Current Status and Trends in Muskegon Lake, Michigan

J. Great Lakes Res. 34:169–188 Internat. Assoc. Great Lakes Res., 2008

Current Status and Trends in Muskegon Lake, Michigan Alan D. Steinman*, Mary Ogdahl, Richard Rediske, Carl R. Ruetz III, Bopaiah A. Biddanda, and Lori Nemeth Annis Water Resources Institute Grand Valley State University 740 West Shoreline Drive Muskegon, Michigan 49441 ABSTRACT. A long-term monitoring program was initiated in 2003 to determine the ecological status of Muskegon Lake, a Great Lakes Area of Concern. This paper presents data generated from the first 3 years of the monitoring program, discusses how the data are being used to establish and justify lake restoration targets, and assesses how water quality conditions have changed over time. Between 1972 and 2005, lake-wide averages of total phosphorus and soluble reactive phosphorus from the water surface have declined from 68 to 27 µg/L and from 20 to 5 µg/L, respectively. In addition, average chlorophyll a concentrations have declined from 25 to 6 µg/L over this period, while Secchi disk depths have increased from 1.5 to 2.2 m. Wastewater diversion, and perhaps dreissenid filtering activity, is most likely responsible for these changes. However, nitrate concentrations have increased from 70 to 270 µg/L over the same time period. During 2003–2005, phytoplankton abundance and fish catch were lower in the spring compared to the summer and fall. Microcystis was the most abundant phytoplankton genus; the fish community generally was dominated by round goby (Neogobius melanostomus) in spring and summer, and sunfishes (Centrarchidae) in the fall. Dreissenid abundance was highly variable over time, but densities were low relative to Saginaw Bay. Approximately 65% of the Muskegon Lake shoreline has been hardened (i.e., physically altered). Overall, the water quality of Muskegon Lake has improved over the past 30 years, but environmental challenges still exist, including contaminated sediments, loss of natural habitat, and invasive species. INDEX WORDS:

Muskegon Lake, Areas of Concern, monitoring, water quality.

thropogenic) stresses in Muskegon Lake and its watershed. Species of particular concern include the walleye (Sander vitreus), Chinook salmon (Oncorhynchus tshawytscha), steelhead (O. mykiss), and threatened lake sturgeon (Acipenser fulvescens). Habitat loss and degradation, invasive species, and changes to food resources have been implicated as factors contributing to fishery concerns (Clapp et al. 2001, MLPAC 2002). To delist an AOC, the U.S. Environmental Protection Agency requires that restoration targets be developed and achieved. Specific guidance for the State of Michigan has been developed by the Michigan Department of Environmental Quality (MDEQ 2006). Delisting targets have been approved by the MDEQ for five of the nine beneficial use impairments (BUIs) in Muskegon Lake. For example, restoration targets for the eutrophication BUI include surface total phosphorus and chlorophyll a concentrations of 30 and 10 µg/L, respec-

INTRODUCTION Muskegon Lake (W 86° 17′ 25.42″, N 43° 13′59.45″), a drowned river-mouth system that connects directly to Lake Michigan through a navigation channel, has a long history of industrial activity on its shoreline. The environmental impairments from these activities resulted in the lake being designated as an Area of Concern (AOC) in the Great Lakes. Although remnants of the industrial legacy still exist, Muskegon Lake is an important recreational and commercial fishery in the region (Alexander 2006). Upstream connections to the Muskegon River’s wetland complex and downstream linkages to the coastal zone of Lake Michigan help account for Muskegon Lake’s fisheries production (LaMP 2004). Currently, a number of important fish are undergoing ecological (likely an*Corresponding

author. E-mail: [email protected]

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FIG. 1. Sampling locations in Muskegon Lake. Circles = 2003–2005 water quality monitoring stations; triangles = 1972 EPA water quality stations (Freedman et al. 1979); closed squares = 2003–2005 fish sampling sites; open square = 2003–2005 Dreissena monitoring site. Inset map shows the location of Muskegon Lake in the lower peninsula of Michigan. tively. Because supporting data based on site-specific monitoring must be collected as an integral part of the delisting process (U.S. EPA 2001), a long-term, comprehensive environmental monitoring program was initiated in Muskegon Lake in 2003 (Steinman and Ogdahl 2004). The information generated from this monitoring effort is providing the scientific basis to assess the degrees to which 1) the restoration targets are being met and 2) the lake’s overall condition has recovered. In this paper, we discuss the initial results from the monitoring work, relate relevant information to existing BUI delisting targets, and assess environmental changes in Muskegon Lake over the past 30 years where data are available. METHODS Site Description and History Muskegon Lake covers ~17 km2 with mean and maximum depths of 7 and 23 m, respectively. Aver-

age hydraulic retention time is 23 days (Freedman et al. 1979). The lake is a drowned river mouth system with the Muskegon River serving as its primary inflow and a navigation channel to Lake Michigan being the primary outflow (Fig. 1). Muskegon Lake has experienced a long history of anthropogenic stress. Lumbering activity reached its peak in the mid 1880s, when 47 sawmills surrounded Muskegon Lake. In 1887, the sawmills cut 665 million board feet (Alexander 2006) and Muskegon became known as the “Lumber Queen of the World.” The lumber industry crashed soon thereafter due to the unsustainable harvesting practices of the time. By the mid-1900s, industrial activity along the lake’s shoreline included foundries, metal finishing plants, a paper mill, and petrochemical storage facilities. Based on the lake’s environmental conditions, the MDEQ and the Muskegon Lake Public Advisory Council (MLPAC) identified nine BUIs for Muskegon Lake (Table 1). A process was devel-

Muskegon Lake Status and Trends TABLE 1.

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Beneficial use impairments (BUIs) identified for Muskegon Lake.

BUI Identified in the 1987 Remedial Action Plan Update* 1. Restrictions on human consumption of fish and wildlife 2. Loss of fish and wildlife habitat 3. Degradation of fish and wildlife populations 4. Degradation of benthos 5. Restrictions on dredging

Rationale Elevated PCBs in carp and mercury in walleye and bass Fish tainting, loss of sport fish, proliferation of rough fish, anoxia, and contaminated sediments Low diversity, low numbers, dominance by oligochaetes, anoxia, and contaminated sediments Contaminated sediments Degradation due to contaminated sediments, poor water quality, and historic filling of the shoreline

Identified between 1994 and 2002 by the Muskegon Lake Public Advisory Council** 6. Degradation of aesthetics Frequent sewer breaks and pump failures 7. Beach closings Abandoned oil wells and Superfund sites 8. Eutrophication or undesirable algae Historic hypereutrophic conditions 9. Restrictions on drinking water consumption Metal debris, floating scum, oil slicks (groundwater contamination) *Wuycheck (1987) **MLPAC (2002)

oped to establish delisting targets that included both public participation and scientific peer-review (Fig. 2). First, criteria were established to help narrow the population of possible delisting target indicators to a manageable level. Examples of criteria that were used included data availability, data quality, and cost of collecting future data. Following this step, a list of indicators was identified by research scientists working with the MLPAC for each BUI (Fig. 2). These indicators were then vetted before members of the MLPAC and invited scientists at a public meeting, where members voted on what delisting target indicators they wanted to include. Using peer-reviewed literature and best professional judgment, scientists then gathered relevant data for each of the indicators (Fig. 2); at this point, the process could follow one of two paths. If there was adequate information available about the indicator, delisting targets were established by scientists and sent for peer review by expert scientists in the field. Alternatively, if inadequate information was available to set targets, information needs were identified so that delisting targets could be set in the future, and ultimately sent for peer review (Fig. 2). In the case of Muskegon Lake, there was sufficient information to set targets for five of the nine delisting targets; data are currently being collected for the remaining four BUIs (Table 2). After peer review, the delisting targets are approved at a MLPAC public meet-

ing, and submitted to the MDEQ for approval (Fig. 2). The five delisting targets were approved by the MDEQ in 2007.

FIG. 2. Flow chart showing process by which delisting target indicators are developed and ultimately approved.

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Delisting targets for beneficial use impairments in the Muskegon Lake Area of Concern.

BUI 1. Restrictions on human consumption of fish and wildlife

Delisting Targets No statistically significant difference in fish tissue contaminant concentrations (that cause fish consumption advisories) between Muskegon Lake and a control site.

2. Loss of fish and wildlife habitat

Not completed

3. Degradation of fish and wildlife populations

Not completed

4. Degradation of benthos

1) for non-contaminated tributaries: MDEQ Procedure #51 scores meet the standard for aquatic life in two successive monitoring cycles; 2) for contaminated tributaries: all remedial actions are completed; 3) for Muskegon Lake: sediment toxicity tests result in > 60% survival for amphipods; Hexagenia is present in river mouth littoral zone; % oligochaetes < 75%; Chironomidae density > 500/m2; Shannon-Weaver diversity > 1.5

5. Restrictions on dredging

Not completed

6. Degradation of aesthetics

Data for two successive monitoring cycles indicate there are no persistent, high levels of turbidity, color, oil films, floating solids, foams, settleable solids, suspended solids, or deposits

7. Beach closings

1) no waterbodies with the AOC are included on the list of impaired waters in the most recent Water Quality and Pollution Control in MI reports; and 2) contact advisories have not been issued due to sewage infrastructure failure for three consecutive years beginning in 2006

8. Eutrophication or undesirable algae

1) no waterbodies with the AOC are included on the list of impaired waters in the most recent Water Quality and Pollution Control in MI reports; and 2) the following average annual concentrations/values are achieved in Muskegon Lake for two consecutive annual monitoring events: surface TP = 30 µg/L; chlorophyll a = 10 µg/L; Secchi disk depth = 2.0 m; trophic state index = 50–55

9. Restrictions on drinking water consumption (groundwater contamination)

Not completed

Sampling Methods Data were evaluated for two periods of record: 1972 (representative of pre-diversion of wastewater to the treatment plant) and 2003–2005. The data from 1972 were based on studies by Freedman et al. (1979). Water samples were collected from one main station (106; Fig. 1), which was intended to be representative of the entire lake. An additional station (103; Fig. 1) was sampled to determine spatial variation in water quality. According to Freedman et al. (1979), dissolved oxygen (DO) was measured by Winkler titration, temperature with a standard mercury thermometer, pH with a standard meter, and specific conductance with a conductivity meter. Secchi disk readings were taken to estimate water

transparency. Chlorophyll a samples were collected and stored in 2-liter amber polybottles containing 5 mL of magnesium carbonate suspension, filtered through a 0.45–µm Millipore filter, and frozen until analyzed with a Turner model 110 fluorometer. Water samples for nutrient analysis were collected with Van Dorn bottles 1 m below the water surface and 1 m above the lake bottom, transferred to 1liter polyethylene containers, transported on ice, and frozen until analysis at the University of Michigan laboratory. Nutrient analyses followed standard methods (APHA 1971). With the exception of chlorophyll a, no systematic biological sampling was reported for this time period. Water quality sampling during 2003–2005 took

Muskegon Lake Status and Trends place during spring, summer, and fall at six sampling sites in Muskegon Lake (Fig. 1): CHAN, near the channel outlet to Lake Michigan; BEAR, near the inflow from Bear Lake; DEEP, at the deepest point in Muskegon Lake; RUDD, near the inflow from Ruddiman Creek; AWRI, just lakeward from the Annis Water Resources Institute; and MUSR, near the inflow from the north branch of the Muskegon River. At each site, a Hydrolab DataSonde 4a was used to measure depth, DO, pH, temperature, and specific conductance at 1 m below the water surface and 1 m above the sediment surface. A Secchi disk was used to measure water transparency. Water samples were filtered (Whatman GF/F) for chlorophyll a analysis and frozen until analysis (within 4 weeks) using a Shimadzu UV1601 spectrophotometer following Standard Methods (APHA 1998). Water samples for nutrient analysis were collected 1 m below the water surface and ~1 m from the lake bottom with a Van Dorn bottle, stored in acid-washed bottles, and put on ice until delivery to the laboratory, always within 5 hr of collection. Soluble reactive phosphorus (SRP) and total phosphorus (TP) were analyzed on a Bran+Luebbe Autoanalyzer III (U.S. EPA 1983). Nitrate was analyzed by ion chromatography on a Dionex DX500 (APHA 1998). Ammonia analysis was conducted by the automated phenate method (APHA 1998) on a Bran+Luebbe AutoAnalyzer III. Chemical constituents with concentrations below detection limits were assigned a value of one-half the detection limit. QA/QC procedures followed method guidelines including 10% method blanks and 10% matrix spikes/matrix spike duplicates (± 15% limits for precision and accuracy). Phytoplankton samples were collected from the near surface at each site using Van Dorn bottles. Subsamples (100 mL) were fixed with either formalin or Lugol’s solution (1%). Phytoplankton species were identified and enumerated utilizing a Nikon Eclipse TE200 inverted microscope (Utermöhl 1958). Most of the identifications were made using magnifications of 450 and 1,000× with phase contrast illumination. In all the samples, 200–300 algal units (cells or filaments) were counted. The cell volume of each species was calculated by applying the appropriate geometric formulae (Hillebrand et al. 1999). For the determination of heterotrophic bacterial abundance, a 5-mL aliquot of each water sample was preserved with 2% final concentration of 0.2 µm-filtered formaldehyde. Sub-samples (0.25–1.0 mL) were acridine orange-stained, filtered onto 0.2

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µm black Millipore polycarbonate filters and frozen (–80°C) until observation by epifluorescence microscopy (Hobbie et al. 1977). Between 20–40 fields of view and a minimum of 300 cells were examined for every sample. Dreissena populations were measured at one site in Muskegon Lake (Fig. 1) using artificial substrates (cf. Nalepa et al. 1995). Artificial substrates were chosen because they facilitate multi-year population density comparisons and because Muskegon Lake is dominated by soft sediments. Twelve unglazed clay tiles (15 cm × 15 cm, with a smooth and a grooved side) were held vertically in the water column with the use of a PVC pipe frame. Vertical orientation of the tiles prevented preferential colonization of the upper surface (Nalepa et al. 1995); tiles were placed at 90° angles to one another to minimize the effect of directional currents on settlement. The structure was placed in Muskegon Lake behind the Annis Water Resources Institute at a depth of ~3 m (Fig. 1). The tiles were randomly assigned to four different year classes, each with three replicates: 1-year, 2-year, 3-year, and 4-year. The 1-year tiles were placed in the water for one full growing season (early spring to late fall, both in 2004 and 2005). The 2-year tiles remained in the water through two full growing seasons (early spring 2004–late fall 2005), including the winter in between. The 3-year tiles were removed from the lake at the end of the growing season in 2006, whereas the 4-year tiles will be removed from the lake at the end of the growing season in 2007. Data from the 3-year tiles are not included in this study. Upon removal, tiles were individually placed in plastic bags and frozen (Nalepa et al. 1995). Mussels were removed from the tiles and counted under 8× magnification. Shell lengths were measured to the nearest 0.01 mm using calipers and placed into size classes based on 1-mm increments. Mussels with shells < 5 mm were not measured and were lumped into one size class. Biomass estimates were made by determining the relationship between shell length and soft tissue weight (Nalepa et al. 1995). From each size class, five mussels were selected and their soft tissues placed in pre-combusted (550°C for 1 h), preweighed aluminum weigh dishes. They were dried at 60°C for at least 48 h, weighed, ashed at 550°C for 1 h, and re-weighed to determine ash-free dry mass (AFDM). The AFDM and shell length data were combined for all tiles collected from 2004–2005, converted to natural logs, and a linear regression equation was determined. Total biomass

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was then determined by calculating the AFDM of the median shell length in a category and multiplying by the number of individuals in that category (Nalepa et al. 1995). Individuals in the < 5 mm size class were assigned a shell length of 2.5 mm. For these data, no attempt was made to differentiate Dreissena species. Fish were sampled at four locations within the littoral zone of Muskegon Lake (Fig. 1) three times per year from 2003–2005 concurrent with water quality sampling. Using the approach described by Ruetz et al. (2007), three fyke nets (4-mm mesh; see Breen and Ruetz [2006] for detailed description of nets) were set during daylight hours at each site in areas with depth < 1 m: one net was positioned perpendicular to the shoreline with the net’s mouth toward shore and two nets were positioned parallel to the shoreline with leads fished end to end. About 25–50 m were between perpendicular and parallel fyke nets, and fyke nets were fished for about 24 h (i.e., one net night). Fish were identified to species, enumerated, measured for total length (TL), and released. A survey was conducted to assess the percent of hardened (i.e., replacement of natural shoreline with seawall or riprap) shoreline of Muskegon Lake. We acquired the latest available aerial imagery of the entire Muskegon Lake shoreline area that was in a digital, ortho-rectified, Geographic Information System (GIS) format for use in ArcView 3.3 GIS software (1998 digital orthophotographs from the National Aerial Photography Program taken during spring season [image acquisition scale: 1:40,000]). The shoreline vector data layer was digitally edited to match the 1998 digital orthorectified Muskegon Lake shoreline profile by superimposing the shoreline vector data layer on top of the 1998 ortho-rectified aerial photographs in ArcView GIS 3.3. Field validation of shoreline structures was performed by boat in August, 2004. The shoreline was divided into 50-m reaches, and classified as either natural or hardened, with the hardened class further subdivided into finer categories. A laptop computer provided real-time access to the GIS laboratory-interpreted shoreline vector data layer and the aerial photographs, to ensure accurate geographic placement of the field survey throughout the survey. One-way analysis of variance (ANOVA) was used to determine if there were significant differences between sampling dates (1972 vs. 2003–2005 or among seasons for 2003–2005 data) for most variables. Phytoplankton and fish data were ana-

lyzed with a three-way ANOVA, using year, season, and taxa as main factors. For the seasonal comparisons (2003–2005), we assumed that sampling events were independent of one another given that hydrologic retention times were shorter than the time intervals of sampling. Where transformation did not successfully result in normal distribution of the data (as determined by the KolmogorovSmirnov test), the Kruskal-Wallis test was used. Statistically significant results were analyzed with either the Tukey (for equal sample sizes) or Dunn (for unequal sample sizes) test to make pairwise comparisons. RESULTS 1972 and 2003–2005 Differences Soluble reactive phosphorus (SRP) was measured only once in 1972 (prior to wastewater diversion to the Muskegon County Waste Water Management System), in November, at station 106 (Fig. 1). SRP concentration declined from 20 µg/L at the nearsurface in 1972 to a mean of 5 µg/L in 2003–2005 (Fig. 3A). Near-bottom SRP concentration declined from 13 µg/L to a mean of 10.5 µg/L between 1972 and 2003–05 (Fig. 3A). The single sampling date in 1972 precludes inferential tests to determine if these declines are statistically significant. Mean total phosphorus (TP) concentrations declined significantly between 1972 and 2003–05. For near-surface waters, mean TP concentrations declined from 58 µg/L to 26 µg/L (Kruskal-Wallis: H = 15.72, P < 0.001; df = 1; Fig. 3B), whereas for near-bottom waters mean TP concentrations declined from 80 µg/L to 26 µg/L (H = 8.59, P = 0.003, df = 1; Fig. 3B). Mean nitrate concentrations at both the near-surface and near-bottom increased 4- to 5-fold between 1972 and 2003–2005, which were both statistically significant (H = 10.27, P = 0.001, df = 1 and F1,58 = 10.05, P = 0.002, respectively; Fig. 4A). Mean ammonia concentrations at the near-surface did not change significantly between 1972 and 2003-2005 (H = 0.01, P = 0.91, df = 1; Fig. 4B), but ammonia levels near the bottom declined ~7-fold, which was highly significant (H = 15.19, P < 0.001, df = 1; Fig. 4B). Mean near-surface water chlorophyll a (chl a) concentrations declined significantly from 24.7 µg/L to 5.9 µg/L between 1972 and 2003–05 (H = 5.4, P = 0.024, df = 1; Fig. 5A). The mean Secchi disk depth increased from 1.5 to 2.3 m (~50%) from 1972 to 2003–2005 (Fig. 6A). Data for individual

Muskegon Lake Status and Trends

FIG. 3. Historical comparison of mean (+SE; except 1972 SRP, n = 1) phosphorus concentrations in Muskegon Lake: 1972 (Freedman et al. 1979) vs. 2003-2005. A) soluble reactive phosphorus. Only one SRP observation was available from 1972, so no inferential statistical analyses were conducted; B) total phosphorus. Results from a Kruskal-Wallis statistical test are provided for differences in TP concentration between time periods. See text for more details.

sites from 1972 were not available, so inferential statistics could not be conducted to determine if this increase in water clarity was statistically significant. 2003–2005 Differences TP concentrations in the near-surface waters during 2003–2005 ranged from below detection to 60 µg/L (Fig. 7A). When pooled across years and analyzed by season, summer surface TP was significantly greater than either spring or fall

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FIG. 4. Historical comparison of mean (+SE) nitrogen concentrations in Muskegon Lake: 1972 (Freedman et al. 1979) vs. 2003–2005. A) nitrate; B) ammonia. Results from Kruskal-Wallis statistical tests are provided for differences in nitrate and ammonia concentrations between time periods. See text for more details. concentrations (H = 28.98, P < 0.001, df = 2), and there was no significant difference between spring and fall. TP concentrations were slightly greater in near-bottom waters than near-surface waters, ranging from 10 µg/L to 80 µg/L (Fig. 7B). When pooled across years and analyzed by season, nearbottom TP was significantly greater in summer than in spring (H = 12.27, P = 0.002, df = 2), and no significant differences were detected between summer and fall, or between spring and fall. Near-surface nitrate concentrations ranged from 20 to 580 µg/L during 2003–2005 (Fig. 8A). When pooled across years and analyzed by season, spring near-surface NO3 was significantly greater than either summer or fall concentrations (H = 32.96,

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FIG. 5. Chlorophyll a concentrations in Muskegon Lake. A) historical comparison of mean (+SE) surface concentrations in 1972 (Freedman et al. 1979) vs. 2003–2005. Results from a Kruskal-Wallis statistical test are provided for the difference between 1972 and 2003–05; B) surface concentrations measured at the 6 long-term monitoring sites from 2003 to 2005. Analysis of variance results are provided for the differences among seasons. * denotes missing data point (AWRI, spring 2004). See text for more details.

Muskegon Lake Status and Trends

FIG. 6. Secchi disk depths in Muskegon Lake. A) historical comparison of mean (+SE) depths in 1972 (Freedman et al. 1979) vs. 2003–2005. Only one Secchi disk depth observation was available from 1972, so no inferential statistical analyses were conducted; B) depths measured at the six long-term monitoring sites from 2003 to 2005. Results from a Kruskal-Wallis statistical test are provided for the differences among seasons. See text for more details.

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FIG. 7. Total phosphorus concentrations in Muskegon Lake from 2003–2005. A) nearsurface concentrations; B) near-bottom concentrations. Results from Kruskal-Wallis statistical tests are provided for the differences among seasons. See text for more details.

Muskegon Lake Status and Trends

FIG. 8. Nitrate concentrations in Muskegon Lake from 2003–2005. A) near-surface concentrations; B) near-bottom concentrations. Results from Kruskal-Wallis statistical tests are provided for the differences among seasons. See text for more details.

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FIG. 9. Ammonia concentrations in Muskegon Lake from 2003–2005. A) near-surface concentrations. Results from a Kruskal-Wallis statistical test are provided for the differences among seasons; B) near-bottom concentrations. Analysis of variance results are provided for the differences among seasons. See text for more details.

Muskegon Lake Status and Trends P < 0.001, df = 2), and there was no significant difference between summer and fall. Nitrate concentrations were slightly greater in near-bottom waters than in near-surface waters, ranging from 10 to 700 µg/L (Fig. 8B). When pooled across years and analyzed by season, near-bottom TP was significantly greater in spring than in summer or fall (Tukey MCT; P < 0.001), and summer was significantly greater than fall (Tukey; P = 0.022). Near-surface ammonia concentrations ranged from below detection to 90 µg/L during 2003–2005 (Fig. 9A). When pooled across years and analyzed by season, there were no statistically significant differences detected among seasons (H = 0.34, P = 0.845, df = 2). Mean (± 1 SE) ammonia concentrations (µg/L) were very similar between the nearsurface (40 ± 3) and near-bottom (42 ± 4) waters (Fig. 9B). There were no significant differences among seasons for near-bottom ammonia concentrations when pooled across years (F2,53 = 0.50, P = 0.612). During 2003–2005, chl a concentrations in the near-surface waters ranged from below detection to 16.0 µg/L (Fig. 5B). Temporal variation was evident; summer chl a values were relatively low in 2004, a period marked by unusually cool temperatures and increased cloud cover. When pooled across years and analyzed by season, both fall and summer near-surface chl a levels were significantly greater than spring concentrations (F2,47 = 16.05, P < 0.001), and there was no significant difference between summer and fall. During 2003–2005, Secchi disk readings ranged from 1.5 to 4.7 m (Fig. 6B). When pooled across years and analyzed by season, there was a marginally significant difference in Secchi disk values (H = 5.81; P = 0.064, df = 2), with median readings in spring (2.6 m) greater than summer (2.25 m) or fall (2.2 m). Mean bacteria density (Fig. 10) ranged from < 0.1 × 10 6 (April 2004 at BEAR) to 1.6 × 10 6 cells/mL (July 2005 at MUSR). When pooled across years, season had no significant difference on bacterial density (H = 0.42; P = 0.812, df = 2). During 2003 to 2005, phytoplankton biovolume (Fig. 11A) collected from the near-surface ranged from 0.32 × 106 (spring 2003 at CHAN) to 12.80 × 10 6 µm 3/mL (fall 2004 at BEAR). When pooled across years, phytoplankton biovolume was significantly lower during spring compared to either summer or fall (F2,53 = 51.50; P < 0.001); there was no significant difference between summer and fall (P > 0.05). Similar seasonal trends were present in

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FIG. 10. Mean (+SD) seasonal bacterial densities in Muskegon Lake from 2003–2005. Results from a Kruskal-Wallis statistical test are provided for the differences among seasons. See text for more details.

the chlorophyll a data (Fig. 5B). Species richness ranged from 13 (fall 2003 at DEEP) to 47 (fall 2005 at DEEP). Similar to biovolume, phytoplankton species richness was significantly lower in spring than either summer or fall (Fig. 11A; H = 11.33; P = 0.003; df = 2). Cyanobacteria were the dominant phytoplankton division in Muskegon Lake during 2003 to 2005 (Fig. 11B). Mean biovolume for cyanobacteria (2.52 × 106 µm3/mL) was 4- to 6-fold greater than diatoms (0.65 × 106 µm3/mL), cryptophytes (0.52 × 106 µm3/mL), and all other taxa lumped together (0.43 × 106 µm3/mL), which was statistically significant (F 3,215 = 3.69; P = 0.013). Among the cyanobacteria, the four most dominant genera were Microcystis, Aphanocapsa, Anabaena, and Aphanizomenon (Fig. 11C). Dominance varied by season, although overall Microcystis was by far the most dominant genus in Muskegon Lake and had more biovolume than any other genus (F3,215 = 88.74; P < 0.001); its mean biovolume was 1.86 × 10 6 µm3/mL, which was an order of magnitude greater than the mean biovolumes of 0.12, 0.49, and 0.08 × 106 µm3/mL for Aphanocapsa, Anabaena, and Aphanizomenon, respectively. Dreissena density and biomass was highly variable between the two years that the mussels were harvested from artificial substrates. Dreissena density declined significantly (F1,5 = 11.99; P = 0.026) from ~53,600 individuals/m2 in 2004 to ~8,000 in-

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FIG. 12. Mean (+SE) density and biomass of Dreissena spp. harvested from artificial substrates in Muskegon Lake in 2004 and 2005. Exposure times for each sampling period are given in parentheses. Analysis of variance results are provided for differences in density and biomass among years. See text for more details.

FIG. 11. Phytoplankton community characterization in Muskegon Lake from 2003–2005. A) mean (+SE) total biovolume and total species richness; B) mean (+SE) algal division composition; C) mean (+SE) Cyanobacteria genera from 20032005. Text boxes show the results of statistical tests. See text for more details.

dividuals/m2 in 2005, equivalent to an 85% reduction (Fig. 12). Biomass followed the same trend, although the decline was only marginally significant (F = 5.94; P = 0.071; df = 1), from 4.5 g AFDM/m2 in 2004 to 1.7 g/m2 in 2005. Substrates that were incubated for 2 years had similar Dreissena densities (~8,500 individuals/m 2 ) but higher biomass

(2.8 g/m2) than those that were incubated for 1 year (Fig. 12). The number of fish caught varied by season. Total catch was significantly lower in spring than fall (Fig. 13A; H = 7.48; P = 0.024; df = 2), but no statistically significant differences were detected between spring and summer total catch and between summer and fall total catch (P > 0.05). Species richness followed the same general pattern as total catch, with spring richness significantly lower than either summer or fall (F2,33 = 13.52; P < 0.001). Mean spring richness ranged from 5 to 7 species and summer/fall mean richness ranged from 10 to 14 species (Fig. 13A). The fish community generally was dominated by round gobies (Neogobius melanostomus) in spring and summer, and sunfishes (Centrarchidae) in the fall, although these abundances were not statistically significant (F3,135 = 0.967; P = 0.412; Fig. 13B). The 2005 catch was an exception, with minnows (Cyprinidae) dominant in the summer and yellow perch (Perca flavescens) in the fall. The dominant sunfishes were (by numbers) pumpkinseed (Lepomis gibbosus), rock bass (Ambloplites rupestris), largemouth bass (Micropterus salmoides), and bluegill (Lepomis macrochirus). The dominant minnows were mimic shiners (Notropis volucellus), bluntnose minnows (Pimephales promelas), and spottail shiners (Notropis hudsonius), although > 99% of mimic shiners were

Muskegon Lake Status and Trends

FIG. 13. Fish community characterization in Muskegon Lake from 2003–2005. A) mean (+ SE) total fish catch and species richness; B) mean (+ SE) fish community composition based on major taxonomic groups. All data represent a sampling effort of 3 net nights; * denotes sampling events where effort was less than 3 net nights. Data for these events were standardized per 3 net nights to facilitate comparisons. Text boxes show the results of statistical tests. See text for more details. captured at the western-most fish sampling site (Fig. 1) during summer 2005. Shoreline Survey A total of 49.99 km of Muskegon Lake shoreline was inventoried. Of that total, 64.93% (32.46 km) was hardened and 35.07% (16.99 km) was natural (Fig. 14). No attempt was made to classify the natural habitat. The three most abundant categories of hardened shoreline were metal seawall (10.31 km), concrete riprap (9.98 km), and rock riprap (3.85 km). The majority of the natural shoreline is located on the lake’s north side; historic and remaining industrial activity is focused on the south shore (Fig. 14).

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DISCUSSION Areas of concern in the Great Lakes are defined in the U.S.-Canada Great Lakes Water Quality Agreement as “geographic areas that fail to meet the general or specific objectives of the agreement where such failure has caused or is likely to cause impairment of beneficial use of the area’s ability to support aquatic life.” Of the 43 AOCs identified by the United States and Canada, only Collingwood Harbour (Ontario) has been officially delisted. According to the U.S. EPA (2001), the critical test for delisting is to ensure that the process is rigorous, scientifically defensible, and allows for full review and comment from interested and affected stakeholders. As part of the delisting process for Muskegon Lake, restoration targets for five of the nine BUIs have been approved by the MDEQ and the other four are currently in the process of being developed (Table 2). The monitoring data collected as part of this study allow us to show whether or not certain restoration targets have been met and maintained through a specified time period, as required by the delisting process (U.S. EPA 2001). For example, improvements are evident with respect to the eutrophication target, as the 2003–2005 near-surface mean TP was 25 µg/L compared to the restoration target of 30 µg/L, the 2003-2005 chlorophyll a mean concentration was 5.9 µg/L compared to the restoration target of 10 µg/L, and the 2003–2005 Secchi disk mean depth was 2.27 m compared to the restoration target of ~2.0 m (Table 2). The general improvement of water quality conditions in Muskegon Lake is attributable, in large part, to wastewater being diverted to the Muskegon County Wastewater Management System instead of being discharged directly to the lake. The county’s 11,000-acre land application system uses extended aeration, lagoon impoundment, slow-rate irrigation and rapid-sand filtration, and is capable of providing tertiary treatment to 42 million gallons per day of wastewater. The treated wastewater is discharged in the Muskegon River, ~16 km upstream of Muskegon Lake. The diversion of wastewater in the 1970s has resulted in water quality improvements not only to Muskegon Lake, but also to Mona Lake and White Lake in Muskegon County (Freedman et al. 1979, Steinman et al. 2006). Of course, the declines in chlorophyll a and TP, and increases in transparency, also may be attributable to the invasion of Muskegon Lake by dreissenids. As filter-feeders, dreissenids can potentially

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FIG. 14.

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Map of Muskegon Lake showing natural and hardened shoreline features.

remove large amounts of phytoplankton from the water column (MacIsaac et al. 1992), which can account for the reductions of chlorophyll and TP in the water column of Muskegon Lake. However, dreissenids also can mineralize nutrients (Arnott and Vanni 1996, Conroy et al. 2005), resulting in higher soluble nutrient levels, which in turn can stimulate algal growth, especially in nutrient-limited systems. The ecosystem-level influences of dreissenids largely depend on their density, condition, and habitat (cf. Makarewicz et al. 1999, Wilson 2003); the strong influence of dreissenids in Lake Erie shortly after their introduction corresponded to periods when their densities reached 340,000 individuals/m 2 (MacIsaac et al. 1991). In Muskegon Lake, dreissenid densities on

artificial substrates were considerably lower (~10,000–50,000 individuals/m2) than the densities found in Lake Erie or Saginaw Bay (Nalepa et al. 1995), so we suspect that wastewater diversion had a more significant influence on water quality than Dreissena invasion. Carter et al. (2006) also concluded that changes in Muskegon Lake’s benthic invertebrate populations between 1972 and 1999 were attributable to water quality improvements related to wastewater diversion. Dreissenids were found in 20 of 27 sites in 1999 (collected with Ponar grab samples) with average densities of < 400 individuals/m 2 at all but three sites. The substrate at the three sites with higher densities was more favorable (wood chips, coarse sand, and gravel) than the soft substrate (silt) found at most of the other sites

Muskegon Lake Status and Trends (Carter et al. 2006). Based on the low dreissenid numbers and predominance of unfavorable, soft substrate, the authors hypothesized that wastewater diversion was the main driver for recovery of the benthic community. Although invasive species are recognized as having potentially profound ecological impacts in AOCs, they are not considered a BUI as their impacts reflect basin-wide problems, and are not related to attributes unique to a particular Area of Concern. Dreissenids also can influence the taxonomic composition of phytoplankton through differential ingestion and egestion (Heath et al. 1995, Makarewicz et al. 1999). The relationship between dreissenids and the cyanobacterium Microcystis aeruginosa is of particular interest because of the ability of Microcystis to produce the hepatotoxin microcystin. Raikow et al. (2004) showed that there is a positive relationship between dreissenid presence and Microcystis abundance in lakes with TP concentrations between 10 and 25 µg/L; a subsequent enclosure study by Sarnelle et al. (2005) found that Dreissena reduced Microcystis biomass when TP concentrations were < 3 µg/L but that Dreissena had a positive influence on Microcystis biomass when the enclosures were enriched in TP to 9 µg/L. Our data from Muskegon Lake are consistent with the findings of Raikow et al. (2004), as the high abundance of Microcystis corresponded to relatively high TP concentrations. It is unlikely that nutrient excretion from dreissenids is responsible for Microcystis abundance in Muskegon Lake given the lake’s relatively high ambient phosphorus levels, but selective feeding may play a minor role. The absence of taxonomic data precluded a comparison of phytoplankton species composition over time in Muskegon Lake. However, the dominance of cyanobacteria in Muskegon Lake is consistent with reports of increasing levels of cyanobacteria in the nearshore waters of southern Lake Michigan (Madenjian et al. 2002) and presumably is related to high nutrient concentrations (Downing et al. 2001). In the offshore waters of Lake Michigan, phytoflagellates were still the dominant algal group (Makarewicz et al. 1998). Three of the four dominant cyanobacteria in Muskegon Lake are known producers of cyanotoxins (Microcystis, Anabaena, and Aphanizomenon); hence, despite the overall reductions in phosphorus and algal biomass in Muskegon Lake, the algal taxonomic composition still presents a potential ecological and human health risk. Phytoplankton taxonomic composition was a proposed indicator that did not pass the vet-

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ting and peer-review process (Fig. 2) because of concerns regarding the cost of analysis, inability to control for bias in making identifications as analysts changed with time, and difficulty in developing appropriate targets. Not all nutrients showed improvement in Muskegon Lake. Nitrate levels increased significantly over the past 30 years; we suspect this increase is a reflection of atmospheric deposition (cf. Vitousek et al. 1997), and not local processes, as human activities have significantly increased the amount of fixed N inputs to terrestrial systems from preindustrial to modern times (Holland et al. 1999). The nitrogen cascade (i.e., the sequence of events resulting from the circulation of anthropogenic nitrogen, and subsequent magnification of its effects in the air, on land, in freshwater and marine systems, and on human health) then results in a substantial fraction of this N reaching our water bodies (Galloway et al. 2003). However, the declines in chlorophyll a over time indicate that the increased nitrate was not associated with greater algal growth; this is not surprising given that algae usually prefer reduced forms of nitrogen over oxidized forms (cf. Dortch 1990). The wide range of heterotrophic bacterial abundances found in Muskegon Lake was comparable to those reported for inland waters, rivers and Lake Michigan. This, in conjunction with measured chlorophyll a concentrations, suggests that Muskegon Lake is typically a mesotrophic body of water that makes seasonal excursions into oligotrophic and eutrophic conditions (cf. Biddanda et al. 2001, Biddanda and Cotner 2002, Kalff 2002). Experimental studies will be needed to determine what environmental factors cause the wide variability in bacterial abundance found in Muskegon Lake; it is possible that both dissolved organic matter quality and quantity may be important in regulating bacterial dynamics in land-margin ecosystems, such as Muskegon Lake, that receive substantial autochthonous as well as allochthonous inputs. Although heterotrophic bacteria were not included as a delisting target, they are included as part of the Muskegon Lake monitoring program because of their critical roles in both nutrient cycling and as an indicator of trophic status. The improvements to water quality in Muskegon Lake appear to be benefiting fish populations. Muskegon Lake has good-to-excellent fishing for several important sport fishes (O’Neal 1997, Hanchin et al. 2007). High total fish catch, species richness, and presence of sunfishes are all indica-

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tors of healthier coastal ecosystems (Uzarski et al. 2005). Additionally, our results suggest good recruitment of yellow perch in 2005 (all individuals were < 14.5 cm and nearly 79% were < 9.0 cm), although yellow perch recruitment is often linked to a host of complex ecological interactions (Clapp and Dettmers 2004). Nevertheless, there also is cause for concern. Non-native round gobies appeared to dominate littoral habitats (Fig. 13B) and their numbers were higher in Muskegon Lake than other drowned river mouth lakes (Cooper et al. 2007). Walleye populations are maintained largely by stocking programs, and several native fishes—including lake sturgeon, white bass (Morone chrysops), and muskellunge (Esox masquinongy)— that were historically abundant are now rare or absent from Muskegon Lake (O’Neal 1997). Many of these issues reflect basin-wide fisheries problems however, and may not be the result of problems unique to Muskegon Lake. The fish consumption BUI in Muskegon Lake was based on elevated PCB concentrations in common carp (Cyprinus carpio) and elevated mercury concentrations in walleye and largemouth bass. Measurements addressing this delisting target are not made as part of the regular Muskegon Lake monitoring program, and will rely on data collected by MDEQ or as part of other externally-funded grants. Delisting targets for the fish habitat and fish population BUIs will likely be based on a fishbased IBI (index of biotic integrity) that combines metrics developed for Great Lakes coastal wetlands (Uzarski et al. 2005). Once finalized, they will be submitted to the MLPAC for approval, and subsequent submission to the MDEQ (Fig. 1). A number of remediation activities have been undertaken in the Muskegon Lake AOC over the past 30 years. Projects of particular significance include: 1) the construction of the Muskegon County Wastewater Management System (mid 1970s), which now receives and treats industrial and municipal discharges that previously were directed to Muskegon Lake; 2) remediation of numerous former industrial sites with contaminated soil and groundwater, including the Story/Ott Superfund site (1993–2002); and 3) remediation of Ruddiman Creek (2005– 2006), an urban tributary to Muskegon Lake, which included the removal of 88,000 m3 of contaminated sediment. These projects have resulted in improved water quality, with restoration targets currently being met for the Eutrophication BUI. Despite the restoration activities in Muskegon Lake, the water body still faces significant environ-

mental challenges. Contaminated sediments have been shown to affect benthic invertebrate communities in Muskegon Lake (Carter et al. 2006), and a recent survey concluded that there are still over 150,000 cubic yards of mercury-contaminated sediment in the lake (M. Tuchman, U.S. EPA, pers. comm.). Another challenge facing Muskegon Lake is the development of shoreline (MLPAC 2002), which often decreases the amount of woody debris in littoral habitats (Christensen et al. 1996), thereby decreasing abundance and individual growth of some fishes (Sass et al. 2006); the hardening of shorelines also can provide better habitats for invasive species such as round gobies and zebra mussels (Jude and DeBoe 1996). Finally, invasive species threaten the ecological integrity of Muskegon Lake. Hemimysis anomala (bloody-red mysid), the latest invader to the Great Lakes, was recently discovered in the channel connecting Lake Michigan to Muskegon Lake (Pothoven et al. 2007). Although the ecological consequences of Hemimysis invasion are still unknown, the impacts of other invaders, such as round gobies, the macroalga Enteromorpha (Lougheed and Stevenson 2004), and dreissenid mussels are well documented. The long-term monitoring program on Muskegon Lake is designed to provide consistent, scientifically rigorous information and to accommodate the information needs of the community through the de-listing process (Steinman and Ogdahl 2004). As such, these monitoring data serve an important scientific and community function (cf. Lovett et al. 2007). Delisting will not be a panacea for Muskegon Lake or for other Great Lakes AOCs, as regional stressors such as invasive species, atmospheric deposition, and global climate change are outside the scope of this process. Ultimately, an integrated approach will be needed to restore the ecological health of these AOCs. This process will involve community involvement, dedicated funding, and additional research and monitoring. ACKNOWLEDGMENTS Funding for this work was provided by The Community Foundation for Muskegon County through their Muskegon Lake Research Endowment Fund and Environmental Committee Fund, The Fremont Area Community Foundation through the Ice Mountain Stewardship Fund, U.S. EPA Great Lakes National Program Office, MDEQ Water Bureau, the Great Lakes Commission, and Grand Valley State University. Field and laboratory support were pro-

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