., Bonnell, R., Bowen, K., Burley, M., Fitzpatrick, M.,
Gerlofsma, J., Niblock, H. 2007. Assessment of lower food web in Hamilton Harbour,
Lake Ontario, 2002 -2004.
As an Area of Concern on the Great Lakes, Hamilton Harbour is an extremely stressed
environment with high nutrient levels and contaminated sediments. Remediation efforts have
greatly improved water quality and encouraged habitat restoration. Fisheries and Oceans
undertook a comprehensive program to examine components of the harbour's food-web, from
microbes up to fish.
Bi-weekly sampling between May and October in 2002, 2003 and 2004 examined
environmental and biological components simultaneously at the same sites. Site location changed
each year. In 2002 and 2003, a nearshore to offshore gradient was examined in the middle (3
sites), or western end of the harbour (2 sites), between 1.5 to 24 m deep. The two offshore sites
were re-examined in 2004. Depth profiles of temperature, oxygen, and light were taken, and
samples collected for water chemistry, chlorophyll a, seston, total and size fractionated primary
production, bacteria, ciliates, phytoplankton, and zooplankton. Benthic samples were collected at
3 seasons only in 2002 and 2003. In 2002, a spatial survey was conducted for most components
at a number of sites in the harbour.
Annual average phosphorous concentrations at the sites ranged from 25 µg l-1 to 34 µg l-1,
above the Remedial Action Plan goal of 17 µg l-1. Anoxic conditions (dissolved oxygen < 1 mg
l-1) were observed at depths > 10 m from late June until October. The phytoplankton community
was variable with peaks of Diatomeae, Dinophyceae, Chlorophyta and Chrysophyceae
throughout the seasons. Primary productivity was dominated by larger phytoplankton (> 20 µm),
except in late spring/early summer when picoplankton (<2 µm) contributed 40% of the size
fractionated production. The shallowest site (1.5 m) supported the highest zooplankton
production (10,188 mg dry m-3). Zooplankton production ranged between 2211 and 4934 mg m-3
at sites >3 m deep. The shallow site supported a diverse cladoceran community dominated by
benthic or plant-associated taxa such as Eurycercus, Sida and Alona. Dominant zooplankton
offshore were Bosmina spp., Daphnia retrocurva, Eubosmina coregoni, and juvenile Cyclopoid
copepods. Dreissena veligers were common in 2002 and 2003. Rotifer biomass was 1-3% that
of zooplankton biomass. Low oxygen restricted the benthic community, so that in mid-harbour,
the composition was almost exclusively tubificid worms (18.8 g m-2 wet). Chironomidae were
common only at shallower sites on the north and west end of the harbour. Dreissena polymorpha
was common at sites along the north shore, but few were present below 8 m depth. This report
provides information about the food web of the harbour.
v
RÉSUMÉ
Dermott, R., Johannsson, O., Munawar, M., Bonnell, R., Bowen, K., Burley, M., Fitzpatrick, M.,
Gerlofsma, J., Niblock, H. 2007. Évaluation des maillons inférieurs du réseau
alimentaire du havre Hamilton (lac Ontario), 2002 -2004.
Le havre Hamilton, un secteur préoccupant des Grands Lacs, est un milieu hautement
perturbé dont les taux en nutriments sont élevés et les sédiments sont contaminés. La qualité de
l’eau s’est beaucoup améliorée et la restauration de l’habitat est favorisée grâce aux efforts
d’assainissement. Pêches et Océans Canada a entrepris un programme intégré de mise à l’étude
des composants du réseau alimentaire du havre Hamilton, du microorganisme jusqu’au poisson.
Un échantillonnage bimensuel, de mai à octobre en 2002, 2003 et 2004 avait pour but
d’étudier, simultanément, des composants environnementaux et biologiques aux mêmes sites.
Les lieux de sites variaient chaque année. En 2002 et 2003, un gradient côtier - extracôtier au
milieu du havre (à 3 sites) ou à l’extrémité ouest du havre (à 2 sites), aux sites où la profondeur
était de 1.5 à 24 m était à l’étude. On a réétudié les deux sites côtiers en 2004. Des profils de
concentration pour la température, l’oxygène et la lumière ont été faits et des échantillons ont été
pris pour la composition chimique de l’eau, la chlorophylle a, le seston, la production primaire
totale et fractionnée en fonction de la taille, les bactéries, les ciliés, le phytoplancton et le
zooplancton. En 2002 et en 2003, des échantillons benthiques ont été pris pendant trois saisons
seulement. En 2002, on a fait une évaluation spatiale à plusieurs sites portant sur certains des
composants nommés ci-dessus.
La concentration moyenne en phosphore aux sites variait de 25 µg l-1 à 34 µg l-1, donc audessus de l’objectif de 17 µg l-1 du Plan de Mesures Correctives. On a observé des conditions
anoxiques (oxygène dissous de < 1 mg l-1) à des profondeurs de > 10 m, à partir de la fin dejuin
jusqu’en octobre. La composition de la population phytoplanctonique était variable : le nombre
de diatomées, dinophycées, chlorophycées et chrysophyées montait à la hausse en toute saison.
Les grandes cellules phytoplanctoniques (> 20 µm) dominaient la production primaire, sauf vers
la fin du printemps et le début de l’été lorsque le picoplancton (< 2 µm) faisait 40 % de la
production fractionnée en fonction de la taille. Le site le moins profond (1,5 m) soutenait la
production la plus importante en zooplancton (10,188 mg secs m-3). La production en
zooplancton variait de 2 211 à 4934 mg m-3 aux sites de >3 m de profondeur. Le site peu
profond soutenait des cladocères divers dont les taxa benthiques ou végétaux Eurycerus, Sida et
Alona dominaient, entre autres. Au large, Bosmina spp., Daphnia retrocurva, Eubosmina
coregoni et des copépodes cyclopoides juvéniles étaient le zooplancton dominant. Des Dreissena
au stade veligère étaient communes en 2002 et 2003. La biomasse en rotifères était de 1 à 3 %
de la biomasse en zooplancton. Un bas niveau d’oxygène compromettait la population benthique
du havre et donc, au milieu du havre, elle était composée presque exclusivement de tubificidés
(18.8 g m-2 humides). Des chironomes étaient communs uniquement aux sites moins profonds
des extrémités nord et ouest du havre. Le dreissena polymorphe était commune aux sites du
littoral nord, mais peu présent au-dessous de 8 m de profondeur. Le présent rapport est au sujet
du réseau d’alimentation du havre.
vi
1
ASSESSMENT OF HAMILTON HARBOUR; BACKGROUND
Ronald Dermott
INTRODUCTION
Hamilton Harbour is a major industrial port at the west end of Lake Ontario. It has had a long
history of pollution from industrial and municipal wastes, culminating in its description as the
largest and most beautiful septic tank in the world (Matheson 1958). Early concerns were high
concentrations of bacteria and phenolic substances in the water. Hypolimnetic oxygen depletion
occurs during thermal stratification from June until late September. Total nitrogen levels had
been between 1 and 2 mg l-1 in 1949 and up to 3 mg l-1 in 1976 (Matheson 1958, Piccinin 1977).
In 1950, total phosphate was 0.04 with levels up to 0.08 mg l-1 measured during the summer
(Matheson 1958). The water quality and biological community of the harbour have been
impaired both from excessive nutrient and ammonia loadings from municipal sewage from the
surrounding cities and toxic contaminants from heavy industries. The harbour was designated in
1987 by the International Joint Commission as one of the Areas of Concern on the Great Lakes
under the terms of the Great Lake Water Quality agreement (IJC 1988).
For 90 years, wastes from the steel industry and associated coking facilities contaminated
the sediments with iron-manganese oxides, heavy metals such as cadmium, copper, and zinc;
coal dust and numerous polyaromatic hydrocarbons (PAHs) and polychlorinated biphenyls
(PCBs) (Poulton 1987, Mayer and Johnson 1994, Fox et al. 1996). In several areas of the
harbour, concentrations in the sediments exceed Ontario's guidelines for dredged sediment
disposal (OME 1985). Randle Reef, situated along the industrial south shore, and the
Windermere basin are the areas with greatest contamination of coal tar-contaminated sediments,
and resuspension of these materials is a major source of genotoxic PAH in the water column
(Marvin et al. 2000).
A Remedial Action Plan (RAP) has been developed and several beneficial impairments
were identified including high nutrient levels, contaminated sediments and impaired planktonic
and benthic communities. Following financial expenditures of over 600 million dollars by
industry and municipalities, remediation efforts including improvements to wastewater treatment
have considerably reduced nutrient and contaminant loadings (Hamilton Harbour RAP 1992).
Improvements to water quality, habitat recovery and the establishment of waterfront parks have
increased both wildlife and public use of the harbour. However, phosphate levels are still above
initial RAP target goals of 0.034 mg l-1 (Charlton and Le Sage 1996), and total nitrogen and
ammonia levels still often exceed provincial guidelines (Barica 1990, OMEE 1994).
Concentrations of unionized ammonia approach the provincial guideline of 0.02 mg l-1 during
2
winter. Water transparency measured as Secchi depth averaged 2.1 m during 1994, still above
the target Secchi depth goal of 3 m (Hamilton Harbour RAP 1992). Continuing progress on load
reductions from municipal sewage plants is ongoing and will take further infrastructure funding
and time before final target goals can be reached.
In addition to the improving conditions in the harbour and increased wildlife habitat
areas, several new introduced species have become established in the harbour. These include the
zebra mussels (Dreissena polymorpha and recently D. bugensis), round goby (Neogobius
melastoma), and the predatory zooplankton (Cercopagis pengoi). These new species are now
living together with the native species and with long established introduced species including the
common carp (Cyprinus carpio) and faucet snails (Bithynia tentaculata). The present biological
communities have had to adapt to both the chemical and biological pollution in the harbour.
Basin Description
Hamilton Harbour, sometimes called Burlington Bay, is separated from Lake Ontario by a
natural sand bar (Fig. 1). The bay, which covers an area of 21 km2, is bowl shaped with a
maximum depth of 24 m and mean depth of 13 m, yet 12 % of the area is less than 2 m deep
(Table 1). Re-development of both industrial docks and parklands has reduced the area from
28 km2 and a maximum depth of 28 m since 1964, when 15 % of the area was less than 2 m deep
(Barica 1990, Johnson and Matheson 1968). The watershed surrounding the bay is small,
tributaries supply about 1.6 m3 s-1 of the inflow. During summer, wastewater can represent 67 %
of the inflow to the bay (4.9 m3 s-1), with wastewater from the city of Hamilton responsible for
75 % of the Sewage Treatment Plant (STP) inflows (Hamblin and He 2003). Industrial users
recycle up to 27 m3 s-1 mainly for cooling purposes. A shipping canal of 9.5 m depth connects
the harbour to Lake Ontario. The west end of Lake Ontario is subject to cold upwellings and
flow through the canal can be multi-layered and bi-directional depending on lake level, wind
direction and internal seiches (Poulton et al. 1988, Hamblin and He 2003). Inflow from Lake
Ontario through the canal can exceed 11.4 m3 s-1 (Hamblin and He 2003).
The shape and depth of the harbour results in summer stratification leading to oxygen
depletion in the hypolimnion for much of the summer. The oxygen demand of the sediment
surface can exceed 0.65 mg O2 l-1 day-1. A demonstration oxygenation experiment required an
aeration rate of 8.5 m-3 (300 cubic feet) compressed air per minute to prevent deoxygenation of
the hypolimnion in summer (Piccinin 1977). Sediment anoxia is responsible for increasing metal
bioavailability in fall-collected sediments (Krantzberg 1995). Cold oxygenated water from Lake
Ontario often enters the hypolimnion or metalimnion via the canal, resulting in complex
temperature and oxygen profiles in the harbour. Internal seiches within the harbour can also
force the low oxygen water up to above the 8 m depth zone (Matheson 1958).
The north shore is mostly residential, whereas the south shore is mostly industrial which
accounts for 46% of the total shoreline (BARC 2006). A secondary sandbar at the west end of
3
the bay, remnant of the higher Lake Iroquois stage of 12,000 years ago separates the harbour
from Cootes Paradise (Rukavina and Versteeg 1996). The northwest end of the harbour and
Cootes Paradise are part of undeveloped Royal Botanical Gardens lands which supports a variety
of wildlife habitats, which are used by a number of rare species including Blanding's turtles
(Emydoidea blandingii) and Prothonotary warblers (Protonotaria citrea).
OBJECTIVES
Less is known about the biological condition of the harbour than is known about the nutrient and
contaminant levels. Fisheries and Oceans has undertaken an ecosystem approach and developed
a comprehensive program to examine all components of the food-web, from microbes up to fish,
as well as information on habitat and aquatic plants in the harbour. This is the first time such a
comprehensive study has been done in the harbour. The aim of this work was to add to the
knowledge of the biology of Hamilton Harbour over all lower trophic levels from microbial to
non-vertebrate predators. This included sampling both for inter-annual changes, and spatial
variability in microbial, phytoplankton, zooplankton, and benthic organisms, their species
composition and biomasses as related to the depth, temperature and oxygen layers in the water
column.
Bi-weekly sampling conducted between May and October in 2002, 2003 and 2004
examined environmental and several biological components simultaneously at the same sites.
The site identifications used follow a number of previously established sampling sites including:
water quality and sediment sites (200 series) of the Ontario Ministry of Environment (Poulton
1987); Environment Canada's water quality sites (900 series) used by Murray Charlton (Charlton
et al. 1992, Halfon 1996); locations used by Fisheries and Oceans electrofishing transects
(Brousseau et al. 2005), as well some of the benthic sampling locations established by Johnson
and Matheson (1968). Locations and depths of the biweekly sample sites are listed in Table 2.
The location and number of sites changed each year. In 2002, three sites were sampled
along a nearshore to offshore gradient, the two nearshore sites 17 (1.5 m depth) and 6 (6 m) were
close to La Salle Park, and the offshore site 258 was at mid-harbour at 23.5 m depth. During
2003, sampling was done in the west-end of the harbour; the nearshore site at Willow Cove in
3.8 m and the other site situated halfway between Willow Point and the south shore at a depth of
14 m (site 908). In 2004, sampling was repeated at the two offshore sites, site 908 and the
mid-harbour site 258 (Fig 2).
On each sampling date, all the physical variables and biological components except
bottom fauna were sampled. Depth profiles of temperature, oxygen, and light profiles were
taken. Samples were collected for water chemistry, chlorophyll a, seston, total and size
fractionated primary production, bacteria, ciliates, phytoplankton, and zooplankton. Bottom
samples were collected on 3 seasons only in 2002 and 2003.
4
A more extensive spatial survey was also conducted during the summer of 2002 at a
number of sites between the canal and Willow Cove. Only some of the biological components
were sampled in the spatial survey, including: microbial loop including bacteria, ciliates, and size
fractionated algae; composition and biomass of the phytoplankton; zooplankton; and bottom
fauna.
This research was supported under the Great Lakes Action Plan. The work falls under
DFO's priority of an ecosystem approach to management of human activities in the harbour,
using several rate processes as ecosystem indicators of human perturbation. DFO priority
research on ecosystem modelling and linkages to habitat productive capacity are also addressed.
The data produced will eventually contribute toward an ecosystem model (ECOPATH) of the
harbour's food web to assess sustainability of the fisheries targets for the harbour. The data will
also be used to evaluate the Beneficial Use Assessments for the plankton and benthic
communities under the Remedial Acrion Plan for Hamilton Harbour.
ACKNOWLEDGEMENTS
The authors would like to thank Scott Millard and Vic Cairns of Fisheries and Oceans Canada
for initiating this research and providing direction and linkage with the Hamilton Remedial
Action Plan Committee. Funding was provided by the Great Lakes Action Plan (GLAP).
Summer students and interns were partly funded by ULEARN (Upper Lakes Environmental
Research Network) and the YMCA Federal Public Sector Youth Internship Program. Peter
Jarvis, Silvina Carou and Rachel Nagtegaal helped with the boat launching and sample
collections in all types of weather. Assisting with the field sampling and laboratory work on the
microbial foodweb, water chemistry and phytoplankton were: Sara Booth, Silvina Carou, Calais
Irwin, Latha Logasundaran and Danielle Tassie. Helping with the zooplankton samples were:
Rachel Nagtegaal, Bianca Radix, and Margaret Vogel. Bianca Radix, Ling Ying Ong and Andrea
Bernard helped collect the smelly mud samples.
REFERENCES
Bay Area Restoration Council 2006. (BARC). Toward Safe Harbours: Progress toward delisting
-toxic substances and sediment remediation. Bay Area Restoration Council, Hamilton.
June 2006, 43 p.
Barica, J. 1990. Ammonia and nitrite contamination of Hamilton Harbour, Lake Ontario. Water
Poll. Res. J. Canada. 25: 359-386.
Brousseau, C.M., Randall, R.G., and Clark, M.G. 2005. Protocol for boat electrofishing in
nearshore areas of the lower Great Lakes: transect and point survey methods for collecting
fish and habitat data, 1988 to 2002. Can. Manuscr. Rep. Fish. Aquat. Sci. 2702: 89 p.
Charlton, M.N., Milne, J.E., Booth, W.G., and Roy, R. 1992. Update on eutrophication section of
Hamilton Harbour RAP Stage 1. National Water Research Institute (NWRI) Contribution
No. 92-58.
5
Charlton, M.N., and Le Sage, R. 1996. Water quality trends in Hamilton Harbour: 1987-95.
Water Qual. Res. J. Canada. 31: 473-484.
Fox, M.E., Khan, R.M., and Thiessen, P.A. 1996. Loadings of PCBs and PAHs from Hamilton
Harbour to Lake Ontario. Water Qual. Res. J. Canada. 31: 593-608.
Halfon, E. 1996. Data animator - software that visualizes data as computer-generated animation
on personal computers: an application to Hamilton Harbour. Water Qual. Res. J. Canada.
31: 609-622.
Hamblin, P.F., and He, C. 2003. Numerical models of the exchange flows between Hamilton
Harbour and Lake Ontario. Can. J. Civ. Eng. 30: 168-180.
Hamilton Harbour Remedial Action Plan (RAP). 1992. Remedial Action Plan for Hamilton
Harbour, Goals, Options and Recommendations. RAP Stage 2 Report, ISBN 0-77780533-2, Canada Ontario Agreement. Nov. 1992. 327 pp.
International Joint Commission (IJC) 1988. Revised Great Lakes Water Quality Agreement of
1978, as amended by Protocol signed Nov. 18, 1987. Consolidated by the International
Joint Commission of United States and Canada. January, 1988. 130 p.
Johnson, M.G., and Matheson, D.H. 1968. Macroinvertebrate communities of the sediments
of Hamilton Bay and adjacent Lake Ontario. Limno. Ocean. 13: 99-111.
Krantzberg, G. 1994. Spatial and temporal variability in metal bioavailability and toxicity of
sediment from Hamilton Harbour, Lake Ontario. Environ. Toxicol. Chem. 13 (10):
1685-1698.
Mayer, T., and Johnson, M.G. 1994. History of anthropogenic activities in Hamilton Harbour as
determined from the sedimentary record. Envir. Poll. 86: 341-347.
Marvin, C.H., McCarry, B.E., Villella, J., Allan, L.M., and Bryant, D.W. 2000. Chemical and
biological profiles of sediments as indicators of sources of contamination in Hamilton
Harbour. Part II: Bioassay-directed fractionation using the Ames Salmonella/microsome
assay. Chemosphere. 41 (2000): 989-999.
Matheson, D.H. 1958. A consolidated report on Burlington Bay. Dept. of Municipal
Laboratories, City of Hamilton, Hamilton, Ontario.
Ontario Ministry of the Environment (OME). 1985. Hamilton Harbour technical summary and
general management options. 125 p.
Ontario Ministry of the Environment and Energy (OMEE). 1994. Water management policies,
guidelines, provincial water quality objectives. ISBN 0-778-8473-9, PIBS 3303E, Ontario
Ministry of Environment and Energy, Toronto, Ontario. July 1994. 67 p.
Piccinin, B.B. 1977. The biological survey of Hamilton Harbour 1976. Dept. of Biology,
McMaster University. Tech. Report Series No. 2. 129 p.
Poulton, D.J. 1987. Trace contaminant status of Hamilton Harbour. J. Great Lakes. Res. 13:
193-201.
Poulton, D.J., Simpson, K.J., Barton, D.R. and Lum, K.R. 1988. Trace metals and benthic
invertebrates in sediments of nearshore Lake Ontario at Hamilton Harbour. J. Great
Lakes. Res. 14: 52-65.
Rukavina, N.A. and Versteeg, J.K. 1996. Surficial sediments of Hamilton Harbour: Physical
properties and basin morphology. Water Qual. Res. J. Canada. 31: 529-551.
6
Table 1. Hamilton Harbour hypsometric data from GIS polygons of the depth contours (from
C. Bakelaar, DFO).
Depth range (m)
Area (km2)
Total area (m2)
0- 2
2- 5
5 - 10
10 - 15
15 - 20
20 +
2.49
1.24
4.27
5.11
5.67
2.20
11.9
5.9
20.3
24.4
27.0
10.5
Sum
20.97
Table 2. Depth and location of sites used for the intensive bi-weekly sampling in Hamilton
Harbour during the years 2002, 2003 and 2004.
Site
Depth (m)
Latitude
Longitude
2002
17
6
258
1.5
5.7
23.8
43° 18.201'
43° 18.133'
43° 17.241'
079° 50.354'
079° 50.300'
079° 50.446'
2003
WC
908
3.2
14.8
43° 17.183'
43° 16.768'
079° 52.268'
079° 52.443'
2004
908
258
14.8
23.5
43° 16.768'
43° 17.241'
079° 52.443'
079° 50.446'
7
Burlington
<2 m
N
10 m
La Salle
Willow
Cove
STP
Lake
Ontario
Canal
CCIW
20 m
Botanical Gardens
<2 m
m
Steel Mills
r
de
in
le
nd
Ra ef
e
R
10 m
W
e
er
Co
e
adis
Par
s
e
ot
Bayfront
<2 m
STP
Hamilton
Figure 1. Hamilton Harbour with depth contours and surrounding features, scale bar = 1800 m.
2002
2003
2004
17
6
WC
CCIW
258
908
Figure 2. Locations of sites used for intensive biweekly sampling in 2002, 2003 and 2004.
8
9
WATER QUALITY AND PHYTOPLANKTON PHOTOSYNTHESIS
Michele Burley
Great Lakes Laboratory for Fisheries and Aquatic Sciences
Fisheries and Oceans Canada
867 Lakeshore Road, P.O. Box 5050
Burlington, Ontario, L7R 4A6
INTRODUCTION
Hamilton Harbour is an embayment with a total area of 2150 hectares, connected to Lake Ontario
by the Burlington Ship Canal across the sandbar that forms the harbour’s eastern edge. Significant
and highly variable exchanges of water occur between Hamilton Harbour and Lake Ontario via the
canal. A 250 hectare area of marsh and shallow open water (Cootes Paradise) discharges into the
Harbour’s western end via the Desjardins Canal. The Harbour’s watershed covers an area of 49,400
hectares with three major tributaries flowing into the Harbour. Spencer Creek drains the north-west
and west portion of the watershed and feeds into the Harbour via Cootes Paradise; it accounts for
the largest tributary inflow of 54%. Redhill Creek (15%) drains the south-east area of the basin and
Grindstone Creek (14%) drains the north central portion (Hamilton Harbour Remedial Action Plan
(RAP) 1992).
Water quality issues plaguing Hamilton Harbour stem from its long history as an industrial
port and receiver of industrial and municipal wastes and watershed runoff. In the 1850s the
harbour was deemed unfit as the drinking water supply due to raw sewage contamination from the
City of Hamilton (RAP 1992). Starting in the early 1900s steel and iron industry wastes
contaminated the sediments with heavy metals, coal tar containing polyaromatic hydrocarbons
(PAHs) and later polychlorinated biphenyls (PCBs) (Poulton 1987). Sewage treatment has greatly
improved over the years, but still results in excessive eutrophication from nutrient loading of
phosphorus, ammonia and suspended solids, as well as a host of other contaminants. Eutrophication
symptoms include offensive algal growths, poor water clarity and depleted oxygen. Hypolimnetic
oxygen concentrations in the summer reach critical lows (0.5-1 mg l-1) due to the oxidation of
ammonia to nitrate by nitrifying bacteria which rapidly depletes the water column of dissolved
oxygen. In summer, about half of all the water entering the harbour is waste water effluent
(Charlton and Milne 2005). Combined sewer overflow (CSO) effluent from the City of Hamilton
and runoff discharged from the eastern part of the harbour contribute to the harbour’s loading of
nutrients, suspended solids and other contaminants (e.g. lead, zinc, PAHs). The inflow of Lake
Ontario water from the Burlington Ship Canal is also a contributor of persistent contaminants to
10
Hamilton Harbour, but also provides essential oxygenated water to the hypolimnion during the
summer months (Hamilton Harbour Remedial Action Plan (RAP) 2003).
The south and east shores of the harbour have been infilled for industrial and marine
activities as well as for railway or highway construction. Together with other developments, 75%
of the wetlands and fish nursery habitat have been eliminated from the Harbour (RAP 1992). Nearly
3.6 million m3 of material had been removed from the south shore between 1951-1962 (Whillans
1979), illustrating the magnitude of wetland removed from this portion of the harbour. Sustained
high water levels, poor water clarity and plant disturbance by carp have contributed to the
disappearance of the formerly extensive marshes in Cootes Paradise and Grindstone Creek (RAP
1992). The wetland at the mouth of Redhill Creek is only a remnant of a considerable marsh in this
area (Holmes and Whillans 1984).
The International Joint Commission in 1987 designated Hamilton Harbour an Area of
Concern with the aim to restore and protect beneficial uses (IJC 1988). The Remedial Action Plan
(RAP) for the harbour set a number of targets for water clarity, phosphorus, ammonia and
chlorophyll concentration (RAP 1992). High nutrient concentrations, algal blooms, suspended
solids, reduced water transparency and low dissolved oxygen are impairments to the beneficial uses
of the harbour. This includes degraded fish and wildlife (iii), degradation of benthos (vi),
eutrophication (viii), degradation of aesthetics (xi), added cost to agriculture or industry (xii) and
degradation of phytoplankton and zooplankton (xiii). In the past two decades, industry and regional
municipalities have spent an estimated $600 million to improve the harbour’s water quality (RAP
1992).
A large reduction in nutrient load from recent wastewater upgrades has improved water
clarity and chlorophyll levels. The concentration of metals has met provincial guidelines for some
time (Charlton and Milne 2005). The Hamilton RAP recommends that nutrient reduction be
sufficient to result in final concentrations of 0.017 mg l-1 for total phosphorus, <0.02 mg l-1 for unionized ammonia and 5-10 µg l-1 for chlorophyll; no final target has been set for suspended solids
for the harbour. The final target for water clarity (Secchi disc depth) is 3 m and the final target for
minimum dissolved oxygen has been set to > 4 mg l-1 (RAP 1992). This chapter reports on the
physical-chemical limnology and present status of phytoplankton productivity based on research
conducted in 2002 through 2004. For the purpose of assessing impairment, comparisons are made
with the Bay of Quinte, another Area of Concern which also has eutrophication issues.
MATERIALS AND METHODS
The 2002-2004 field sampling was conducted by scientific personnel from Fisheries and Oceans
Canada (Burlington) using Boston whalers docked at the Canada Centre for Inland Waters (CCIW).
In all years, thirteen cruises were conducted on alternate weeks starting in early May and ending in
late October (Table 1). In 2002, stations 17, 6 and 258 were sampled. Stations WC and 908 were
sampled in 2003 followed by stations 908 and 258 in 2004 (Table 1).
11
All laboratory work including photosynthesis experiments, chlorophyll filtration and
nutrient processing were conducted at CCIW. The following measurements and collections
occurred at each station for all dates. Depth profiles of temperature, dissolved oxygen, pH and
specific conductivity were obtained using a Hydrolab minisonde 4a (Hach Environmental,
Loveland, Colorado). Stratification status and mixing depth were determined utilizing specialized
density gradient analysis software (DENS). Secchi depth was recorded to determine water
transparency. Light extinction coefficients (400-700 nm - εpar) were determined using a Li-Cor
underwater quantum sensor (Li-Cor, Lincoln, Nebraska). Composite water samples were collected
for phytoplankton production, nutrients, seston and chlorophyll a. A Van Dorn sampler was used to
collect water at evenly spaced depths within the water column. At stratified stations, composite
water sampling was limited to the epilimnion. For unstratified or shallow stations whole water
column samples were taken. For a detailed description of the methods see MacDougall et al. 2001.
Phytoplankton production was determined using a 14C-uptake technique (Fee 1990). The
photosynthetic parameters derived were used as input for the FEE program that computes depth and
time-integrated photosynthetic rates. Seasonal (May 1-Oct 31) areal (g C m-2) and volumetric
(g C⋅m-3) rates of phytoplankton photosynthesis were calculated.
Composite water samples were processed for water chemistry following the National
Laboratory for Environmental Testing (NLET) guidelines (Environment Canada 1995). This lab
analyzed all nutrient samples for the study. Chlorophyll a was processed as per MacDougall et al.
(2001) following methodology of Strickland and Parsons (1972). Analysis for seston parameters
was determined by filtering aliquots of sample water on pre weighed GF/C filters (Whatman Co.).
Total suspended solids were determined by weighing samples after drying at 60° C. Ash content
was determined after processing in a muffle furnace at 450° C. Two filters were processed per
station and averaged to determine final weights.
Eutrophication issues also occur in the elongated Bay of Quinte, another Area of Concern
located in the north eastern section of Lake Ontario. The same sample collection and analysis
techniques were conducted in both Hamilton Harbour and the Bay of Quinte. As a result, data for
both locations are directly comparable. The upper Bay of Quinte station is the most heavily
eutrophied and shallow, averaging 5 m deep. The mid bay station HB (Fig. 1) has an average depth
of 12 m and stratifies; thus it is similar to the Hamilton station 908 which has an average depth of
14.6 m. Conway, in the Bay of Quinte, is a deep station with average bottom depth of 31 m, which
is deeper than the mid harbour station 258 (mean depth 23.4 m; Table 2).
RESULTS
Hydrolab Profiles
Stations 258 (sampled in 2002, 2004) and 908 (sampled in 2003, 2004) were the only harbour
stations in the study deep enough for persistent stratification to occur. The mixed (epilimnetic)
depths at stations 258 and 908 are shown in Fig. 2. For the purpose of analysis, sampling events
12
were divided seasonally according to stratification status and mixing depth temperatures. The first
three cruises from early May through mid June (Table 1) were designated as spring cruises because
they were unstratified with a mean mixing depth (epilimnetic) temperature ranging from 10.3 to
17.0 °C. Cruises from late June through September (cruises 4-11) were designated as summer
cruises because they were stratified with mean mixing depth temperatures ranging from
16.5 – 24.1 °C. Fall cruises encompassed the last two sampling dates in mid October until early
November (cruises 12 and 13) and were unstratified with mean mixing depth temperatures ranging
from 10.2 – 15.7 °C.
The mean summer epilimnetic depth at station 258 was 6.3 m in 2002 and 8.0 m in 2004; at
station 908 the mean summer epilimnetic depth in 2002 was 5.2 m compared to 8.2 m in 2004.
Summer epilimnetic temperatures at station 258 in 2002 averaged 21.9 °C and 21.5 °C in 2004; at
station 908 mean summer temperature in 2003 was 19.8 °C and in 2004 was 20.3 °C.
At the offshore stations mean spring temperatures ranged from 12.2 to 14.6 °C and fall temperatures
ranged from 12.6 to 14.2 °C for all study years (Fig. 3). Epilimnetic temperatures were similar to
that at Hay Bay (HB; Fig. 3).
The Hamilton Harbour Remedial Action Plan (RAP) has set an initial goal for minimum
oxygen concentration in the harbour to be greater than 1 mg l-1 and a final goal of greater than
4 mg l-1 oxygen. Oxygen profiles were analyzed for occurrence of concentrations below these
targets. Oxygen depletion occurred consistently at station 258 and 908 (Fig. 4) during the stratified
period from late June until September. At station 258 dissolved oxygen values at or below 4 and 1
mg l-1 occurred at 69% and 54% of the cruises respectively for both study years. At station 908, the
average occurrence of dissolved oxygen at or below 4 mg l-1 was 69% and 61.5% of the cruises, and
38% and 31% for the 1 mg l-1 concentration in 2003 and 2004 respectively.
At station 258 in 2002, the spring and fall whole water column oxygen levels averaged
9.9 mg l-1 and 6.2 mg l-1 respectively. During stratification, mean hypolimnetic oxygen ranged from
a high of 3.8 mg l-1 in late June to four occurrences of mean hypolimnetic oxygen below 1.0 mg l-1
from August to early October. At station 258 in 2002 the overall mean hypolimnetic oxygen
concentration was 1.3 mg l-1. Similar dissolved oxygen trends were observed at station 258 in 2004.
Dissolved oxygen and temperature at station 258 were plotted against depth for sampling
dates in 2002 (Fig. 5), to illustrate the irregularity of the oxygen pattern that can occur at this station.
Surface oxygen levels were high but could decrease throughout the epilimnion. Oxygen
concentrations plunge in the metalimnion and hypolimnion, but upward surges in oxygen
concentration could occur in the hypolimnion. At station 908 the oxygen concentrations were
typically acceptable in the epilimnion but plunged throughout the metalimnion and hypolimnion.
Dissolved oxygen concentrations and temperatures are plotted against depth for station 908 for
sampling dates in 2004 (Fig. 6). At station 908, the mean whole-water column dissolved oxygen
levels were range from 9.0 to 9.1 mg l-1 in spring and 7.7 to 8.2 mg l-1 in the fall (2003, 2004).
Epilimnion dissolved oxygen concentrations for both years (summer) ranged from 6.0 to 6.4 mg l-1.
Hypolimnetic mean dissolved oxygen ranged from 0.25 to 2.9 mg l-1 with an overall stratified
13
hypolimentic mean of 1.6 mg l-1 in 2004. For both study years (2003, 2004) hypolimnetic oxygen
concentrations at 908 were as high as 5.6 and as low as 0.17 mg l-1. Oxygen depletion in the harbour
was typically associated with stratification but there were incidents of low oxygen during
unstratified conditions mainly in the early fall (Fig. 5).
There were also occurrences of oxygen depletion at the nearshore stations. Station 6 had
four occurrences where oxygen dipped below 4 mg l-1. These were during summer between early
July until early August. There was only one cruise (July 8, 2002) where oxygen decreased below
1 mg l-1 and this was at the station bottom (6.7 m). Oxygen depletion also occurred at the nearshore
Willow Cove (WC) station with one occurrence of oxygen below 4 mg l-1 on July 8, 2003, when a
low of 1 mg l-1 was also measured at the station bottom (3.2 m). Unlike Hamilton Harbour, the Bay
of Quinte does not experience significant depleted oxygen at any of the sampled stations.
Nutrients
Seasonal data is defined as that data for the spring, summer or fall following the conventions
previously outlined in regards to stratification. Whole season data represents the mean of all data
collected over the year of sampling (i.e. May through October). Whole season water chemistry
results are summarized in Table 3. Nutrient levels exhibited a high degree of variation both
spatially and temporally in the harbour (Figures 7-10). With the exception of station 258 during
spring 2002, the seasonal trend for the offshore stations was high total phosphorus levels in the
spring followed by stepwise decreases in the summer and fall (Fig. 10). In 2002 seasonal means of
total phosphorus peaked in the summer at all stations sampled. Comparing all stations, total
phosphorus levels were highest in the spring at station 908 in 2003 (42.6 ug l-1) and lowest in the
fall at station 258 (2004= 21.8 ug l-1; Fig. 10). A gradient from nearshore to offshore was also
observed with the annual mean total phosphorus at the nearshore stations ranging from 35.8 - 37.9
ug l-1 compared to 25.5-30.8 ug l-1 for those offshore (Table 3).
The Hamilton Harbour RAP has set an initial harbour goal for total phosphorus
concentration to be 34 ug l-1; the final goal is 17 ug l-1. Total phosphorus values were below the
initial target on all but one occasion at station 258 in 2004 (May 26; Fig. 7); in 2004 there were also
three cruises where the total phosphorus values were below the final target listing. On a seasonal
basis the initial goal was met 57% of the time over the course of the study, mainly at the offshore
stations (Fig. 10). The final goal of 17 ug l-1 for the seasonal means was not met at any station.
Total phosphorus levels less than 17 ug l-1 were measured on seven dates during the study at the
offshore stations. In comparison, mean whole season total phosphorus in the offshore of Hamilton
Harbour during 2004 (25.5, 28.5 ug l-1) are comparable to the means observed in the upper and mid
Bay of Quinte stations B (26.5 ug l-1) and HB (24.4 ug l-1) over the same time period.
The whole season mean total phosphorus versus chlorophyll ratios for the harbour was
plotted in relation to the Bay of Quinte ratios for 2002-2004 (Fig. 11). As the Bay of Quinte data
were collected with the same methodology as for Hamilton Harbour, direct comparisons between
these two compromised ecosystems is valid. Nearshore station 17 exhibited the highest phosphorus
14
concentration and resulting chlorophyll density. The harbour’s offshore areas cluster mainly with
the Belleville (B) and Hay Bay (HB) stations in the Bay of Quinte. The deep station in the Bay of
Quinte (Conway (C) – 23.0 m) had the lowest total phosphorus versus chlorophyll ratio (Fig. 11)
Concentrations of soluble reactive phosphorus (SRP, orthophosphate) exhibited a high
degree of spatial and temporal variation throughout the year (Fig. 7), reaching a maximum of 6
ug l-1 at station 258 (2002) in mid September. At all stations the whole season mean SRP ranged
from 1.0 to 2.4 ug l-1 (Table 3a). The maximum seasonal mean of SRP occurred at station 258
(2002) in the spring (2.9 ug l-1). Spring and fall peaks were observed at stations 258 (2004), 908
(2004) and 17 (2002).
Ammonia concentrations reported in this study are dissolved values and not total. Ammonia
levels peaked in the spring and began to drop in late June (Fig. 8). Concentrations declined
considerably by late July and reached their lowest values in mid-August to early September. The
highest concentration observed was at Willow Cove in mid May (2003 = 1150 ug l-1). Average
ammonia levels in the spring were greater than 600 ug l-1 for all stations (Fig. 10). At all stations
there was a considerable decrease from the spring to summer ammonia levels. Summer ammonia
levels were up to eight times lower than those occurring in the spring (stations 258, 908, WC). Both
the highest seasonal mean (spring 2003 = 1060 ug l-1) and lowest (summer 2004 = 94 ug l-1)
occurred at station 908. The Bay of Quinte ammonia levels ranged from 12 to 14 ug l-1, drastically
less than the harbour values. Ammonia concentrations were calculated to un-ionized ammonia
using pH and temperature conversion factors from Emerson et al. (1975). This involved calculation
of seasonal means of pH and whole water column temperatures for each station for all years.
Conversion factors were determined from these means and the data converted accordingly (Fig. 10).
The Hamilton Harbour RAP has set a target for the un-ionized ammonia concentration to be less
than 20 ug l-1.
The un-ionized ammonia values are derived from dissolved values and not total. As a result,
the occurrence of exceedance of this contaminant above target levels is likely higher than reported
in this study. Seasonally, un-ionized ammonia concentrations exceeded the target concentration at
four stations in the spring. These were in 2003 (908 = 22.3 and WC = 20.4 ug l-1) and in 2004 (258
= 60.9 ug l-1 and 908 =57.5 ug l-1). Un-ionized ammonia levels in summer ranged from 1.2 to
2.4 ug l-1 and in the fall 0.8 to 1.7 ug l-1. As expected, seasonal means of un-ionized ammonia
followed the same pattern as measured ammonia with high levels in spring followed by a swift drop
in the summer and fall.
Nitrate-nitrite levels (dissolved) were high throughout the harbour and at all seasons (Fig.
10). Values reported indicate both nitrate and nitrite but generally the nitrate component represents
the bulk of the reported value as nitrite is rare in aquatic systems (Keeney 1972). However Barica
(1990) reported that nitrite levels in the harbour were a concern and therefore their portion in the
nitrate-nitrite results cannot be discounted. Levels over 2900 ug l-1 were observed at station WC
(2003) in June and again in early July. The lowest observed value occurred at station 17 (2002) at
854 ug l-1 (Fig. 8). With the exception of station 908 in 2004, the nitrate-nitrite levels for spring
15
versus summer were similar (spring mean range 1863–2343 ug l-1; summer mean range 1646-2243
ug l-1. For all stations nitrate-nitrite levels dropped considerably in the fall (mean range 1390–1776
ug l-1). Highest seasonal mean of nitrate-nitrite was in the nearshore (WC-2003 spring; 2343 ug l-1)
and the lowest was at station 258 in the fall (2002-1390 ug l-1). The highest mean nitrate-nitrite
concentration for 2002 - 2004 in the Bay of Quinte occurred at Conway (250 ug l-1); this is
considerably less than the mean concentrations observed in Hamilton Harbour for the same
timeframe (1673-2147 ug l-1).
As with other nutrients, silica concentrations were variable between stations and sampling
periods (Fig. 9) ranging from 0.06 to 1.6 mg l-1. Whole season silica means for the nearshore
stations (range 0.52-0.80 mg l-1) were comparable to those offshore (range 0.55-0.82 mg l-1; Table
3). Highest silica concentration (1.6 mg l-1 ) occurred at station 258 (2002) in mid October. On a
seasonal basis, lowest mean silica concentrations were observed in the spring (Fig. 10). The lowest
spring values occurred at all stations in 2002 (6 = 0.08, 17 = 0.12, 258 = 0.12 mg l-1). The only time
silica concentrations exceeded 1.0 mg l-1 was in the fall. This fall upsurge of silica occurred at all
stations but remained below 1.0 mg l-1 at station 908 in 2003 (0.89 mg l-1). In the Bay of Quinte,
station B had higher whole season (2002 through 2004) silica concentrations ranging from 2.7 to
3.3 mg l-1. Conway, the deepest and least enriched Bay of Quinte station had annual silica
concentrations ranging from 1.2 to 1.3 mg l-1 for the same timeframe. The spring mean at Conway
for 2002-2004 was 1.9 mg l-1.
Mean annual total suspended solids (TSS) ranged from 3.4 to 4.7 mg l-1 at the offshore
stations and 4.4 to 8.0 at nearshore stations (Table 2). Comparing seasonal means, the highest level
of 9.8 mg l-1 occurred in the summer at station 17 (2002) and a low of 2.7 mg l-1 was observed at
stations 258 (2004) and WC (2003) in the fall (Fig. 12).
Chlorophyll a
The highest whole season mean chlorophyll a concentration in the study (22.9 ug l-1) was observed
at the shallowest station (17-depth 1.3 m; Table 2). At the offshore stations, lower whole season
concentrations were observed in 2004 (258 = 11.4 ug l-1, 908 = 11.7 ug l-1) than in 2002 (15.7 ug l-1)
or in 2003 (16.4 ug l-1). These latter means are comparable to the whole season means observed at
stations 6 (16.4 ug l-1) or WC (13.4 ug l-1). For the most part chlorophyll a concentrations peaked in
the summer. Station 17 had the highest seasonal means for the study in the summer (25.0 ug l-1) and
fall (28.2 ug l-1) (Fig. 13).
The RAP has set an initial goal of 15-20 ug l-1 chlorophyll a concentration and a final goal of
5-10 ug l-1. At the central station 258, chlorophyll a concentrations were below 20 ug l-1 at all
sampling events in 2004 and all but three dates in 2002. Station 908 had lower chlorophyll a
concentrations in 2004 when all but two dates had levels below 20 ug l-1. Stations 6 and 17
exceeded 20 ug l-1 chlorophyll a concentrations on four occasions and WC was in exceedance of 20
ug l-1 on two occasions (Fig. 9). For all stations, chlorophyll a concentrations were less than 20 ug l-1
at 77% of the time over the duration of the study. Chlorophyll a concentrations below the final target
16
of 10 ug l-1 occurred 35% of the time during the study. The station which most consistently met this
target was 258 in 2004 when 9 of the 13 samples were below 10 ug l-1.
In the Bay of Quinte, whole season mean chlorophyll a concentrations from 2002-2004 for
B, HB and C were 12.9, 12.0 and 3.2 ug l-1 respectively (Table 2). In 2003, HB had unusually high
chlorophyll a levels, excluding 2003 the mean for this station was 9.6 ug l-1, which was more typical
for this station (Burley and Millard 2006). These levels reflect the gradient in nutrient enrichment
which is high in the upper bay (Belleville) and decreases downstream to Conway. The whole
season means for the offshore harbour stations in 2004 (258=11.4 ug l-1, 908= 11.7 ug l-1) and WC
(2003= 13.4 ug l-1) are comparable to the chlorophyll levels observed at B and HB for the same
timeframe.
Light extinction
The Hamilton Harbour RAP set initial and final goals for Secchi disk transparency to be 2 and 3 m
respectively. Station 17 in 2002 was excluded from the comparisons as its station depth was only
1.3 m; so the bottom was always visible at this station. On a whole season basis all but two stations
met the initial goal criteria of 2.0 m; stations WC (2003) and 908 (2003) were close at 1.8 and 1.9 m
respectively (Table 2). None of the station’s annual mean Secchi depth met the final 3 m goal. Mid
harbour station 258 had the highest whole season mean of 2.4 m in both 2002 and 2004; station WC
(2003) had the lowest at 1.8 m (Table 2).
There were occurrences when Secchi depth met the final goal of 3 m. In 2003, Secchi depths
were at or greater than the final RAP target once at both station 908 and WC (Fig. 14). There were
three occurrences when Secchi depth was at or greater than 3 m at station 258 in both sampling
years, and at station 908 in 2004.
Seasonally lowest transparencies occurred in the spring where all but one station (258-2002;
2.5 m) were below the 2.0 m target. Generally transparency increased stepwise from spring to
summer to fall. The fall mean Secchi depths exceeded the 2.0 m target at all of the stations. The
lowest seasonal mean Secchi depth occurred at station 908 in spring 2003 at 1.5 m. This station
also had the maximum Secchi depth mean for the study at 2.9 m in the fall of 2004 (Fig. 12).
The light extinction coefficients (εpar) increase as transparency decreases. Lowest observed
whole season εpar mean (highest transparency) occurred at station 258 in 2004 (0.709 m-1; Table 2).
At the offshore stations higher εpars were associated with the spring. A drop in εpar from summer to
fall was observed at stations 258 in 2002 and 908 in 2003 and at all the nearshore stations (Fig. 12).
Both the light extinction coefficients and Secchi depth indicated that greatest water clarity
on a whole season basis occurred at station 258. Station 908 and nearshore stations 6 (2002) and
WC (2003) were similar to each other with εpars ranging from 0.76 to 0.77 m-1 (Table 2). Despite
visually being able to identify the bottom at shallow station 17, it had the highest whole season εpar
mean of 1.06 m-1. While the Secchi disk data indicates lowest water transparency occurred in the
17
spring, the εpar data indicates that the spring or summer season could have the lowest water
transparency. Spring highs in εpar (lowest water transparency) were observed at stations WC, 258 in
2004 and 908 in both study years which is congruent with Secchi depths which were lowest in
spring at these stations (Fig. 12). The light extinction coefficient was most notable at 908 in 2003
where the spring εpar (1.264 m-1) was almost double the summer mean (0.656 m-1). Summer highs
in εpar were observed nearshore at stations 6 and 17 (2002) and offshore at station 258 in 2002.
In the Bay of Quinte, comparative whole season means for Secchi depth for 2002, 2003 and
2004 at the B, HB and C were 1.9, 2.2 and 6.0 respectively. Whole season mean light extinction
coefficients for B, HB and C for 2002 through 2004 were 0.957, 0.749 and 0.359 m-1) respectively
(Table 2, Fig. 12). This reveals that transparency increases from the uppermost shallow station
Belleville to the deep station Conway near the mouth of the bay which exchanges water with Lake
Ontario.
Phytoplankton Photosynthesis
Seasonal areal phytoplankton photosynthesis (SAPP) was calculated for all stations and years. The
lowest SAPP observed in the study was at station 17 at 156 g C m-2. Offshore stations 258 and 908
had lower SAPP values in 2004 (190, 215 g C m-2 respectively) compared to 2002 (258= 282 g C
m-2) and 2003 (908= 346 g C m-2) (Table 4, Fig. 13). This latter SAPP value at 908 was the highest
observed in the study. The SAPP percentage of cloudless for 2004 (258-65%, 908-66%) was low in
comparison to the other years in the study (70-77%). In the Bay of Quinte, average SAPP for
Quinte station B ranged from 189-289 g C m-2 and 136-359 g C m-2 at HB. Conway’s SAPP values
were much lower at 119-188 g C m-2 from 2002-2004 (Table 4). Total phosphorus concentrations
versus SAPP values were analyzed comparing Hamilton Harbour and the Bay of Quinte (Fig. 11).
DISCUSSION
During the study period (2002 through 2004), the Harbour was stratified by late June and remained
so until early to mid October. At station 258 the mixed depth became deeper in mid-September and
conditions were isothermal by mid October. Station 908 is not as deep (14.6 m) as 258 (23.3 m) and
was isothermal by mid September in both study years. The Ministry of Environment (1985)
reported isothermal conditions persisting in the harbour from mid-October through May. There are
complex and seasonally variable exchanges of harbour and Lake Ontario water via the Burlington
Ship Canal. During the stratified period, cold well oxygenated Lake Ontario water can travel along
the bottom of the harbour providing oxygen to the hypolimnion. These lake water intrusions are
reported to spread up to two-thirds of the way to the Stelco property and up to 1.5 km laterally
(MOE 1985). Despite this periodic influx, stratification results in oxygen depletion in the
hypolimnion because epilimnetic oxygen remains trapped in the upper layer resulting in oxygen
depletion in the metalimnion and hypolimnion.
Oxygen depletion in the harbour remains a persistent issue and is a serious barrier to long
term ecosystem recovery. The main factor causing oxygen depletion is bacterial oxidation of the
18
ammonia from sewage effluent to nitrate by nitrifying bacteria in the water column, which accounts
for 35 to 45 percent of the oxygen demand (RAP 1992, Charlton and Milne 2005). Oxygen is
further depleted through bacterial oxidation in the water column and sediments of reduced carbon,
nitrogen and sulphur present in the effluents and oxygen demand from phytoplankton decay (RAP
2003). Decomposing algae may represent 30 to 35 percent of the oxygen demand during the
summer (RAP 1992). Reaeration, photosynthesis and inflows from Lake Ontario all contribute to
the oxygen present in the harbour (MOE 1985).
The Ministry of Environment (1985) reported that from 1976 to 1980 the harbour exhibited
a rapid decrease in oxygen in May and June and hypolimnetic concentrations were close to zero. In
the current study oxygen levels remained sufficient (9-10 mg l-1) until stratification set up in mid to
late June. In the 1976 to 1980 time period, hypolimnetic means during stratification ranged from 1
to 2 mg l-1 , which are comparable to the overall hypolimnetic means of 1.3 to 1.6 mg l-1 observed in
the current study. Epilimnetic dissolved oxygen in 1976 to 1980 averaged between 6.0 to 8.0 mg l-1;
the current study had epilimnetic means lower than 6.0 mg l-1 at station 258 but they were still above
the 4.0 mg l-1 RAP target. The study sites sampled were likely beyond the range of the main
exchange of Lake Ontario water via the Burlington Canal (RAP 2003).
Water quality goals for the harbour were adopted by the Hamilton Harbour RAP committee
in 1992 (Charlton and Milne 2005). The final RAP goal of a minimum of 4 mg l-1 dissolved oxygen
concentration is the provincial water quality objective set by the Ministry of Environment for warm
water biota living at 20 to 25 degrees Celsius. The objectives for cold water biota at zero degrees
Celsius range from 4 to 8 mg l-1 dissolved oxygen. The initial RAP goal for dissolved oxygen is a
minimum of 1 mg l-1 dissolved oxygen concentration. Below 1 mg l-1 conditions are hypoxic and
unable to sustain most biota. The persistence of hypoxia at the stratified stations is a critical issue
limiting biota from these areas. The periodic occurrence of depleted oxygen at the nearshore
stations puts additional stress on the biota of the harbour. Currently the harbour fish populations are
dominated by species that tolerate low oxygen such as carp, bullhead and gizzard shad. The return
of viable cold water populations of lake trout, whitefish, herring and sturgeon is unlikely and
remediation efforts are focusing on restoring warm water species such as pike, bass, perch, crappies,
as well as rainbow and brown trout and Pacific salmon (RAP 1992).
Davis (1975) reported negative physiological and/or behavioural responses on the desired
warm water species such as largemouth bass, rainbow trout and Pacific salmon at dissolved oxygen
levels less than 4.5 mg l-1. Seager et al. (2000) reported greater than 95 percent mortality of rainbow
trout at 1.6 mg l-1 dissolved oxygen over a short term exposure and implicated duration of the
exposure as a critical factor. The high occurrence of oxygen levels below 4 mg l-1 in the deep
stations will hamper establishment of healthy population of these warm water fishes in the harbour.
Dermott (this volume) indicates that the benthic invertebrate community living beyond nine metres
is also restricted due to the oxygen depletion across the harbour bottom.
Nutrient loading to the harbour is a causative agent of the oxygen depletion and
eutrophication occurring in the harbour. To promote the reduction of the nutrients, the RAP
19
committee has set initial (34 ug l-1) and final (17 ug l-1) targets for total phosphorus concentration in
the harbour. Phosphorus is considered to be a major factor controlling phytoplankton biomass and
resulting trophic status (Dillon and Rigler 1974, Schindler, D.W. 1974). Charlton and Milne (2005)
report total phosphorus concentrations peaking in summer which corresponds to the observed
seasonal means at the nearshore stations and at station 258 in 2002. Spring peaks of total
phosphorus were also observed at the offshore stations 258 (2004) and 908 (2003, 2004) in this
study and this pattern has also been reported by Charlton and Milne (2006). There is variability in
the nutrient dynamics of the harbour which is evident in the biweekly sampling values.
Phosphorus concentrations are high due to inputs from four sewage treatment plants,
combined sewer overflows, agricultural and urban runoff and the steel industry. Loadings to the
harbour have been reduced from 1,200 kg/day in 1967 to less than 10 kg/day in 1989 (RAP 1992).
The MOE (1985) reported that in 1975 to 1983 annual means of total phosphorus ranged from 56
to 104 ug l-1. During 2002 to 2004, the harbour whole season means for total phosphorus ranged
from 25.5 to 37.9 ug l-1 which represents a two and a half fold decrease from the previous study.
Values in the current study did not exceed 72.9 ug l-1 and Charlton and Milne (2005) reported
annual means below 80 ug l-1 since 1990. Reducing loads to the harbour have been effective and in
this study the initial target was met on a seasonal basis over half of the time at the stations studied.
This fact combined with the occurrence of total phosphorus values below the final target on some
dates over the course of the study is encouraging. Continued reductions of this nutrient are an
important element in further limiting phytoplankton growth and increasing water transparency.
Seasonal means for total phosphorus versus chlorophyll were plotted to compare Hamilton
Harbour data with the Bay of Quinte. Both areas are embayments of Lake Ontario requiring
rigorous nutrient enrichment controls to abate eutrophication. The Bay of Quinte has undergone
anthropogenic eutrophication beginning with European colonization; remediation efforts have been
undertaken since the early 1970s. The linear regression of total phosphorus to chlorophyll indicates
that for the timeframe of 2002 through 2004 there is no difference in response to total phosphorus
inputs in terms of chlorophyll production between the two Areas of Concern. Comparing both
systems the nearshore stations are the most nutrient enriched; station 17 in Hamilton Harbour
exhibits the highest chlorophyll to total phosphorus ratio. The offshore areas of Hamilton Harbour
closely resemble the enriched upper (B) and mid bay (HB) area of the Bay of Quinte and are far
removed from the low chlorophyll to total phosphorus ratios observed in the deep station in the Bay
of Quinte (station C).
Soluble reactive phosphorus (SRP, orthophosphate) is the inorganic form of phosphate and
is a measure of biologically available phosphorus readily uptaken by bacteria and phytoplankton
during photosynthesis. The current results concur with Harris et al. (1980) who reported peak
concentrations of SRP associated with spring and fall overturn. Summer lows are likely associated
with depletion of SRP under high algal densities. Charlton and Milne (2005) reported possible
regeneration of SRP from the sediments triggered by extended periods of anoxic conditions.
The phosphorous loading to Hamilton Harbour has dramatically been reduced but continuing to
reduce loading is required in order to further shift the system from its eutrophic state. From 1984 to
20
1986 SRP averaged 13 ug l-1; improvements to the current levels in the harbour (1.0 - 2.4 ug l-1) is a
factor controlling algal abundance (RAP 2003).
The 1992 Remedial Action Plan (RAP 1992) stated that while industrial loadings of
ammonia have decreased substantially in the last forty years there has been little change in the
ammonia loadings from municipal sewage treatment plants. The latter contributes 80 % of the
ammonia to the harbour. Annual ammonia concentrations of 1130 to 1950 ug l-1 during 1975 to
1983 were reported by MOE (1985). The current concentrations of 218 to 320 ug l-1 ammonia
indicate a drastic reduction over the last twenty five years but these values are dissolved, not total
ammonia. Compared with station B, the most enriched station in the Bay of Quinte study, ammonia
levels are sixteen to twenty three times higher in Hamilton Harbour. Ammonia concentrations are
highest in the spring because of a build-up of ammonia over winter. During these months sewage
effluent enters the harbour but nitrifying bacteria are inactive in the cold temperatures (Fletcher
1979). As water temperature rises, oxidation of ammonia to nitrate by nitrifying bacteria increases
to the point where hypolimnetic oxygen concentrations in the summer reach critical lows
(0.5-1 mg l-1) as nitrification rapidly depletes the dissolved oxygen from the bottom water layer.
Another issue with ammonia is its un-ionized form which is the most toxic form to aquatic
life; the RAP committee has adopted the Provincial guideline of 20 ug l-1 as the target for un-ionized
ammonia in the harbour. The percent of ammonia present in the un-ionized form is dependent on
the pH and water temperature, and the subsequent conversion of the ammonia data indicated levels
exceeded the target in the spring both at the offshore stations as well as at Willow Cove. The
highest concentration un-ionized ammonia in the current study was 60.9 ug l-1 as a result of
increased pH. The Harbour RAP also reported concentrations exceeding the target in the late spring
and summer with a high over 120 ug l-1 reported in the 1987 and 1988 data (RAP 1992). Barica
(1990) reported values in 1987 to 1988 that exceeded the targets for most of the spring and summer.
The current study suggests a possible reduction in un-ionized ammonia concentrations, but the
number of sampling sites was limited and results were based on dissolved ammonia, not total.
Charlton and Milne (2005) reported fewer occurrences of un-ionized ammonia concentrations above
the target in 2000 to 2005, due to greatly improved removal at the Burlington Skyway wastewater
treatment plant. Further controls to reduce ammonia in the harbour are an essential component for
improving the oxygen regime in the harbour.
In 1977, the Ministry of the Environment (MOE 1981) indicated the peak in seasonal nitrate
coinciding with the time of maximum harbour BOD; maximum surface nitrate values (in July)
averaged 2610 ug l-1. This relates to the nitrification of ammonia in the spring which requires
oxygen. A further review (MOE 1985) listed seasonal means for nitrates from 1975-1983 ranging
from 1410 to 2180 ug l-1; this is similar to the current concentrations from 2002 through 2004 which
ranged from 1673 to 2147 ug l-1. In the current study, peaks occurred in a similar timeframe
generally in early July. The maximum of 2960 ug l-1 in early June indicates high levels of nitratesnitrites still occur in the harbour despite reductions in ammonia loadings (RAP 1992). Seasonal
means over 2000 ug l-1 were observed at both offshore stations (908-2003, 258-2004) and nearshore
at station WC (2003). Generally nitrate-nitrite levels dropped by early September and lowest
21
means occurred in the fall. This is likely related to denitrification of nitrates during the anoxic
summer period and gassing off as atmospheric nitrogen (MOE 1981). Harris et al. (1980) found
rapid nitrification of ammonia in June and July but the higher nitrate values they reported in the fall
did not occur in this study. It does not appear there has been a meaningful reduction in nitratesnitrites in Hamilton Harbour since 1980.
The emphasis is on reductions in un-ionized ammonia, and there is currently no target set for
nitrates or nitrites. Harris et al. (1980) reported nitrite levels in 1976 and 1977 and Barica (1990)
reported nitrite concentrations frequently exceeding acute toxic levels of 250 ug l-1 and chronic
toxicity thresholds (30 ug l-1) in 1987-1988. Scott and Crunkilton (2000) reported nitrate toxicity in
fish fry and zooplankton but exposure levels far exceeding what was observed in the harbour. The
lowest toxic concentration of nitrates reported was 6250 ug l-1 which resulted in sublethal effects on
fry of lake trout and lake whitefish (McGurk et al. 2006). The nitrate-nitrite levels in the harbour
are considerably higher than in the Bay of Quinte. Concurrent reductions of nitrates and nitrites
associated with ammonia reductions would be positive for Hamilton Harbour to reduce potential
toxicity and overall nutrient enrichment to the ecosystem.
The RAP report (1992) indicated about one third of the suspended solid loadings enter the
harbour via Cootes Paradise; combined sewer overflows are the second greatest contributor at 19
percent. Industry and sewage treatment plants have reduced suspended solids loading in
conjunction with pollutant reductions. The Ministry of the Environment (MOE 1985) found
suspend solid concentrations in the harbour of 4.7 to 5.8 mg l-1 from 1975 to 1983 which are
comparable to the current findings of 3.4 to 8.0 mg l-1. With the exception of station 17, the whole
season concentrations of suspended solid in Hamilton Harbour are similar to that in the upper (B)
and mid bay (HB) station in the Bay of Quinte. Further reductions in suspended solids will be
beneficial to water clarity and reduced oxygen demand in the harbour.
Silica can be a limiting nutrient for phytoplankton growth since diatoms require large
quantities of silica for their cell walls (Goldman and Horne 1983). Levels were low at all stations in
Hamilton Harbour over the duration of the study compared to the Bay of Quinte. Station B had
whole season means four to six times higher than Hamilton Harbour for the same timeframe. The
spring seasonal means at station C, the least enriched and deepest station in the study were sixteen
times greater than spring means in Hamilton Harbour. The difference between the two Areas of
Concern is unclear but the lows in the spring are related to depletion by spring diatom blooms and
silica may be a limiting nutrient for diatoms in the harbour at this time of year.
Harris et al. (1980) reported silica in the Harbour ranging from 0.2 to 1.5 mg l-1 in 1975
which is comparable to the range in the current study (0.06 to 1.6 mg l-1). Thus current silica levels
in the harbour are comparable to mid 1970s values. Goldman and Horne (1983) indicate a release
of silica under anoxic sediments. In the harbour silica may be released from sediments into the
oxygen depleted hypolimnion during summer stratification and mixed into the whole water column
after fall turnover. In combination with the low diatom levels observed in the fall, this may partially
explain the high silica levels observed in the fall at all stations. Harris et al. (1980) reported mid and
22
bottom water silica concentrations rising during the stratified period but found decreases in silica
during the fall.
Water transparency measured as Secchi depth was lowest in the spring and typically highest
in the fall. Transparency in the harbour has been mainly related to chlorophyll a levels, suspended
silt and dissolved substances (RAP 2003, Harris et al. 1980). The spring transparency may be
related to turbidity from suspended solids entering the harbour during spring runoff. Total
suspended solids were highest in the spring at stations 908 and WC. These stations are located in
the western portion of the harbour near Cootes Paradise and Grindstone Creek. This turbidity may
reach as far as the central station 258 which exhibited high spring turbidity in 2004. The light
extinction coefficient is a direct measure of photosynthetically available radiation and consequently
is a more accurate tool to assess water transparency than Secchi depth. The light extinction
coefficients revealed that lowest transparency occurred in summer at nearshore stations 6 and 17.
Harris (1980) reported that chlorophyll a was a significant factor controlling water transparency in
the water column. Chlorophyll a levels were high at stations 6 and 17 in the summer and are likely
related to the seasonally low water transparency (high εpars) during this season. In 1986 the
average Secchi disc transparency in the harbour was 1.4 m (RAP 2003); the current overall mean of
2.1 m indicates a 50 % increase in transparency in the past two decades.
The mean whole season Secchi depth for the harbour met or came very close to the initial
RAP committee target for Secchi depth of 2.0 m. Charlton and Milne (2005) reported increased
Secchi depth since the late 1990s and better than usual Secchi disc transparency readings in 2005
(Charlton and Milne 2006). With the exception of station 17, the εpar in the current study ranged
from 0.71 to 0.77 which indicates a marked improvement in water clarity from a 1975 study by
Harris et al. (1980) who reported extinction coefficients ranging from 0.8 to 1.6 m-1.
Offshore station 258 had the highest water transparency in the study considering both εpar
and Secchi depth, but its εpar was still two times greater than that observed in the deep station
Conway in the Bay of Quinte. On a positive note, both stations 258 and 908 had occurrences where
Secchi depth exceeded the final goal of 3.0 m. Comparing light extinction coefficients, nearshore
station 17 is similar to the upper Bay of Quinte (station B) while the remaining nearshore stations
and station 908 most closely resemble the middle Bay of Quinte (station HB). The offshore station
258 has slightly greater water transparency than the mid bay (HB) station.
Chlorophyll a concentrations are an indicator of phytoplankton density and were highest at
the shallowest station 17. At the offshore stations in 2004, chlorophyll a levels were lower than in
2002 and 2003, likely related to the lower total phosphorus concentrations observed in 2004
Charlton and Milne (2006) also reported lower than usual chlorophyll a concentrations at station
258 in 2005. In the offshore, chlorophyll a concentrations peaked in summer; this also occurred at
station 6 which was the deepest nearshore station at 7 m. Charlton and Milne (2005) reported high
summer chlorophyll a in the harbour which loosely paralleled phosphorus concentrations; it is
notable that spring highs in total phosphorus were also observed in the current study. The RAP
targets for chlorophyll a concentration are 15-20 ug l-1 for the initial phase and 5-10 ug l-1 as the
23
final goal. On a whole season basis, this initial target was met at all stations except 17 which was
slightly higher at 22.9 ug l-1. In 2004, central station 258 was compliant with the initial goal on
every sampling occasion. Charlton and Milne (2005) reported a decrease in peak chlorophyll a
concentrations since 2000 and a trend for more values are within the range of the final goal.
The offshore whole season chlorophyll a levels for 2004 were comparable to the upper bay
station B (12.9 ug l-1). The offshore whole season chlorophyll a concentrations for 2002 as well as
nearshore stations 6 and WC were elevated compared to station B, ranging from 15.7 to 16.4 ug l-1;
station 17 has considerably higher chlorophyll a levels than B. Charlton and LeSage (1996)
reported only a few occurrences of compliance to the final chlorophyll a target for transects along
the north shore of the harbour in 1994. The highest seasonal mean observed in the current study
was 28.2 ug l-1 which was a marked reduction in chlorophyll a since 1975 when Harris et al. (1980)
found chlorophyll a fluctuated between 30 to 60 ug l-1 in the harbour. The current 77% and 35%
compliance rate for all sampling occurrences to initial and final goal concentration targets
respectively is encouraging.; but further reductions in phosphorus loading to the harbour would be
beneficial in limiting algal growth.
Seasonal areal phytoplankton photosynthesis (SAPP) in the offshore harbour stations was
approximately one and a half times higher in 2002 and 2003 than in 2004. This relates to higher
total phosphorus and chlorophyll a levels observed in the earlier years. The low percentage of
cloudless in 2004 (Table 4) was atypical and a result of the low incident solar irradiance that year.
This same trend was noted in the Bay of Quinte for 2004 resulting in low SAPP values in both the
Bay of Quinte and Hamilton Harbour in 2004 (Burley and Millard 2006). Atypically, low solar
irradiance in 2004 was mainly responsible for the lower production values observed at both
locations in that year. Station 17 had the lowest SAPP in the study despite having high chlorophyll a
levels; depth truncation of the calculated rates explains in part the low value observed at this site;
reduced water transparency is another cause for the low SAPP. The highest SAPP was observed in
2003 at station 908; although water transparency was lower at 908 in 2003 than 2004 it was not to
the same degree as was observed at station 17. The nearshore stations 6 and WC had SAPP values
comparable to the central station 258.
Both the Ministry of the Environment (MOE 1985) and Harris et al. (1980) reported lower
than predicted phytoplankton production in Hamilton Harbour in the mid to late 1970s. This was
attributed to light limitation due to low water transparency caused by high levels of suspended
solids and dissolved organics in the harbour (MOE 1985, Harris et al. 1980, Piccinin 1976). Light
extinction coefficients in the current study indicate improved water transparency to a range similar
to that in the Bay of Quinte. The euphotic zone is the depth range in which solar irradiance
penetrates to permits photosynthesis. It is generally considered to be from the surface to a depth
with 1 % of the surface solar irradiance. The mean euphotic depth at station 258, 908 and Quinte
station HB were 6.8 m, 6.7 m and 6.7 m respectively, indicating water clarity is now comparable
between these two locations.
24
The Ministry of the Environment (MOE 1985) implicated variable mixing depths as another
major factor limiting algal production. They reported that due to inputs from Lake Ontario, wind
and other effects, the thermocline depth, and hence the mixing depth, is variable. Phytoplankton are
often mixed below the euphotic zone dampening production. In the current study stratification
appeared stable and the mixing depth did not alter radically over the course of the summer at either
offshore station. The stability may be related to the location of the offshore stations, which were far
enough from the canal to limit the influence of intruding Lake Ontario water. Mixing depth did
exceed the euphotic zone depth during isothermal conditions. This occurred at station HB as well
but mainly in the latter part of the field season. Toxicity of chemical contaminants to phytoplankton
in the harbour has also been suggested as having an inhibitory effect on production (RAP 2003) but
MOE (1985) reported this was secondary when compared with constraints due to physical factors
discussed previously.
The range of SAPP observed in Hamilton Harbour are comparable to those observed in the
Bay of Quinte. A linear regression of total phosphorus concentrations versus SAPP indicates that
for the 2002 to 2004 time period, the ecosystem response of SAPP to total phosphorus inputs was
comparable for the two study locations. Phytoplankton production in the harbour is not dampened in
comparison to the Bay of Quinte. Increased water transparency in the harbour and thermal stability
are contributing factors to the equivalency of phytoplankton production observed in these two Areas
of Concern. This suggests that continuing to reduce phosphorus loading to the harbour is a key
factor controlling algal growth, as water quality continues to improve through the remediation
efforts.
Hamilton Harbour’s long history as an industrial port and dumping site for industrial and
municipal wastes makes it a complicated Area of Concern to successfully remediate and restore the
beneficial uses. In addition to chemical contamination of its waters and sediment, the
eutrophication of the harbour has compounded the water quality issues. Great strides have been
made to improve water quality. Concentration of metals currently meet provincial guidelines and
reductions in nutrient loads have improved water clarity and reduced phytoplankton biomass. A
focus to further reduce nutrient loading, in particular ammonia and phosphorus, will improve the
oxygen regime in the harbour. The current gains in water quality observed since the 1970s is
testimony to the commitment and resolve of the Hamilton Harbour RAP and community.
ACKNOWLEDGEMENTS
The author would like to thank Ron Dermott, Marten Koops, Jocelyn Gerlofsma and Mark
Fitzpatrick for their assistance in preparing this report.
25
REFERENCES
Barica, J. 1990. Ammonia and nitrite contamination of Hamilton Harbour, Lake Ontario. Water
Pollution Research Journal of Canada, 25(3): 359-386.
Burley, M. and Millard, S. 2006. Photosynthesis, chlorophyll a, and light extinction. In: The Big
Cleanup: Project Quinte Annual Report 2004, Bay of Quinte Remedial Action Plan,
Kingston, ON. pp. 23-39.
Charlton, M.N. and Le Sage, R., 1996. Water quality trends in Hamilton Harbour: 1987 to 1995.
Water Qual. Res. J. Canada. (31): 473-484.
Charlton, M. and Milne, J. 2005. Eutrophication indicators in Hamilton Harbour 2004. Hamilton
Harbour Remedial Action Plan Research and Monitoring Report; 2004 Season. ISSN 17034043, Nov. 2005. 2-12.
Charlton, M. and Milne, J. 2006. Water quality update 2005. Hamilton Harbour Remedial Action
Plan Research and Monitoring Report; 2005 Season. ISSN 1703-4043, Dec. 2006. 2-6.
Davis, J.C. 1975. Minimal dissolved oxygen requirements of aquatic life with emphasis on
Canadian species: a review. J. Fish. Res. Board Can. 32(11): 2295-2332.
Dillon, P.J. and Rigler, F.H. 1974. The phosphorus-chlorophyll relationship in lakes. Limnol.
Oceanogr. 19, 767-773.
Environment Canada. 1995. Manual of analytical methods: Major ions and nutrients. Vol 1.
N.W.R.I. Water Quality Branch, Ottawa, Ontario. 340 pp.
Fee, E.J. 1990. Computer programs for calculating in situ phytoplankton photosynthesis. Can.
Tech. Rep. Fish. and Aquat. Sci. 1740, 27 pp.
Fletcher, M. 1979. The aquatic environment. In: J.M. Lynch and N.J. Poole (Eds.), Microbial
Ecology: A Conceptual Approach. Wiley, New York
Goldman, C.R., Horne, A.J. 1983. Limnology. McGraw-Hill, Inc., New York, NY.
Harris, G.P., Piccinin, B.B., Haffner, G.D., Snodgrass, W., Polak, J. 1980. Physical variability and
phytoplankton communities: I. The descriptive limnology of Hamilton Harbour. Arch.
Hydrobiol. 88(3): 303-327.
Hamilton Harbour Remedial Action Plan (RAP). 1992. Remedial Action Plan for Hamilton
Harbour, Goals, Options and Recommendations. RAP Stage 2 Report, ISBN 0-7778-0533-2,
Canada Ontario Agreement. Nov. 1992. 327 pp.
Hamilton Harbour Remedial Action Plan (RAP). 2003. Remedial Action Plan for Hamilton
Harbour: Stage 2 Update 2002, ISBN 0-9733779-0-9, Canada Ontario Agreement. June
2003. 247 pp.
26
Holmes, J.A. and Whillans, T.H. 1984. Historical review of Hamilton Harbour fisheries. Can. Tech.
Rep. Fish. Aquat. Sci. 1257, i-x 117 pp.
International Joint Commission (IJC). 1988. Revised Great Lakes Water Quality Agreement of
1978, as amended by Protocol signed Nov. 18, 1987. Consolidated by the International Joint
Commission of United States and Canada. January, 1988. 130 pp.
Keeney, D.R. 1972. The fate of nitrogen in aquatic ecosystems. Eutrophication Information
program, University of Wisconsin Water Resources Center, Madison, Wisconsin. Literature
Review 3, 1-59.
MacDougall, T.M., Benoit, H.P., Dermott, R., Johannsson, O.E., Johnson, T.B., Millard, E.S.,
Munawar, M. 2001. Lake Erie 1998: assessment of abundance, biomass and production of the
lower trophic levels, diets of juvenile yellow perch and trends in the fishery. Can. Tech. Rep.
Fish. Aquat. Sci. 2376, 190 pp.
McGurk M.D., Landry, F., Tang, A., Hanks, C.C. 2006. Acute and chronic toxicity of nitrate to
early life stages of lake trout (Salvelinus namaycush) and lake whitefish (Coreognus
clupeaformis). Environ.Toxicol. Chem. 25(8): 2187–2196.
Ministry of Environment of the province of Ontario (MOE). 1981. Hamilton Harbour study 1977:
volume 1. 320 pp.
Ministry of Environment of the province of Ontario (MOE). 1985. Hamilton Harbour technical
summary and general management options. 125 pp.
Piccinin, B.B. 1976. The biological survey of Hamilton Harbour, 1975. Hamilton: McMaster
University, Department of Biology. Submitted to the Ontario Ministry of Environment.
Poulton, D.J. 1987. Trace contaminant status of Hamilton Harbour. J. Great Lakes Res. 13(2):193201.
Schindler, D.W. 1974. Eutrophication and recovery in experimental lakes: implications for lake
management. Sci. 184, 897-899.
Scott, G. and Crunkilton, R.L. 2000. Acute and chronic toxicity of nitrate to fathead minnows
(Pimephales promelas), Ceriodaphnia dubia, and Daphnia magna. Environ. Toxicol. Chem.
19(12): 2918-2922.
Seager, J., Milne, I., Mallett, M. and Sims, I. 2000. Effects of short-term oxygen depletion on fish.
Environ. Toxicol. Chem. 19(12): 2937-2942.
Strickland, J.D.H. and Parsons, T.R. 1972. A practical handbook of sweater analysis. 2nd ed. Bull.
Fish. Res. Board Can. 167, 310 pp.
Whillans, T.H. 1979. Historic transformation of fish communities in three Great Lakes Bays. J.
Great Lakes Res. 5(2): 195-215.
27
Table 1. Hamilton Harbour sampling dates for all stations and years sampled.
Cruise
17
2002
6
2002
WC
2003
258
2002
258
2004
908
2003
908
2004
1
15-May
15-May
14-May
15-May
13-May
14-May
13-May
2
28-May
28-May
27-May
28-May
26-May
27-May
26-May
3
13-June
13-June
10-June
12-June
8-June
10-June
8-June
4
26-June
26-June
24-June
25-June
24-June
24-June
24-June
5
8-July
8-July
8-July
9-July
6-July
8-July
6-July
6
24-July
24-July
22-July
23-July
19-July
22-July
19-July
7
7-Aug
7-Aug
6-Aug
6-Aug
3-Aug
6-Aug
3-Aug
8
21-Aug
21-Aug
20-Aug
20-Aug
17-Aug
20-Aug
17-Aug
9
4-Sep
3-Sep
3-Sep
4-Sep
31-Aug
3-Sep
31-Aug
10
18-Sep
18-Sep
16-Sep
17-Sep
14-Sep
16-Sep
14-Sep
11
1-Oct
1-Oct
29-Sep
1-Oct
27-Sep
29-Sep
27-Sep
12
16-Oct
16-Oct
14-Oct
15-Oct
15-Oct
14-Oct
15-Oct
13
6-Nov
6-Nov
27-Oct
6-Nov
27-Oct
27-Oct
27-Oct
Table 2. Comparison of whole season means for station depth, epilimnetic mixed depth, total
suspended solids (TSS), light extinction εpar (m-1), Secchi depth disc (m), and chlorophyll (µg l-1).
*
TSS
εpar
Secchi
Chl
1.3
8.0
1.060
bottom
22.9
6.7
6.7
4.5
0.759
2.0
16.4
2003
3.4
3.4
4.4
0.772
1.8
13.4
258
2002
23.5
6.3
4.0
0.727
2.4
15.7
258
2004
23.2
8.0
3.4
0.709
2.4
11.4
908
2003
14.6
5.2
4.7
0.773
1.9
16.4
908
2004
14.6
8.2
3.9
0.758
2.3
11.7
B
2002-2004
5.2
5.2
4.41
0.957
1.9
12.9
HB
2002-2004
12.3
5.1, 5.4, 6.3
3.83
0.749
2.2
12.0
C
2002-2004
31.1
13.0
1.04
0.359
6.0
3.2
Year
Station Depth
Epi Depth
17
2002
1.3
6
2002
WC
Station
Nearshore
Offshore
Bay of Quinte
* for HH nearshore stations and B epi depth = mean bottom depth, for remaining stations epi depth = mean summer epilimnion depth
28
Table 3a. Whole season means for water chemistry including unfiltered phosphorus P-total (µg l-1),
dissolved phosphorus P-flt (µg l-1), particulate phosphorus P-part (µg l-1), soluble reactive phosphorus
SRP (µg l-1), dissolved total nitrogen N-flt (µg l-1), dissolved ammonia NH3 (µg l-1) and dissolved
nitrate/nitrite NO3/NO2 (µg l-1). A dash indicates no samples processed for these parameters.
Station
Year
P-total
P-flt
P-part
SRP
N-flt
NH3
NO3/NO2
17
2002
36.8
14.8
-
1.7
-
279
1673
6
2002
37.9
13.1
-
1.8
-
301
1854
WC
2003
35.8
17.0
28.7
1.4
2649
320
2147
258
2002
30.8
13.2
27.3
2.4
2437
304
1751
258
2004
25.5
11.1
20.5
1.0
2881
310
2115
908
2003
35.5
13.6
31.6
1.5
2684
218
2063
908
2004
28.5
12.6
24.1
1.3
2752
283
1987
B
2002-04
26.5
11.4
22.9
2.1
437
14
56
HB
2002-04
24.4
11.9
20.3
2.3
413
14
74
C
2002-04
11.4
7.6
9.2
1.7
476
12
250
Nearshore
Offshore
Bay of Quinte
Table 3b. Whole season means for water chemistry including dissolved inorganic carbon DIC (mg l-1),
dissolved organic carbon DOC (mg l-1), dissolved chloride CL (mg l-1), dissolved silica SIO2 (mg l-1),
dissolved sulphate SO4 (mg l-1), particulate organic carbon POC (mg l-1) and particulate organic nitrogen
PON (mg l-1). A dash indicates no samples processed for these parameters.
Station
Year
DIC
DOC
CL
SIO2
SO4
POC
PON
17
2002
25.0
5.6
77.7
0.60
49.4
-
-
6
2002
26.0
5.2
73.8
0.80
47.3
-
-
2003
25.7
4.9
-
0.52
-
1.17
0.19
258
2002
26.0
5.0
73.4
0.75
48.0
1.18
0.18
258
2004
26.5
4.8
-
0.74
-
0.95
0.15
908
2003
26.0
4.8
106.4
0.55
53.1
1.25
0.22
908
2004
26.6
4.9
-
0.82
-
0.97
0.16
B
2002-04
24.1
7.6
12.3
3.10
10.8
0.98
0.16
HB
2002-04
23.7
6.8
14.0
2.22
14.2
0.95
0.15
C
2002-04
22.2
4.2
19.3
1.24
21.9
0.28
0.05
Nearshore
WC
Offshore
Bay of Quinte
29
Table 4. Seasonal areal phytoplankton photosynthesis (g C m-2) in Hamilton Harbour 2002-2004.
Location
Year
Empirical Irradiance
Cloudless Irradiance
% Cloudless
17
2002
156
202
77
6
2002
263
378
70
WC
2003
252
343
74
2002
282
398
71
2004
190
290
65
2003
346
470
74
2004
215
324
66
2002
248
312
80
2003
289
386
75
2004
189
282
67
2002
282
358
79
2003
359
469
77
2004
136
200
68
2002
124
154
81
2003
188
245
77
2004
119
165
72
Nearshore
Offshore
258
908
Bay of Quinte
B
HB
C
30
Fig. 1. Bay of Quinte and location of sample stations: Belleville (B), Hay Bay (HB), Conway (C).
31
Station 258 2004
Station 258 2002
mixed depth
0
5
d e p th (m )
de pth (m )
mixed depth
euphotic depth
10
15
20
25
euphotic depth
0
5
10
15
20
25
1 2 3 4 5 6 7 8 9 10 11 12 13
1 2 3 4 5 6 7 8 9 10 11 12 13
cruise
cruise
Station 908 2004
Station 908 2003
mixed depth
mixed depth
euphotic depth
5
10
15
euphotic depth
0
d e p th (m )
d e p th (m )
0
5
10
15
20
20
1 2 3 4 5 6 7 8 9 10 11 12 13
1 2 3 4 5 6 7 8 9 10 11 12 13
cruise
cruise
Station Hay Bay
2002-2004
mixed depth
euphotic depth
d ep th (m )
0
5
10
15
1 2 3 4 5 6 7 8 9 10 11 12 13
cruise
Fig. 2. 2.Mixed
epilimnetic
and euphotic
zone depths
for Hamilton
Harbour
and Hay
in 2002Figure
Mixed
and euphotic
zone depths
for Hamilton
Harbour
and Hay
Bay Bay
2002-2004.
2004.
32
Station 258 mixed depth temperatures
2003
2002
25
HB
20
2004
15
10
5
te m p (°C )
te m p (°C )
Station 908 mixed depth temperatures
25
HB
20
2004
15
10
5
0
0
1 2 3 4 5 6 7 8 9 10 11 12 13
1 2 3 4 5 6 7 8 9 10 11 12 13
cruise
cruise
Fig. 3. Mixed epilimnetic temperatures for station 258 (2002, 2004) and station 908 (2003, 2004)
compared to that at Hay Bay.
-1
-1
Station 258 depth of DO ≤ 4 mg l
Station 908 depth of DO ≤ 4 mg l
2003
2004
0
5
10
15
20
25
2004
0
depth (m )
d ep th (m )
2002
5
10
15
1 2 3 4 5 6 7 8 9 10 11 12 13
1 2 3 4 5 6 7 8 9 10 11 12 13
cruise
cruise
-1
-1
Station 258 depth of DO ≤ 1 mg l
Station 908 depth of DO ≤ 1 mg l
2003
0
5
10
15
20
25
2004
0
2004
depth (m )
d ep th (m )
2002
5
10
15
1 2 3 4 5 6 7 8 9 10 11 12 13
1 2 3 4 5 6 7 8 9 10 11 12 13
cruise
cruise
Fig. 4. Initial depth where low dissolved oxygen concentrations of 4 or 1 mg l-1 were recorded
in Hamilton Harbour 2002 – 2004.
33
TTemp
e m p (°C)
(°C )
00 .4
5
10
15
Depth (m)
00.4
3
D e p th
(m )
25
M ay 29
15
20
DO
T
Jun 12
7
DO
13
T
15
0
5
0
5
T1 e0 m p ( 1° C5- 1 )
D O1 0 ( m g 1l5 )
20
25
20
25
Jul 9
2 3 .5
0 .4
0
5
0
5
6 .1
T1 e0 m p (1° C5 )
D O1 0 ( m g 1l -51 )
20
25
0 .5
Jul 23
2 .2
D e p th
(m )
223
0 .5
0
5
0
5
10
15
T e m p ( ° C- 1)
D 1O0 ( m g 1 l5 )
A ug 20
2
4 .2
6
7 .9
D e p th
(m )
8 .9
1 1 .1
DO
1 4 .9
0
5
0
5
T
10
15
T e m p (°C )
D O1 0 ( m g 1l -51 )
1 9 .9
2 2 .2
2 0 .7
25
0
5
20
25
0
5
0 .4
S ep 4
2
3 .9
5 .8
7 .8
D e p th 10
( m ) 1 2 .2
14
1 5 .8
DO
1 7 .1
19
23
20
25
0
5
25
Oct 1
0
5
T1e0m p ( °1C5 )
D 1O0 ( m g 1l5- 1 )
25
0
5
25
0
5
14
DO
T
D e p th
(m )
5
1DO
0 (m1 g5 l-1) 2 0
D O ( m g l -1 )
25
25
20
25
20
25
1DO
0 (m1 g
5 l-1) 2 0
25
T
T1 e0 m p ( 1° C5 )
D 1O0 ( m g 1l -51 )
S ep 17
10
1 1 .9
1 4 .1
T
DO
T
1 6 .1
2 0 .9
10
15
T e m p (°C )
D O1 0 ( m g l1- 15 )
20
25
0
5
20
25
0
5
0 .3
10
15
T e m p (°C )
D 1O0 ( m g 1l 5- 1 )
N ov 6
4 .3
D e p th
(m )
8 .7
1 2 .2
DO
6 .5
1 0 .1
1 4 .6
T
DO
T
1 9 .2
22
0
25
20
15
8
2 0 .3
22
23
20
10
T e m p (°C )
D 1O0 ( m g 1l 5- 1 )
1 .9
1 8 .6
18
25
T
3 .9
1 6 .1
1 6 .1
0 .4
2
7 .7
D e p th
(m )
5
20
6 .1
D e p t h 1 2 .1
(m )
5
20
4 .3
9 .9
25
20
DO
DO
O ct 15
2 .2
7
20
Jun 25
A ug 6
1 2 .1
T
20
20
25
8 .1
1 5 .7
0 .4
2 .9
DO
1 6 .9
1 6 .2
20
9 .9
14
T
15
6
1 2 .5
DO
10
4 .5
D e p t h 1 0 .8
(m )
10
5
0
2 .1
8 .2
8 .1
0
25
0 .2
1 .3
2 .7
4 .5
7 .9
D e p t h 8 .9
( m ) 1 1 .9
1 4 .5
2 0 .1
2 0 .9
2 2 .1
23
0
20
25
9
1 4 .4
Depth (m)
10
5
D e p th
(m )
7 .1
00. 4
5
3 .2
5 .1
1 9 .1
23
2 3 .4
0
1
12
Depth (m)
0 .4
5
00.3
T e mTemp
p ( ° C(°C)
)
T Temp
e m p ((°C)
°C )
20
Depth (m)
1
3
5
4
6
.9
D e p th
(m )
9
1 1 .5
1 5 .1
1 9 .2
2 2 .2
223
3 .8
0
0
5
-1
1 DO
0 (m1 5g l ) 2 0
-1
D O (m g l )
2 2 .1
25
0
5
D O ( m g l -1 )
Fig. 5. Temperature and dissolved O2 concentration profiles versus depth for station 258 in 2002.
34
Temp
T e m p(°C)
(°C )
00 .3
0
5
10
T eTemp
m p ( °(°C)
C)
Temp
T e m p (°C)
(°C )
15
20
25
0
0 .1
M ay 13
5
10
15
20
25
Depth (m)
3 .9
6 .9
T
DO
D e p th
(m )
7 .1
25
20
25
20
25
6
7 .9
0
5
0
5
T1e0m p ( °1C5 )
D 1O0 ( m g 1l 5- 1 )
20
25
20
25
Jul 6
00.5
T
DO
0
5
0
5
T1 e0 m p ( °1C5)
D 1O0 ( m g 1l 5- 1 )
20
20
1 4 .1
14
0
5
0
5
0 .4
0 .8
1
2
2 .6
3 .5
4 .5
D e p t h 5 .5
6 .6
(m )
7 .5
8 .6
9 .6
1 0 .4
1 1 .5
1 2 .5
14
0
25
25
0 .5
0
A ug 17
1 .2
T1 e0m p ( °1C5 )
D1O0 ( m g 1l5- 1 )
25
0
5
20
25
0
5
10
15
T e m p (°C )
D 1O0 ( m g 1l -51 )
A ug 3
3
5
D e p t h 7 .1
(m )
9
T
DO
T
DO
11
13
5
5
T1 e0 m p ( 1° C5 )
D 1O0 ( m g 1l -51 )
20
20
25
25
A ug 31
0 .5
0
5
0
5
T1 e0 m p ( °1C5 )
D1O0 ( m g 1l -51 )
20
25
20
25
20
25
20
25
20
25
S ep 14
1 .9
3 .6
5 .1
4
7 .2
5 .1
D e p th
(m )
5 .9
6 .7
9
D e p th
( m ) 1 0 .1
8 .1
1 0 .9
8
10
T
DO
10
DO
1 1 .6
T
14
14
0
5
0
5
T1 e0 m p ( °1C5 )
D1 O0 ( m g 1l -51 )
20
25
20
25
1 4 .1
0 .5
14
0
5
0
5
S ep 27
10
15
T e m p (°C )
D1 O0 ( m g 1l5- 1 )
20
25
20
25
1 4 .6
0 .6
1 .9
6 .2
5
4
D e p th
(m )
9 .1
7
D e p th
(m )
DO
10
T
13
14
DO
10
15
D O ( m g l -1 )
DO (m g l-1)
20
25
6 .1
1 0 .2
T
DO
T
1 2 .7
12
5
T1e0m p ( °1C5)
D1 O0 ( m g 1l5- 1 )
8
8
0
5
5
O ct 27
3
1 1 .1
0
0
O ct 12
3
7 .2
T
DO
1 2 .4
1 2 .4
12
1 0.2
1
Jul 19
3 .2
D e p th
(m )
T
20
2 .1
5
DO
1 2 .1
1 2 .1
Depth (m)
00.4
2
3 .2
4 .4
5
5 .9
7 .2
7 .9
D e p t h 8 .6
9
(m )
9 .4
9 .9
10
1 0 .3
1 0 .9
11
1 1 .4
114
2 .4
10
T
DO
1 1 .2
1 2 .1
14
Depth (m)
20
9 .3
8 .2
Depth (m)
15
4
6
D e p th
(m )
10
Jun 24
1 .9
3 .9
D e p th
(m )
0 .3
5
M ay 26
1
2
D e p th
(m )
0
0
5
10
15
DDO
O ( m(mg gl - 1l)-1)
20
25
0
5
10
15
DDO
O (m
g l -1 )
(m g l-1)
Fig. 6. Temperature and dissolved O2 concentration profiles versus depth for station 908 in 2004.
35
Total Phosphorus - 258
Soluble Reactive Phosphorus - 258
8
70
2002
60
2004
2002
6
50
4
40
initial
30
2
20
final
0
10
1
2
3
4
5
6
7
8
9
10
11
12
1
13
2
3
4
5
6
7
8
9
10
cruise
cruise
Total Phosphorus - 908
Soluble Reactive Phosphorus - 908
80
11
12
13
14
8
70
2003
60
2004
2003
2004
6
u g /l
u g /l
50
4
40
initial
30
2
20
final
0
10
1
2
3
4
5
6
7
8
9
10
11
12
1
13
2
3
4
5
6
7
8
9
10
11
cruise
cruise
Total Phosphorus - Nearshore
Soluble Reactive Phosphorus - Nearshore
80
70
12
6-2002
14
17-2002
6
WC-2003
60
13
8
17-2002
50
6-2002
WC-2003
u g /l
u g /l
2004
u g /l
u g /l
80
4
40
initial
30
20
final
2
0
1
10
1
2
3
4
5
6
7 8
cruise
9
10
11
12
13
2
3
4
5
6
7
8
9
10
11
12
13
14
cruise
Fig. 7. Total phosphorus and soluble reactive phosphorus concentrations for Hamilton
Harbour 2002 – 2004 dark lines indicating RAP initial and final targets for total phosphorus.
36
Nitrate/Nitrite - 258
Ammonia - 258
3500
1200
2002
2002
1000
2004
2004
2500
u g /l
800
u g /l
3000
2000
600
400
1500
200
1000
0
500
1
2
3
4
5
6
7
8
9
10
1
11 12 13
2
3
4
5
6
8
9
10
11
13
Nitrate/Nitrite - 908
Ammonia - 908
1200
1000
2003
3500
2004
3000
2003
2004
2500
u g /l
800
2000
600
400
1500
200
1000
0
500
1
2
3
4
5
6
7
8
9
10 11
1
12 13
2
3
4
5
6
7
8
9
10
11
12
13
cruise
cruise
Am monia - Nearshore
1200
1000
Nitrate/Nitrite - Nearshore
17-2002
3500
6-2002
3000
17-2002
6-2002
WC-2003
800
2500
WC-2003
u g /l
ug /l
12
cruise
cruise
u g /l
7
600
2000
400
1500
200
1000
0
500
1
2
3
4
5
6
7
8
cruise
9 10 11 12 13 14
1
2
3
4
5
6
7
8
9
10
11
12
13
cruise
Fig. 8. Dissolved ammonia and dissolved nitrate/nitrites concentrations for Hamilton Harbour
2002 – 2004.
37
Silica
258
Chlorophyll
258
2.0
50
1.5
40
2004
0.5
30
2002
20
2004
u g /l
u g /l
2002
1.0
10
0.0
0
1
2
3
4
5
6
7
8
9
10 11 12 13
1
2
3
4
5
6
cruise
7
8
9 10 11 12 13
cruise
Silica
908
Chlorophyll
908
2.0
50
1.5
40
2004
0.5
30
2003
20
2004
u g /l
u g /l
2003
1.0
10
0.0
0
1
2
3
4
5
6
7
8
9
10 11 12 13 14
1
3
4
5
6
7
8
9 10 11 12 13
cruise
Silica
Nearshore
Chlorophyll
Nearshore
2.0
50
17-2002
17-2002
40
6-2002
1.5
2
cruise
6-2002
WC-2003
WC-2003
u g /l
u g /l
30
1.0
20
0.5
10
0.0
0
1
2
3
4
5
6
7
cruise
8
9
10
11
12
13
1
2
3
4
5
6
7
8
9
10 11 12 13
cruise
Fig. 9. Soluble reactive silica and chlorophyll concentrations for Hamilton Harbour 2002 – 2004.
38
Nitrate/Nitrite-seasonal means
Total phosphorus-seasonal means
2500
45
40
35
initial
30
25
20
15
2000
spring
summer
1500
fall
final
spring
summer
10
5
0
fall
258- 2582002 2004
908- 9082003 2004
1000
500
0
6- 17- WC2002 2002 2003
258- 2582002 2004
908- 9082003 2004
6- 17- WC2002 2002 2003
ammonia-seasonal means
soluble reactive phosphorus-seasonal means
4
1200
1000
3
spring
summer
2
fall
spring
800
summer
600
fall
400
1
200
0
0
258- 2582002 2004
908- 9082003 2004
258- 2582002 2004
6- 17- WC2002 2002 2003
Silica-seasonal means
908- 9082003 2004
6- 17- WC2002 2002 2003
unionized ammonia-seasonal means
1.5
80
spring
1.0
summer
fall
60
summer
fall
µ g l- 1
spring
40
0.5
target
20
0
0.0
258- 2582002 2004
908- 9082003 2004
6- 17- WC2002 2002 2003
258- 2582002 2004
908- 9082003 2004
6- 17- WC2002 2002 2003
Fig. 10. Seasonal nutrient concentrations (µg l-1) for Hamilton Harbour 2002 – 2004. Dark lines
indicate RAP initial and final targets.
39
Total Phosphorus (TP) versus Chlorophyll 2002-2004
whole season mean Chl (ug l -1)
25
20
15
10
HH
QUINTE
5
0
10
15
20
25
30
35
40
35
40
-1
whole season mean TP (ug l )
Total Phosphorus versus SAPP 2002-2004
400
350
SAPP (g C m-2)
300
250
200
150
100
HH
50
QUINTE
0
10
15
20
25
30
whole season mean TP (µg l-1)
Fig. 11. Whole season mean total phosphorus versus chlorophyll a concentrations and SAPP
for Hamilton Harbour and the Bay of Quinte (B, HB, C) for 2002 – 2004.
40
Total suspended solids-seasonal means
10.0
mg l-1
8.0
spring
summer
6.0
f all
4.0
2.0
0.0
2582002
2582004
9082003
9082004
62002
172002
WC2003
Se asonal mean Secchi disc depths
m
258- 258- 908- 90802
04
03
04
6
17
WC
B
HB
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
4.0
C
spring
summer
f all
w hole
initial
f inal
6m
Seasonal mean epar
spring
summer
fall
whole
1.4
1.2
m-1
1.0
0.8
0.6
0.4
0.2
0.0
258-02 258-04 908-03 908-04
6
17
WC
B
HB
C
Fig. 12. Seasonal mean total suspended solids, Secchi disc depths and εpar for Hamilton Harbour
2002 – 2004 and the Bay of Quinte. Dark lines indicate RAP initial and final targets.
41
Seasonal mean chlorophyll
35
spring
30
summer
µg l-1
25
fall
20
whole
15
10
5
0
258-02 258-04 908-03 908-04
6
17
WC
B
HB
C
g C m -2
SAPP
400
350
300
250
200
150
100
50
0
258- 258- 908- 90802 04 03 04
6
17
WC
B-02 B-03 B-04 HB- HB- HB02 03 04
Fig. 13. Seasonal mean chlorophyll concentrations and average areal phytoplankton photosynthesis
(SAPP) for Hamilton Harbour, 2002 – 2004.
42
Secchi Disc depths
258
cruise 1
0
2
3
4
5
6
7
8
9
10
11
12
13
2002
1
2004
2
m
initial
final
3
4
5
Secchi Disc depths
908
cruise
0
1
2
3
4
5
6
7
8
9
10
11
12
13
2003
1
2004
initial
m
2
final
3
4
5
cruise1
Secchi Disc depths
Nearshore
2
3
4
5
6
7
8
9
10 11 12 13
0
6-2002
WC-2003
1
m
2
3
initial
final
4
5
Fig. 14. Secchi disc depths for Hamilton Harbour 2002 – 2004. Dark lines indicate RAP initial
and final targets.
43
AN INTEGRATED ASSESSMENT OF THE MICROBIAL AND
PLANKTONIC COMMUNITIES OF HAMILTON HARBOUR
Mohiuddin Munawar & Mark Fitzpatrick
Great Lakes Laboratory for Fisheries and Aquatic Sciences
Fisheries and Oceans Canada
867 Lakeshore Road, P.O. Box 5050
Burlington, Ontario, L7R 4A6
INTRODUCTION
Studies of the microbial and planktonic food webs of aquatic ecosystems have proven to be
important in increasing our understanding of ecosystem structure and function. Taken together,
the microbial and planktonic communities form the pelagic component of the lower food web
and play an important role in transferring autochthonous energy to higher trophic levels (e.g.
Munawar and Weisse 1989, Munawar et al. 2005, Fitzpatrick et al. 2007). Thus, healthy
fisheries and healthy ecosystems are dependant upon the relative health of the lower food web.
With respect to Hamilton Harbour, adverse health effects termed “Beneficial Use
Impairments”, have been known to persist for some time and were responsible for it’s
designation as an “Area of Concern” (International Joint Commission (IJC) 1989). These
Beneficial Use Impairments (BUIs) included “water quality and bacterial contamination” and
“fish and wildlife impacts” as a result of municipal and industrial wastes (Environment Canada
2005). Fisheries & Oceans Canada undertook a temporally extensive (May – October) survey of
the bay from 2002–2004 in order to provide a baseline assessment of the health of the microbial
and planktonic food webs.
The structure of the Hamilton Harbour food web was assessed, including bacteria,
autotrophic picoplankton, phytoplankton, heterotrophic nanoflagellates and ciliates, using
standard techniques. Size fractionated primary productivity was included as a functional
measurement. Furthermore, a preliminary analysis of phytoplankton and zooplankton
interactions was undertaken to provide insights into predator – prey interactions and energy
transfer at the lowest trophic levels. The results of this study will provide scientists and
managers with important information on this highly complex and stressed ecosystem.
44
MATERIALS AND METHODS
Two stations within Hamilton Harbour were sampled for this study and are shown in Figure 1.
Station 258 was sampled from May to October 2002 and 2004, and station 908 was sampled
from May to October 2003 and 2004. Microbial loop samples (bacteria, autotrophic
picoplankton and heterotrophic nanoflagellates) were fixed with 1.6% formaldehyde and
enumerated using DAPI staining (Porter and Feig 1980) under epi-fluorescence microscopy
(Munawar and Weisse 1989). Freshweight biomass was estimated as 0.91 pg cell-1 for bacteria,
1.82 pg cell-1 for autotrophic picoplankton and 127 pg cell-1 for heterotrophic nanoflagellates
(Sprules et al. 1999). Ciliate samples were preserved with Lugol’s iodine and enumerated
following the Quantitative Protargol Staining technique (Montagnes and Lynn 1987).
Phytoplankton samples were fixed immediately with Lugol’s iodine. Identification and
enumeration followed the Utermöhl (1958) inverted microscope technique (see Munawar et al.
1987). Zooplankton data from Gerlofsma et al. (this volume) was also incorporated in this
analysis.
Size fractionated primary productivity was determined for three size categories of
phytoplankton (<2 μm, 2-20 μm and >20 μm) following the standard protocol of Munawar and
Munawar (1996). Whole water samples were spiked with Na14CO3, incubated for 2 - 4 hours at
surface temperatures and irradiance levels equivalent to Popt (Fee 1969). After incubation, size
classes were determined via filtration of the sample water through polycarbonate filters and
radioactivity was determined by liquid scintillation counting.
RESULTS AND DISCUSSION
Hamilton Harbour is a highly variable physical, chemical and biological system. As with other
areas in the Great Lakes, phosphorus abatement was introduced in the 1970s in order to control
eutrophication and ultimately improve the health of the ecosystem. Target total phosphorus
concentrations of 34 μg l-1 (interim) and 17 μg l-1 were set under the Great Lakes Water Quality
Agreement. Previous reports have indicated that the interim target total phosphorus
concentration was met by the late 1980s (e.g. Charlton and Le Sage 1996), however further
reductions are still to be realized (see Burley, this volume). Fisheries & Oceans Canada
implemented a holistic monitoring program of Hamilton Harbour in 2002 in order to assess the
current health of the lower trophic levels.
Size Fractionated Primary Production
Size fractionated primary productivity for 2002 – 2004 at both stations is summarized in Figures
2 and 3. At station 258, net plankton (>20 μm) productivity (SWM) declined from 33.5 – 13.5
mg C m-3 h-1 between 2002 and 2004. Nanoplankton (2 – 20 μm) productivity declined very
slightly from 18.7 – 17.1 mg C m-3 h-1 during the same time frame, and picoplankton (<2 μm)
45
decreased from 22.3 – 5.4 mg C m-3 h-1. SWM primary productivity also revealed a decreasing
trend at station 908 from 2003 to 2004. Net plankton productivity fell slightly from 15.0 – 12.9
mg C m-3 h-1, net plankton from 38.1 – 22.8 mg C m-3 h-1 and picoplankton from 10.7 – 5.7 mg C
m-3 h-1. Primary productivity was high and dominated by larger net (>20 μm) and nano (2-20
μm) plankton which is generally characteristic of a eutrophic system.
Microbial Loop
The seasonal distribution of bacteria, autotrophic picoplankton, heterotrophic nanoflagellates and
ciliates for stations 258 and 908 are shown in Figures 4 and 5, respectively. The seasonally
weighted mean (SWM) bacterial biomass at 258 declined from 451.5 mg m-3 in 2002 to
418.6 mg m-3 in 2004. Similar trends were apparent in picoplankton, declining from 98.9 to 15.4
mg m-3, heterotrophic nanoflagellates from 155.6 to 104.9 mg m-3, and ciliates from 89.6 to 65.1
mg m-3 between years. At station 908, SWM microbial loop biomass also declined between
years, with bacteria falling from 545.7 to 452.1 mg m-3, picoplankton from 78.6 to 29.6 mg m-3,
nanoflagellates from 212.9 – 166.1 mg m-3 and ciliates from 93.2 – 62.8 mg m-3. The microbial
loop was dominated by bacteria at both sites, with seasonally weighted mean biomass ranging
from 420 – 550 mg m-3, and was typically 4-5 times greater than each of the other components.
Phytoplankton Biomass and Composition
The seasonal distribution of the phytoplankton community including biomass and % composition
(by biomass) is shown in Figures 6-9. At station 258, SWM biomass declined from 2034.6 mg
m-3 in 2002 to 1819.3 mg m-3 in 2004. Dominant taxonomic groups (>50% of biomass)
included: Diatomeae, Chlorophyta, Cryptophyceae, and Dinophyceae during 2002. During
2004, the dominant phytoplankton groups included Chrysophyceae, Chlorophyta and
Dinophyceae. At station 908, SWM phytoplankton biomass showed a very small increase from
1448.9 to 1477.1 mg m-3 from 2003 to 2004. Dominant phytoplankton groups throughout 2003
included Diatomeae, Cryptophyceae, Chlorophyta and Dinophyceae. In 2004, dominant taxa
included Chrysophyceae, Diatomeae, Chlorophyta and Dinophyceae. A detailed listing of
species contributing greater than 5% to the total biomass of any sample is given in Table 1 for
station 258 and Table 2 for station 908.
Phytoplankton biomass was quite high in the harbour (SWM: 1450 – 2040 mg m-3)
indicating mesotrophic conditions based on the trophic index of Munawar and Munawar (1982).
High biological variability was evident in the phytoplankton community at both sites where
Diatomeae, Dinophyceae, Chlorophyta and Chrysophyceae were all dominant at different times
throughout each year. The rapidly changing phytoplankton community might be a consequence
of the physical disturbances (e.g. wind, cargo ships) that Hamilton Harbour is subject to. These
physical disturbances have previously been shown to limit the overall size of standing crop
(Haffner et al. 1980) and could be expected to have a bottom-up effect on the zooplankton
community. A detailed discussion of phytoplankton and zooplankton interactions follows.
46
TROPHIC INTERACTIONS IN HAMILTON HARBOUR
The extreme variability observed in the phytoplankton community of Hamilton Harbour, both in
terms of biomass and composition could be expected to have a bottom-up influence on the
zooplankton community. We therefore consider the interactions of zooplankton and
phytoplankton in an attempt to gain some understanding of predator-prey interactions at the base
of the food web. This is a preliminary analysis of the phytoplankton – zooplankton relationship
styled after Munawar et al. 2007. Other factors which could influence this relationship, such as
nutrient dynamics affecting phytoplankton growth and planktivore predation on zooplankton are
not directly considered here, but are important nonetheless.
Station 258, 2002
Spring
Zooplankton biomass was initially low during spring of 2002 (≈ 0.1-0.6 g m-3 fresh weight)
before surging to 2.8 g m-3 as part of an upward trajectory that would continue into the summer
(Fig. 6). The zooplankton community was initially dominated by Cyclopoids, but rapidly shifted
towards Cladocera. Phytoplankton biomass was initially high (3.2 g m-3) but declined to
0.9 g m-3 by late spring. The composition of the phytoplankton community also showed a rapid
shift from Diatomeae to Chlorophyta in this period. Interestingly, almost 50% of the
zooplankton biomass was carnivorous during the spring, which would normally be associated
with low biomass, however this proportion held even as biomass increased. The phytoplankton
community was then dominated by largely edible species of Stephanodiscus, Scenedesmus and
Coelastrum, which likely helped sustain the herbivorous zooplankton and would in turn provide
food for the carnivorous zooplankton.
Summer
Zooplankton biomass showed a bimodal pattern during summer, soaring to 10 g m-3 then falling
to 2 g m-3 and rising again to 10 g m-3 before dampening out in the 2 – 3 g m-3 range (Fig. 6).
Cladocera dominated the zooplankton biomass until the late summer when Cyclopoids became
more prevalent. A peak of dreissenid veligers was also observed in mid summer. Phytoplankton
biomass showed considerable variability throughout the summer and peaked at 4.1 g m-3 in early
September. Phytoplankton composition was quite variable with Chlorophyta, Cryptophyceae
and Dinophyceae overwhelmingly dominating the biomass at different periods throughout the
season. The proportion of carnivorous zooplankton decreased to 20 – 30% of the biomass as
largely edible species of phytoplankton dominated the biomass including Cryptomonas reflexa,
Oocystis lacustris, Scenedesmus braziliensi, and Coelastrum reticulum, although the inedible
Ceratium furcoides was prevalent in late summer. The large biomass of herbivorous
zooplankton observed throughout the summer may have helped create conditions for the
observed dinoflagellate bloom by reducing the standing crop of edible algae.
47
Fall
Zooplankton biomass declined in the fall from its late summer peak of 3.4 g m-3 to 0.6 g m-3 (Fig.
6). Zooplankton composition was almost equally split between Cyclopoids and Cladocera, along
with a smaller component of Calanoids. Phytoplankton biomass also declined during fall from a
late summer peak of 4.1 g m-3 to 0.5 g m-3 and was tightly coupled with zooplankton biomass.
Fall biomass was overwhelmingly dominated by species of Chlorophyta including Coenochloris
pyrenoidosa and Coelastrum reticulatum. The close relationship between phytoplankton and
zooplankton biomass observed during fall may be a result of the large proportion of herbivorous
zooplankton observed and the correspondingly large population of edible algae.
Station 258, 2004
Spring
Zooplankton biomass at station 258 increased during the spring of 2004 from 0.7 to 2.9 g m-3 as
the composition shifted from Cyclopoids to Cladocera (Fig. 7). Phytoplankton biomass
increased from 0.8 g m-3 to 2.2 g m-3 but then began to wane in the late spring and early summer.
Phytoplankton was composed mainly of Cryptophyceae followed by Diatomeae during this
period. Slightly more than half of the zooplankton community was herbivorous and the
phytoplankton community contained mostly edible species including Rhodomonas minuta,
Cryptomonas reflexa and Stephanodiscus niagarae.
Summer
Zooplankton reached its first maxima in early summer of 4.1 g m-3 before declining to 1.7 g m-3
in mid summer and increasing to 3.5 g m-3 by late summer (Fig. 7). Zooplankton communities in
early and late summer were dominated by Cladocera while a mid summer peak of Cyclopoids
was observed. Phytoplankton biomass showed an increasing trend peaking at 3.6 g m-3 in mid
summer before declining somewhat to 2.8 g m-3 in late summer. Phytoplankton was composed
of Chlorophyta in the early to mid summer period, while Cryptophyceae and Dinophyceae
became more prevalent in the late summer period. The proportion of herbivorous zooplankton
peaked in mid July at 80% of the biomass but generally ranged from 50-60% of the total
biomass. Phytoplankton contained some edible species including Cryptomonas reflexa and
Lagerheimia ciliate but also inedible species including Dinobryon divergens and Ceratium
furcoides.
Fall
Zooplankton biomass peaked in the early fall at 3.5 g m-3 and then declined to 0.4 g m-3.
Cladocera dominated the early fall period but then Cyclopoids became more prevalent (Fig. 7).
Phytoplankton biomass continued its downward trend falling from 2.1 to 0.6 g m-3. The
phytoplankton community was dominated by Dinophyceae in early fall but the late fall contained
48
a mixture of Chlorophyta, Diatomeae and Cryptophyceae. The proportion of herbivorous
zooplankton declined from 90% to 70% of the biomass and phytoplankton consisted of inedible
Ceratium furcoides in early fall but edible algae, particularly Stephanodiscus niagarae, was
prevalent later in the fall.
Station 908, 2003
Spring
Zooplankton biomass at station 908 was high, though variable during spring, ranging from 2.4 –
3.8 g m-3 and was dominated by Calanoids, with a significant amount of Cladocera being
observed (Fig. 8). Phytoplankton biomass was initially quite high (≈ 2.0 g m-3)but dropped to
0.3 g m-3 in late spring. Diatomeae followed by Cryptophyceae were the dominant
phytoplankton. The zooplankton community was almost evenly split between herbivores and
carnivores, and the phytoplankton community contained inedible (Fragilaria crotenensis) and
edible (Rhodomonas minuta) forms of algae.
Summer
Zooplankton biomass displayed a bimodal pattern with an early summer peak of 5.4 g m-3
declining to 1.2 g m-3 and rising to its secondary peak of 3.3 g m-3 and continued to be variable
through to the end of the summer (Fig. 8). Zooplankton was dominated by Cladocera throughout
the summer although a significant amount of dreissenid veligers were observed mid summer and
Cyclopoids became more prevalent in late summer. Phytoplankton biomass also showed a
bimodal distribution declining from a maximum of 2.3 g m-3 in early summer to a minimum of
0.9 g m-3 in mid summer and increasing to its peak of 2.8 in late summer before declining again
into fall. Phytoplankton composition was highly variable, proceeding from Cryptophyceae to
Chlorophyta and then to Dinophyceae and Cyanophyta. The proportion of carnivorous
zooplankton ranged from 25 – 50% of the zooplankton biomass in the summer. Potential
phytoplankton prey for the zooplankton contained edible species including Rhodomonas minuta,
Cryptomonas reflexa and Coelastrum pseudomicrosporum, although the inedible Ceratium
furcoides became prominent in late summer.
Fall
Zooplankton biomass at Stn 908 declined in the fall from 0.7 g m-3 to 0.4 g m-3 and was
composed of a mixture of Cyclopoids and Cladocerans although some Calanoids were also
present (Fig. 8). Phytoplankton biomass was very similar to zooplankton biomass in this period
and followed the same trend declining from 0.5 – 0.1 g m-3. Phytoplankton composition was
quite variable and no single group was dominant. The proportion of predator biomass that was
carnivorous ranged from 25 - 40% and the potential prey was composed of edible species of
Cryptomonas (C. reflexa; C. restriformis), but also various inedible forms of Cyanophyta as well
as Ceratium furcoides.
49
Station 908, 2004
Spring
Zooplankton biomass displayed an increasing trend during the spring of 2004 rising from 0.6 –
2.3 g m-3 and Cladocera dominated the zooplankton community (Fig. 9). Phytoplankton biomass
was very similar to zooplankton biomass, increasing from 0.5 – 2.1 and displaying the same
upward trend. Phytoplankton was comprised of Cryptophyta and Diatomeae. During spring
predators were split almost evenly between herbivores and carnivores and the prey contained
edible species including Rhodomonas minuta and Stephanodiscus niagarae.
Summer
Zooplankton biomass displayed a bimodal pattern during summer 2004, soaring to a peak of
5.3 g m-3 in early summer, and then declining to 1.2 g m-3 by mid summer before rising to a
second peak of 5.3 g m-3 in late summer (Fig. 9). Cladocerans dominated the zooplankton
biomass from early to mid summer, while Cyclopoids became more prevalent during late
summer. Phytoplankton biomass reached its lowest level of 0.8 g m-3 in early summer as
zooplankton biomass peaked, but gradually increased over the mid summer period to a high of
2.9 g m-3 as zooplankton levels declined. Phytoplankton biomass declined again in late summer
to 0.8 g m-3 as zooplankton biomass increased. During the early part of the summer,
phytoplankton was comprised of a mixture of Chlorophyta and Diatomeae followed by
Chlorophyta in the mid summer and Dinophyceae in the late summer. The ratio of carnivorous
to herbivorous zooplankton was highly variable during the summer, with 45% of the zooplankton
being carnivorous in the early summer, followed by a sharp decline to 8% in mid summer and
ranging from 20 – 30% throughout the rest of the summer. The rapid shifts in zooplankton
feeding are associated with rapid shifts in phytoplankton composition. Edible phytoplankton
species included Stephanodiscus niagarae in the early summer, Coelastrum pseudmicroporum
and Pediastrum boryanum in mid summer, and Cryptomonas reflexa in late summer. However,
a significant proportion of inedible species including Dinobryon divergens and Ceratium
furcoides were present throughout the summer.
Fall
Zooplankton biomass was still high in the early fall, 4.3 g m-3, but declined rapidly to less than
0.5 g m-3 for the remainder of the sampling season (Fig. 9). Cladocera dominated the
zooplankton biomass in the early fall, but was later replaced by a mixture of Cyclopoids and
calanoids. Phytoplankton biomass increased from 0.8 – 2.0 g m-3 at the end of the sampling
season and composition shifted from Dinophyceae to Cryptophyceae. Approximately 70 – 80%
of the zooplankton was herbivorous during the fall, and the potential prey shifted from largely
inedible Ceratium furcoides in early fall to edible Cryptomonas reflexa.
50
SUMMARY AND CONCLUSION
Some interesting observations on the lower food web of Hamilton Harbour were made during
our study. The first is that phytoplankton biomass was lower than expected, given that the
harbour is classified as eutrophic. While other factors, including high total phosphorus
concentrations and extended periods of hypoxia in the lower thermal stratum (Burley, this
volume) were indicative of eutrophic conditions, phytoplankton biomass was indicative of
mesotrophic conditions. The reduced biomass suggests that phytoplankton is responding to
reductions in phosphorus loadings. However, it also raises the question of whether or not this
response is enough to alleviate eutrophication in the harbour.
The second observation was that phytoplankton composition was highly variable.
Chlorophyta, Diatomeae, Cryptophyta and Dinophyceae were all dominant groups and different
times throughout the sampling season and their respective dominance did not always follow
expected seasonal patterns. In part, this may be due to the extreme physical disturbances,
including ship traffic and wind and storm events that the harbour is constantly subject to. The
constant change observed in species composition would be expected to exert some bottom-up
pressure on potential predators including heterotrophic nanoflagellates, ciliates and especially
zooplankton.
The third observation is that zooplankton biomass is typically higher than phytoplankton
biomass. We would expect the amount of zooplankton predators to be restricted by the
availability of phytoplankton prey. Our findings suggest, however, that there may be strong top
down control of phytoplankton by zooplankton. We therefore need to consider the feeding
ecology of zooplankton which leads to our final observation.
The fourth observation is that the proportion of carnivorous zooplankton to herbivorous
zooplankton was relatively high, typically accounting for 25 – 50% of the zooplankton biomass.
What these findings suggest is that the standing crop of phytoplankton was sufficient to support a
large biomass of herbivorous zooplankton and this in turn was sufficient to support a large
biomass of carnivorous zooplankton. Other components of the microbial food web including
ciliates and heterotrophic nanoflagellates would also be expected to provide an additional food
resource for carnivorous zooplankton. However, it is less than clear how the energy
requirements of the zooplankton community is being met and whether or not planktivory is
limiting the size of the zooplankton community.
Future research needs to be directed towards understanding energy transfer within the
complete microbial and planktonic food web. The high proportion of secondary to primary
producers observed in Hamilton Harbour suggests that autochthonous production may not likely
be sufficient to sustain the food web. However, other sources of autochthonous energy including
benthic algae and macrophytes need to be considered as do allochthonous sources of energy.
Furthermore, more integrative research needs to be directed to understanding how the available
energy is utilized by the fishery.
51
Hamilton Harbour continues to be a mesotrophic - eutrophic ecosystem subject to strong
physical disturbances. The highly variable nature of the ecosystem was evident in our bi-weekly
monitoring of phytoplankton, microbial loop and primary productivity as well as our holistic
assessment of the planktonic food web. More intense and continuous sampling strategies need to
be deployed to capture short term fluctuations and understand their implications. Continuous
monitoring could be achieved with deployment of new technologies including Fluoroprobe,
FlowCAM and the Laser Optical Plankton Counter. Energy flow in Hamilton Harbour is driven
by primary producers, and research is needed into the transfer of energy from lower to higher
trophic levels to understand how fisheries are affected and sustained.
ACKNOWLEDGEMENTS
We wish to thank H. Niblock, S. Carou and J. Gerlofsma who assisted with sample collection,
processing and data analyses. Phytoplankton was identified and enumerated by H. Kling, ciliates
by D. Lynn and microbial loop by T. Weisse and L. Logasundaram.
52
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54
Table 1. A list of phytoplankton species contributing 5% or more to total biomass of any
sample from Stn 258, Hamilton Harbour.
2002
2004
Cyanophyta
Aphanizomenon flos-aquae
Aphanizomenon sp
Chroococcus limneticus
Lyngbya birgei
Microcystis viridis
Cyanophyta
Lyngbya birgei
Microcystis aeruginosa
Pseudanabaena mucioli
Chlorophyta
Closteriopsis longissima
Closterium dianae
Coccomonas orbicularis
Coelastrum asteroideum
Coelastrum reticulatum
Coenochloris pyrenoidosa
Oocystis lacustris
Oocystis sp
Scenedesmus braziliensis
Sphaerocystis schroeteri
Tetraedron minimum
Westella botryoides
Chlorophyta
Coelastrum pseudmicroporum Kors
Coelastrum reticulatum
Cosmarium cf margarinatum
Lagerheimia ciliata
Pandorina morum
Staurastrum gracile
Tetraedron minimum
Chrysophyceae
Ochromonas sp
Chrysophyceae
Dinobryon divergens
Diatomeae
Actinocyclus normanii
Fragilaria crotonensis
Stephanodiscus binderanus
Stephanodiscus niagarae
Diatomeae
Fragilaria capucina
Fragilaria crotonensis
Stephanodiscus niagarae
Cryptophyceae
Cryptomonas erosa
Cryptomonas marssonii
Cryptomonas reflexa
Cryptomonas sp
Rhodomonas lacustris
Rhodomonas lens
Rhodomonas minuta
Cryptophyceae
Cryptomonas marssonii
Cryptomonas reflexa
Cryptomonas rostratiformis
Rhodomonas lens
Rhodomonas minuta
Dinophyceae
Gymnodinium helveticum
Ceratium furcoides
Dinophyceae
Ceratium furcoides
Ceratium hirundenella
Gymnodinium spp
55
Table 2. A list of phytoplankton species contributing 5% or more to total biomass of any
sample from Stn 908, Hamilton Harbour.
2003
2004
Cyanophyta
Anabaena crassa
Aphanizomenon flos-aquae
Cyanophyta
Microcystis botrys
Microcystis wesenbergi
Woronichinia naeglianum
Chlorophyta
Chlamydomonas gracilis
Chlamydomonas sp
Coelastrum pseudmicroporum Kors.
Coelastrum reticulatum
Monoraphidium contortum
Pediastrum boryanum
Chlorophyta
Chlamydomonas sp
Coelastrum pseudmicroporum Kors.
Coelastrum reticulum
Cosmarium margarinatum
Pediastrum boryanum
Staurastrum gracile
Chrysophyceae
Ochromonas sp
Chrysophyceae
Dinobryon divergens
Diatomeae
Actinocyclus normanii
Fragilaria capucina
Fragilaria crotonensis
Diatomeae
Stephanodiscus niagarae
Cryptophyceae
Cryptomonas erosa
Cryptomonas marssonii
Cryptomonas reflexa
Cryptomonas rostratiformis
Rhodomonas minuta
Cryptophyceae
Cryptomonas erosa
Cryptomonas marssonii
Cryptomonas reflexa
Cryptomonas rostratiformis
Rhodomonas minuta
Dinophyceae
Ceratium furcoides
Ceratium hirundenella
Glenodinium spp
Gymnodinium helveticum
Dinophyceae
Ceratium furcoides
Ceratium hirundenella
Gymnodonium spp
56
258
908
0 0.5 1
2 km
Fig. 1. Map of Hamilton Harbour, Lake Ontario showing planktonic sampling locations
57
-3
mg C m h
-1
>20 μm
80
60
40
20
0
May May Jun
15
28
12
2-20 μm
Jun Jul 8 Jul
25
23
2002
-1
80
Oct Nov
15
6
2004
60
-3
mg C m h
Aug Aug Sep Sep Oct
6
21
3
17
1
40
20
0
May May Jun Jun Jul
15 28 12 25
8
<2 μm
2002
-1
80
2004
60
-3
mg C m h
Jul Aug Aug Sep Sep Oct Oct Nov
23
6
21
3
17
1
15
6
40
20
0
May May Jun Jun Jul
15 28 12 25
8
Jul Aug Aug Sep Sep Oct Oct Nov
23
6
21
3
17
1
15
6
2002
2004
Fig. 2. Size fractionated primary productivity at Station 258 during 2002 and 2004.
58
>20 μm
50
mg C m-3 h-1
40
30
20
10
0
May May Jun Jun Jul
14 27 10 24
7
2-20 μm
Jul Aug Aug Sep Sep Sep Oct Oct
22
6
20
3
16 29 14 27
2003
2004
150
-3
mg C m h
-1
120
90
60
30
0
<2 μm
May May Jun Jun Jul
14 27 10 24
7
Jul Aug Aug Sep Sep Sep Oct Oct
22 6 20 3 16 29 14 27
May May Jun Jun Jul
14 27 10 24
7
Jul Aug Aug Sep Sep Sep Oct Oct
22
6
20
3
16 29 14 27
mg C m-3 h-1
50
40
30
20
10
0
2003
2004
Fig. 3. Size fractionated primary productivity at Station 908 during 2003 and 2004.
59
a) Stn 258, 2002
b) Stn 258, 2004
2500
1200
2000
1000
800
m g m -3
m g m -3
1500
1000
600
400
500
200
0
0
May 15 June 12
July 9
Aug 6
Sept 4
Oct 2
Nov 6
May May Jun Jun Jul 6 Jul Aug Aug Aug Sep Sep Oct Oct
13 26 8 24
19 3 17 31 14 27 12 27
2004
100%
2002
100%
40%
80%
Ciliates
HNF 60%
APP
Bacteria 40%
20%
20%
80%
60%
0%
May15
June12
July9
Aug6
Sept 4
Oct 2
1
Ciliates
HNF
APP
Bacteria
0%
May May June June July July Aug Aug Aug Sept Sept Oct Oct
13 26 8 24 6 19 3 17 31 14 27 12 27
Nov 6
Fig. 4. Biomass and relative composition of microbial loop communities (Bacteria,
Autotrophic Picoplankton (APP), Heterotrophic Nanoflagellates (HNF) and Ciliates at
station 258 during a) 2002 and b) 2004.
a) Stn 908, 2003
2000
b) Stn 908, 2004
1000
1200
600
m g m -3
800
m g m -3
1600
800
400
400
200
0
0
May May Jun Jun Jul 8 Jul Aug Aug Sep Sep Sep Oct Oct
22 6 20 3 16 29 14 27
14 27 10 24
May May Jun Jun Jul 6 Jul Aug Aug Aug Sep Sep Oct Oct
13 26 8 24
19 3 17 31 14 27 14 27
2004
2003
100%
100%
80%
80%
40%
Ciliates
60%
HNF
APP
40%
Bacteria
20%
20%
60%
0%
May May Jun Jun Jul 8 Jul 22 Aug 6 Aug Sep 3 Sep Sep Oct Oct
14 27 10 24
20
16 29 14 27
Ciliates
HNF
APP
Bacteria
0%
May May Jun 8 Jun Jul 6 Jul Aug Aug Aug Sep Sep Oct Oct
13 26
24
19 3 17 31 14 27 14 27
Figure 5. Biomass and relative composition of microbial loop communities (Bacteria,
Autotrophic Picoplankton (APP), Heterotrophic Nanoflagellates (HNF) and Ciliates at
station 908 during a) 2003 and b) 2004.
60
a) Phytoplankton and Zooplankton Biomass (mg m-3)
12000
10000
mg m
-3
8000
6000
4000
2000
0
May May Jun
15
28
13
Jun Jul 9 Jul
25
23
Aug Aug Sep Sep Oct
6
20
4
17
1
Phytoplankton
Oct Nov
15
6
Zooplankton
b) Phytoplankton Composition (% Biomass)
100%
Dinophyceae
Cryptophyceae
Diatomeae
Chrysophyceae
Chlorophyta
Cyanophyta
80%
60%
40%
20%
0%
May May
15
28
Jun
13
Jun Jul 9
25
Jul
23
Aug
6
Aug
20
Sep
4
Sep Oct 1 Oct
17
15
Nov
6
c) Zooplankton Feeding Ecology (% biomass)
100%
80%
Carnivorous Zooplankton
Herbivorous Zooplankton
Veligers
60%
40%
20%
0%
May
15
May
28
Jun
13
Jun
25
Jul 9
Jul
23
Aug
6
Aug
20
Sep
4
Sep Oct 1 Oct
17
15
Nov
6
d) Zooplankton Composition (% Biomass)
100%
80%
Cyclopoids
Cladocera
Calanoids
Veligers
60%
40%
20%
0%
May
15
May
28
Jun
13
Jun
25
Jul 9
Jul
23
Aug
6
Aug
20
Sep
4
Sep Oct 1 Oct
17
15
Nov
6
Fig. 6. A representation of the planktonic food web at station 258 during 2002 including a)
Phytoplankton and Zooplankton Biomass, b) Phytoplankton composition, c) Zooplankton
Feeding Ecology and d) zooplankton composition.
61
-3
Phytoplankton and Zooplankton Biomass (mg m )
8000
mg m-3
6000
4000
2000
0
May May Jun
13
26
8
Jun Jul 6 Jul Aug Aug Aug Sep Sep Oct
24
19
3
17
31
14
27 12
Phytoplankton
Zooplankton
Oct
27
Phytoplankton Composition (% Biomass)
100%
80%
Dinophyceae
Cryptophyceae
Diatomeae
Chrysophyceae
Chlorophyta
Cyanophyta
60%
40%
20%
0%
May May Jun
13
26
8
Jun Jul 6 Jul
24
19
Aug Aug Aug Sep Sep Oct
3
17
31
14
27
12
Oct
27
Zooplankton Feeding Ecology (% biomass)
100%
80%
60%
Carnivorous Zooplankton
Herbivorous Zooplankton
Veligers
40%
20%
0%
May May Jun
13
26
8
Jun Jul 6 Jul
24
19
Aug Aug Aug Sep Sep Oct
3
17
31
14
27
12
Oct
27
Zooplankton Composition (% Biomass)
100%
80%
Cyclopoids
Cladocera
Calanoids
Veligers
60%
40%
20%
0%
May May Jun
13
26
8
Jun Jul 6 Jul
24
19
Aug Aug Aug Sep Sep Oct
3
17
31
14
27
12
Oct
27
Fig. 7. A representation of the planktonic food web at station 258 during 2004 including a)
Phytoplankton and Zooplankton Biomass, b) Phytoplankton composition,
c) Zooplankton Feeding Ecology and d) zooplankton composition.
62
Phytoplankton and Zooplankton Biomass (mg m-3)
6000
5000
mg m-3
4000
3000
2000
1000
0
May May Jun
14
27 10
Jun Jul 8 Jul Aug Aug Sep Sep Sep Oct
24
22
6
20
3
16
29 14
Phytoplankton
Zooplankton
Oct
27
Phytoplankton Composition (% Biomass)
100%
80%
Dinophyceae
Cryptophyceae
Diatomeae
Chrysophyceae
Chlorophyta
Cyanophyta
60%
40%
20%
0%
May May Jun
10
14
27
Jun Jul 8 Jul
24
22
Aug Aug Sep Sep Sep Oct
6
20
3
16
29
14
Oct
27
Zooplankton Feeding Ecology (% Biomass)
100%
80%
Carnivorous Zooplankton
Herbivorous Zooplankton
Veligers
60%
40%
20%
0%
May May Jun
14
27
10
Jun Jul 8 Jul
24
22
Aug Aug Sep Sep Sep Oct
6
20
3
16
29
14
Oct
27
Fig. 8. A representation of the planktonic food web at station 908 during 2003 including a)
Phytoplankton and Zooplankton Biomass, b) Phytoplankton composition,
c) Zooplankton Feeding Ecology and d) zooplankton composition.
63
Phytoplankton and Zooplankton Biomass (mg m-3)
6000
mg m-3
5000
4000
3000
2000
1000
0
May May Jun Jun Jul 6 Jul Aug Aug Aug Sep Sep Oct
8
24
19
3
17 31 14 27 14
13 26
Phytoplankton
Zooplankton
Oct
27
Phytoplankton Composition (% Biomass)
100%
80%
Dinophyceae
Cryptophyceae
Diatomeae
Chrysophyceae
Chlorophyta
Cyanophyta
60%
40%
20%
0%
May May Jun
13
26
8
Jun Jul 6 Jul
19
24
Aug Aug Aug Sep Sep Oct
3
17
31
14
27
14
Oct
27
Zooplankton Feeding Ecology (% Biomass)
100%
80%
Carnivorous Zooplankton
Herbivorous Zooplankton
Veligers
60%
40%
20%
0%
May May Jun
13
26
8
Jun Jul 6
24
Jul
19
Aug Aug Aug Sep Sep
3
17
31
14
27
Oct
14
Oct
27
Fig. 9. A representation of the planktonic food web at station 908 during 2004 including a)
Phytoplankton and Zooplankton Biomass, b) Phytoplankton composition,
c) Zooplankton Feeding Ecology and d) zooplankton composition.
64
65
ZOOPLANKTON IN HAMILTON HARBOUR 2002-2004
Jocelyn Gerlofsma, Kelly Bowen and Ora Johannsson
INTRODUCTION
The Hamilton Harbour Remedial Action Plan (RAP) identified the degradation of zooplankton
as one of the beneficial use impairments (Hamilton Harbour RAP 1992). It stated zooplankton
abundance was high reflecting eutrophication and high productivity. Also, the mean size of
zooplankton was small indicating heavy fish predation by a population dominated by
planktivores.
During the 1990s, much effort has been spent to reduce phosphorus inputs to the
Harbour in order to lower the level of eutrophication. Additionally, DFO and the RAP have
improved fish habitat around the edge of the harbour and reduced carp breeding area by
installing a fishway into Cootes Paradise for improvement of the native fish community. These
efforts are expected to affect the zooplankton community. In order to assess the impact of these
improvements on the aquatic ecosystem as a whole and reassess targets for the RAP, DFO has
undertaken a long-term study of the Hamilton Harbour aquatic ecosystem starting in 2002.
This is an interim report examining zooplankton species composition, size structure,
abundance, biomass and productivity in 2002, 2003 and 2004. The harbour is a large,
hydrodynamically complex system with a range of habitats for zooplankton. The sampling
stations were chosen to allow examination of the dominant spatial gradients specifically from
inshore/shallow to offshore/deep and from regions near the major inflows in the west to the
more mixed central basin.
METHODOLOGY
Zooplankton
Biweekly zooplankton sampling was carried out from the beginning of May to the end of
October. From 2002 to 2004, five stations were examined with a different set of stations
sampled each year (Table 1, Figure 1). All stations, except station 17, were sampled using a 41
litre Schindler-Patalas trap fitted with 64µm mesh. Samples were collected at discrete depths
(Table 1) and preserved individually using a 4% sugar buffered formalin solution. For analysis,
a single composite sample was constructed for each station-date by combining 50% of the
sample from each depth.
66
The shallowest site, Station 17, was a 100m transect parallel to shore following the
1.5m depth contour, along which the ends and mid-point were sampled. At each of these three
points, a Guzzler® diaphragm hand pump and 25mm diameter hose were used to collect 10L of
water from 0.5m and 1.0m depths for a total of 60L. The samples were pooled and concentrated
using a 64µm mesh net, and preserved as above.
A minimum of 400 individual zooplankters and all loose eggs within a subsample
aliquot were enumerated from each sample. At least 100 individuals of each major group were
included. Cercopagis pengoi, a predatory cladoceran which invaded Lake Ontario in 1998
(MacIsaac et al. 1999), cannot be accurately enumerated from subsamples because their hooked
caudal spines become entangled, and form clumps. Therefore the larger organisms from each
sample were captured on a 400-μm mesh and all the C. Pengoi were removed and counted.
Seasonally-weighted mean (SWM) abundance, biomass and production of zooplankton
were calculated over the May 1 to October 31 sampling season. Lengths of cladocerans were
measured from the top of the helmet to the base of the tail spine, copepods from the anterior
end of the cephalothorax to the end of the caudal rami, and veligers across the widest section of
the shell. Body mass (mg dry weight) was estimated from length-weight regressions from the
literature, summarized in Johannsson et al. (2000). Regression equations for Cercopagis are
given in Grigorovich et al. (2000). Production was estimated by the egg-ratio method of
Paloheimo (1974) as described in Cooley et al. (1986) where cyclopoid or calanoid nauplii and
copepods were assigned to species according to the relative abundance of the adults. When a
species’ seasonally-weighted mean biomass was <50 mg m-2, production was estimated using
P/B relationships as described in Johannsson et al. (2000). However in several cases (e.g.
cyclopoid copepods), the egg-ratio production estimate was zero or close to zero, so the P/B
production estimate was used despite a SWM biomass >50 mg m-2. Based on the Bay of Quinte
data, it appears that P/B values estimated from the literature relationships overestimate
production. Therefore, the P/B production estimates were amended using the correction
equations for Belleville in the Bay of Quinte (Bowen and Johannsson 2005) as follows:
Cladocerans:
ln (egg-ratio prod.) = -0.301 + 1.026 ln (P/B prod.)
N = 19; r2 = 0.89
Cyclopoids:
ln (egg-ratio prod.) = -0.696 + 1.010 ln (P/B prod.)
N = 10; r2 = 0.93
67
Rotifers
Samples were collected using a Guzzler® diaphragm hand pump and 16mm diameter hose. The
hose was lowered to the bottom sample depth and then raised through the water column at a set
rate (1 pump stroke per 0.5 m (stations 17, 908 and 258) or 1 stroke per 0.25 m (stations 6,
WC) (Table 2). The residual volume of water held by the hose was accounted for during the
sampling.
For preservation, 6 L of sample water was sieved through 20µm mesh, rotifers were
narcotized with soda water and preserved with 4% sugared buffered formalin. For analysis a
May to October seasonal composite for each station was made by pooling 50% of the sample
from each date.
To enumerate and measure the rotifers a subsample was removed and placed in a
Sedgewick-Rafter chamber at 100 x magnification. A minimum of 200 organisms were counted
in a sample dominated by one to three species. For more diverse samples, a minimum of 400
individuals were counted. At least 20 individuals of each species were measured. For
numerically dominant species, 50 or more individuals were measured. A maximum of 20% or
25% of the sample, by volume, was entirely analysed. Biomass, except for the Polyarthra
species, was estimated according to the formula of Ruttner-Kolisko (in McCauley 1984). The
formulae determined for Lake Erie Polyarthra were also applied in Hamilton Harbour
(Johannsson et al. 2000). For P. dolichoptera, v = 0.205a3 was used where v = volume in μ3 (or
wet body mass in μg x 10-6 assuming a density of one) and a = longest dimension in microns,
while for P. vulgaris, P. major and P. remata, v = 0.158a3.
RESULTS
Zooplankton Density
In 2002, the seasonally weighted mean (SWM) areal density ranged from 0.69 x106
indviduals·m-2 at the 1.5 m deep station to 7.66 x106 indviduals·m-2 at the deepest site, 24 m.
There was an offshore to nearshore gradient with the average (volumetric) density increasing
towards shallower depths (Table 3).
Species composition at the shallowest station (17) varied relative to the offshore stations
(6 and 258) (Table 3). There were a total of 23 taxa at station 17 compared to 18 at station 6
and 19 at station 258. Herbivorous cladocerans were more diverse in the nearshore with more
macrophyte-associated species (e.g. Sida crystallina, Camptocercus rectirostris) and benthicassociated species (e.g. Alona sp., Eurycercus sp.) and harpacticoid copepods (Balcer et al.
1984) (Table 3). The offshore stations, especially station 258, supported more Eubosmina
coregoni, Daphnia galeata mendotae, D. retrocurva and dreissenid veligers. At all three
stations Bosmina longirostris had the highest SWM areal density – 4.19 x106 indviduals·m-2.
68
In 2003, both areal density and density increased towards the offshore, opposite to the
pattern in 2002 where density decreased towards the offshore (Table 3). The densities in 2003
were also lower than even the density at the deepest site in 2002. A similar distribution of
species occurred at both stations WC and 908 in 2003; 17 taxa were identified at each.
Bosmina was the most abundant taxon, followed by cyclopoid naupli and copepodids, and
veligers (Table 3). Most of the cladocerans, including Bosmina, Eubosmina, D. retrocurva, D.
galeata mendotae, and Ceriodaphnia lacustris had lower densities than in 2002. Cyclopoid
densities in 2003 were similar to those in 2002, but calanoid densities were lower. Over the
three years studied, Cercopagis pengoi was most abundant in 2003.
Sampling in 2004 revisited the offshore stations 258 (24 m in depth) and 908 (14 m in
depth). Both stations had lower densities compared to the prior years. Zooplankton at station
258 had a higher areal density but lower density (4.20 x106 indviduals m-2, 0.193 x106
indviduals m-3 ) than zooplankton at station 908 (2.71 x106 indviduals m-2, 0.175 x106
indviduals m-3) (Table 3). Station 908 had 18 taxa, and 258 had only 15. Although community
composition at the two stations was similar in 2004, there were some differences compared to
the previous years. Bosmina was still the most common taxon, but numbers were low. Veliger
densities also fell drastically in 2004. Densities of other cladocerans varied year to year,
depending on the station. The cyclopoids Diacyclops thomasi and Mesocyclops edax were
higher in 2004, but densities of juvenile cyclopoids were intermediate.
Zooplankton Biomass
SWM volumetric biomass was higher in 2002 than 2004, as determined by comparing
biomasses at station 258 between years, and was similar or slightly higher in 2004 than 2003,
as determined by comparing biomasses at station 908 between years (Table 4).
In 2002, areal SWM biomass had a strong nearshore to offshore gradient, with an offshore
biomass at station 258 10 times greater than the nearshore station (7682 mg m-2 vs. 748 mg m-2;
Table 4). More than 75% of the biomass was comprised of herbivorous cladocerans
(bosminids, Daphnia and others; Table 4). Bosmina was the most dominant taxon at the
offshore stations 258 and 6 (Figure 2). Daphnia was also important at stations 258 and 6, but
not at 17 where other herbivorous cladocerans dominated. Copepod contribution was highest at
258 and lowest at 17 (Table 4). Cyclopoids composed 6.6% - 14.2% of the biomass and
calanoids 1.2 % -3.0 % with the greatest percentage of copepods at 258 and the lowest at 17
(Table 4). Harpacticoids were noteworthy only at 17, contributing 2.5% of the biomass.
Veliger biomass contribution was highest at station 258 (7.8%) compared to station 6 (2.4%)
and station 17 (1.5%).
In 2003, SWM areal biomass again had an offshore to nearshore gradient with offshore
station 908 substantially higher than nearshore station WC (3252 mg m-2 vs. 500 mg m-2; Table
4). However, the community composition at the two stations was similar. Cladocerans made up
69% of the biomass; the main contributors being bosminids (54.8%) and Daphnia sp. (10.6%)
(Table 4). Approximately 25% of the biomass came from cyclopiods.
69
In 2004, SWM areal biomass at station 908 (3452 mg m-2) was similar to the previous
year (Table 4). At station 258, the areal biomass was lower in 2004 (biomass = 5433 mg m-2)
than in 2002. Cladocerans were still dominant, but Daphnia, not bosminids, were the main
contributors (Figure 2). Cylcopoid and calanoid biomass values were 25.5% and 5%
respectively. Veliger biomass was very low in 2004 (0.1%) relative to the previous two years.
Zooplankton SWM areal biomass values were compared to stations from the Bay of
Quinte over the same time period (Figure 2). Belleville (B) and Napanee (N) are both 5 m
deep, unstratified stations located in the eutrophic upper bay. The 2002-2004 mean biomass
values at B (527 mg m-2) and N (653 mg m-2) were substantially lower than the nearshore
Hamilton Harbour station 6 (2357 mg m-2 ) (Bowen and Johannsson 2006). Hay Bay (HB),
located in the mesotrophic middle bay, is 12 m deep and somewhat comparable to the offshore
harbour stations 908 and 258. Mean biomass was also lower at HB (2197 mg m-2) than 908 or
258. Conway (32 m deep) is positioned at the mouth of the bay and represents mesooligotrophic conditions. No shallow nearshore stations were sampled in the Bay of Quinte from
2002 to 2004 that were comparable to 17 and WC. However, mean biomass in a weedy
embayment in the upper bay was 358 mg m-2 in 2001, a value substantially lower than station
17 (748 mg m-2) in 2002 (Bowen et al. 2003).
Seasonal Biomass Trends
Seasonal biomass patterns varied at the different stations over the three years (Figures 3 and 4).
However, Bosmina consistently peaked early in the season around the end of June, crashed and
was then replaced by a small population of E. coregoni until the end of September or early
October. Daphnia generally started appearing around mid-July when Bosmina began their
decline. D. retrocurva peaked in August at stations 258, 908 and WC in 2002 and 2003. At the
shallow stations in 2002 it did not develop large summer populations. In 2004, D. retrocurva
started increasing earlier in July and remained relatively steady until it peaked in September.
D. retrocurva biomass then dropped quickly as D. galeata mendotae peaked at the end of
September. Predatory cladocerans comprised a very small portion of the community, but
Leptodora kindtii had a strong presence at 908 in August 2003. C. pengoi generally had a low
and short-lived presence just prior to L. kindti.
Cyclopoid biomass varied from year to year. In 2002, cyclopoids were most common
from late summer to mid-autumn. In 2003, a spring cyclopoid bloom ended in late June.
Cyclopoids remained at low levels until the end of October with only one small peak midSeptember at 908. In 2004, cyclopoids had a strong presence in the zooplankton population
from early spring until October. Generally, spring populations were composed of D. thomasi
and later populations were mainly M. edax. Never dominant, calanoids generally appeared in
late summer or early fall. Harpacticoids were only seen at the shallow station 17, where they
formed a small peak in early October 2002.
70
Veligers also varied from year to year. A short bloom of veligers occurred in early
August 2002. In 2003, they peaked sharply in late August at 908, and at WC from late August
to mid-September. Veligers in 2004 were very low in abundance and made no significant
contribution to biomass at any point in the season.
Zooplankton Production
Seasonal (May-October) areal production increased from inshore to offshore with station depth
(Figure 5), and during our study, was highest at station 258 in 2004 (107.1 g m-2) and lowest at
station WC in 2003 (7.7 g m-2). In 2003, seasonal production at the offshore station 908
(45.5 g m-2) was lower than the offshore station 258 in 2002, but higher than the nearshore
stations (6 and 17). In 2004, the production at station 908 (51.5 g m-2) was similar to the
previous year at the same station (Table 5, Figure 5). Production at station 258 (61.7 g m-2)
was lower than that found in 2002.
The relative contributions of the major zooplankton groups to total seasonal production
were similar across stations within a year, but differed between years. In 2002, the majority of
the production was from herbivorous cladocerans, ranging from 96% (32.9% bosminids, 1.3%
Daphnia and 61.8% others) at station 17 to 81.5% at station 258 (Figure 5). At the stations 6
and 258, Bosmina and Daphnia contributed more than any other species to zooplankton
production (Figure 6). Cyclopoid production ranged from 3.8 % at station 17 to 7.4 % at
station 258, whereas calanoids represented less than 1% of the total (Table 5). Veliger
production was high at 258 (10.4%) compared to stations 6 (3.1%) and 17 (1.4%).
In 2003, Bosmina was again dominant, representing 67% of production at WC and 57%
at 908. Predatory cladocerans, mainly L. kindtii, were noteworthy at Station 908 (3.4%).
Cyclopoids contributed proportionately more to the production (13.3-16.9 %) than in 2002, but
the total production (average=3.7 g m-2) was similar to 2002 (Table 5). Calanoids production
was very low (average = 0.06 g m-2), adding less than 1 % of the total. Veliger production
contributed more to the nearshore (WC, 6.4%) than to offshore (908, 3.8%).
In 2004, the percent contribution of the different zooplankton groups was similar
between the two stations, 258 and 908 (Table 5, Figure 6), with the areal production higher at
station 258. In contrast to previous years, cladoceran production at both stations was dominated
by Daphnia (56%), not Bosmina (24%). Also, veligers crashed in 2004 with only 76 mg m-2
production at station 908 and 127 mg m-2 at 258.
Cladoceran Length
Small bosminids dominated the system in 2002 and 2003, resulting in a consistently low mean
cladoceran length ranging from 313µm to 349µm (Figure 7). The higher relative abundance of
Daphnia increased the cladoceran mean size the following year (423µm to 430µm). Compared
to the Bay of Quinte, harbour cladocerans tended to be smaller in 2002 and 2003, but larger or
similar in size in 2004.
71
Oxygen and Zooplankton
At station 258 in 2002, zooplankton samples taken from various depths in the water column
were analyzed discretely on four different dates. These depths were chosen to represent a
gradient in dissolved oxygen, and to show how zooplankton biomass changed when the oxygen
dropped to hypoxic (1-4 mg l-1) and near-anoxic (<1 mg l-1) levels. On July 9th and 23rd, the
oxygen levels gradually decreased from > 8 mg l-1 at the surface to < 1 mg l-1 at the bottom of
the metalimnion (7-9 m deep) (Figure 8). In the epilimnion and at the top of the metalimnion,
zooplankton biomass was high (July 9: 420-479 mg l-1 and July 23: 230-360 mg l-1), but
dropped (July 9: 162 mg l-1 and July 23: 160 mg l-1) near the bottom of the metalimnion where
the oxygen levels fell below 1 mg l-1. Zooplankton biomass remained low in the hypolimnetic
samples (July 9: 130 mg l-1 and July 23: 191 mg l-1). The hypolimnetic oxygen increased to 2-3
mg L-1 between 11-14 m and then decreased back down to about 1 mg l-1. In September, the
oxygen profile indicated that oxygen remained high in the epilimnion (Sept. 4: 10.5 mg l-1 and
Sept. 17: 7 mg l-1), dropped rapidly over the 4-5 m of the metalimnion to <0.5 mg l-1 (11-13 m
deep) and remained there through the hypolimnion. Zooplankton biomass was the highest
(Sept. 4: 529 mg l-1 and Sept. 17: 159 mg l-1) in the metalimnion at oxygen concentrations of
6 mg l-1 to 0.5 mg l-1. Zooplankton biomass in the hypolimnon was higher than in the
epilimnion, even with low oxygen levels, on both dates (Figure 8). On all four dates the percent
composition of the major zooplankton groups did not vary with depth, except for C. pengoi. It
was only in one sample - July 23 at 5 m.
Rotifers
Each year, a single seasonal composite sample was analyzed for rotifers at each station. A total
of 27 rotifer taxa were identified in Hamilton Harbour between 2002 and 2004, most to the
species level. Ten families were represented. Between 15 and 18 taxa were found at each
station in 2002 and 2004, and between 12 and 14 taxa in 2003 (Table 6). Seven taxa were
found in all seven samples, whereas 5 taxa were found in only one sample.
In 2002, rotifer areal and volumetric density increased from the nearshore to the
offshore (Table 6). In 2003 and 2004, areal density again increased, but volumetric density
decreased along this nearshore offshore gradient. In 2003, rotifer areal density at nearshore WC
(2.54 x105m-2) was higher than the nearshore stations 17 and 6 in 2002. In 2004, when stations
908 and 258 were re-sampled, densities were lower than the previous years.
Numerically, Keratella cochlearis was the most dominant taxon at all stations, where it
comprised between 52 and 77% of the total rotifer community. Another dominant species was
Polyarthra dolychoptera (4-17% of the total). These are both fairly small rotifers, averaging
114 and 87 µm, respectively. Several taxa were more abundant at the shallow macrophtye
station 17, including Filinia terminalis, Lecane spp. and Trichocerca porcellus, whereas
Kellicottia bostoniensis and Pompholyx sulcata tended to be associated with the offshore
(Table 6).
72
Although seasonal mean rotifer and zooplankton density values were in the same range,
mean rotifer biomass was generally only 1-3% that of zooplankton biomass because of their
small size. One exception occurred at the WC station in 2003, where rotifer biomass was 9%
of the SWM zooplankton biomass. During the three years, the areal biomass was highest at the
deep stations and lowest at the shallow stations, ranging between 8.1 mg m-2 at Station 17 and
136.5 mg m-2 at Station 258 in 2002 (Table 6; Figure 8). Values between 2003 and 2004 at
station 908 were similar (82.2 mg m-2 in 2003; 78.6 mg m-2 in 2004), but there was a drop in
the rotifer biomass from 2002 to 2004 at station 258 (136.6 mg m-2 in 2002; 85.4 mg m-2 in
2004). Rotifer biomass in Hamilton Harbour was generally two to three times greater than in
the Bay of Quinte over this same time period (Figure 8).
Due to its relatively large size (390 µm), Asplanchna priodonta was generally the most
dominant rotifer by biomass in the harbour. This taxon usually comprised between 43.9 and
61.8% of total biomass at most stations (Figure 8). One exception was 908 in 2004, which
supported a more diverse rotifer assemblage. Other dominant rotifer genera by biomass were
Keratella (10.2%-30.4%), Polyarthra (2.8%-35.6%) and Trichocera (0.6%-19.2%). In the Bay
of Quinte, Polyarthra, followed by Asplanchna and Trichocera were the most dominant
rotifers.
DISCUSSION
Hamilton Harbour remains a polluted, eutrophic body of water, although much improved
through the remediation actions undertaken over the past thirty years. Total phosphorus has
declined from 54 µg l-1 to 33 µg l-1, chlorophyll a from 33 µg l-1 to 11.4 µg l-1 and Secchi depth
has increased from 1.5 m to 2.4 m (MOE 1981, Painter et al. 1990, Burley this volume). Nearanoxic conditions (O2 < 1 mg l-1) in the hypoliminion remain a problem, but are less persistent
than in the 1970s (MOE 1981, Burley this volume). Has the zooplankton community responded
to these improvements?
Little historical information is available on the zooplankton of Hamilton Harbour. The
first studies were conducted by Harris (1976), Piccinin (1977) and Piccinin and Harris (1980)
during the 1975-1979 period at the height of eutrophication, before remediation commenced.
A brief study was also conducted in 1990 when Koenig (1992) examined copper and cadmium
levels in Hamilton Harbour plankton. These studies provide some background with which to
compare the current status of the zooplankton community. The zooplankton community in
Hamilton Harbour in 2002-2004 can also be compared to the upper Bay of Quinte (2002-2004:
TP = 31-46 µg l-1), another eutrophic embayment in Lake Ontario (Nicholls and Millard 2006).
The present study also provides an opportunity to look at inter-annual variation and inshoreoffshore gradients in zooplankton community dynamics.
The zooplankton community in the late 1970s was primarily dominated by large rotifers
and the cladoceran Bosmina longirostris. There were very few other cladoceran species and
only two copepod species (Harris 1976, Piccinin 1977, Piccinin and Harris 1980). The
dominant rotifers were Keratella quadrata, Brachionus angularis, Filinia terminalis and
Trichocera cylindrical. These species are all indicative of eutrophic conditions in the Great
73
Lakes (Gannon and Stemberger 1978). Unfortunately the studies in 1975-1979 do not provide
abundance or biomass data comparable to the 2002 -2004 data. In 1990, the zooplankton
community was still dominated by cladocerans and rotifers. Cladocerans included Bosmina
(86% biomass), Daphnia sp. (12%) and Leptodora kindtii (2%) (Koenig 1992). The dominant
rotifers were Polyarthra sp., Synchaeta, Keratella and Pompholyx. Although it was one of the
most abundant species in the 1970s, only a few B. angularis were present in August 1990.
Shifts in composition have continued into the early 2000s. Numerically, the dominant species
was still a rotifer, K. cochlearis, followed by a cladoceran B. longirostris, but the presence of
Daphnia sp. and copepods continued to grow (Table 3 & 6). The rotifers B. angularis and T.
cylindrical were not present at any of the stations in the 2000s, and F. terminalis was found in
very low densities at station 258 in 2004 (Table 6). The present zooplankton community still
indicates that Hamilton Harbour is eutrophic. Higher abundances of cladocerans and
cyclopoids compared to calanoids is a good indicator of eutrophic waters (Gannon and
Stemberger 1978). Also, Patalas (1972) found that zooplankton communities dominated by B.
longirostris, E. coregoni, D. retrocurva, D. galeata mendotae, Mesocyclops edax, Diacyclops
thomasi indicated more eutrophic conditions in the Laurentian Great Lakes, and these are all
present in high numbers in Hamilton Harbour. The decline in rotifers may be associated partly
with the decline in total phosphorous; however increased predation by the growing populations
of copepods, and competition and interference mortality associated with the increasing
presence of Daphnia spp. are also likely involved (Gilbert 1988, MacIsaac and Gilbert 1991).
The change to a zooplankton community less dominated by rotifers and with higher
biodiversity reflects an improvement in Hamilton Harbour waters, its foodweb structure, and
energy flow to higher trophic levels.
In the 1970s anoxic and near anoxic (O2 < 1 mg l-1) conditions in Hamilton Harbour’s
hypolimnion limited most of the zooplankton biomass to the epilimnion from June to
September. There was even one occurrence in 1979 where a stable period in July caused anoxic
conditions to rise within 4 m of the surface (Piccinin and Harris 1980). Bosmina, the most
abundant zooplankton, disappeared from the hypolimnion when near-anoxia to anoxia
occurred, but rotifers were still present (Harris 1976, Piccinin and Harris 1980). Bottom
aeration experiments during the late 1970s increased the zooplankton biomass of Bosmina and
Filinia sp. and their presence was observed in the hypolimnetic waters (Harris 1976, Piccinin
1977). Hypolimnetic hypoxia was still a problem during this present study (mean hypolimnetic
for 2002-2004 = 1.3 mg l-1 to 1.6 mg l-1; Burley this volume). In 2002, over the stratification
period the mean hypolimnetic oxygen was 1.3 mg l-1. There were four occurrences of mean
hypolimnetic oxygen levels below 1.0 mg l-1 from August to early October, whereas in the 1970s
the hypolimnion was persistently anoxic or near-anoxic (O2 = 0-1 mg l-1) from June to September
(Harris 1976, MOE 1981, Burley this volume). The discrete depth zooplankton samples from
this study showed that zooplankton biomass (200 mg l-1 – 500 mg l-1) was quite strong in the
metalimnion even where oxygen dropped to <1 mg l-1 (Figure 8). Zooplankton also occupied
the hypolimnion ranging in biomass from 94-291 mg l-1 on the 4 dates measured. In July,
elevated oxygen in the hyplomnion indicated the water came in from Lake Ontario, possibly
bringing the zooplankton with it (Hamblin and He 2003). Yet, in September the oxygen was
< 1 mg l-1 over the whole hypolimnion (Figure 8). Zooplankton may migrate into the lower
oxygen areas to avoid fish predation. In Irondequoit Harbour, a refuge from planktivorous fish
was created in the metalimnion by bring the oxygen levels to no higher than 2 mg l-1 using
74
hypolimnetic oxygen injection (Klumb et al. 2004). Some zooplankton such as D. galeata
mendotae, are tolerant of low oxygen levels and can migrate to layers of low oxygen to avoid
fish (Heberger and Reynolds 1977).
Strong predation of zooplankton by fish is a concern in the harbour as indicated by the
low mean cladoceran length (Figure 7). The mean length of cladocerans is interpreted as an
indicator of the level of planktivory as fish preferentially consume larger zooplankton
individuals (Cooley et al. 1986). Mills et al. (1987) have proposed that mean zooplankton
community length can serve as an indicator of balance between piscivores and planktivores
within the fish community. The same should be true of cladocerans and a single group may
provide a more consistent measure as it is not affected by the relative abundance of copepods to
cladocerans which are very different in shape. However, as of yet, no optimum mean size of
cladocerans has been determined to best reflect the status of the fish community. The high
chlorophyll a/total phosphorus (CHLa/TP) ratios in the harbour (0.41 to 0.62 during the present
study) confirm that the pelagia in the harbour is an ‘odd-linked’ system dominated by
planktivores (Mazumder 1994, Dahl et al. 1995, MacDougall et al. 2001, Nicholls and Millard
2004). Cooley et al. (1986) indicated that high planktivory might mask improvements in
eutrophic waters, since predation keeps larger zooplankton taxa from becoming dominant.
Larger cladocerans can strongly depress algal abundance, which is an indicator of trophic
condition and contributes to the CHLa/TP ratio. In many systems, dreissenids usurp a
significant proportion of pelagic productivity and route it through the benthos (e.g. Johannsson
et al. 2000, Johannsson and Nicholls 2003). Hamilton Harbour is one of the few shallow Great
Lakes systems where dreissenids are not abundant (Dermott and Bonnell this volume), and the
CHLa/TP ratio is still an indicator of the relationship between plankton and fish. In the
Harbour, cladoceran mean length ranged from 320-425 µm over this three year study (Figure
6). These measures tended to be lower than those in the Bay of Quinte over the same time
period, with the exception of 2004 when Daphnia were abundant (Figure 4). Overall,
planktivory can still be considered high in the harbour. The harbour is a nursery area for many
young fish and high levels of planktivory might be expected to be the norm, not an impairment.
Food web models can help to define optimum zooplankton composition and cladoceran size in
the harbour.
In 2002 and 2003, the zooplankton community was examined for inshore to offshore
trends. In both years, based on areal measurements, there was an inshore to offshore gradient
with the greatest density, biomass and production of zooplankton occurring in the offshore.
There was also a distinctive zooplankton community at the shallowest station, 17 (1.5 m deep).
In the spring, the zooplankton community at this shallow station was similar to that at the
offshore sites (6 and 258) with a high abundance of Bosmina and very few other species
(Figure 2). As the submergent macrophytes began growing in early summer, the community
started to include many benthic-associated species (e.g. Alona sp., Eurycercus sp. and
harpacticoids (Balcer et al. 1984)) and macrophyte-associated species (e.g. Sida crystallina
(Fairchild 1981)), rarely seen in the offshore.
The two offshore stations 908 and 258 were each visited in two years, and both sampled
in 2004. This allowed for a comparison between the deeper, mid-harbour region which is
strongly influenced by incursions of water from Lake Ontario (station 258), and the shallower
75
flats in the western end of the harbour which are closer to the major natural inputs (station 908).
Zooplankton areal density, biomass and production were about 1.5 times higher at station 258
than 908. Thus the deeper hypolimnion is supporting a larger zooplankton community,
although the seasonal areal phytoplankton production at the two sites is similar
(258 = 190 g C m-2 and 908 = 215 g C m-2). There are at least two possible reasons for this
observation and both may be operating. First, it may suggest that the shorter hypolimnion at
station 908 does not provide as good protection from predation, both vertebrate and
invertebrate, as does the deeper hypolimnion at station 258. Second, it may suggest that the
micro- and macro-zooplankton at 258 are more effectively recycling the settling organic
material as it falls through the hypolimnion because it remains in the water column for a longer
period of time, and therefore, the zooplankton in the hypolimnion can live and metabolize at
these low oxygen concentrations. The zooplankton data provide less support for the former
than the latter. Cladoceran mean length was very similar between the two stations and
zooplankton community composition was also very similar. These two observations indicate
that predation did not favour one station over the other. If that is true, then zooplankton are
effectively using hypolimnetic food sources despite the lower oxygen levels.
The two deep water stations also give us the opportunity to start assessing the degree of
inter-annual variability in zooplankton community structure, biomass and productivity. The
main differences occurred between 2004 and 2002 at station 258. Although Daphnia were more
abundant in 2004, which can be associated with an increase in zooplankton biomass and
productivity (Johannsson et al. 2000, Johannsson and Nicholls 2003), the relative proportion of
cladocerans in the population decreased while the relative proportion of cyclopoids increased.
This resulted in biomass and production levels 45% and 42% lower in 2004 than in 2002,
respectively. The Harbour is known for its high spatial and temporal variability in water
movement, and temperature and oxygen patterns. Harris (1976) found significantly different
seasonal trends in zooplankton and phytoplankton between the west and central regions of the
Harbour in the mid 1970s. Therefore, we were surprised to find conformity in community
structure across the harbour within all three years. This suggests that some of the source
promoting the biological gradient is either annually variable or has changed – a subject that
needs further consideration.
In summary, the zooplankton of Hamilton Harbour still reflects a highly productive and
eutrophic system that is being dominated by planktivores. Compared to the mid-1970s the
zooplankton community has changed to one less dominated by rotifers and with higher
biodiversity which indicate substantial improvement. Overall, the Hamilton Harbour
zooplankton community is quite dynamic, varying in composition from the very nearshore to
the inshore, and in biomass and productivity from the inshore to the offshore. Thankfully, from
a monitoring perspective, greater variability was observed between years than within years.
There are many influences on the zooplankton community including a variety of water inflows
-creeks, sewage treatment plants and Lake Ontario, the presence of reed and macrophyte beds,
and an abundant and relatively diverse fish community to list a few. Monitoring of the
zooplankton along with the other lower trophic levels (microbial loop, phytoplankton and
benthos) and fish should continue for a better understanding of this dynamic system and allow
for modelling efforts to better define optimal conditions.
76
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79
Table 1. Hamilton Harbour zooplankton stations and sample depths for 2002-2004.
Station
17
6
WC
908
258
Year(s)
Sampled
2002
2002
2003
2003, 2004
2002, 2004
Sample Depths
(m)
0.5, 1
1, 3, 5
0.5, 1.5, 2.5
1, 3, 5, 7, 9, 11, 13
1, 3, 5, 7, 9, 11, 15, 19, 22
Station Depth
(m)
1.5
6.0
3.5
14
24
Table 2. Hamilton Harbour rotifer sampling depth and pump interval for 2002-2004.
Station
17
6
WC
908
258
Sample Depth
(m)
0-1
0-5
0-2.5
0-13
0-22
Sample interval
depth (m)
0.50
0.25
0.25
0.50
0.50
Hose length
(m)
3
6
6
24
24
Table 3: Seasonal mean densities (individuals·m-2) of zooplankton taxa in Hamilton Harbour from 2002-2004.
Species
Group
HH17
2002
HH6
HH258
2003
HHWC
2004
HH908
HH908
HH258
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
Herb. Clad
363 841
1 655
1 648
880
0
122 283
245
2 834
15 126
10 695
1 418
9 122
0
24 075
0
1 682 367
92 232
83 905
30 593
366
2 387
0
21 007
5 970
0
0
0
0
124
120
4 191 419
568 871
336 719
61 340
0
10 509
0
134 047
14 136
0
0
11
0
2 632
0
340 169
10 528
21 712
3 696
0
219
0
15 769
338
0
0
0
23
73
0
1 946 628
130 580
93 062
22 187
0
337
0
73 651
3 365
0
0
0
0
616
0
1 091 014
162 692
203 274
72 359
9 761
0
147
28 029
18 903
0
0
0
0
155
0
Leptodora kindtii
Polyphemus pediculus
Cercopagis pengoi
Pred. Clad
Pred. Clad
Pred. Clad
12
44
4
1 169
0
300
1 638
0
498
56
0
204
3 187
0
2 371
1 106
0
318
1 658 812
158 073
319 271
129 275
1 524
0
0
34 833
25 522
0
0
0
0
0
0
1 696
0
305
Diacyclops thomasi
Cyclops vernalis
Tropocyclops extensus
Mesocyclops edax
Eucyclops agilis
Cyclopoida nauplii
Cyclopoida copepodids
Skistodiaptomus oregonensis
Leptodiaptomus siciloides
Eurytemora affinis
Calanoida nauplii
Calanoida copepodid
Cyclopoid
Cyclopoid
Cyclopoid
Cyclopoid
Cyclopoid
Cyclopoid
Cyclopoid
Calanoid
Calanoid
Calanoid
Calanoid
Calanoid
44
1 058
0
44
4 688
31 808
25 142
0
0
321
2 483
3 676
2 482
879
0
3 825
0
105 333
86 810
56
3 138
0
34 582
9 796
14 995
7 225
0
44 318
89
473 815
408 194
671
21 983
0
154 304
30 283
2 945
824
0
886
0
155 911
58 054
0
165
0
9 330
1 909
23 239
1 903
0
9 264
0
735 364
387 872
32
3 016
0
42 779
8 813
44 235
614
102
51 754
517
480 354
423 185
78
17 861
0
57 499
25 457
34 417
2 914
338
80 143
906
824 356
716 204
102
37 280
0
106 647
42 363
Harp.
Harp.
3 706
35 934
0
56
0
18
37
89
299
713
0
149
0
0
Dreissenid
20 996
683 779
455 853
122 210
2 289 708
381 618
1 183 954
7 661 668
319 236
69 801
692 740
197 926
234 334
3 723 610
265 972
15 505
2 705 067
193 219
28 445
4 203 428
175 143
Harpacticoida nauplii
Harpacticoida adults
Dreissenia veligers
Total (individuals·m-2)
Total (individuals·m-3)
80
80
Bosmina longirostris
Eubosmina coregoni
Daphnia retrocurva
Daphnis galeata mendotae
Other Daphnia sp.
Ceriodaphnia lacustris
Ceriodaphnia sp.
Chydorus sphaericus
Diaphanosoma birgei
Eurycercus sp.
Sida crystallina
Camptocercus rectirostris
Ilyocryptus spinifer
Alona sp.
Chydorus piger
81
Table 4. Percent distribution of zooplankton seasonal biomass (May 1- October 31, 2002-2004)
amongst the taxonomic groups in Hamilton Harbour.
2002
2003
2004
17
6
258
WC
908
908
258
88.2
87.7
75.0
69.0
69.0
69.3
69.5
Herbivorous
88.2
87.2
74.6
68.3
66.4
68.3
68.5
Carnivorous
0.0
0.5
0.3
0.7
2.6
1.0
0.9
Cyclopoids
6.6
7.8
14.2
25.6
26.8
26.3
25.1
Calanoids
1.2
2.1
3.0
0.0
1.3
4.2
5.3
Harpacticoids
2.5
0.0
0.0
0.0
0.0
0.0
0.0
Veligers
1.5
2.4
7.8
5.4
2.9
0.1
0.1
Total (mg m-3)
498.8
351.8
331.1
142.8
232.3
247.0
234.2
Total (mg m-2)
748
2 357
7 682
500
3 252
3 458
5 433
Cladocerans
Table 5. Percent distribution of zooplankton seasonal biomass (May 1- October 31, 2002-2004)
amongst the taxonomic groups in Hamilton Harbour.
2002
2003
2004
17
6
258
WC
908
908
258
96.0
90.0
81.5
76.7
82.7
84.6
82.5
Herbivorous
95.9
89.3
81.0
75.4
79.3
83.3
81.3
Carnivorous
0.0
0.6
0.5
1.2
3.4
1.4
1.2
Cyclopoids
3.8
6.3
7.4
16.9
13.3
14.4
16.1
Calanoids
0.2
0.6
0.5
0.1
0.2
0.8
1.2
Veligers
1.4
3.1
10.6
6.4
3.8
0.1
0.2
Cladocerans
Total (mg m-3)
-2
Total (mg m )
10 334
4 934
4 462
2 207
3 250
3 676
2 569
15 501
29 602
107 100
7 724
45 493
51 462
61 658
82
Table 6. Rotifer total biomass (areal and volumetric), total number of taxa and density
(indivuduals·m-2) of common rotifers in Hamilton Harbour
2002
Taxa Name
HH17
-2
Areal Biomass (mg·m )
-3
Volumetric Biomass (mg·m )
Total No. Taxa
2003
HH6
HH258
HHWC
2004
HH908
HH908
HH258
8.06
64.59
136.56
44.62
82.23
78.59
85.36
5.37
18
10.76
16
5.69
15
12.75
12
5.87
14
5.61
16
3.56
17
6 154
61 538
164 103
35 897
47 863
11 966
81 239
3 077
30 769
41 026
35 897
79 772
35 897
54 159
-2
Density of Common Taxa (ind·m )
Asplanchna priodonta
Conochilus unicornis
Filinia terminalis
Kellicottia bostoniensis
Kellicottia longispina
Keratella cochlearis
Keratella cochlearis tecta
Keratella quadrata
Lecane spp
Ploesoma hudsonii
Ploesoma truncatum
Polyarthra dolychoptera
Polyarthra major
Polyarthra vulgaris
Pompholyx sulcata
10 769
0
0
21 538
0
0
27 080
0
7 692
287 179
0
0
0
108 319
0
0
107 692
54 159
0
82 051
166 154
1230 769 6 358 974
1 550 769 4 371 510
2 811 966 4 292 122
3 077
23 077
164 103
57 436
159 544
119 658
54 159
67 699
4 615
61 538
451 282
57 436
207 407
215 385
15 385
0
0
0
0
0
0
3 077
7 692
0
0
0
0
0
4 615
38 462
123 077
0
47 863
0
0
43 077
292 308
328 205
603 077
526 496
921 368
230 177
0
7 692
0
0
179 487
27 080
9 231
0
0
0
127 635
442 735
230 177
0
0
0
0
590 313
179 487
148 938
Synchaeta kitina
6 154
84 615
205 128
35 897
191 453
47 863
0
Synchaeta sp.
7 692
30 769
246 154
50 256
15 954
47 863
40 619
0
23 077
0
0
0
0
1 538
53 846
41 026
57 436
0
191 453
121 858
23 077
15 385
41 026
28 718
31 909
0
0
6 154
15 385
123 077
28 718
79 772
95 726
54 159
Synchaeta stylata
Trichocerca multicrinis
Trichocerca porcellus
other taxa
-2
Total Areal Density (ind·m )
313 846 1 984 615 8 656 410
-3
Total Volumetric Density (ind·m ) 209 231
330 769
360 684
2 563 077 6 477 493
732 308
462 678
5 408 547 5 591 944
386 325
232 998
83
2002
2003
17
2004
6
CCIW
WC
258
908
Fig. 1. Zooplankton sampling Stations in Hamilton Harbour for 2002, 2003, and 2004.
84
9000
A
Veligers
Harp
Calanoids
Cylopoids
Pred Clad
Other Herb Clad
Daphnia
Bosminids
8000
Biomass (mg m-2)
7000
A
6000
5000
4000
A
3000
A
Bay of Quinte
A
B
2000
B
1000
B
0
258
6
2002
17
908 WC
258 908
2003
2004
B
N
HB
C
2002-2004
Fig. 2. The seasonal weighted mean areal biomass in Hamilton Harbour for the dominant
zooplankton groups in 2002, 2003 and 2004. Mean biomass and standard errors are also
shown for the Bay of Quinte stations from 2002 to 2004 (B = Belleville, N = Napanee,
HB = Hay Bay and C = Conway). ‘A’ and ‘B’ above the bars denote comparable stations
between Hamilton Harbour and the Bay of Quinte.
85
2002
Veligers
Harpacticoids
Calanoids
Cylopoids
Pred Clad
Other Herb Clad
Daphnia
Bosminids
-3
Biomass (mg.m )
2500
Station 258
2000
1500
1000
500
30-Oct
16-Oct
2-Oct
18-Sep
4-Sep
21-Aug
7-Aug
24-Jul
10-Jul
26-Jun
12-Jun
29-May
15-May
0
-3
Biomass (mg.m )
2500
Station 6
2000
1500
1000
500
30-Oct
16-Oct
2-Oct
18-Sep
4-Sep
21-Aug
7-Aug
24-Jul
10-Jul
26-Jun
12-Jun
29-May
15-May
0
Station 17
-3
Biomass (mg.m )
2500
2000
1500
1000
500
30-Oct
16-Oct
2-Oct
18-Sep
4-Sep
21-Aug
7-Aug
24-Jul
10-Jul
26-Jun
12-Jun
29-May
15-May
0
Date
Fig. 3. May to October 2002 seasonal trends in volumetric dry biomass at stations 258,
6, and 17 in Hamilton Harbour. Bosminids include Bosmina and Eubosmina, Daphnia
includes D. galeata mendotae and D. retrocurva. “Other Herb Clad” represent the
remaining herbivorous cladocerans. “Pred Clad” are the predatory cladocerans.
.
86
2004
2003
Cylopoids
Pred Clad
Other Herb Clad
Daphnia
14-Oct
30-Sep
16-Sep
2-Sep
19-Aug
5-Aug
22-Jul
8-Jul
600
Date
14-Oct
30-Sep
16-Sep
2-Sep
19-Aug
5-Aug
22-Jul
8-Jul
24-Jun
10-Jun
27-May
0
Station 258
13-May
0
15-Oct
100
1-Oct
100
17-Sep
200
3-Sep
200
20-Aug
300
6-Aug
300
23-Jul
400
9-Jul
400
25-Jun
500
11-Jun
500
86
Station WC
28-May
24-Jun
2004
600
14-May
10-Jun
Bosminids
2003
Biomass (mg.m-3)
Calanoids
Station 908
13-May
0
15-Oct
0
1-Oct
100
17-Sep
100
3-Sep
200
20-Aug
200
6-Aug
300
23-Jul
300
9-Jul
400
25-Jun
400
11-Jun
500
27-May
Station 908
28-May
Harpacticoids
600
500
14-May
-3
Biomass (mg.m )
600
Veligers
Date
Fig. 4. May to October 2003 and 2004 seasonal trends in volumetric dry biomass at stations 908, WC and 258 in Hamilton
Harbour. Bosminids include Bosmina and Eubosmina, Daphnia includes D. galeata mendotae and D. retrocurva. “Other Herb
Clad” represents the remaining herbivorous cladocerans. “Pred Clad” are the predatory cladocerans.
-2
May-October Zooplankton Areal Production (g.m )
87
120
100
80
60
40
2002
2003
2004
20
0
0
5
10
15
20
25
Bottom Depth (m)
Fig. 5. Areal zooplankton production at different depths in Hamilton Harbour in 2002, 2003,
and 2004.
88
Production (x 103 mg m-2)
120
Veligers
Calanoids
Cylopoids
Pred Clad
Other Herb Clad
Daphnia
Bosminids
100
80
60
40
20
0
258
6
2002
17
908 WC
2003
258 908
2004
Fig. 6. Seasonal (May-October) areal production for the dominant zooplankton groups
in Hamilton Harbour at the stations sampled in 2002, 2003 and 2004.
B of Q
Mean Cladocern Length ( μ m)
500
B of Q
450
400
HH
500
500
450
450
400
350
300
300
300
250
250
250
200
200
200
150
150
150
100
100
100
50
50
50
350
0
258 6
17
2002
B
HB
B of Q
400
HH
350
0
HH
0
908 WC B
2003
HB
258 908 B
HB
2004
Fig. 7. June 1- October 6 time-weighted mean length of cladocerans at the Hamilton
Harbour (HH) stations (17, 6, 258, WC and 908) and upper Bay of Quinte (B of Q)
(Belleville = B and Hay Bay = HB).
89
Fig. 8. Vertical distribution of zooplankton biomass at station 258 plotted with oxygen levels
on July 9 and 23, September 4 and 17 2002 in Hamilton Harbour. Zooplankton samples not
taken at all depths - bars indicate depths analyzed.
90
Other
Pompholyx
Synchaeta
Trichocerca
Polyarthra
Keratella
Asplanchna
160
Biomass (mg m-2)
140
A
120
100
A
A
80
Bay of Quinte
B
60
A
40
B
20
B
0
258 6
2002
17
908 WC
2003
258 908
2004
B
N HB C
2002 - 2004
Fig. 9. Seasonal mean biomass (mg·m-2) of dominant rotifer genera in Hamilton Harbour
from 2002 to 2004. Mean biomass and standard errors are also shown for the Bay of Quinte
stations from 2002 to 2004 (B = Belleville, N = Napanee, HB = Hay Bay and C = Conway).
‘A’ and ‘B’ above the bars denote comparable stations between Hamilton Harbour and the
Bay of Quinte.
91
BENTHIC FAUNA IN HAMILTON HARBOUR: 2002 - 2003
Ronald Dermott and Robert Bonnell
Great Lakes Laboratory for Fisheries & Aquatic Sciences
Department of Fisheries & Oceans
867 Lakeshore Road, Box 5050
Burlington, Ontario. L7R 4A6
INTRODUCTION
Hamilton Harbour has had a history of poor water quality from eutrophication, low dissolved
oxygen levels, and contaminated sediments due industrial and municipal wastewater. In 1981 the
harbour was named an Area Of Concern having significant environmental degradation and
severe impairment of beneficial uses. Prior to 1964, untreated sewage was discharged into the
harbour. Secondary sewage treatment was begun in 1973, expanded in 1979 and improved in
1996 in order to removal much of the nitrate and phosphorus. Industrial loadings were also
reduced between 1967 and 1983, with phosphorus (P) and ammonia loadings from the major
steel plants declining by 93 % and 94 % respectively (Ontario Ministry of Environment (OME)
1985). A Remedial Action Plan (RAP) to restore the harbour has been developed by a
consultative process between government agencies, municipalities and local industries. One of
the goals is to restore the benthic community to one that does not diverge from unimpacted sites
of comparable physical characteristics, and also to reduce sediment-associated contaminants so
that toxicity is not higher than in control sediments (BARC 2006)
Since 1913, benthic fauna have been used as indicators of water quality and oxygen
concentrations. These relatively stationary organisms, over a time period of months to years,
integrate the chemical changes in the sediments and water column above them (Brundin 1958,
Weiderholm 1980). In 1964, Johnson and Matheson (1968) surveyed the macro-invertebrate
community and sediments of Hamilton Harbour, when they found no macro-invertebrates
present in approximately 2 km2 nearest the steel mills. Elsewhere, the benthic community was
dominated by the pollution tolerate oligochaete worm Limnodrilus hoffmeisteri, (Milbrink 1983)
and the sewage worm Tubifex tubifex was abundant at depths beyond 12 m. In 1984, the same
sites used by Johnson and Matheson were resampled by the Department of Fisheries & Oceans
(DFO) to examine what changes had occurred following the initial environmental improvements
to the harbour (Portt et al. 1988 unpublished report). At that time, no area of the harbour was
devoid of benthic macro-invertebrates, but oligochaetes remained the only invertebrates in the
deepest part of the harbour. Hanna (1993) found that in spite of the improvement in abundance
and species composition, the benthic community in 1989 continued to reflect eutrophic
conditions, with only oligochaetes and a few Crustacea present at sites deeper than 12 m during
summer.
92
A re-survey of the benthic community in the harbour was undertaken in 2002 and 2003
to examine what changes have occurred in response to the improved water clarity and reduced
nutrient concentrations following the implemented remedial actions. Also, invading zebra
mussels (Dreissena spp.) have colonized the docks and breakwalls in the harbour since 1990,
increasing water clarity and altering the community. In addition to resampling the soft bottomed
community at sites sampled previously, the current work also examined the benthic community
in shallow sandy habitats at depths < 3 m in conjunction with research on the shallow nearshore
fish habitat.
METHODS
A benthic survey of Hamilton Harbour was conducted in conjunction with plankton sampling in
2002 and 2003, no benthic samples were collected during the plankton surveys in 2004. In 2002,
the sites chosen had an offshore gradient from the shallow habitat, along the relatively
undeveloped north shore at La Salle Park, to the deep central part of the harbour (naturalized
beach - industrial harbour). The three plankton sites were at the DFO electrofishing transect site
17 in 1.5 m depth, site 6 at 6.0 m depth near La Salle Park, and at mid harbour Ontario Ministry
of Environment (OME) site 258 at a depth of 24 m (OME 1981, Poulton 1987), which is the
same as Johnson & Matheson's site 19. Benthic sampling was done at only two of these three
sites (site 17 and 258) in May, July and late September. A nearby site 10 at 7.4 m depth was
substituted for the plankton site 6 because historical benthic data existed at site 10 from 1964 and
1984 (Johnson and Matheson 1968, Portt et al. 1988, DFO unpublished data). The bottom fauna
was also sampled at an additional 6 nearshore sites during May, July and September 2002 to
examine the community at depths < 3 m (Fig. 1). These nearshore sites were between La Salle
Park and the Burlington sewage treatment plant (STP), of these, sites 10, 32 and 33A had been
sampled by Johnson and Matheson (1968). In September 2002, a spatial survey sampled another
5 deeper sites > 9 m with one replicate analyzed from these sites. These 5 sites had been sampled
by Johnson and Matheson (1968), and were between Willow Cove and along the harbour's south
shore including the heavily industrialized steel mills at the east end of the harbour. Thus in
September 2002, a total of 13 sites were sampled to give a spatial view of distribution and
biomass of the benthic community in the harbour (Table 1).
In 2003, only two sites were sampled for plankton and benthos. Both sites were located at
the west end of the harbour; the shallow nearshore site <4m depth was at Willow Cove (WC),
and the deeper Environment Canada site 908 at 14 m was halfway between Willow Cove and the
piers on the south shore. Benthic samples were collected at site 908 in May 2003, but they were
contaminated with oil. Site 908 is in a no anchor area (submerged oil pipeline), so it was decided
to shift the bottom sampling further west to the location of spatial site 3 that was used in 2002.
The new site was satisfactory and resampled again in July and October 2003.
The coordinates (based on GPS) for the benthic sites sampled in 2002-2003 are given in
Table 1. Wherever possible, site locations and sampling methods were matched to those used in
1964 (Johnson and Matheson 1968) to allow comparisons with historical densities. A large 9
93
inch Ekman (area 0.05 m-2) was used for sites with a soft bottom; a smaller 6 inch mini-Ponar
(area 0.023 m2) was used at sites with depths <5 m having sandy substrates. Between 1 to 3
replicates were collected at each site. During 2002, only one replicate was analyzed from each
site during each season. In 2003, all three replicate samples collected were analyzed each season
(Table 1). The collected sediment was screened through a 580 µm screening bucket (US
standard #30 mesh), as used by Johnson and Matheson (1968). Screened residues were
transferred into Mason jars and preserved using neutralized formalin (37% formaldehyde) with
added CaCO3. Based on the volume in the jars, the required volume of formaldehyde (100 %
neutral formalin) was added to result in a final concentration of 8 to 10 % formalin. Formalin
was used to fix and harden the oligochaete and flatworms in order to reduce their fragmentation
especially in the sandy sediments.
Within 2 weeks after collection, the preservative was changed to alcohol for long term
storage. The formalin was decanted through a 180 µm brass sieve (US standard #80 mesh), and
the sample rinsed with cold tap water. This finer mesh ensured no small organisms retained on
the 580 micron screen in the field would be lost during re-screening in the lab. The retained
residue, including organisms, was returned to the original field jar and re-preserved with 50%
isopropyl alcohol to which a small amount of Rose Bengal was added to stain organisms in order
to assist sorting.
Later when analyzed, the alcohol was poured through an 180 µm mesh screen and the
sample rinsed with tap water. This insured that alcohol-water convection currents would not
interfere with the sorting process. To separate the benthos into its constituents, 5-10 ml aliquots
were examined under a stereo dissecting microscope at 6X power. Any organisms present were
removed, identified to taxa (usually Family), and their numbers tallied using a 9 unit laboratory
counter.
The biomass (wet weight including shells) for each taxa was measured directly as
blotted wet weight after rinsing in distilled water. (Dermott and Paterson 1974, Dermott 1979)
The enumerated organisms, including fragments, were transferring into a drop of distilled water
on a clean Petri dish. A separate drop was used for each taxa that could be identified under the
dissecting microscope. The organisms were counted when being sorted. If the organisms could
not be identified under the dissecting scope they were placed in alcohol in covered dish until
examined in more detail. After counting, the organisms in the water drops were weighed. They
were removed from the water drop with forceps, blotted on filter paper to remove excess water,
transferred onto a tared weighing pan and weighed on an analytical balance to the nearest
0.05 mg. Surface tension in the water drop made picking up the clump of small organisms, such
as ostracods and nematodes, easier than if they were submerged in a vial or dish of water after
being counted. Distilled water was added to the drops to keep them from dehydrating until the
sample was weighted. Weights of small specimens whose combined weight was less than
0.05 mg, were calculated from average weights in samples which had sufficient numbers so that
their pooled weight was measurable. Density and total wet weight (including shells) of each
94
identified taxa were recorded. Biomass is reported as wet weight including shells except for the
areal weighted calculations which used shell-free biomass.
Organisms requiring further identification under a compound microscope, such as
oligochaetes and chironomids, were weighted to the level of family or subfamily prior to
mounting on microscope slides. Wet mounts in water were usually sufficient to identify the
chironomids to genus. Dreissena polymorpha and Dreissena bugensis were separated into
species based on shell shape. Newly settled Dreissena mussels <1 mm were identified only to
genus. After weighing, the organisms were then placed into labelled vials in 70 % ethanol that
contained about 1 % glycerine to prevent complete desiccation.
For each site, average density per sample, and wet biomass were calculated for the
organisms. Average annual density at the sites and standard errors were calculated where
multiple replicates or seasons were sampled. The average benthic community diversity was
calculated for each site based on the Margalef (1958) diversity index D calculated as: number of
species -1 / natural log (total number of individuals). It was assumed that a minimum of two
oligochaete taxa (Limnodrilus hoffmeister and immature with hairs) were present at all sites
where oligochaetes were found but not identified.
Area weighted biomass was calculated based on bottom area (103 m2) of the depth
contour zones of the harbour listed in chapter 1, and the annual average wet biomass during 2002
-2003 for the major taxa, oligochaetes, chironomids, Sphaeriidae, and Dreissena. The wet shellon biomass of the molluscs was converted to shell-free wet tissue by multiplying by the average
percent shell-free ratio for each family, 30 % for Sphaeriidae and 56 % for Dreissena. Total
Biomass in grams was then summed for all the contour zones in the harbour. Area weighted
biomass (wet shell-free g m-2) was then calculated by dividing Total Biomass by total area of the
harbour. Standard error and average were used to calculate the low error estimate (Ave. - S.E.).
RESULTS
The four nearshore transect sites 13A, 15A, 17 and 18 were located on sandy sediments at depths
ranging between 1.4 and 2.3 m (Fig. 1, Table 1). Sites 17 and 18, located near La Salle Park,
were well protected by the constructed habitat islands, and the north shore. Thus substrates at
these two sites contained more plant debris and macrophytes than did the substrate at exposed
sites 13A and 15. Substrates at the deeper sites were finer silts with higher organic content. A
large deposit of dead zebra mussel shells occurred at depths between 6 and 7.4 m off site 10.
Visible oil and tar residues occurred in the sediment grabs at the deepest site 19 in the middle of
the harbour and at sites 14 and 34 along the south industrial shore (Fig. 1). Due to the oil, benthic
sampling at site 908 was relocated to nearby site 3.
The highest density of benthic macroinvertebrates occurred at nearshore site 17 at 1.5 m
depth, located near La Salle Park. Lowest density in 2002 occurred at site 3 in the west end, and
95
density was also low at site 14 near the south shore. Total density ranged from 2,440m-2 to
38,835 m-2. The bottom fauna of Hamilton Harbour remained dominated by oligochaete worms
which were the only invertebrates found at several of the deeper sites. Oligochaete densities
ranged from 2,440 m-2 at site 3 to 29,233 m-2 at site 33 to (Table 2). Average density at site 908 /
3 increased to 27,182 m-2 in 2003 compared to only 2400 m-2 in 2002 but again oligochaetes
formed over 97 % of the invertebrates present. In 2003, both Pisidium sp. and chironomids were
present at low density at site 908 / 3, while in September 2002 they were absent at site 3 (Table
2).
Total wet biomass of the invertebrates ranged from 0.38 g m-2 at site 14 near the south
shore, to 120.9 g m-2 at site 10 (Table 3, Fig. 2). In spite of the large biomass of the shelled zebra
mussels, the non-Dreissena biomass in the harbour remained greater than 55 % of the total
benthic biomass present at all sites except 17, 18 and especially at site 10 where the biomass of
non-Dreissena invertebrates represented only 5 % of the total wet shell-on biomass. (Fig. 2). Site
908 / 3 had the next highest benthic biomass of 21.8 g m-2, which was due to the high numbers of
oligochaetes.
After oligochaetes, chironomids were the next most common invertebrates collected
from the harbour. Chironomid density was greatest at site 17, while biomass was greatest at site
WC (4.7 g m-2 wet). This biomass exceeded that of the oligochaetes at this site (Table 3).
Although seasonally chironomid density was greatest at WC in October (Tables 4 to 7 and 5),
their biomass was very constant over the summer, ranging from 4.78 g m-2 wet in May, 2003 to
4.79 g m-2 wet in October. As with zebra mussels, chironomid density was greatest in water less
than 7 m, the midges were rare or absent at sites deeper than 9 m. Seasonal biomass of the
chironomids was highest at 7 m in the spring before the larger midges (Chironomus plumosus
and C. attenuatus) emerged (Fig. 3). The chironomid biomass in the shallows was more
consistent over the summer as the smaller species present in the shallows often have two or more
generations a year, compared to one synchronous emergence in deeper water.
In 2003, total density and wet biomass were much greater at site 908 / 3 in July than at
nearshore site WC (Tables 4 to 7). As with other sites > 9 m in the harbour, tubificid
oligochaetes dominated the benthic fauna at site 908. Chironomids dominated the benthic fauna
at site WC, but oligochaetes were as important in the biomass of the October 2003 samples at
site WC (Tables 4 and 5). A few snails and amphipods were present at this 3.5 m deep nearshore
site. No insects other than chironomids were collected at site WC. Seasonally, biomass was
greatest in the fall at site WC (11.48 g m-2 ), but greatest in spring (26.10 g m-2) at the deeper site
908 / 3 (Tables 5 and 7). The higher biomass in May 2003 at site 908 was due to large maturing
tubificid oligochaetes, and a few Chironomids, especially Procladius sp. which mature and
emerge in July.
Zebra mussels (Dreissena polymorpha) dominated the Dreissena in the harbour during
2002. Quagga mussels (D. bugensis) represented less than 0.6 % of the mussels collected and
were present only at sites 10 and WC. The abundance of zebra mussels was inversely related to
96
depth with few sites beyond 8 m supporting live mussels (Fig. 4). However, mussel biomass was
greatest at site 10 in 7.4 m of water. At this site, average seasonal shell-on wet mussel biomass
exceeded 115.2 g m-2, in spite of the low density (306 m-2), because very large mussels were
attached to the dead shells present at that site (Tables 2 and 3). The estimated wet tissue of the
mussels at site 10 was 64.5 g m-2 (wet shell-free), assuming the ratio of wet shell-free tissue to
total wet weight including shells was 56 % as in eastern Lake Erie (Dermott et al. 1993). The
shell-free dry biomass equivalent of this wet tissue would be 7.80 g m-2 or 12.1% of the wet
shell-free biomass (Dermott et al. 1993). Wet shell-on biomass of the Dreissena at shallow sites
13A, 15, 17, and 18 averaged 3.23 g m-2 during 2002. This would represent an average of
1.81 g m-2 of wet tissue (shell-free).
The numerous small mussels at these four nearshore sites (< 2 m depth) were those that
had settled during the summer. The exposed nature of these shallow sites, unstable sandy
sediments, and risk of heavy ice scour at depths less than 2 m during winter would make these
sites unfavourable mussel habitat except during summer. The number of small mussels increased
rapidly over the summer at the shallow inshore sites (# 13A to 18), however weight gain was
limited (Fig. 5). At more protected site WC (3.2 m), the majority of the mussels settled between
July and October 2003. However, there was little increase in Dreissena biomass following the
settlement (Table 4 and 5). At the same time, a few Dreissena settled at site 908 / 3 (Tables 6 and
7). The 7 m depth zone (sites 10 and 32) had the greatest average wet weight, but the high
Dreissena biomass at site 10 in September 2002 was due to one sample containing a clump of
very large mussels (Fig. 5). Average Dreissena biomass in the harbour was much less than in
Lake Erie or similar habitats in the Bay of Quinte (Table 8).
Mussel biomass was greatest along the north shore and almost absent on the soft silty
bottom of the harbour (Fig. 6). The few Dreissena at site 19 in 2002 (Table 2) likely had fallen
from the mooring float at that site, rather than settled and grown on the soft anoxic sediments
during the summer. Although not sampled in 2002, populations of mussels occur on the rocky
fill along the west and east sides of the harbour, and along the docks and breakwalls of the
harbour. Unpublished data from midsummer 2002 indicated that zebra mussels were very
common along the dock at Canada Centre for Inland Waters (CCIW). During October 2002,
density of mussels which had settled in fish cages along the dock at CCIW over the summer was
7525 m-2, with a wet biomass including shells of about 1050 g m-2, representing settlement and
growth between June 1 and October 3, 2002. In mid summer 2003, mussels were very rare along
the rocky shoreline at the entrance to CCIW (43° 18.12'N: 079° 48.05'W) due to extensive
mortality during the previous cold winter. In November 2003, density of mussels (all less than 10
mm length) was about 100 m-2 on the rocks at the entrance to CCIW. Variability of the
settlement was high between sites and between years. In October 1991, a year after the arrival of
mussels in the harbour, density at the entrance to CCIW ranged from 71 to 3033 m-2, with a wet
biomass ranging from 117.6 to 434 g m-2 (including shells). In 1991, mussel density on the rocky
causeway over the outlet from the Burlington Sewage Treatment Plant (STP) was 29,890 with a
wet biomass of 6847 g m-2. In December 1992, this rocky causeway supported a Dreissena
population of 265,081 m-2 having a total wet biomass of 8694 g m-2. Of these, the density of
97
adult size mussels with shell length greater than 10 mm was only 1079 m-2, with a wet biomass
(including shells) of 2346 g m-2.
The amphipod Gammarus fasciatus, was collected only at shallow sites of less than
3.5 m depth (Table 2). No Echinogammarus ischnus were collected in sediment samples from
2002 or 2003, but they were abundant on rocky substrates and among the zebra mussels on the
rocks near CCIW (DFO unpublished). Specimens of the Lepidoptera Acentria were found only
at site 13A in July 2002. These were the only non-chironomid insects collected in the benthic
grabs during 2002 and 2003.
In addition to the worms, only a few small fingernail clams (Sphaeriidae) and encysted
Harpacticoida were present in the sediments from the hypolimnion of the harbour. The isopod
Caecidotea sp. and the Platyhelminthes were collected only at site 10, possibly associated with
the zebra mussel shells, although isopods and flatworms Dugesia sp., and Hydrolimax grisea are
common on the rocks near shore. Gastropods (Valvata sincera, and Pleurocera acuta) were
present at only 2 sites, #32 and WC, both < 8 m deep. Although gastropods are present on the
rocks along the shore (Physella sp., and Bithynia tentaculata), they were absent in the samples
on the sandy substrates from all sites < 3 m deep. Hydra were found only at sites less than 7 m
depth in 2002 but not at site WC in 2003 (Table 2 and 4). Water mites (Hydracarina) were
present to 9 m depth, but their small size added little to the macroinvertebrate biomass. One
specimen of the Hirudinea Mooreobdella fervida was collected at site 33A in May 2002. Like
the flatworms, the leeches Dina sp. and Glossiphonia sp. are present on the rocks along the shore
(DFO unpublished).
The calculated area weighted shell-free biomass as g m-2 is displayed in Table 9. As 62 %
of the harbour area is below 10 m, the average biomass present in the hypolimnion has a large
effect on the total biomass in the harbour. Not surprising, oligochaetes had the greatest area
weighted biomass in the harbour at 11.2 g m-2 (Table 9). However, Dreissena were the next most
important invertebrate with an area weighted wet biomass of 5.1 g m-2 shell-free. This area
weighted shell-free biomass would be about 9.1 g m-2 wet with shells, slightly less than the
estimated numeric average from all the samples of 10.1 g m-2 wet with shells (Table 8).
Chironomids, with a high density were minor components to the overall harbour biomass
(0.6 g m-2; Table 2 and 9). However, variability of Dreissena biomass was very high (range 0.33
to 9.90 g m-2), so their low estimate was less than the low estimate of the chironomid biomass
(0.35 g m-2).
In 2002 - 2003, the average Margalef (1958) diversity index was greatest at site WC
(2.55). Diversity at site 908 / 3 (0.99) was higher than at most of the other sites deeper than 9 m
in the harbour (Table 2). This consistent low diversity at the deeper sites reflects the limitations
put on the community by the low hypolimnetic oxygen levels, perhaps in combination with high
metal and polyaromatic hydrocarbons (PAH) levels in the sediments.
98
Historical Comparisons
Data from the same sites during September 1964 and 1984 (Johnson and Matheson 1968, Portt et
al. 1988) indicated that total macroinvertebrate density (pre-Dreissena) was lower in September
2002 than in 1984 at sites: 3, 6, 10, 14, and 19 (Table 10). This was due to greatly reduced
density of oligochaetes at all these sites, as well as site 34. Most sites had an increased density of
invertebrates between 1964 and 1984, followed by a general reduction by 2002. In 1964,
oligochaetes were the only invertebrates collected at several sites, while no invertebrates were
present at site 34 in 1964. Composition of the oligochaete species was not examined in 2002,
preventing comparisons of changes in species abundance since 1984. No worms of the family
Naididae were identified in 1964, but they were present in 1984 (Table 10). The Naididae are
more sensitive to low oxygen than are the pollution tolerant Tubificid worms. In 1984, almost all
the invertebrates collected were oligochaetes or chironomids, which remained the most abundant
benthic taxa in 2002. A comparison of the oligochaete abundance indicated a reduction in
density in the deeper parts of the harbour between 1984 and 2002 (Fig. 7). The number of
chironomid taxa increased between 1984 and 2002 at sites: 6, 10, 14, 19 and 32. In 1964, only 4
chironomid taxa were identified in the harbour, 13 taxa were present in 2002. For sites sampled
in 1964, 1984 and 2002, the largest number of chironomids were present at site 10 (173 per
Ekman).
Spatial Survey
Zebra mussels (Dreissena; with their shells on), represented more than 50 % of the total
invertebrate biomass at 3 sites along the north shore (10, 17, and 18). Chironomids were
important near the north and northeast shore of the harbour (sites 13, 15, 17, 18 and 32), and at
Willow Cove in the west. They were rare in the samples from the middle and southwest end of
the harbour. Oligochaetes formed over 90 % of the biomass at all sites beyond 8 m depth except
site 6 at 9.6 m deep located east of Willow Cove. The benthic community at all sites in the
middle of the harbour and along the industrial south shore was almost exclusively oligochaetes,
with a biomass above 18 g m-2 near the steel mills (Fig. 8).
Species diversity measured as the Margalef (1958) index D was below 2 at most sites in
the harbour. In 2002 - 2003, the highest diversity index was at site WC (2.55) in the west,
diversity was also high at sites near La Salle Park and the northeast corner of the harbour.
Diversity at site 908 / 3 (0.99) was higher than at most of the other sites deeper than 9 m in the
harbour (Table 2, Fig. 9). Most sites deeper than 9 m had diversity values less than 1 (Tables 2
and 10), indicating a restricted benthic community. This consistent low diversity at the deeper
sites reflects the limitations put on the community by the low hypolimnetic oxygen, perhaps in
combination with high metal and PAH levels in the sediments.
99
DISCUSSION
The community present at depths beyond 9 m suggests the benthic species are still restricted by
severe oxygen limitation during part of the year in spite of improved environmental conditions in
the harbour. Oligochaete density in 1964 was considerably lower than in 1984 or 2002. In 1964,
Johnson and Matheson (1968) found no organisms present at 6 sites. That year was a time of
high eutrophication and high metal levels, up to 25 % Fe2O3 in sediments of the harbour's southeast part, which prevented survival of even the sewage worms Tubifex tubifex and Limnodrilus
hoffmeisteri (Johnson and Matheson 1968). Following sewage treatment after 1973, oligochaete
populations increased in all parts of the harbour resulting in the highest populations in the time
series examined. Further improvements to the sewage treatment since 1984 and filtering by the
zebra mussels have reduced the amount of organic matter settling into the deeper parts of the
harbour. By 2002, this has reduced the density of oligochaete worms, which feed on the bacteria
and organic matter in the deeper sediments.
The increased number of chironomid taxa between 1984 and 2002 at 5 sites suggests
some improvements in water quality. The tribe Tanytarsini and many of the small Chironomini
species such as Cladopelma and Polypedilum appeared at several sites in 2002. These small
genera have higher oxygen requirements than most of the larger Chironomini genera, such as
Chironomus or Cryptochironomus (Brundin 1958). A total of 4 chironomid taxa were found in
1964, 8 in 1984 and 15 were present in 2002. No amphipods or gastropods were collected at any
of these common sites in any of the survey years, but all these 9 sites were deeper than 6 m.
During 1984, no samples other than in Windermere Basin were collected from depths
shallower than 4 m, so no comparison can be made for the benthic community along the northeast shore of the harbour where oxygen is rarely limited. In 1989, Hanna (1993) collected up to
15 invertebrate species from the area near La Salle at depths less than 5 m. This number was
similar to the number of taxa present in the summer of 2002 on the firmer nearshore substrates in
the same area. Hanna (1993) found that oligochaetes were the only class of invertebrates present
at depths beyond 12 m from June to August 1989. Ten oligochaete species were identified in
1964, 9 were found in 1984 and Hanna (1993) found only 7 with Limnodrilus claparedeanus,
L. udekemianus, Spirosperma ferox and Potomothrix moldaviensis being absent in 1989. No
worms of the family Naididae were identified in 1964, 1989 nor 2002, but they were identified in
1984. Worms in the family Naididae are less tolerant of eutrophication than are most of the
genera in the family Tubificidae.
In 1989, Hanna (1993) found that the Margalef species diversity averaged 1.4 at shallow
sites (< 8 m deep), and 0.5 at deeper sites in Hamilton Harbour. In comparison, the diversity of
the benthic community inhabiting the 5 to 7 m depth in the Big Bay portion of the eutrophic Bay
of Quinte averaged 1.11 (SE= 0.05) during 1966, 1.14 (SE=0.18) in 1985, and increased to 3.19
(SE= 0.06) in 2001. Benthic diversity in the Bay of Quinte at 20 m depth (Glenora site) was 2.99
(SE= 0.12) in 2001 (DFO unpublished data). Insects, gastropods, and amphipods were very rare
100
in the Hamilton samples including those from littoral sites < 2.5 m depth. Hanna (1993) also
found only one non-chironomid insect species in sites near the north shore, and that amphipods
and gastropods represented only 0.02 % and 0.01 % of the invertebrates respectively. The
periodic low oxygen episodes that occur in the nearshore as shallow as 3.5 m (Burley, this
volume) would restrict the distribution of the Ephemeroptera and Trichoptera in the nearshore
compared to the more typical community of the sublittoral zone as occurs in the Bay of Quinte.
Added to the limitations by poor water quality, intensive fish predation in the limited suitable
habitat may be reducing the benthic community in the shallows. Gobies (Neogobius
melanostomus) are the most common fish in the harbour and they have been accused of reducing
benthic populations in Lake Erie (Barton et al. 2005).
The composition of the benthic fauna in Hamilton Harbour remains constrained at depths
beyond 8 m, mainly due to the summer anoxia of the hypolimnion (Burley, this volume). Oxygen
concentrations below 1 mg l-1 restrict not only the fish but also the benthic invertebrates that can
survive in the middle of the harbour (Warren et al. 1973). In addition, levels of Cr, Cu, Hg, Pb
and especially Zn are elevated in the mid harbour and deeper west-harbour sediments (OME
1981, Krantzberg 1994, Jackson et al. 1995). These mid harbour sediments were shown to
reduce survival of the amphipod Hyalella in sediment assays (Munawar et al. 1999), and the
PAH contaminated sediments in the harbour induce strong genotoxic responses (Marvin et al.
2000). With time, contaminated sediments from 20 years ago should become buried under less
contaminated particles following the improvements to the waste water and storm drain
management. However continuing anoxic conditions at the sediment surface and re-suspension
of contaminants from shipping and storm disturbance (Rukavina and Versteeg 1996) may
continue to make the harbour sediments an unfavourable environment for the re-establishment of
a normal benthic community.
Mussel density in the harbour was less than densities in Lake Erie and the Bay of Quinte.
The large area of unsuitable, soft bottomed habitat beyond 8 m in the harbour limited the average
biomass in the harbour to about 1/10 the wet biomass that existed on comparable substrates and
depths in Lake Erie (Jarvis et al. 2000), in spite of the lower algal biomass in eastern Lake Erie.
The benthic community living above 9 m is also very limited in species composition.
Amphipods, gastropods and Turbellaria were very rare in samples from this part of the harbour.
For the littoral and sublittoral zones between of 1.4 and 5 m, the community is also severely
restricted. The only insect collected was the Lepidoptera Acentria, which was also collected by
Hanna (1993). Other insects, such as Trichoptera, Ephemeroptera, Odonata, and Coleoptera were
absent in any of the benthic samples collected in 2002 and 2003. As a result, average diversity in
the 1 to 8 m zone of the harbour was 1.98 (S.E = 0.2). Comparative Margalef diversity values of
over 2.5 exist in eastern Lake Erie (Dermott 1994), and up to 5.1 at 1 m depth in relatively
pristine Batchawana Bay (Dermott 1984). Even at 6 m depth in the Bay of Quinte, another Area
of Concern with eutrophication problems, the Margalef (1958) diversity index had increased
from less than 1 in 1982 to above 3.0 by 2000 (Dermott unpublished, Bay of Quinte Annual
101
Report 2002). In comparison, the total benthic biomass in the Bay of Quinte at Big Bay was less
than 5 g m-2 wet with shells, yet above 10 g m-2 in Hamilton (Fig. 2).
SUMMARY
A seasonal survey of the benthic fauna was conducted at 7 sites along the north side of the
harbour in 2002 (May, July, September), with additional samples collected along the south side
in September Oligochaetes dominated the fauna with densities from 23,880 at 1.7 m depth to
19,690 at 23 m in the middle of the harbour. Chironomids ranged from 410 to 6800 m-2 above 9
m depth, but were rare below 9 m. Zebra mussels on sand and silts above 9 m depth ranged from
7 to 6000 m-2. Other invertebrate types were rare in the anoxic sediments below 8 m depth.
Benthic Diversity ranged from 0.2 along the south side to a maximum of only 2.5 off the north
shore at 3 m depth. The benthic community present beyond 9 m still suggests severe oxygen
limitation during mid summer.
ACKNOWLEDGEMENTS
We wish to thank Peter Jarvis, Bianca Radix, Ling Ying Ong, and Andrea Bernard for assistance
with collecting the samples. Bill Morton processed most of the samples
102
REFERENCES
Barton, D.R., Johnson, R.A., Campbell, L., Petruniak, J., and Patterson, M. 2005. Effects of
round gobies (Neogobius melanostomus) on dreissenid mussels and other invertebrates in
eastern Lake Erie, 2002-2004. J. Great Lakes Res. 31 (Suppl. 2): 252-261.
Bay Area Restoration Council (BARC) 2006. Toward Safe Harbours 2006, Progress toward
delisting. Bay Area Restoration Council, Hamilton, Ontario. ISBN 0-9736190-2-3. June
2006, 50 p.
Brundin, L. 1958. The bottom faunistical lake type system and its application to the southern
hemisphere. Moreover a theory of glacial erosion as a factor of productivity in lakes and
oceans. Verh. Internat. Ver. Limnol. 13: 288-297.
Dermott, R.M. 1984. Benthic fauna assemblages in Batchawana Bay, Lake Superior. Can. Tech.
Rep. Fish. Aquat. Sci. 1265: 17 p.
Dermott, R. 1994. Benthic invertebrate fauna of Lake Erie 1979: Distribution, abundance and
biomass. Can. Tech. Rep. Fish. Aquat. Sci. 2018: 82 p.
Dermott, R. and Paterson, C.G. 1974. Determining dry weight and percentage dry matter of
chironomid larvae. Can. J. Zool. 52: 1243-1250.
Dermott, R., Mitchell, J., Murray, I. and Fear, E. 1993. Biomass and production of zebra
mussels (Dreissena polymorpha) in shallow waters of northeastern Lake Erie. p.399-413.
In: T.F. Nalepa and D.W.Schloesser (Eds.), Zebra Mussels: biology, impacts, and control.
Lewis Publishers. Boca Raton, Florida. 810 p.
Hanna, M. 1993. Benthic macroinvertebrate community structure in Hamilton Harbour from
June to August 1989. National Water Research Institute, Environment Canada, NWRI
Contrib. # 93-06: 46 p.
Jackson, M., Milne, J., Johnston, H. and Dermott, R. 1995. Assays of Hamilton Harbour
sediments using Diporeia hoyi (Amphipoda) and Chironomus plumosus (Diptera). Can.
Tech. Rep. Fish. Aquat. Sci. 2039: 20 p.
Jarvis, P., Dow, J., Dermott, R. and Bonnell, R. 2000. Zebra (Dreissena polymorpha) and quagga
mussel (Dreissena bugensis) distribution and density in Lake Erie, 1992 - 1998. Can.
Tech. Report of Fish. and Aquatic Science. No 2304. 46 p.
Johnson, M.G. and Matheson, D.H. 1968. Macroinvertebrate communities of the sediments of
Hamilton Bay and adjacent Lake Ontario. Limnol. Oceanogr. 13: 99-111.
103
Krantzberg, G. 1994. Spatial and temporal variability in metals bioavailability and toxicity of
sediments from Hamilton Harbour, Lake Ontario. Environ Sci Technol. 13: 1687-1698.
Margalef, R. 1958. Information theory in ecology. Gen Syst. 3: 36-71.
Marvin, C.H., McCarry, B.E., Villella, J., Alan, L.M., and Bryant, D.W. 2000. Chemical and
biological profiles of sediments as indicators of sources of contamination in Hamilton
Harbour. Part II: Bioassay-directed fractionation using the Ames Salmonella/microsome
assay. Chemosphere 41: 989-999.
Milbrink, G. 1983. An improved environmental index based on relative abundance of
oligochaete species. Hydrobiologia 102: 89-97.
Munawar, M., Dermott, R., McCarthy, L.H., Munawar, S.F. and van Stam, H.A. 1999. A
comparative bioassessment of sediment toxicology in lentic and lotic ecosystems of the
North American Great Lakes. Aquat. Ecosystem Health & Manag. 2(4): 367-378.
Ontario Ministry of Environment (OME) 1981. Hamilton Harbour Study 1977, Vol. 1. Water
Resources Branch, Ontario Ministry of the Environment. Toronto, 320 p.
Ontario Ministry of Environment (OME) 1985. Hamilton Harbour technical summary and
general management options. Great Lakes Section, Water Resources Branch, Ontario
Ministry of the Environment. Toronto, 125 p.
Portt, C.B., Cairns, V.W. and Minns, C.K. 1988. Benthic macroinvertebrates and sediment
characteristics of Hamilton Harbour in 1984. Great Lakes Laboratory of Fisheries and
Aquatic Sciences, unpublished report, Fisheries & Oceans, Burlington, Ontario. 31 p.
Poulton, D.J. 1987. Trace contaminant status of Hamilton Harbour. J. Great Lakes Res. 13:
193-201.
Rukavina, N.A. and Versteeg, J.K. 1996. Surficial sediments of Hamilton Harbour: physical
properties and basin morphology. Water. Qual. Res. J. Can. 31: 529-551.
Weiderholm, T. 1980. Use of benthos in lake monitoring. J. Water Poll. Control Fed. 52:
537-547.
Warren, C.E., Doudoroff, P. and Shumway, D.L. 1973. Development of dissolved oxygen
criteria for freshwater fish. US EPA, Ecol. Res. Ser. Report. EPA-R3-73-019. Washington
DC. 7 p.
104
Table 1. Site location, number of seasons sampled, and sampling devices used to examine
benthic fauna in Hamilton Harbour, 2003-2003. Spring, summer and fall seasons were May,
July and late September, with several sites sampled only in September 2002. Site 3 was
2
substituted for site 908 in July and September 2003. A 9 inch Ekman (0.05 m ) or 6 inch
2
mini-Ponar (0.023 m ) were used depending on substrate.
Site
Depth (m)
13A
15A
17
18
10
33A
32
4
6
3
34
14
19
-1.4
-1.9
-1.5
-2.3
-7.4
-5.6
-7.8
-9.2
-9.6
-12.2
-13.5
-14.0
-23.5
Site
Depth (m)
WC
908
03
-3.2
-14.6
-14.2
Sampling Year - 2002
Device
Seasons
mini-Ponar
mini-Ponar
mini-Ponar
mini-Ponar
Ekman
Ekman
Ekman
Ekman
Ekman
Ekman
Ekman
Ekman
Ekman
3
3
3
3
3
3
3
1
1
1
1
1
3
Sampling Year - 2003
Device
Seasons
mini-Ponar
Ekman
Ekman
3
1
2
Latitude
Longitude
43° 18.666'
43° 18.756'
43° 18.201'
43° 18.133'
43° 17.725'
43° 18.540'
43° 18.221'
43° 16.559'
43° 17.181'
43° 16.840'
43° 17.104'
43° 16.665'
43° 17.172'
079° 48.481'
079° 48.626'
079° 50.354'
079° 50.405'
079° 51.203'
079° 48.460'
079° 48.612'
079° 52.046'
079° 51.883'
079° 52.530'
079° 48.624'
079° 50.944'
079° 50.224'
Latitude
Longitude
43° 17.183'
43° 16.867'
43° 16.768'
079° 52.268'
079° 51.883'
079° 52.443'
105
Table 2. Average density (no. m-2) and S.E. of benthic invertebrates in Hamilton Harbour during 2002 (May, July, September) and 2003 for sites WC and
908 (May, July, October). Samples without S.E. were only sampled in late September, depths are in meters.
Non-Dreissena
Density
S.E.
14880
5683
32880 12176
13410
1279
15885
5786
31040
5890
19627
5164
20647
3969
12400
na
12868
na
2440
na
31120
na
4740
na
20007
6855
TOTALS
Site Depth Density
S.E.
13A
1.4
16035
6244
17
1.5
38835 14786
15A
1.9
13470
1277
18
2.3
18480
7865
33A
5.6
31047
5891
10
7.4
19933
5131
32
7.8
20680
3982
4
9.2
12400
na
6
9.6
15928
na
3 12.2
2440
na
34 13.5
31140
na
14 14.0
4740
na
19 23.5
20013
6854
3.2
14.2
Site Depth
13A
1.4
17
1.5
15A
1.9
18
2.3
33A
5.6
10
7.4
32
7.8
4
9.2
6
9.6
3 12.2
34 13.5
14 14.0
19 23.5
WC
908
3.2
14.2
15566
27182
Gastropoda
Density
0
0
0
0
0
0
7
0
0
0
0
0
0
6
0
4027
2534
13117
27180
Gammarus
Density S.E.
15
15
75
54
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
5
0
5
0
2641
2534
0
0
0
0
Isopod Harpacticoida
Density Density S.E.
0
90
90
0
0
0
0
30
30
0
45
45
0
627
323
13
5407
3326
0
3260
2609
0
420
na
0
148
na
0
0
na
0
14860
na
0
0
na
0
7
6.7
0
0
44
0
26
0
0
0
8414
26524
1775
2466
2449 1533
2
2
2376
0
73
2
Sphaeriidae
Density S.E.
0
0
0
0
45
26
0
0
140
76
73
55
227
87
0
60
na
0
500
na
0
287
66
733
462
242
218
Ave.
Chironomidae Acarina Lepidoptera H ydra Platyhelminthes No.
Species Diversity
Density S.E. Density
Density Density S.E.
Density
3075 1548
210
210
30
15
0
18
1.92
6795 3259
840
0
720
357
0
16
1.90
6405 1724
210
0
45
26
0
12
1.69
4065 2038
375
0
270
135
0
17
2.03
933
483
27
0
40
31
0
11
1.33
1460 1001
60
0
80
53
7
11
1.37
413
66
40
0
0
0
0
12
1.31
0
na
20
0
0
na
0
4
0.47
600
na
860
0
0
na
0
13
1.80
0
na
0
0
0
na
0
2
0.21
0
na
0
0
0
na
0
6
0.68
20
na
0
0
0
na
0
3
0.37
20
20
7
0
0
0
0
7
0.49
3652
67
726
15
5
2
0
0
0
0
0
0
0
0
19
9
2.55
0.99
105
WC
908
Nematoda Hirudinea Oligochaeta Dreissena spp.
D.
D.
Density S.E. Density Density S.E. Density S.E. bugensis polymorpha
165
123
0
11085 3997 1155 919
0
1155
570
60
0
23880 10813 5955 2984
0
5955
270
135
0
6405 2162
60
40
0
60
120
30
0
10695 3631 2595 2116
0
2595
33
18
7
29233 5220
7
7
0
7
0
0
0
12537 2757
307
33
0
226
13
13
0
16687 1177
33
18
80
33
0
na
0
11960
na
0
0
0
0
na
0
11200
na
3060
na
0
3060
0
na
0
2440
na
0
0
0
20
na
0
15740
na
20
na
0
20
0
na
0
4720
na
0
0
0
0
0
0
19687 6799
7
7
0
7
106
Table 3. Average wet biomass (mg m-2 with shell) and S.E. of benthic invertebrates in Hamilton Harbour during 2002 (May, July, September) and 2003 for
sites WC and 908 (May, July, October). Samples without S.E. were only sampled in late September, depths are in meters.
Site Depth
13A
1.4
17
1.5
15A
1.9
18
2.3
33A
5.6
10
7.4
32
7.8
4
9.2
6
9.6
3 12.2
34 13.5
14 14.0
19 23.5
3.2
14.2
Site Depth
13A
1.4
17
1.5
15A
1.9
18
2.3
33A
5.6
10
7.4
32
7.8
4
9.2
6
9.6
3 12.2
34 13.5
14 14.0
19 23.5
WC
908
3.2
14.2
9728
21805
Gastropoda
Biomass
0
0
0
0
0
0
33
0
0
0
0
0
0
68
0
1298
2429
8346
21804
Gammarus
Biomass S.E.
2
2
29
22
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
0
1
0
Nematoda Hirudinea Oligochaeta Dreissena spp.
D.
D.
Biomass S.E. Biomass Biomass S.E. Biomass S.E. bugensis polymorpha
3
2
0
2111
692 2445 2087
0
2445
6
2
0
3492
294 6714 3585
0
6714
3
2
0
1686
607
249
242
0
249
5
0
0
2349
738 3509 3300
0
3509
1
1
382
10263 2565
408
408
0
408
0
0
0
4661 1080 115241 104531
329
114912
1
1
0
6857 1847
45
41
0
45
0
na
0
7326
na
0
0
0
0
na
0
3622
na
2646
na
0
2646
0
na
0
2544
na
0
0
0
2
na
0
17286
na
32
na
0
32
0
na
0
360
na
0
0
0
0
0
0
18598 6175
2
2
0
2
980
2430
0
0
0
0
Isopod Harpacticoida
Biomass Biomass y S.E.
0
2
2
0
0
0
0
2
2
0
2
2
0
5
3
4
46
26
0
29
27
0
4
na
0
8
na
0
0
na
0
188
na
0
0
na
0
1
1
0
0
1
0
1
0
0
0
3132
21188
450
2475
1382
1
443
1
570
0
812
1
Sphaeriidae
Biomass S.E.
0
0
0
0
78
76
0
0
147
73
102
98
545 186
0
138
na
0
1160
na
0
286
56
444
449
136
133
Biomass
Chironomidae Acarina Lepidoptera H ydra Platyhelminthes No.
Species Diversity
Biomass S.E. Biomass Biomass mg m-2 S.E.
Biomass
896
442
84
146
3 2
0
18
2.39
1571
616
333
0
60 29
0
16
2.28
1655 1271
36
0
3 2
0
12
2.50
1643
613
192
0
23 14
0
17
2.48
1005
434
4
0
1 1
0
11
1.52
840
430
13
0
3 2
1
11
1.26
1664 1094
4
0
0 0
0
12
1.48
0
na
0
0
0 na
0
4
0.51
142
na
0
0
0 na
0
13
2.05
0
na
0
0
0 na
0
2
0.21
0
na
0
0
0 na
0
6
0.73
24
na
0
0
0 na
0
3
0.68
11
11
1
0
0 0
0
7
0.50
4700
155
726
46
3
1
0
0
0
0
0
0
0
0
19
9
-
106
WC
908
Non-Dreissena
Biomass
S.E.
3245
1173
6137
1129
3463
1678
4223
1128
11809
2128
5670
683
9133
1035
7334
na
4278
na
2544
na
18636
na
384
na
18896
6204
TOTALS
Biomass S.E.
5690
3058
12851
4551
3712
1914
7732
4120
12217
2505
120911 104021
9178
1031
7334
na
6924
na
2544
na
18668
na
384
na
18898
6203
107
Table 4. Average seasonal and annual density (m-2) of benthic fauna at site WC in western
Hamilton Harbour, 2003.
Site
Depth (m)
Date
WC
3.2
May 14
Ave.
S.E.
WC
3.7
July 22
Ave.
S.E.
WC
3.9
Oct 09
Ave.
S.E.
Average
Density
2003
Ave.
S.E.
TOTALS
non Dreissenids (total)
5485
5441
2769.6
2744.2
12158
12129
381.3
388.9
28336
21061
6344.6
3031.7
15326
12877
3938.4
2554.2
OLIGOCHAETA (total)
MOLLUSCA:
DREISSENIDAE:
Dreissena bugensis
Dreissena polymorpha
SPHAERIIDAE:
Pisidium sp.
Musculium securis
GASTROPODA:
Amnicola limosa
CRUSTACEA:
AMPHIPODA:
Gammarus fasciatus
CLADOCERA:
Ilyocryptus
HARPACTICOIDA:
2875
1457.0
8199
102.7
14168
1937.5
8414
1775.0
44
0
44
25.4
0.0
25.4
29
0
29
14.7
0.0
14.7
7275
7128
147
3389.8
3359.4
38.8
2449
2376
73
1553.4
1533.6
23.2
29
0
29.3
0.0
1320
0
415.9
0.0
851
0
322.7
0.0
733
0
242.3
0.0
15
14.7
0
0.0
0
0.0
5
5.5
15
14.7
0
0.0
0
0.0
5
4.9
0
0
0.0
0.0
59
0
38.8
0.0
0
132
0.0
50.8
19.6
44.0
14.9
26.4
CHIRONOMIDAE (total)
CHIRONOMINI:
Chironomus sp.
Chironomus anthracinus
Chironomus atritibia
Chironomus plumosus
Cladopelma
Cryptochironomus
Dicrotendipes
Endochironomus subtendens
Glyptotendipes
Microchironomus
Polypedilum halterales
Tribelos jucundus
TANYTARSINI:
Paratanytarsus
Tanytarsus
TANYPODINAE:
Procladius
2508
1298.6
2552
254.0
5896
889.8
3652
725.7
44
0
0
132
0
44
0
25.4
0.0
0.0
50.8
0.0
25.4
0.0
59
704
0
689
0
29
0
38.8
250.2
0.0
172.9
0.0
14.7
0.0
1980
1071
233
0
147
176
15
268.8
588.9
63.9
0.0
77.6
50.8
14.7
694
592
235
234
49
83
5
331.0
242.5
63.9
117.7
33.2
28.8
4.9
0
59
997
0
0.0
38.8
651.8
0.0
15
191
117
15
14.7
102.7
29.3
14.7
0
0.0
616
59
225.8
58.7
5
83
577
24
4.9
42.4
236.5
19.6
15
249
14.7
139.9
29
44
29.3
25.4
103
103
63.9
63.9
49
132
24.8
54.4
968
462.2
660
88.0
1393
271.6
1007
189.5
0
0.0
0
0.0
15
14.7
4.9
4.9
11.7
2.51
1.3
0.18
15.0
2.49
0.6
0.10
18.0
2.64
0.6
0.01
14.9
2.54
1.0
0.06
MISCELLANEOUS TAXA
HYDACARINA:
Number of species
Diversity Index
108
2
Table 5. Average seasonal and annual wet biomass (g m- wet+shells) of benthic fauna at site
WC in western Hamilton Harbour, 2003.
Site
Depth (m)
Date
WC
3.2
May 14
Ave.wt
S.E.
WC
3.7
July 22
Ave.wt.
S.E.
WC
3.9
Oct 09
Ave.wt.
S.E.
Average
Biomass
2003
Ave.wt. S.E.
TOTALS
non Dreissenids (total)
7.487
6.907
3.259
2.841
10.202
8.457
1.763
0.823
11.484
9.664
1.527
0.930
9.724
8.342
1.298
0.980
OLIGOCHAETA (total)
MOLLUSCA:
DREISSENIDAE:
Dreissena bugensis
Dreissena polymorpha
SPHAERIIDAE:
Pisidium sp.
Musculium securis
GASTROPODA:
Amnicola limosa
CRUSTACEA:
AMPHIPOD
Gammarus fasciatus
CLADOCERA:
Ilyocryptus
HARPACTICOIDA:
1.924
0.889
3.238
0.526
4.234
0.149
3.132
0.450
0.581
0.000
0.581
0.481
0.000
0.481
1.745
0.000
1.745
0.941
0.000
0.941
1.820
1.710
0.110
0.872
0.835
0.038
1.382
0.570
0.812
0.443
0.373
0.390
0.019
0.000
0.019
0.000
0.694
0.000
0.229
0.000
0.619
0.000
0.178
0.000
0.444
0.000
0.136
0.000
0.180
0.180
0.0
0.0
0.0
0.0
0.068
0.068
0.002
0.002
0.0
0.0
0.0
0.0
0.001
0.001
0.0
0.0
0.0
0.0
0.003
0.0
0.001
0.0
0.0
0.004
0.0
0.001
0.001
0.001
0.001
0.001
CHIRONOMIDAE (total)
CHIRONOMINI:
Chironomus sp.
C. anthracinus
C. atritibia
C. plumosus
Cladopelma
Cryptochironomus
Dicrotendipes
Endochironomus
Glyptotendipes
Microchironomus
Polypedilum halterales
Tribelos jucundus
TANYTARSINI:
Paratanytarsus
Tanytarsus
TANYPODINAE:
Procladius
4.78
2.22
4.522
0.920
4.796
0.743
4.700
0.726
0.378
0.0
0.0
2.687
0.0
0.109
0.0
0.218
0.0
0.0
1.283
0.0
0.059
0.0
0.116
1.753
0.0
1.858
0.0
0.067
0.0
0.066
0.642
0.0
0.527
0.0
0.035
0.0
1.355
0.974
1.253
0.0
0.132
0.324
0.021
0.327
0.279
0.266
0.0
0.070
0.095
0.021
0.616
0.909
1.253
1.515
0.044
0.167
0.007
0.221
0.324
0.266
0.564
0.030
0.052
0.007
0.0
0.028
0.239
0.0
0.0
0.020
0.158
0.0
0.026
0.065
0.028
0.029
0.026
0.044
0.006
0.044
0.000
0.000
0.145
0.119
0.054
0.119
0.009
0.031
0.137
0.049
0.009
0.014
0.057
0.039
0.003
0.150
0.003
0.082
0.015
0.015
0.015
0.007
0.038
0.035
0.024
0.027
0.019
0.066
0.010
0.032
1.173
0.606
0.550
0.105
0.400
0.142
0.708
0.217
0.0
0.0
0.0
0.0
0.010
0.010
0.003
0.003
MISCELLANEOUS TAXA
HYDACARINA:
109
Table 6. Seasonal average and annual density (m-2) of benthic fauna at site 908 (03) in western
Hamilton Harbour, 2003.
Site
Depth (m)
Date
HH-908
14.6
May 14
Ave.
S.E.
HH-03
14.2
July 22
Ave.
S.E.
HH-03
14.2
Oct 09
Ave.
S.E.
HH-03/908
Average Density
2003
Ave.
S.E.
TOTALS
non Dreissenids (total)
23060
23060
1518.8
1518.8
32487
32487
2219.0
2219.0
25753
25747
6822.2
6819.9
27100
27098
2538.9
2538.5
OLIGOCHAETA (total)
MOLLUSCA:
DREISSENIDAE:
Dreissena bugensis
Dreissena polymorpha
SPHAERIIDAE:
Pisidium sp.
Musculium securis
GASTROPODA:
Amnicola limosa
CRUSTACEA:
AMPHIPODA:
Gammarus fasciatus
CLADOCERA:
Ilyocrytus sp.
HARPACTICOIDA:
22807
1464.5
32233
2245.0
24533
6373.9
26524
2466.4
0
0
0.0
0.0
0
0
0.0
0.0
0
7
0.0
6.7
0
2
0.0
2.2
120
0
41.6
0.0
93
0
24.0
0.0
1173
0
433.5
0.0
462
0
217.9
0.0
0
0.0
0
0.0
0
0.0
0
0.0
0
0.0
0
0.0
0
0.0
0
0.0
20
0
11.5
0.0
100
0
30.6
0.0
7
0
6.7
0.0
42
0
17.5
0.0
CHIRONOMIDAE (total)
CHIRONOMINI:
Chironomus sp.
C. anthracinus
C. atritibia
C. plumosus
Cladopelma
Cryptochironomus
Dicrotendipes
Endochironomus
Glyptotendipes
Microchironomus
Polypedilum
Tribelos jucundus
TANYTARSINI:
Paratanytarsus
Tanytarsus
TANYPODINAE:
Procladius
113
24.0
60
11.5
27
17.6
67
15.6
13
6.7
40
20.0
0
0.0
18
8.5
0
0.0
0
0.0
20
11.5
7
4.7
0
0.0
0
0.0
0
0.0
0
0.0
0
0.0
0
0.0
7
6.6
2
2.2
0
0.0
0
0.0
0
0.0
0
0.0
13
0
6.7
0.0
0
0
0.0
0.0
0
0
0.0
0.0
4
0
2.9
0.0
87
26.7
20
11.5
0
0.0
36
15.6
0
0.0
0
0.0
7
6.7
2
2.2
9
1.134
0.0
0.011
8
0.947
0.6
0.075
7
0.883
8
0.988
0.5
0.058
MISCELLANEOUS TAXA
HYDRACARINA:
Number of species
Diversity Index
1.2
0.134
110
Table 7. Seasonal average and annual wet biomass (g m-2 wet+shells) of benthic fauna at site
908 (03) in western Hamilton Harbour, 2003.
Site
Depth (m)
Date
HH-908
14.6
May 14
Ave.wt
S.E.
HH-03
14.2
July 22
Ave.wt
S.E.
HH-03
14.2
Oct 09
Ave.wt
S.E.
TOTALS
non Dreissenids (total)
26.105
26.105
0.796
0.796
25.710
25.710
2.518
2.518
13.574
13.570
3.601
3.599
21.796
21.795
2.427
2.427
OLIGOCHAETA (total)
MOLLUSCA / BIVALVIA:
DREISSENIDAE:
Dreissena bugensis
Dreissena polymorpha
SPHAERIIDAE: (total)
Pisidium sp.
Musculium securis
GASTROPODA:
Amnicola limosa
CRUSTACEA:
AMPHIPODA:
Gammarus fasciatus
CLADOCERA:
Ilyocryptus
HARPACTICOIDA:
25.706
0.829
25.294
2.470
12.565
3.308
21.188
2.476
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.004
0.000
0.004
0.004
0.000
0.004
0.001
0.000
0.001
0.001
0.000
0.001
0.193
0.000
0.041
0.000
0.273
0.000
0.069
0.000
0.881
0.000
0.254
0.000
0.449
0.00
0.133
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.001
0.000
0.001
0.000
0.002
0.000
0.000
0.000
0.001
0.000
0.001
0.000
0.001
0.000
0.000
0.000
CHIRONOMIDAE (total)
CHIRONOMINI:
Chironomus sp.
C. anthracinus
C. atritibia
C. plumosus
Cladopelma
Cryptochironomus
Dicrotendipes
Endochironomus
Glyptotendipes
Microchironomus
Polypedilum halterales
Tribelos jucundus
TANYTARSINI:
Paratanytarsus
Tanytarsus
TANYPODINAE:
Procladius
0.205
0.086
0.141
0.046
0.119
0.119
0.155
0.046
0.092
0.050
0.115
0.054
0.000
0.000
0.069
0.028
0.000
0.000
0.000
0.000
0.101
0.101
0.034
0.034
0.000
0.000
0.000
0.000
0.019
0.019
0.006
0.006
0.009
0.006
0.000
0.000
0.000
0.000
0.003
0.002
0.104
0.046
0.027
0.014
0.000
0.000
0.044
0.021
0.000
0.000
0.000
0.000
0.004
0.004
0.001
0.001
MISCELLANEOUS TAXA
HYDRACARINA:
HH-03/908
Average Biomass
2003
Ave.wt
S.E.
111
Table 8. Comparative average density (no. m-2) of Dreissena spp. and wet biomass (g m-2 with
shells).
Density
Ave.
S.E.
Biomass
Ave.
S.E.
Location
Year
No. of sites
Hamilton Harbour
2002
13
1015.2
503.8
10.1
8.8
Lake Erie
1998
36
5731.0 1398.0
958
292
Upper Bay of Quinte *
1998
2000
64
96
4712.5
38865.0
n.a
n.a
981 402.1
2438.3 767.7
* area weighted average for Upper Bay of Quinte between Trenton and Telegraph Narrows.
112
Table 9. Area of depth zones (km2); number of samples; annual Average Biomass and Standard
Error (g m 2); calculated Total Biomass per area (g) and low error (Ave - S.E.); and Area Weighted
-2
Biomass (g m ) of major benthic groups in Hamilton Harbour 2002-2003. Values are all wet shellfree biomass without the shells of the mollusks.
Depth
Area
n
Average Biomass
gm-2
S.E.
Total Biomass
Ave x Area
S.E.-Low
Weighted Biomass
Ave
S.E.-Low
Oligochaetes
0-2
2-5
5 - 10
10 - 15
15+
Sum
2.4617
1.2369
4.2664
5.1119
7.8728
20.9496
9
6
11
6
3
2.409
2.741
6.546
10.345
18.598
0.386
0.392
1.154
5.215
6.175
5.9311
3.3900
27.9279
52.8802
146.4182
236.5473
4.9808
2.9056
23.0053
26.2236
97.8059
154.9210
11.291
7.395
Chironomids
Depth
0-2
2-5
5 - 10
10 - 15
15+
Area
2.4617
1.2369
4.2664
5.1119
7.8728
Sum
20.9496
n
9
6
11
6
3
gm-2
1.441
3.171
0.730
0.448
0.011
S.E.
0.183
1.528
0.303
0.373
0.011
Ave x Area
3.5478
3.9229
3.1153
2.2916
0.0840
S.E.-Low
3.0977
2.0326
1.8218
0.3870
0.0000
12.9617
7.3390
Ave x Area
0.0144
0.0824
0.7953
0.6167
0.6755
S.E.-Low
0.0000
0.0000
0.3972
0.1968
0.5430
2.1843
1.1369
Ave x Area
4.4514
1.7022
100.9772
0.0238
0.0088
107.1635
S.E.-Low
2.5969
0.9658
3.2836
0.0012
0.0000
6.8474
Ave
0.619
S.E.-Low
0.350
Sphaeriidae
Depth
0-2
2-5
5 - 10
10 - 15
15+
Area
2.4617
1.2369
4.2664
5.1119
7.8728
Sum
20.9496
n
9
6
11
6
3
gm-2
0.006
0.067
0.186
0.121
0.086
S.E.
0.006
0.067
0.093
0.082
0.017
Ave
0.104
S.E.-Low
0.054
Dreissena
Depth
0-2
2-5
5 - 10
10 - 15
15+
Sum
Area
2.4617
1.2369
4.2664
5.1119
7.8728
20.9496
n
9
6
11
6
3
-2
gm
1.808
1.376
23.668
0.005
0.001
S.E.
0.753
0.595
22.898
0.004
0.001
Ave
5.115
S.E.-Low
0.327
113
-
Table 10. Benthic density (per Ekman grab, 0.05 m 2) at 9 sites in Hamilton Harbour sampled in
late September 1964, 1984 and 2002 (also 2003 at Site 3). Depths are those in 1984. An Ekman
and 580 micron screen (#30 mesh) were used each year.
Site 3
NEMATODA:
HIRUDINEA
OLIGOCHAETA (total)
NAIDIIDAE
TUBIFICIDAE:
Immature with hairs
Immatures without hairs
Limnodrilus cervix
Site 4
1964
na
2003
0
1964
na
8.0 m
1984 2002
na
0
580
na
na
na
na
880
na
na
na
na
1326
na
na
na
na
220
0
560
18
0
90
20
48
380
23
110
122
na
na
na
na
L. claparedianus
Limnodrilus hoffmeisteri
Tubifex tubifex
Quistradrilus multisetosus
DREISSENIDAE (total)
Dreissena bugensis
Dreissena polymorpha
SPHAERIIDAE (total)
Pisidium sp.
Musculium partumeium
GASTROPODA:
AMPHIPODA:
HARPACTICOIDA:
CHIRONOMIDAE (total)
CHIRONOMINI:
Chironomus attenuatus
Chironomus atritibia
Cladopelma
Cryptochironomus
Endochironomus subtendens
Glyptotendipes polytomus
Polypedilum halterales
TANYTARSINI:
Cladotanytarsus
Paratanytarsus
Tanytarsus stellatus
ORTHOCLADIINAE:
Thienemanniella
TANYPODINAE:
Procladius
HYDRACARINA:
Number of species
598
na
na
na
na
1964
na
10.0 m
1984 2002
na
0
530
1140
11
0
240
0
388
670
18
28
28
90
4
36
0
7
200
4
560
na
na
na
na
0
0
0
0.1
0
28
0
0
0
9
0
153
0
0
0
0
0
0
0
0
0
0.1
23.1
23.1
0
0
0
0
0
0
0
0
0
0
0
0
0
3
3
153
3
3
0
0
0
0
0
0
0
3.3
0
0
0
0
21
0
0
3
0
1
7
30
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0.3
0.9
0
0
0.1
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
3
0
0
0
0
0
0
0
0
3
0
1
7
0
1
16
0
0
0
0
0
0
0
0.2
0
0
0
0
0
0
0
0
0
1
1
0
0
0
0
0
0
0
1.8
0.1
0
0
0
0
0
1
0
0
0
0
1
43
2
2
2
8
4
8
4
5
10
13
0.147
0.208
0.988
0.637
1.277
1.796
880
122
1353
533
1144
796
Diversity Index 0.157
TOTALS
Site 6
11.0 m
1984 2002
na
0
580
0.556 1.106 0.467
220
560
620
Continued on next page
114
Table 10 . Continued. Benthic density per Ekman grab (0.05 m-2) in 1964, 1984 and 2002.
NEMATODA:
HIRUDINEA:
OLIGOCHAETA (total)
NAIDIIDAE:
TUBIFICIDAE:
Immature with hairs
Immatures without hairs
Limnodrilus cervix
L. claparedianus
Limnodrilus hoffmeisteri
Tubifex tubifex
Quistradrilus multisetosus
DREISSENIDAE (total)
Dreissena bugensis
Dreissena polymorpha
SPHAERIIDAE (total)
Pisidium sp.
Musculium partumeium
GASTROPODA:
AMPHIPODA:
HARPACTICOIDA:
CHIRONOMIDAE (total)
CHIRONOMINI:
Chironomus attenuatus
Chironomus atritibia
Cladopelma
Cryptochironomus
Site 10
Site 14
Site 19
6.0 m
13.0 m
23.0 m
1964
na
1984
na
2002
0
1964
na
1984
na
2002
0
1964
na
1984
na
2002
0
230
na
na
na
na
970
na
na
na
na
403
na
na
na
na
100
na
na
na
na
1320
na
na
na
na
236
na
na
na
na
76
0
1480
0
0
0
17
0
0
0
0
4
5
0
65
2
0
0
258
1063
0
4
76
22
57
0
861
na
na
na
na
0
0
0
0
0
0
0
0
17
9
9
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
2
0
2
1
11
11
0
0
2
0
2
236
173
0
0
0
0
0
1
0
0
0
0
1
3
2
2
155
0
0
0
0
0
3
0
0
1
0
0
0
0
0
0
0
0
8
0
0
0
0
0
0
0
0
2
0
0
0
0
0
0
0
0
1
0
0
0
0
0
0
0
0
0
0
6
7
0
0
0
0
1
0
0
0
0
0
0
1
1
Endochironomus subtendens
Glyptotendipes polytomus
Polypedilum halterales
TANYTARSINI:
Cladotanytarsus
Paratanytarsus
Tanytarsus stellatus
ORTHOCLADIINAE:
Thienemanniella
TANYPODINAE:
Procladius
HYDRACARINA:
Number of species
Diversity Index
TOTALS
3
3
10
2
3
3
5
7
7
0.367
0.291
1.368
0.217
0.278
0.366
0.924
0.821
0.885
232
972
845
100
1320
237
76
1488
878
Continued on next page
115
-2
Table 10. Continued. Benthic density per Ekman grab (0.05 m ) in 1964, 1984 and 2002.
1964
NEMATODA:
0
HIRUDINEA:
OLIGOCHAETA (total)
170
NAIDIIDAE
na
TUBIFICIDAE:
na
na
Immature with hairs
na
Immatures without hairs
na
Limnodrilus cervix
L. claparedianus
na
Limnodrilus hoffmeisteri
na
Tubifex tubifex
Quistradrilus multisetosus
DREISSENIDAE (total)
0
Dreissena bugensis
0
Dreissena polymorpha
SPHAERIIDAE (total)
0
0
Pisidium sp.
0
Musculium partumeium
GASTROPODA:
AMPHIPODA:
HARPACTICOIDA:
0
CHIRONOMIDAE (total)
0
CHIRONOMINI:
0
Chironomus attenuatus
Chironomus atritibia
Cladopelma
0
Cryptochironomus
Site 32
Site 33
7.8 m
6.0 m
Site 34
12.0 m
1984
na
2002
2
1964
0
1984
0
2002
3
1964
0
1984
0
2002
1
na
na
na
952
na
na
na
na
56
0
1640
206
0
0
0
30
13
1847
na
na
na
na
1590
na
na
na
na
787
na
na
na
na
3
0
258
996
45
13
32
32
58
0
3
20
20
0
0
0
0
0
422
27
7
6
0
0
0
0
0
0
0
0
1
0
0
0
0
0
1
1
0
0
0
0
0
0
1
0
1
1
25
25
0
0
5
0
17
56
95
0
0
0
0
743
0
22
1
5
16
87
0
0
0
0
0
0
1
0
0
0
0
0
0
0
1
0
0
0
0
0
0
1
0
0
0
0
na
na
Endochironomus subtendens
Glyptotendipes polytomus
Polypedilum halterales
TANYTARSINI:
Cladotanytarsus
Paratanytarsus
Tanytarsus stellatus
ORTHOCLADIINAE:
Thienemanniella
TANYPODINAE:
Procladius
HYDRACARINA:
Number of species
Diversity Index
TOTALS
0
0
na
4
5
0
0
0
0
6
4
0
0
0
0
0
0
2
0
10
6
10
10
0
3
6
0.195
0.00
1.239
1.216
1.214
1.184
0.00
0.271
0.680
170
na
1431
61
1657
2006
0
1591
1557
116
15
17
La Salle
13
33 STP
32
18
10
6
258 19
WC
3 908
34
14
4
N
Bayfront
500
0
500
1000
meters
Biomass g m-2 (wet +shells)
Fig. 1. Site locations of benthic sampling during 2002 – 2003.
STP
Dreissena
other Inverts.
depth m
Site #
Big Bay
2002
Fig. 2. Average total wet benthic biomass (of all species) and non-Dreissena wet biomass
(with shells) in Hamilton Harbour between May and September 2002 -2003.
117
2.4
2.0
o
o
-2
Wet
wet weight
wt g mg2m
1.6
- oo- < <3m
3m
- x - 7m
7m
··++
·· 23m
23 deep
m deep
o
1.2
Chironomid wt.
0.8
0.4
0
May
July
Sept
35
Oligochaete wt.
30
25
20
15
10
5
o
0
o
May
July
o
Sept
Month
Fig. 3. Chironomid and oligochaete wet biomass (g m-2) at three depth zones in Hamilton
Harbour from samples collected in May, July and September 2002.
6000
z
Zebra mussel density 2002 - 2003
5000
Mussels m -2
4000
3000
zz
2000
1000
0
z
z
z
z
z z
z zz z
z
Depth m
Fig. 4. Depth distribution of Dreissena mussels (Z) in Hamilton Harbour during 2002 - 2003.
118
number m-2
Dreissena Density
0
0
0x
x
x
Wet Biomass g m-2
Biomass
0
x
x
x
0
0
x
0
Fig. 5. Seasonal density and wet biomass (g m-2 with shells) of Dreissena spp. in three depth
zones of Hamilton Harbour from samples collected in May, July and September 2002.
115.24 gm
2.44
6.71
0.25
3.51
0.41
0.045
2.64
1.41
0.45
0
0.002
0.03
0
Fig. 6. Average wet biomass (g m-2 with shells) of Dreissena spp. at the benthic sites between
May and September 2002.
119
Oligochaete Density
x
x
x
0
Olig. Worms m -2
x
0
x
x
x
0
0
0 0
0
0
0
+
0
x+
+ +
+
x
0
0
+
+
+
+
Depth m
Fig. 7. Average density of oligochaetes (May, July, September) in Hamilton Harbour during
2002, and their density at the same sites in 1964 and 1984.
Dreissena
5.69
3.71
Oligochaetes
Chironomids
12.17
12.85
Others
7.73
9.18
120.9
18.89
6.92
18.67
9.95
0.38
2.54
7.33
Fig. 8 Benthic biomass (g m-2 wet with shells) and composition of the major taxa groups in
Hamilton Harbour September 2002.
120
1.7 1.9
2.0
1.9
1.3
1.3
1.3
2.0
1.8
0.5
0.2
0.7
0.3
0.5
Fig. 9 . Average species diversity of the benthic fauna at the benthic sites in Hamilton Harbour
sampled between May and October 2002.
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