Changes in the benthic communities of Muskegon Lake, a Great Lakes Area of Concern

Journal of Great Lakes Research 39 (2013) 7–18

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Changes in the benthic communities of Muskegon Lake, a Great Lakes Area of Concern Whitney A. Nelson ⁎, Alan D. Steinman 1 Annis Water Resources Institute, Grand Valley State University, 740 West Shoreline Drive, Muskegon, MI 49441, USA

a r t i c l e

i n f o

Article history: Received 9 September 2011 Accepted 26 December 2012 Available online 26 January 2013 Communicated by David Barton Keywords: Macroinvertebrate community structure Muskegon Lake Long-term monitoring Great Lakes Area of Concern

a b s t r a c t Sediment contamination resulting from the direct discharge of industrial and municipal wastes contributed to the designation of Muskegon Lake (Michigan) as a Great Lakes Area of Concern. To assess the changes occurring in the sedient-dwelling invertebrate communities since wastewater diversion began in 1973, benthic samples were collected three times per year (spring, summer, fall) between 2004 and 2010 from six sites and compared to historic samples. The density and diversity of invertebrate populations were analyzed to: 1) identify spatial and temporal patterns in the community structure; 2) determine if community structure patterns were related to environmental variables; and 3) assess the recovery of Muskegon Lake's benthic community following wastewater diversion. Our results revealed that invertebrate community structure changed on both annual and spatial scales, while seasonal differences were shown to be modest between 2004 and 2010. The environmental variables with the greatest explanatory power included dissolved oxygen, pH, and depth. Overall, recovery of benthic invertebrate community structure was evident based on multiple lines of evidence, including increased densities of all major taxonomic groups and species diversity since wastewater diversion, decreases in both the oligochaete–chironomid ratio (0.92 in 1972; 0.69 in 2010) and the proportion of oligochaetes, and declining sediment metal concentration over time. However, comparisons of present-day and historic sampling sites must be viewed with caution because sampling locations and protocols varied among years. Significant changes in benthic invertebrate composition and water quality metrics since 1972 suggest improved environmental conditions and the continued recovery of Muskegon Lake from historic pollution. © 2013 International Association for Great Lakes Research. Published by Elsevier B.V. All rights reserved.

Introduction Benthic invertebrate community composition is widely used to evaluate aquatic ecosystems (Reynoldson et al., 1997). Though used most often in monitoring stream conditions, invertebrates also can be used to infer lake condition (Nalepa et al., 1998; Wiederholm, 1980). Occurring in habitats receiving a continual flux of inorganic and organic compounds through the sediment–water interface, lentic invertebrate communities integrate the effects of contaminant stressors as well as autotrophic and heterotrophic processes (Burton, 1991; Wiederholm, 1980). Species assemblages and the presence or absence of “indicator” species can be used to assess a lake's ecological status (Milbrink, 1983). Although benthic communities can provide insight into the condition of an ecosystem, they are snapshots of the lake's ecological

⁎ Corresponding author at: University of Arkansas, Department of Entomology, 319 Agriculture Building, Fayetteville, AR 72701, USA. Tel.: +1 616 331 3979. E-mail addresses: [email protected] (W.A. Nelson), [email protected] (A.D. Steinman). 1 Tel.: +1 616 331 3979.

condition only at the time of collection (Evans, 1992). While the results of a single sampling event cannot “forecast” future communities, continuous long-term monitoring data are useful in assessing trends in trophic state and environmental condition (Nalepa et al., 2000). As lake-wide experiments are not always ethical or feasible, comparison of current samples with historical records has been an effective method of quantifying changes in aquatic systems (Karr, 1993; Nalepa et al., 2000). By providing a background against which more recent studies can be evaluated, historical records can be used to determine the rate of improvement or decline of habitat in a particular lake or following a particular event (Schloesser et al., 1995; Wiederholm, 1980). Muskegon Lake, a coastal drowned river mouth lake located on the eastern shore of Lake Michigan, has a long history of anthropogenic stress. The region's timber resources were harvested during the logging era in the 1800s, which resulted in altered shorelines and severe erosion (Carter et al., 2006). Industrial expansion in the 1900s produced a shoreline with foundries, paper mills, metal finishing plants, and petrochemical storage facilities (Steinman et al., 2008). Prior to the 1973 installation of a wastewater management system, untreated domestic and industrial wastes were released directly into the lake (Evans, 1992). Reports by the Michigan Department of Natural Resources (MDNR) detailed water quality problems relating to nutrient enrichment

0380-1330/$ – see front matter © 2013 International Association for Great Lakes Research. Published by Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jglr.2012.12.016

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W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18

(e.g., fish tainting, algal blooms) as well as high levels of sediment contaminants including heavy metals, oil slicks, aromatic hydrocarbons, and anoxia (Evans, 1992; Wuycheck, 1987). As a result of these severe environmental impairments related to historic discharges, Muskegon Lake was designated an Area of Concern (AOC) in 1987 (Carter et al., 2006). A 1999 benthic community survey found significant increases in the density of all major taxonomic groups between 1972 and 1999 (Carter et al., 2006), and attributed the improvement in the community to the effects of wastewater diversion. Though industrial discharges are now diverted to the Muskegon County wastewater management system, sediment contamination from the historic discharge of pollutants continues to persist in some areas of Muskegon Lake (Evans, 1992; Rediske et al., 2002). In 2003, a long-term monitoring program was initiated to assess the ecological status of Muskegon Lake and track trends in lake recovery (Steinman and Ogdahl, 2004). The ultimate goals of the long-term monitoring are to establish a record of environmental conditions in the lake, detect patterns, and generate hypotheses related to factors affecting the lake's condition. Degraded benthic communities and fish habitat is 1 of 9 Beneficial Use Impairments (BUIs) associated with the Muskegon Lake AOC. Hence, improvements in the benthic community can be used to evaluate the progress of the lake's Remedial Action Plan (RAP), and play a role in the delisting of Muskegon Lake as an AOC (Carter et al., 2006). The objectives of the current study were to 1) assess spatial and temporal patterns in Muskegon Lake's benthic invertebrate assemblages over a 7-year monitoring period (2004–2010); 2) determine if invertebrate community patterns were related to environmental variables; and 3) assess the recovery of Muskegon Lake's benthic community since wastewater diversion in 1973 by comparing 2004–2010 sampling to historic benthic samples (1972, 1999).

Materials and methods Study area Muskegon Lake is a coastal drowned river mouth lake (~ 17 km 2), connecting to Lake Michigan via a navigational channel. Average hydraulic retention time is 23 days (Freedman et al., 1979), with mean and maximum depths of 7 and 23 m, respectively. Details on the lake, surrounding watershed, and site history are provided in Steinman et al. (2008). Current land use in the Muskegon Lake watershed is 18% agriculture, 14% urban/developed, 55% forested, and 13% other (Cooper et al., 2007). As of 2006, ~ 65% of Muskegon Lake's shoreline was hardened, mostly on the southern shore, and is associated with urban and past industrial development (Steinman et al., 2008). On-going shoreline restoration activities are estimated to reduce the percentage of hardened shoreline to ~ 55% by 2013. 2004–2010 field sampling and lab analysis Benthic samples for community structure analysis were collected from 6 sites in Muskegon Lake (Fig. 1) during the spring, summer, and fall (May, July, and September, respectively) from 2004 through 2010. At each site, 1 sample was collected with a Ponar grab (area = 0.046 m 2), resulting in a total of 126 samples (6 sites × 3 seasons × 7 years). All of the material collected in the grab was rinsed through a 0.5 mm steel mesh screen. Material and organisms retained were preserved in 70% ethanol containing rose bengal stain. The final dataset consisted of 121 samples because 5 benthic samples were lost: 3 from summer 2008 and 2 from fall 2009. In 2010, 4 additional grabs were taken at each site with a petite Ponar (area = 0.030 m 2) for sediment analysis of metals, organic matter, and grain size. The

Fig. 1. Muskegon Lake sampling locations for invertebrate community and heavy metal analysis. Circles (n = 6) indicate 2004–2010 sampling stations for both macroinvertebrate community structure and sediment collection for heavy metals (metals sampled in July, 2010 only). Squares (n = 15) indicate 1972 and 1999 sampling stations for both invertebrate community structure and sediment collection for heavy metals (metals sampled at M-1, M-10, and M-15) (Carter et al., 2006; Evans, 1992). Triangles (n = 3) indicate 1972 sampling stations for sediment collection for heavy metals (H-1, H-2, and H-3) (Evans, 1992).

W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18

4 grabs were combined to form a single, composite sample from which samples were taken for sediment analysis. Dissolved oxygen (DO), pH, temperature, total dissolved solids (TDS) and specific conductance (SpC) were measured at 1 m above the lake bottom; between 2004 and 2006, a Hydrolab DataSonde 4a was used and between 2007 and 2010, a YSI 6600 Sonde was used. Intercalibration of instruments was conducted to avoid bias between datasets. Soluble reactive phosphorus (SRP), chloride, sulfate, total phosphorus (TP), total nitrogen (TN) (USEPA, 1983), nitrate, and ammonia (APHA, 1998) were measured at 1 m above the lake bottom for each sampling event following procedures detailed by Steinman et al. (2008). In the laboratory, retained benthos was transferred to a white enamel pan where all organisms were removed and sorted into major taxonomic groups. All organisms were identified to the lowest practical taxonomic level. Chironomids and oligochaetes were mounted on slides using CMC-10 mounting medium prior to identification. The keys used for identifying the various taxonomic groups were: Oligochaetes (Kathman and Brinkhurst, 1998), Chironomidae (Epler, 2001), Ephemeroptera (Burks, 1953), Chaoborinae (Cook, 1956), Bivalvia (only bivalves alive at time of preservation were counted [Peckarsky et al., 1993]), and Trichoptera (Wiggins, 1977). All other identifications were made using keys from Thorp and Covich (2001). Sediment samples collected in 2010 were analyzed for Target Analyte List (TAL) metals using SW-846 methods (USEPA, 1994): inductively coupled plasma-atomic emission spectroscopy (ICP-AES) for aluminum, calcium, iron, magnesium, manganese, nickel, potassium, sodium, zinc (Method 6010C); antimony, arsenic, barium, beryllium, cadmium, chromium, cobalt, copper, lead, selenium, silver, thallium, vanadium using ICP-AES (Method 6020A), and mercury by coldvapor atomic absorption spectroscopy (Method 7471A). Quality control protocols for TAL metals for EPA 6000/7000 Series Methods were followed. Organic matter (OM) was analyzed using a well-mixed subsample. Subsamples were dried for 24 h at 105 °C to determine dry weight, then combusted for 25 h at 550 °C to determine the ashfree loss of carbon from the sample. The OM for each site was averaged from triplicate subsamples (%OM). Particle size distributions were measured in triplicate by wet sieving (USEPA, 2003) and the mean was reported in phi (Φ) units. Larger Φ indicates finer substrate. The following mesh sizes were used: 2 mm (granule), 1 mm (very coarse sand), 0.5 mm (coarse sand), 0.25 mm (medium sand), 0.125 mm (fine sand), 0.063 mm (very fine sand), and b0.063 mm (coarse silt and smaller). Measured metal concentrations from sediment samples were compared to two consensus-based sediment quality guidelines: a threshold effect concentration (TEC) and a probable effect concentration (PEC) (MacDonald et al., 2000). The TEC is the concentration below which adverse effects are unlikely to occur and the PEC guideline indicates the concentrations above which adverse effects are expected to occur (MacDonald et al., 2000). These guidelines are designed to reflect the toxicity of sediment contaminants to the biota when contaminants occur in mixtures with other contaminants and are aimed at classifying sediments as toxic or not toxic to benthic invertebrates. Data analysis Variation in community density at different spatial and temporal scales was examined using an analysis of variance (ANOVA) model, which examined the significance of year, site, season nested within year, and site × year interactions. Density values were square-roottransformed prior to analysis to aid in approximations of normality and to downplay the weight of abundant species. Statistically significant differences were analyzed with Tukey honestly significant different tests (Tukey HSD) for multiple comparisons. Spatial and temporal patterns in benthic community abundances were examined using non-metric multidimensional scaling (NMDS),

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which is robust for non-monotonic biotic variables and data with frequent zero values. After square-root transformation, a Bray–Curtis similarity coefficient was used as a distance measure (Bray and Curtis, 1957). The Bray–Curtis coefficient is less influenced by the abundance of taxa than other standardization coefficients (Jackson, 1993), and was used in an attempt to minimize potential artifacts relating to taxa abundance in the dataset. Results from the similarity matrix were ranked, coded by the variable of interest, and plotted. Each point plotted represents a single sampling event (n= 121). A stress value was estimated between the calculated and actual distances among the samples to evaluate the goodness-of-fit. In search of the optimal solution with the lowest stress value, the NMDS function was specified to run with 20 random starts. To test if environmental variables explained significant community composition, hypothesized gradients were overlaid onto the NMDS plots without disturbing the configuration of the original ordinations using the linear fitting function “envfit” (Oksanen et al., 2009). This fitting function finds the vector averages of the environmental variables and fits them in the NMDS plot defined by the species data. Overlaying environmental variables on an ordination using “envfit” generates an R 2 goodness-of-fit statistic and a significance value (p) based on the probability that random permutations of the environmental variables would yield a higher degree of fit than the true environmental variables. It is important to note that the environmental variables overlaid with “envfit” are independently modeled effects; therefore, one effect cannot be evaluated as having explained the variation of others (Oksanen et al., 2009). Environmental variable summaries can be found in Gillett and Steinman (2011). To assess if benthic communities had significantly different species compositions, samples were analyzed with the non-parametric analysis of similarity (ANOSIM; Clarke, 1993), using dissimilarity values calculated with the Bray–Curtis similarity coefficient. The ANOSIM was run separately on 1) six sites, 2) three seasons, and 3) seven years. Statistical significance (α = 0.05) was evaluated with 999 permutations. If significant differences were found in the community composition among groups, an indicator species analysis (Dufrene and Legendre, 1997) was used to find the taxa most responsible for the differences in community composition. All data analyses were performed with R (R Development Core Team, 2011). Historical comparisons The changes in benthic community composition after wastewater diversion were assessed by comparing data collected between 2004 and 2010 to data collected in (1) June 1972, the year before wastewater diversions began (Evans, 1976), and (2) October and November of 1999 (Carter et al., 2006). In the 1972 study, 15 sites were sampled (Fig. 1) in triplicate with a petite Ponar, with samples rinsed through a 0.6 mm mesh screen. In 1999, benthic samples were collected from the same 15 sites in triplicate with a petite Ponar and rinsed through a 0.5 mm mesh sleeve. Differences in invertebrate community structure may be attributable to sampling sites and methods in the current study that do not correspond exactly with sites and methods from past studies. To minimize the effects of these potential differences, comparisons of densities were limited to coarser taxonomic groups. The following metrics were calculated for each year: Shannon diversity index (hereafter referred to as diversity), oligochaete–chironomid (O/C) ratio (calculated as [oligochaete density] / [oligochaete density + chironomid density]), taxa richness (number of individual taxa), proportion of oligochaetes in total benthos, chironomid trophic condition index (C-TCI), and chironomid herbivore and predator counts. Macroinvertebrate metric values for 1972 and 1999 were from Carter et al. (2006). The oligochaete–chironomid ratio generally reflects the tendency for tolerant oligochaete species to increase their abundance relative to sedentary chironomids in conditions of nutrient enrichment (Wiederholm, 1980). As a result, higher ratios suggest decreased water quality (Evans,

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W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18

1992). The C-TCI is a weighted average of the abundance of individual chironomid taxa by their tolerance to organic enrichment (Carter et al., 2006). The C-TCI is calculated on a scale ranging from 0 to 2, with higher values indicating greater nutrient enrichment. Values b 0.6 suggest oligotrophic conditions, values 0.6–1.0 suggest mesotrophic conditions, and values >1.0 suggest eutrophic conditions, while values close to 2.0 indicate gross organic pollution (Milbrink, 1983). Trophic tolerance was calculated using the tolerance values of Winnell and White (1985). Taxa that were not classified in Winnell and White (1985) were classified using the biotic index by Hilsenhoff (1987) or tolerance values from Barbour et al. (1999). The classification schemes of both Hilsenhoff (1987) and Barbour et al. (1999) were on a 10-point scale, and were partitioned among the ecological groups within the TCI. The C-TCI was used to make comparisons with the 1972 (Evans, 1976) and 1999 (Carter et al., 2006) data. One-way ANOVA was used to assess differences in community metrics between sampling periods. Statistically significant differences were analyzed with Tukey HSD tests for multiple comparisons. Muskegon Lake sediments collected in 2010 were compared to those collected and analyzed in 1972 by Evans (1976) and in 1999 by Carter et al. (2006). In 1972, stations H-1, H-2, and H-3 (Fig. 1) were sampled by removing the top 2 cm of sediment using an Ekman grab, while in 1999, stations M-1, M-10, and M-15 (Fig. 1) were sampled with a petite Ponar grab and a well-mixed subsample was removed for analysis (Carter et al., 2006). Similar to 1999 sampling methods, 2010 samples were collected with a petite Ponar grab and a well-mixed subsample was removed for analysis. Based on the penetration depth of the petite Ponar (~ 12 cm) and the estimated sedimentation rate for Muskegon Lake of 1 cm·yr −1 (Rediske et al., 2002), the 1999 samples were generally representative of mean metal concentration deposited in sediments deposited from ~ 1987 to 1999 (Carter et al., 2006), while 2010 samples were considered representative of mean metal deposition to the sediments from ~ 1998 to 2010. A one-way ANOVA was used to determine if there were significant temporal differences in mean metal concentrations. Statistically significant differences were further analyzed using Tukey HSD tests for multiple comparisons.

Results Sediment properties Percent OM at Muskegon Lake sites ranged from 1.1 to 22.7%, and mean grain size (Φ) ranged from 1.5 (medium to fine sand) to 4.3 (coarse silt or finer) in 2010 (Table 1). The mean grain size distributions are probably related to the historic industrial development along the shoreline; much of the shoreline was filled with foundry sand during the 1930s–1950s (Rediske et al., 2002). Erosion of the fill material likely resulted in sand deposits in the nearshore and

high erosion areas (e.g. AWRI, CHAN, RUDD). However, the recent history of shore stabilization and eutrophic conditions in the lake has resulted in the deposition of finer-grained material in the lake (i.e. BEAR, MUSR, DEEP). Concentrations of metals among sites in 2010 ranged from b0.050 to 150 mg kg − 1 (dry weight). Higher concentrations of heavy metals and % OM were measured at sites BEAR and DEEP (Table 1) than at other sites in 2010. In contrast, the CHAN site had the lowest concentrations of heavy metals and % OM.

Taxa abundance and composition There were annual fluctuations in benthic community composition during the 7 year monitoring period, but the dominant taxa remained consistent; chironomids were the most abundant group, with total mean yearly densities ranging from 11,741 to 39,672 m−2 (Appendix A). This group accounted for 1–100% of all organisms collected at a given site and exceeded 50% density at 64 of 121 samples. Twentyseven chironomid taxa were identified, with Chironomus, Procladius, and Tanytarsus as the most common (Appendix A). Oligochaetes were the second most abundant group, with total mean annual densities ranging from 13,664 to 25,893 m−2. Eighteen oligochaete species were identified, with Limnodrilus hoffmeisteri, Limnodrilus udekemianus, and Quistadrilus multisetosus being the most common. In 2004 and 2010, oligochaetes accounted for 36 (±5) % and 38 (±6) % of all organisms collected, respectively. Interestingly, no oligochaete taxa were observed during the sampling years 2007–2009 (see Discussion). Bivalves were found in 62 of 121 samples, with dreissenid mussels making up a notable fraction of the samples across years (Appendix A). Dreissena rostriformis bugensis was found in only 29 of 121 samples, but its mean overall density was 3025±1418 m−2, which was second only to Chironomus. The remaining benthic groups: Chaoboridae, Ceratopogonidae, Ephemeroptera, Trichoptera, Polychaeta, Gastropoda, Bivalvia (excluding dreissenids), Isopoda, Amphipoda, and Zygoptera (collectively referred to as “others”), were generally found in lower densities and/or occurred infrequently. Significant differences (p ≤ 0.05) were seen in mean benthic density for sampling years and seasons nested within year (Table 2). Post-hoc pairwise comparisons revealed that the years 2007, 2008, and 2009 had significantly lower mean densities than 2004, 2005, and 2010, and that the year 2006 had a significantly lower mean density than 2005 and 2010 (Fig. 2A). There was a seasonal effect for benthic density, though this was significant only when season was nested within year (Table 2). The seasonal effect (Fig. 2B) was most apparent for the summer of 2005, which was significantly different from both the fall and spring of 2007, 2008, and 2009 (Fig. 2B). In 2009, summer was significantly different from the fall. No significant differences were observed for site (Fig. 2C, p = 0.31) or the interaction between site and year (p = 0.25) (Table 2).

Table 1 Concentrations of heavy metals (mg kg−1, dry weight), organic matter (OM, % dry weight), and mean grain size (Φ) in sediments at 6 sites in Muskegon Lake, July 2010. Values for metals are single measurements. Grain size for each site is a mean of Φ values (n = 3), with the grand mean represented (±standard error). The OM for each site is an average of a triplicate subsample, with the grand mean represented (±standard error) in the table. For site locations, see Fig. 1. Site

Φ

OM

Arsenic

Cadmium

Chromium

Copper

Nickel

Lead

Zinc

Mercury

AWRI BEAR CHAN DEEP MUSR RUDD Mean (±standard error)

2.8 4.3 2.4 4.3 4.3 1.5 3.3 (0.49)

2.6 22.7 1.1 15.1 11.9 8.9 10.4 (3.3)

2 12 1.9 11 4.2 4.3 5.9 (1.8)

0.35 1.8 0.075 1.9 0.72 0.87 0.95 (0.31)

7.8 32 3.5 54 17 25 23.2 (7.5)

7.3 41 1.6 41 16 21 21.3 (6.8)

3.1 13 1.4 16 8.6 8.7 8.5 (2.3)

8.4 39 2.5 61 19 29 26.5 (8.8)

29 130 11 150 59 74 75.5 (22.4)

b0.050 0.15 b0.050 0.19 0.082 0.098 0.10 (0.02)

W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18 Table 2 Summary of ANOVA tests assessing the variation of community density at different nested spatial and temporal scales, with presentation of mean square (MS) and degrees of freedom (d.f.). Values in bold indicate significance at p≤0.05. Nested nominal values are represented as group (subgroup) [e.g. season nested within year is represented as: season (year)]. Two-factor nominal interaction variables are represented as group×group (e.g. an interaction term between year and site is represented as year×site). Scale

MS

d.f.

F

p

Year Site Season (year) Year × site Residuals

18,735 3912 9613 3963 3243

6 5 14 30 65

5.78 1.21 2.96 1.22

b0.001 0.316 0.001 0.246

Mean annual invertebrate densities ranged from 7349 to 40,083 m−2 over the sampling period, with the maximum and minimum mean annual densities measured in 2005 and 2008, respectively (Table 3). A total of 78 taxa were collected, and the mean overall richness was 8 taxa per site (range 1–23). Mean densities were highest in the summer season (Table 3). The AWRI site had the highest mean density (26,327 m − 2) while the CHAN site had the lowest mean density (16,494 m − 2), though the differences were not statistically significant. It is unclear if oligochaetes were truly absent during the 2007–2009 time period or if sample processing resulted in organism fragmentation and the resulting failure to observe them. Because of this concern, NMDS plots were constructed that included all observed taxa (Figs. 3A–C) and with all oligochaete data removed (Figs. 3D–F). There was no distinct separation among samples when coded by season (Figs. 3A, D). In contrast, sites did show a clear separation, with CHAN, DEEP and MUSR distinctly clustered from the other three sites (Fig. 3B). This cluster of the 3 sites remains distinct in the plot without oligochaete taxa (Fig. 3E). The effect of year was distinct, with 2007, 2008, and 2009 separating into one cluster, and 2004, 2005, and 2010 separating into another group (Fig. 3C). 2006 was the only year that did not appear to exhibit any sort of clustering effect. Again, this grouping by years is seen in the plot without oligochaete taxa (Fig. 3F). Environmental variables that correlated highly with the ordination space defined by all collected taxa (Figs. 3A–C) included depth, pH, SRP, DO, total nitrogen (TN), and chloride (Cl) (Table 4). The environmental variables fitted to the ordination plots without oligochaete taxa (Figs. 3D–F) included depth, pH, SRP, and DO. The R 2 values calculated for the envfit of environmental variables plots without oligochaete taxa differed very little from plots including all taxa, and are not detailed here. By visual inspection, the arrows for Cl and TN concentrations are pointed in the same direction, showing a similar shift in species composition across the gradient (Fig. 3A, Table 4). Sites with high TN and Cl concentrations were dominated by chironomid taxa (Chironomus, Cryptochironomus, Procladius). The second envfit dimension had high values for depth and SRP (Table 4), under which conditions oligochaete taxa (largely immature tubificids) were most abundant. High concentrations of SRP and depth were negatively correlated with DO. Sites with high values for DO had high densities of chironomids (Chironomus, Procladius, Cryptochironomus, and others). Overall, the highest ranked environmental factors were found to be depth and pH. The indicator species analysis run on the results of the ANOSIM for species composition (on all taxa) revealed that the predatory chironomid Procladius was associated with the RUDD sampling site, along with the oligochaetes Dero digitata and Aulodrilus limnobius (Table 3). In contrast, the mayfly Hexagenia rigida was an indicator species at MUSR. With respect to seasonality, the chironomids Procladius and Tanytarsus were indicator species of spring sampling, while the oligochaete Aulodrilus pluriseta was an indicator species of fall sampling

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dates. Chironomid and oligochaete taxa were most often selected as indicator species for individual sampling years, with no consistent associations among years (Table 3). Historical comparisons: 1972, 1999, and 2010 Overall benthic community (excluding Dreissena) density increased 10-fold from pre-wastewater diversion values (2858 m−2) to 2010 (20,630 m −2) (p≤ 0.05; Table 5). The densities of most major groups increased significantly and consistently among sampling years, the only exceptions being Sphaeriidae and Dreissena, which increased from 1972 to 1999, but not between the 1999 and 2010 sampling periods. Community metrics also showed changes among sampling years (Table 5). Taxa richness increased significantly between 1972 and 1999 (p b 0.001; Table 5), but not between 1999 and 2010 (p = 0.585). Diversity showed significant increases from 1972 to 1999 to 2010 (p b 0.001), whereas the O/C ratio, proportion of oligochaetes, and the chironomid-TCI all declined significantly between the 1972 and 2010 sampling dates (p b 0.001). Because historical comparisons were made with samples collected from different seasons and different sites in Muskegon Lake, it is possible that community comparisons over time were biased. To minimize this possibility, we paired specific sites from 1972 (H-2) and 1999 (M-3) with 2010 sampling sites whose spatial locations were in close proximity to each other (AWRI and RUDD, respectively). Comparisons between 1972 and 2010 were made only for spring samples and comparisons between 1999 and 2010 were made only for fall samples, as those were the only seasons when there was sampling overlap. Invertebrate density data were consistent (i.e., not significantly different) with the results presented in Table 5. Invertebrate diversity showed significant differences between the 1972 and 2010 spring comparisons, but not between the 1999 and 2010 fall comparisons. While there is no guarantee that comparisons between historical data and 2004–2010 data from the other sites and seasons will not introduce bias, the results of historical comparisons using matched sites and seasons provide some evidence that these other comparisons appear reasonable, especially for the invertebrate density, whereas the diversity data should be viewed with greater caution. Mean heavy metal concentrations declined lake-wide by 13 to 82% between 1999 and 2010, with a decrease of 67 to 92% in metal concentration from pre-wastewater diversion values compared to 2010 concentrations (Fig. 4). Chromium declined from 274 to 23.4 mg kg − 1 between 1972 and 2010, and lead concentrations declined from 165 to 26.6 mg kg − 1 between 1972 and 2010. Declines were significant for all metals among all years, except for mean arsenic concentrations between 1999 and 2010. To ensure that comparisons across years were not biased by comparison of sites that were not at the exact same locations, we conducted comparisons among paired sites, similar to what was done for the invertebrate data. Metal concentration data were consistent (i.e., not significantly different) with the results presented in Table 4, suggesting comparisons of present-day data with historical data are reasonable, but again should be viewed with caution. As of 2010, the mean metal concentrations of the six sites sampled did not exceed PEC guidelines, and concentrations exceeding TEC guidelines (MacDonald et al., 2000) were found for only cadmium and mercury at some site locations (Fig. 4). Discussion Significant progress has been made to control point-source discharge of toxic substances in the Great Lakes (IJC, 2000), but contaminated sediments remain a challenge in the recovery of many Great Lakes AOCs. Environmental variables collected over the 7-year

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a

Individuals per m-2 (sq. root)

A

Year b ac a bc b b

2004

2005

2006

2008

2007

2009

2010

B Individuals per m-2 (sq. root)

Season

ab a

b

Spring

Summer

Fall

C Individuals per m-2 (sq. root)

Site

AWRI

BEAR

CHAN

DEEP

MUSR

RUDD

Fig. 2. Total benthic community density sorted by year (A), season (B), and site (C). Different letters above boxplots denote significant differences (ANOVA, p ≤ 0.05). Boxes show lower and upper quartiles with medians inside, whiskers extend 1.5 times the interquartile range, and points represent outliers. Note square-root scale of y-axis.

W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18 Table 3 Benthic community variables and results from the indicator species analysis summarized by their means (ranges) for the whole dataset (All), for each sampling year (2004–2010), for each season (spring, summer, fall), and by site. The (n) in the row headings represents total number of samples summarized. Only up to three of the taxa with the highest indicator values and statistical significance at p ≤0.05 for specific clusters are listed in decreasing order of indicator values. Grouping (sample size) Year: All years (121) 2004 (18) 2005 (18) 2006 (18) 2007 (18) 2008 (15) 2009 (16) 2010 (18) Season: Spring (42) Summer (39) Fall (40) Site: AWRI (21) BEAR (20) CHAN (21) DEEP (20) MUSR (20) RUDD (19)

Variable name (units) Density (m−2)

21,090 (172–145,711)

Species richness

Top indicator species

8 (1–23)

23,864 (861–89,045)

11 (3–20) Pisidium ferrugineum, Dreissena polymorpha, Bithynia 40,083 (1550–145,711) 13 (2–23) Dreissena, Probezzia, Gammarus 17,271 (1033–50,465)

7 (1–15) Ablabesmyia annulata

11,176 (689–32,036)

4 (2–7)

–

7349 (172–20,496)

4 (1–9)

Paralauterborniella

22,455 (1206–123,837)

4 (1–7)

Chironomus

23,290 (6028–53,910)

19,844 (689–61,660)

14 (7–21) Chaoborus, Aulodrilus pluriseta, Limnodrilus hoffmeisteri

14,911 (517–89,046)

9 (2–23) Procladius, Tanytarsus, Ilyodrilus templetoni 9 (1–21) Chironomus, Dero digitata, Valvata lewisi 7 (1–23) Aulodrilus pluriseta

26,327 (172–145,711)

9 (1–21) Oecetis

20,505 (3961–69,583)

8 (1–17) Chironomus

16,494 (689–137,961)

7 (2–14) Monodiamesa

24,483 (3100–53,909)

7 (1–15) Chaoborus

28,768 (172–145,711)

18,515 (2584–53,910) 20,133 (2411–55,632)

10 (3–21) Coelotanypus, Hexagenia rigida, Paratendips 10 (3–23) Procladius, Dero digitata, Aulodrilus limnobius

sampling period in the current study were generally indicative of good overall ecosystem health, and were consistent with overall improving water quality conditions in Muskegon Lake (Steinman et al., 2008), but water quality data alone are unable to fully characterize an ecosystem (Reynoldson et al., 1997). Our benthic community dataset provided additional information on the status of Muskegon Lake, and also offered the opportunity to compare current benthic data to previous benthic data, in an attempt to track the recovery of the lake's ecological condition. Spatial and temporal patterns from 2004 to 2010 There were clear differences among years, for both species composition and density. Overall, we found that benthic community density and species composition were influenced by sampling year and site, while seasonality had a limited influence. For example, there was a distinct clustering of the years 2007–2009 resulting from both statistically lower densities and the apparent differences in species composition during those years compared to the other sampling years. This 2007–2009 benthic clustering is likely related, at least in part, to the absence of oligochaetes during those 3 years. It is unclear if oligochaetes

13

were truly absent during this time period or if their non-detection was an artifact of sample processing (e.g., excessive fragmentation) or insufficient preservative addition, as different individuals were responsible for initial preservation and sorting, although the same individual (WAN) performed all identifications. However, failure of a sewage line in March 2007 resulted in the release of 25.5 million gallons of raw sewage into Muskegon Lake (DEQ, 2008), which likely impacted the benthic community and presumably accounts for at least a portion of the community differences seen from 2007 to 2009. Additional sampling of benthos in Muskegon Lake in 2009, 2010, and 2011 by NOAA scientists measured abundances of oligochaetes and total macroinvertebrates that were substantially lower in 2009 compared to 2010 and 2011 (S. Lozano, NOAA, unpublished data). These data are consistent with lingering impacts of the sewage spill on the benthic community in 2009, followed by recovery in 2010, although the NOAA data showed impairment beyond just the oligochaete community. The phytoplankton community in Muskegon Lake also showed similar clusters to what we observed with macroinvertebrates, with the years 2008 and 2009 separating from other sampling years (Gillett and Steinman, 2011). These authors found a high algal biovolume in 2008 (Gillett and Steinman, 2011), which is opposite of what was seen in the 2008 invertebrate populations. Cyanotoxin production by large populations of Microcystis in 2008 may have had deleterious effects on invertebrates (cf. Lindsay et al., 2006), although we did not measure this interaction in Muskegon Lake. Although it is not possible to determine the reason for the absence of oligochaetes in the 2007–2009 samples, we re-analyzed our 2004–2010 data excluding oligochaetes from all years, and compared the results with those from our original analyses, to determine if their absence influenced the results. We found no statistically significant differences between the sets of analyses for the ANOVA and indicator species analysis (pb 0.05). Additionally, running NMDS plots without oligochaete taxa did not show any different clustering than plots with oligochaete data. For these reasons, we do not report the analyses without the oligochaete taxa, although we recognize that their absence in 2007–2009 introduces uncertainty in the interpretation of the invertebrate community structure data. Spatial variation in invertebrate species composition was evidenced by the separation of the sites MUSR, CHAN, and DEEP. While this separation of sites was most pronounced in the ordination plots coded by site with all taxa, they remained in plots without oligochaete taxa. The persistence of this separation suggests either that the influence of oligochaetes was modulated by the remaining benthic taxa found in Muskegon Lake or that the entire system was affected in a similar manner by a disturbance. Because oligochaetes have the ability to tolerate high levels of pollution and reduced DO, they have historically been a large component in the overall density of the benthic community in Muskegon Lake — 85% in 1972 (Carter et al., 2006). The most pronounced environmental gradient was observed between depth and DO concentration, with oligochaete taxa abundant in deeper water and low DO, and chironomid taxa populating the gradient with high DO values. While the absolute densities of oligochaetes in 2010 have increased almost 4-fold when compared to the absolute densities recorded in 1972, the relative density of oligochaetes was 38%. While 38% may seem comparatively large, it is significantly lower than pre-diversion relative oligochaete density in the population, and is apparently not sufficient to overwhelm the influence of non-oligochaete taxa in separating sites. The presence of the mayfly H. rigida at the MUSR site is most likely because the site is located near the mouth of the Muskegon River, which is an area where this burrowing mayfly is abundant (Rediske et al., 2002). Monodiamesa is known to prefer sandy substrates, which is most likely why it is more likely to be found at the sandier CHAN site. The spring presence of the predatory chironomid Procladius may be attributable to its mobility, which allows it to move in and out of areas with lower oxygen values during the spring lake turnover.

W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18

A) Season

2

3D stress: 13.7

3D

1

stress: 16.2

-1 -2

-2

-1

0

1

2

3D

B) Site

-1

0

1

E) Site

2

3

3D stress: 16.2

0

0

-2

-2

-1

-1

NMDS II

1

1

stress: 13.7

-2

2

-2 2

D) Season

0

0 -1

NMDS II

1

2

14

-1

0

1

2

C) Year

3D

-2

2

-2

0

-1

1

F) Year

2

3

3D

stress: 13.7

-2

-1

0

NMDS II

1

stress: 16.2

-2

-1

0

1

2

NMDS I

-2

-1

0

1

2

3

NMDS I

Fig. 3. Non-metric multidimensional scaling plot of benthic invertebrate community taxa samples (n = 121) from Muskegon Lake. Plots are presented with taxa per sample by season, site, and year. No oligochaete taxa were observed during 2007–2009, so plots are presented with all observed taxa (A–C) and with oligochaetes removed from all years (D–F) to compare influence of oligochaete presence/absence on community patterns. Each symbol on plots represents a single sampling event. Vectors indicate the significant environmental variables (p ≤ 0.05) and their direction of largest change. Variable abbreviations are in Table 4. The three-dimensional plots are illustrated in two dimensions for ease of representation.

Historic comparisons Table 4 Select results from the environmental vector fitting in the ordination space of the NMDS plot with variable scores along two dimensions of the vectors produced by the vector fitting, goodness-of-fit statistic (R2) and significance (p-values based on 1000 permutations). Ranking of environmental variables (all taxa), with separation is based on goodness-of-fit (R2) from function “envfit” (R Development Core Team, 2011). See text for location of variable summaries. Variable name (abbreviation)

Dimension 1

Dimension 2

R2

p-Values

All taxa Depth pH Chloride (Cl) Soluble reactive phosphorus (SRP) Dissolved oxygen (DO) Total nitrogen (TN)

−0.01 −0.48 0.44 0.02 −0.02 0.84

0.99 −0.88 −0.90 0.99 −0.99 −0.55

0.25 0.21 0.17 0.15 0.11 0.09

0.001 0.001 0.001 0.002 0.002 0.004

The increase in invertebrate densities over time is potentially biased by the differences in the sampling time of year among years and the size of mesh used (e.g., 0.6 mm in 1972 vs. 0.5 mm in 1999 and 2010). We believe these differences are small relative to the significant trends of improvement seen across the sampling period. In the Great Lakes region, increased densities of oligochaetes have been considered indicative of degraded conditions or nutrient enrichment (Nalepa et al., 1998; Winnell and White, 1985). So at first appearance, the greater absolute density of the oligochaete L. hoffmeisteri, which is generally thought to be indicative of gross organic pollution (Hilsenhoff, 1987), in recent collections (2004–2010) compared to past sampling dates is inconsistent with our conclusion. However, the absolute occurrence and density of chironomids increased more than those of oligochaetes,

W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18 Table 5 Mean (±standard error) density and community index values for sampling sites in Muskegon Lake, Michigan for 1972, 1999, and 2010. 1972 and 1999 values are for 15 sites × 1 sampling event (from Carter et al., 2006) and 2010 values are for 6 sites × 3 seasons. Samples from 1972 were collected in the spring, 1999 samples were taken in the fall, and 2010 samples span spring, summer, and fall sampling seasons. Averages for taxa/sample are presented for 2010 because the seasonal differences were not significant for 2010. We do recognize that because of this, the 2010 numbers may be biased by this larger sampling size. Community percentages used to calculate proportion of Oligochaetes exclude Dreissena population. Diversity was calculated with oligochaetes combined to a single group. Significant differences among metrics for sampling years were determined using one-way ANOVA (p ≤ 0.05). Letters indicate significant differences in means among years.

The reduced amount of heavy metals in Muskegon Lake since the 1999 samples may be partially due to the removal of contaminated sediments from Ruddiman Creek, a potential source area that flows directly to Muskegon Lake. From 2005 to 2006, ~ 60,000 m 3 of sediment, contaminated with high concentrations of cadmium, chromium, lead, and organics (Nederveld, 2009; Rediske et al., 2002) were removed from Ruddiman Creek. The decrease of metal concentrations at the 2010 sampling sites also may be a result of burial of historically contaminated sediments through the deposition of new non-contaminated sediments. Dreissenid mussels: impact and future

Year

Number of samples (n) Density (m−2) Sphaeriidae Dreissena Oligochaeta Chironomidae Amphipoda Others Total (excl. Dreissena) Metrics Taxa richness Diversity Oligochaete/chironomid Proportion of oligochaetes Chironomid-TCI

15

1972

1999

2010

15

15

18

40 (14)a 0a 2558 (717)a 196 (47)a 13 (7)a 238 (48)a 2858 (710)a

582 (133)b 3200 (2714)b 4562 (644)b 980 (79)b 179 (56)b 28 (86)b 6452 (598)b

1703 (637)b 2660 (697)b 8928 (299)c 3914 (517)c 57 (86)b 6028 (823)c 20,630 (253)c

3 (0.27)a 0.68 (0.11)a 0.92 (0.02)a 0.85 (0.03)a 1.85 (0.13)a

13 (0.36)b 1.66 (0.17)b 0.84 (0.02)a 0.68 (0.04)a 1.97 (0.01)a

14 (1.04)b 3.56 (0.10)c 0.69 (0.05)b 0.38 (0.06)b 1.37 (0.02)b

resulting in significant reductions in the O/C ratio, overall proportion of oligochaetes, and chironomid-TCI values across sampling years, which supports our conclusions about lake recovery. Krieger and Ross (1993) found increases in the number of Chironomidae and Sphaeriidae taxa and reduced relative densities of oligochaetes in Cleveland Harbor (Lake Erie). While the absolute density of oligochaete worms also increased in this system (similar to our findings in Muskegon Lake), the authors concluded that these changes were indicative of improved environmental conditions (Krieger and Ross, 1993). However, a more detailed evaluation of the nutrient and organic composition of Muskegon Lake sediments would be needed to determine if changes in community composition in Muskegon Lake were related to increased organic inputs, contamination by heavy metals or synthetic organics in the water or sediments, or some combination of the factors. The legacy effects of sediment contamination is a concern because concentrations of heavy metals and organic pollutants may be “re-circulated” through the sediments (Reynoldson, 1987), especially in a well-mixed drowned river mouth such as Muskegon Lake. This mixing could result in a prolonged exposure of the invertebrate communities to toxic levels of pollutants. Historically, high concentrations of oils were found in organic sediments throughout the lake with higher concentrations measured near industrial discharges along the south shoreline (Peterson, 1951). Mean sediment oil concentrations taken from 13 sites in 1972 were 3400 mg/kg, with southeast nearshore concentrations as high as 6104 mg/kg (Evans, 1992). In 1999, concentrations of lead, chromium, copper, cadmium, zinc, and mercury were elevated in the South Central (SC) area of Muskegon Lake, often exceeding PEC guidelines (Carter et al., 2006), which may have resulted in the lower benthic density and diversity from this area. The sites in the SC sampling area correspond with the 2010 “AWRI” site, which had highly reduced concentrations of metals when compared with the findings of Carter et al. (2006) from 1999. Though there is an element of spatial uncertainty associated with the comparison of historic metal concentrations because of the shifting nature of the sediments, as of 2010 all metals are well below PEC guidelines and largely below TEC guidelines.

The impacts of dreissenid mussels on water quality and benthic community structure and function in the Great Lakes are well known (e.g. Fahnenstiel et al., 1995; Haynes et al., 1999; Nalepa et al., 2006). Carter et al. (2006) concluded that the impacts of dreissenid mussels in Muskegon Lake are likely minor compared to the influences of wastewater diversion, riverine inputs, and sediment metal concentrations, while our study suggests that the presence of dreissenid mussels has not prevented an overall increase in benthic invertebrate density and diversity. The influence of dreissenids on the benthic community is not straightforward (cf. Idrisi et al., 2001; Stewart et al., 1998). The overall mean density of dreissenid mussels in Muskegon Lake did not change significantly between 1999 and 2010 (1972 was pre-invasion of the mussels). However, the sampling sites used for this study, as well as in 1999, consisted largely of soft substrates. In 2010, the highest densities of dreissenid mussels in Muskegon Lake were found in areas with harder substrate (wood chips). Comparisons with dreissenid mussel densities colonizing hard artificial substrates (cf. Steinman et al., 2008) show higher values than those found on natural soft substrates at the six long-term monitoring sites. This suggests that Dreissena densities in Muskegon Lake may be limited by substrate type. While studies have shown substrate limitation to be a factor for Dreissena polymorpha, substrate is not a limiting factor for D. bugensis (Berkman et al., 1998; Dermott and Munawar, 1993; Mellina and Rasmussen, 1994), and there is observational evidence that the dreissenid population in Lake Michigan is undergoing a shift from D. polymorpha to D. bugensis (Nalepa et al., 2009), and anecdotal evidence suggests that a similar shift may be occurring in Muskegon Lake. This population shift may further change the dynamic of the benthic community. Continued monitoring to track shifts in the effects of the mussels on the benthic population and the overall lake energy dynamics, combined with experimental manipulations to determine if dreissenids are having direct or indirect impacts on specific taxa or invertebrate guilds in Muskegon Lake, will be important for future lake management. Summary Past studies have shown that long-term monitoring can aid in the detection of trends that may not be identified in a shorter sampling timeframe (Stow et al., 1998; Urquhart et al., 1998). In systems such as Muskegon Lake that have histories of anthropogenic stresses and the resulting legacy effects of contaminants, continuous long-term monitoring is an essential element in assessing lake status. This study found spatial and temporal patterns from 2004 to 2010 in the invertebrate community, though the annual variation was relatively weak. However, recent studies on the status of Muskegon Lake environmental trends (Steinman et al., 2008) suggest that the overall condition of the lake is improving when compared to pre-wastewater diversion conditions. The overall positive trend in the benthic community we observed in the present study also suggests an improvement in the lake's environmental health, presumably resulting from wastewater diversion (cf. Carter et al., 2006) and ongoing remediation efforts. While we cannot single out an

16

W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18

mg kg-1, dry weight

35 30 25 20

a

15 10

b

5 0

1972

1999

b 2010

mg kg-1, dry weight

12

c 1972

1999

2010

1.0 0.8 0.6

a

6 4

0.4

b

2

c 1972

1999

2010

350

mg kg-1, dry weight

b

a

8

0

300

b c

0.2 0.0

1972

1999

2010

60

a

50

250

40

a

200 30

b

150

b 20

100 50 0

c 1972

1999

2010

c

10 0

1972

1999

2010

600

160

mg kg-1, dry weight

a

1.2

14

10

200 180 160 140 120 100 80 60 40 20 0

140

500

a

120 100

400

a

300

80

b

60 40

c

20 0

1972

1999

b

200

2010

c

100 0

1972

Year

1999

2010

Year

Fig. 4. Mean heavy metal concentrations (mg kg−1, dry weight) in sediments at 6 sites in Muskegon Lake, Michigan in 2010, and sediments from three sites in Muskegon Lake in 1972 (H-1, H-2, H-3) and 1999 (M-1, M-10, M-15). Significant differences in mean concentrations among years are indicated by letters (ANOVA, p ≤ 0.05). Probable effect concentrations (PEC) and threshold effect concentrations (TEC), from MacDonald et al. (2000), are indicated by dashed bold lines and solid lines, respectively (see text for explanation).

event that caused the lower benthic abundances in 2007–2009, it is notable that the invertebrate community in Muskegon Lake did appear to recover from a negative event, as seen by the recovery of the abundance values recorded for 2010. This suggests that the system is healthy enough to recover from impacts as severe as the sewage spill in 2007 and the duration of the hypoxia that likely resulted. The overall improvement in benthic community indices also may be indicative of a decreased legacy effect from contaminated sediments via burial or removal, or from a shift in system structure and function via Dreissena invasion. Managing systems in which multiple system stressors are occurring can be difficult. For this reason, long-term monitoring, combined with experimental manipulations and ecosystem modeling, will be essential elements in assessing lake status and overall trends, as both historic and future stressors (e.g., climate change) will continue to affect Muskegon Lake.

Acknowledgments We are indebted to the many AWRI staff that assisted in the collection and preservation of benthic samples and analysis of water samples in Muskegon Lake, including (but not limited to) Mary Ogdahl, Bopi Biddanda, Maggie Weinert, Brian Scull, and Scott Kendall. Additional thanks are owed to Geraldine Nogaro, Carl Ruetz, and Nadia Gillett for their help with statistics and Program R, to Mary Ogdahl for maintaining the Muskegon Lake database, and to Kurt Thompson for making the map. Allen Burton, Rick Rediske, Eric Snyder, and two anonymous reviewers provided helpful suggestions that improved this paper. Funding was provided by the Community Foundation for Muskegon County as a part of the Muskegon Lake Research Endowment Fund, and from the Grand Valley State University Presidential Research Grant.

W.A. Nelson, A.D. Steinman / Journal of Great Lakes Research 39 (2013) 7–18

17

Appendix A Taxa collected at 6 sites (3 seasons) and 7 years in Muskegon Lake, Michigan (n = 121). Density is the mean ± SE for each taxon and n is the number of samples where the taxon was found

Taxon Oligochaeta Lumbriculidae spp. Naididae Dero digitata Tubificidae Aulodrilus americanus Aulodrilus limnobius Aulodrilus pigueti Aulodrilus pluriseta Branchiura sowerbyi Ilyodrilus templetoni Isochaetides freyi Limnodrilus cervix Limnodrilus hoffmeisteri Limnodrilus maumeensis Limnodrilus udekemianus Monopylephorus helobius Quistadrilus multisetosus Rhyacodrilus Tubifex tubifex Immatures w/o hair chaetae Polychaeta Manayunkia speciosa Gastropoda Amnicola limosa Bithynia Bithynia tentaculata Menetus Promenetus Physa Valvata Valvata lewisi Valvata tricarinata Viviparus Bivalvia Corbicula fluminea Musculium Pisidium Pisidium ferrugineum Pisidium ventricosum Sphaerium Dreissena Dreissena polymorpha Dreissena rostriformis bugensis Isopoda Caecidotea Amphipoda Gammarus

Density, no. m−2 (n) 258 ± 25 (2) 172 ± 16 (1) 241 ± 25 (5) 230 ± 22 (3) 345 ± 31 (2) 360 ± 48 (11) 896 ± 120 (10) 172 ± 16 (2) 459 ± 66 (15) 172 ± 16 (1) 249 ± 23 (9) 2811 ± a 285 (47) 172 ± 16 (1) 459 ± 46 (27) 172 ± 16 (1) 457 ± 51 (23) 172 ± 16 (1) 624 ± 66 (8) 6674 ± 712 (52) 172 ± 16 (1) 344 ± 38 (11) 379 ± 41 (5) 689 ± 63 (1) 1137 ± 197 (5) 316 ± 31 (6) 172 ± 16 (2) 459 ± 49 (3) 574 ± 72 (6) 431 ± 47 (6) 172 ± 16 (1) 431 ± 45 (8) 172 ± 16 (2) 3191 ± 422 (34) 2708 ± 389 (18) 1493 ± 156 (3) 3617 ± 688 (11) 7788 ± 2264 (23) 1253 ± 214 (40) 13,951 ± 3013 (29) 498 ± 68 (9) 700 ± 152 (16)

References APHA, American Public Health Association, 1998. Standard Methods for the Examination of Water and Wastewater, 20th ed. American Public Health Association, Washington, D.C. Barbour, M.T., Gerritsen, J., Snyder, B.D., Stribling, J.B., 1999. Rapid Bioassessment Protocols for Use in Streams and Wadeable Rivers: Periphyton, Benthic Macroinvertebrates, and Fish, second edition. U.S. Environmental Protection Agency; Office of Water, Washington, D.C.(EPA 841-B-99-002). Berkman, P.A., Haltuch, M.A., Tichich, E., Garton, D.W., Kennedy, G.W., Gannon, J.E., Mackay, S.D., Fuller, J.A., Liebenthal, D.L., 1998. Zebra mussels invade Lake Erie muds. Nature 393, 27–28. Bray, J.R., Curtis, J.T., 1957. An ordination of the upland forest communities of southern Wisconsin. Ecol. Monogr. 27, 325–349. Burks, B.D., 1953. The mayflies, Ephemeroptera, of Illinois. Bull. Ill. Nat. Hist. Surv. 26, 1–216. Burton Jr., G.A., 1991. Assessing the toxicity of freshwater sediments. Environ. Toxicol. Chem. 10, 1585–1627. Carter, G.S., Nalepa, T.F., Rediske, R.R., 2006. Status and trends of benthic populations in a coastal drowned river mouth lake of Lake Michigan. J. Great Lakes Res. 32, 578–595. Clarke, K.R., 1993. Non-parametric multivariate analysis of changes in community structure. Aust. J. Ecol. 18, 117–143.

Taxon Diptera Ceratopogonidae Bezzia Probezzia Sphaeromias Chaoboridae Chaoborus Chironomidae Chironominae Chironomus Cladopelma Coelotanypus tricolor Cryptochironomus Dicrotendipes Harnischia Pagastiella Parachironomus Paralauterborniella Paratanytarsus Paratendips Polypedium halterale grp. Polypedium Polypedium trigonus Pseudochironomus Saetheria Stempellinella Stictochironomus Tanytarsus Orthocladiinae Orthocladius Prodiamesinae Monodiamesa Tanypodinae Ablabesmyia annulata Ablabesmyia Clinotanypus Coelotanypus Labrundina Procladius Ephemeroptera Brachycercus Caenis Hexagenia rigida Zygoptera Nehalennia Trichoptera Oecetis Neureclipsis Polycentropus

Density, no. m−2 (n)

344 ± 31 (1) 391 ± 45 (22) 306 ± 33 (9) 3389 ± 706 (46)

4797 ± 823 (107) 344 ± 31 (1) 172 ± 16 (1) 344 ± 31 (1) 172 ± 16 (3) 172 ± 16 (1) 287 ± 30 (3) 344 ± 35 (2) 230 ± 22 (3) 2743 ± 618 (13) 172 ± 16 (1) 404 ± 54 (23) 172 ± 16 (1) 431 ± 46 (2) 172 ± 16 (1) 517 ± 47 (1) 215 ± 20 (4) 1267 ± 191 (56) 172 ± 16 (3) 297 ± 28 (11) 301 ± 30 (4) 588 ± 73 (17) 517 ± 47 (1) 1967 ± 226 (31) 172 ± 16 (1) 2244 ± 236 (99) 172 ± 16 (1) 197 ± 18 (7) 313 ± 32 (11) 172 ± 16 (1) 207 ± 20 (5) 586 ± 92 (5) 172 ± 16 (1)

Cook, E.F., 1956. The Nearctic Chaoborinae (Diptera: Culicidae). Univ. Minn. Agr. Exp. Sta. Tech. Bull. 218, 1–102. Cooper, M.J., Uzarski, D.G., Burton, T.M., 2007. Macroinvertebrate community composition in relation to anthropogenic disturbance, vegetation, and organic sediment depth in four Lake Michigan drowned river-mouth wetlands. Wetlands 27, 894–903. DEQ. Michigan Department of Environmental Quality Water Bureau, 2008. Combined Sewer Overflow (CSO) & Sanitary Sewer Overflow (SSO) 2007 Annual Report. Lansing, MI (michigan.gov/documents/deq/deq-wb-crorroreport07_243725_7.pdf). Dermott, R., Munawar, M., 1993. Invasion of Lake Erie off-shore sediments by Dreissena, and its ecological implications. Can. J. Fish. Aquat. Sci. 50, 2294–2304. Dufrene, M., Legendre, P., 1997. Species assemblages and indicator species: the need for a flexible asymmetrical approach. Ecol. Monogr. 67, 345–366. Epler, J.H., 2001. Identification Manual for the Larval Chironomidae (Diptera) of North and South Carolina. North Carolina Department of Environmental and Natural Resources, Raleigh, NC. Evans, E., 1976. Final Report of the Michigan Bureau of Water Management's Investigation of the Sediments and Benthic Communities of Mona, White, and Muskegon Lakes, Muskegon County, Michigan, 1972. Michigan Bureau of Water Management, Lansing, MI.

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