Effects of invasive macrophytes on littoral-zone productivity ... - BioOne

J. N. Am. Benthol. Soc., 2005, 24(2):300–320 䉷 2005 by The North American Benthological Society

Effects of invasive macrophytes on littoral-zone productivity and foodweb dynamics in a New Zealand high-country lake DAVID J. KELLY1 National Institute of Water and Atmospheric Research, PO Box 8602, Riccarton, Christchurch, New Zealand

IAN HAWES2 National Institute of Water and Atmospheric Research, PO Box 11-115, Hamilton, New Zealand Abstract. Invasion of littoral zones by adventive macrophyte species can facilitate major changes in the ecology of lakes. In Lake Wanaka, a large alpine New Zealand lake, the macrophytes Lagarosiphon major and Elodea canadensis (Hydrocharitaceae) have invaded parts of the lake where they form tall dense plant beds throughout mid-depths (2–7 m) of the littoral zone. We investigated differences in plants, benthic invertebrates, fish, and food webs characterizing native and exotic plant beds in mid-depths of the littoral zone. The 3⫻ higher plant biomass and 2⫻ higher plant surface area in exotic than in native plant beds (quillworts, milfoils, and charophytes) contributed to greater standing stocks and productivity of epiphyton in the exotic plant beds. Invertebrate communities were less dense (1890/m2 vs 4030/m2) and less diverse (richness ⫽ 9 vs 12) in native than in exotic plant beds because of differences in biomass, productivity, and physical structure of native and exotic plant communities. Invertebrate communities in native beds were dominated by snails, oligochaetes, and nematodes, whereas chironomids, snails, and caddisflies were dominant in exotic beds. Stable isotope signatures (13C and 15N) and dietary analyses indicated that Potamopyrgus antipodarum, the dominant invertebrate taxon in both bed types, consumed mostly epiphyton. In native beds, consumption of sedimentary fine benthic organic matter by oligochaetes and nematodes made significant contributions to C flow, whereas, in exotic beds, consumption of epiphyton by grazers (e.g., snails, caddisflies, and chironomids) was an important pathway for C flow. Macrophytes made only small contributions to C flow in either bed type. The dominant native fish in Lake Wanaka, the bully Gobiomorphus cotidianus, was more abundant in exotic than in native beds, but bully predation rates on snails were significantly lower in exotic than in native beds. Invasion by adventive macrophyte species can cause significant shifts in lake productivity, species composition, and foodweb dynamics. Key words: littoral zone, productivity, food web, benthic invertebrates, adventive macrophytes, benthic fish, stable isotopes, Lagarosiphon major, Elodea canadensis, Gobiomorphus cotidanus, Potamopyrgus antipodarum.

Macrophytes are a key component of many lakes and are involved in a diverse array of ecosystem processes, many of which relate to the complex 3-dimensional architecture characteristic of most macrophyte beds (Loeb and Reuter 1981). The productivity of macrophyte beds is a major avenue of energy flow, particularly in oligotrophic lakes, and can be a useful correlate of fish productivity and biomass (Hanson and Leggett 1982, Rasmussen and Kalff 1987, Vander Zanden and Vadeboncoeur 2002). Macrophytes are important primary producers, and they form a surface for the growth of epiphytic algae that is important food for invertebrate grazers (Loeb and Reuter 1981, Suren and Lake 1 2

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1989, James et al. 2000a). Plant beds provide refuge from predators (Gotceitas 1990, Connolly 1994, Valley and Bremigan 2002), and dense vegetation may be sought specifically by prey species, many of which live exclusively in plant beds (Stoner 1980, Uiblein et al. 1996). The structure of plant canopies also affects movement of water through the littoral zone and influences environmental conditions such as dissolved O2 and dissolved inorganic C concentrations, pH, temperature, and light, and settling rates of phytoplankton and seston (Carter et al. 1991, Schwarz and Howard-Williams 1993). The native flora of New Zealand lakes consists of a distinctive group of littoral macrophyte species distributed along a series of depth and water-clarity zones (Hawes et al. 2003a). Typically, New Zealand littoral macrophytes are

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MACROPHYTES AND LITTORAL FOOD WEBS

low-growing, and an archetypal zonation would extend from a diverse turf-forming community of angiosperms and pteridophytes at depths of 0 to 2 m, through an Isoetes-dominated zone growing to 20 cm height in water up to 10 m deep, to a species-rich characean community growing to 20 to 30 cm in height in water up to 40 m deep. Over the past century, many New Zealand lakes have undergone rapid invasion by adventive macrophyte species (Howard-Williams et al. 1987). Many of these species are tall canopy-forming hydrocharitacean species that are able to out-compete the lower-growing native flora. The tall, dense, and oligospecific canopies of these adventives are quite different from the lower, more open, and species-rich canopies of native plants. The differences in flora may affect the use of macrophyte beds by littoral invertebrate and fish species. A combination of field surveys and experiments was used to examine differences between native plant beds and beds that have been invaded by hydrocharitacean adventives in Lake Wanaka, a high-country lake in New Zealand. We hypothesized that changes in the macrophyte host species from native low-growing forms to exotic tall stands of adventives would have significant effects on the architecture of the plant canopy, productivity of plant beds, resources available to consumers, and energy transfer through the littoral food web. The structure of the plant canopies; biomass and productivity of macrophytes and epiphytes; zoobenthic density, diets, and community composition; fish abundance; and energy flow as assessed by stable isotopic ratios of biota were compared in native and exotic plant beds. We speculated that invasive macrophytes might increase littoral habitat, thereby increasing productivity of benthic primary and secondary producers and availability of food to littoral fish. However, increased density and height of plant canopies might negate the effect of increased food availability by impairing foraging capacity. Potential effects of differences in plant canopy architecture on fish predation were tested in laboratory foraging trials in aquaria with different macrophyte species/density combinations. Trials were conducted using the common bully (Gobiomorphus cotidanus), the dominant native benthic fish in Lake Wanaka, and the common snail Potamopyrgus antipodarum. Our study integrated components from a whole foodweb per-

TABLE 1. Water-quality variables during the 2001 study period at 6 sites on Lake Wanaka, New Zealand. SD ⫽ standard deviation.

Secchi depth (m) Light attenuation (Kd, /m) Planktonic chlorophyll a (mg/m3) Dissolved reactive P (mg/m3) Total dissolved N (mg/m3)

Mean

SD

12.2 0.21 0.45 0.28 78.2

1.3 0.15 0.08 0.08 5.9

spective, so a survey approach was taken to examining our hypotheses, rather than a rigorous detailed examination of a particular component of the food web. Methods Study site Lake Wanaka is particularly suited for studying changes to the ecology of lakes after macrophyte invasions because hydrocharitacean weeds currently are confined to certain parts of the lake. The rest of the lake is inhabited by native plants growing along shores with aspect, exposure, and slope comparable to those invaded by the adventives. Lake Wanaka (lat 44.5⬚S, long 169⬚E) is a large (180 km2), deep (maximum depth ⫽ 311 m), monomictic, oligotrophic lake in the Otago high country on the South Island of New Zealand. The catchment consists mainly of natural tussock grassland (56%) and beech forest (14%), with a low proportion of land used for farming and urban development (⬍5%). The lake has low concentrations of dissolved nutrients, low phytoplankton biomass (mean chlorophyll a [chl a] ⫽ 0.45 mg/m3), and high water transparency (Kd(PAR) ⫽ 0.21/m) (Table 1). Like other large glacial lakes in New Zealand, the lake has low seasonal variation in temperature (surface temperature ⫽ 8.5–16.9⬚C, bottom temperature ⫽ 8.8–9.0⬚C) and, in summertime, has a deep seasonal thermocline (30– 40 m). Further limnological details are reported by Livingstone et al. (1986) and Hawes et al. (2003a, b). The littoral zone of Lake Wanaka is typical of most large oligotrophic lakes in New Zealand. It consists of 3 habitat zones based on plant community composition (Wells et al. 1998): 1) a shallow mixed-species zone in water from 0 to

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3 m deep, 2) a tall vascular zone in water from 3 to 8 m deep, and 3) a characean meadow zone in water from 9 m to the maximum depth of the plant zone (typically ⬃17 m in Lake Wanaka). The hydrocharitacean species Elodea canadensis and Lagarosiphon major have invaded the tall vascular zone in parts of Lake Wanaka. Both of these canopy-forming species tend to completely displace native plants, forming tall (sometimes reaching to the surface) dense monospecific or mixed-bispecific stands in water between 2 and 7 m deep. Lagarosiphon major was first detected in the southeastern part of the lake in 1972 and is now found throughout 3 sub-basins in the southern part of the lake, including Paddock–Parkins Bay, Roys Bay, and Stevensons Arm (Fig. 1). No initial record of E. canadensis in Lake Wanaka exists, but the species has been in the lake for considerably longer and has a much wider distribution than L. major. Native species in the tall vascular zone in Lake Wanaka are predominantly low-growing Isoetes alpinus intermixed with some taller Potamogeton cheesemanii, Myriophyllum triphyllum, and some Nitella and Chara spp. The 6 sites, 3 native beds and 3 exotic beds, selected for our study were sheltered from the predominantly northwest and southwest winds. Sites had similar shoreline profiles and the same sand–silt substrate and were in water between 3 and 5 m deep (Fig. 1). At each site, a 4 ⫻ 4 m area was delineated with weighted transect lines from which all surveys of littoral biota occurred. All sites were clear of any inflows, but the Beacon Point site was close to the main outflow (the Clutha River). Adventive hydrocharitaceans were found near each of the native sites, suggesting that all sites could have supported exotics and that the sites were comparable. Plant and epiphyton abundance Macrophyte cover and biomass were estimated during a single survey (July 2001) using 2 methods. Percent cover and canopy height in 1m2 quadrats (ⱖ10 quadrats/site) were measured by SCUBA divers, and dry mass (DM) was determined from plants collected using diver-operated Wisconsin grabs (250-␮m mesh size, grab area ⫽ 0.09 m2, 3 grabs/site). Plants from Wisconsin grabs were sorted by species, dried overnight (60⬚C), and weighed. Plants were identified using the keys in Johnson and Brooke

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(1989). Surface area (SA) of plants per unit area of the littoral zone was calculated for each site by averaging the SA of plants contained in the 3 grabs and dividing by the sampling area. Macrophyte SA/DM was calculated from plant cuttings (ⱖ20 cuttings/species) that were flattened and measured using a Licor LI3100 leafarea analyser (Licor, Lincoln, Nebraska), and then dried (60⬚C) and weighed. Macrophyte SA/DM was determined for each species by individual linear regressions. Total SA of macrophytes in each grab was estimated by multiplying the DM of each species by the SA to DM ratio (SA/DM), and summing across all the species present. Epiphyton biomass at each site was estimated from chlorophyll a (chl a) biomass on artificial plants of known surface area, and converted to /m2 littoral zone by multiplying chl a biomass by the estimated macrophyte leaf area/unit area littoral zone. Artificial plants, constructed from clear plastic strips attached to acrylic rods, had one of 2 phenotypes that mimicked the morphology of either L. major (exotic sites) or I. alpinus (native sites). Artificial plants were placed within plant beds in July 2001 and sampled after 2 and 5 mo. Triplicate plastic strips were collected from each site and analyzed separately. Chl a was extracted by boiling in ethanol for 10 min and measured as absorbance at 750, 665, and 664 nm on an Ocean Optics USB 2000 Spectrophotometer equipped with a LS1 tungsten halogen lamp. Chl a concentration was corrected for phaeophytin content by measuring absorbance of extracts after acidifying with 1 N HCL. Concentrations were used to calculate chl a as mg/m2. Samples of macrophytes and epiphytes for stable isotope analyses (see below) were taken from plants collected with Wisconsin grabs. Epiphyton was removed from live macrophytes (⬎5 plants/grab) using a soft bristle brush, and examined under 400⫻ magnification to verify that the material contained little or no macrophyte tissue. Benthic primary productivity Macrophyte and epiphyton primary productivity were measured over a 2-d period (10–11 March 2002) at 2 sites, 1 exotic and 1 native, using standard in situ 14C incubation techniques (Wetzel and Likens 1991). Plant cuttings (1st

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FIG. 1. Study area in Lake Wanaka, New Zealand, showing the locations of transects in the native and exotic plant beds (3 sites in each bed type).

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day) and epiphyton-covered plastic strips (2nd day) were incubated in the morning (0900–1200 h) and afternoon (1300–1600 h) at each site. Incubations were deployed in triplicate at 3 depths within the canopy (top, middle, and bottom) at the exotic site; only one depth was necessary at the native site where canopy height was the same as the height of the incubation bottle. Plant materials consisted of an ⬃4-cm clipping of either L. major (exotic site), I. alpinus (native site), or epiphyton-covered plastic strips (both sites) collected from the particular site and appropriate canopy position. Plants were cleaned of epiphyton with a soft bristle brush prior to incubation. Plants were incubated in 500-mL glass bottles of lake water from the particular plant bed. Bottles were inoculated with 1 mL of NaH14CO3 (2.5 mCi/mL) and deployed at collection depths on poles staked in the macrophyte canopy. Single dark bottles were deployed with plants to serve as blanks in the analyses. After incubation, plants were removed from the bottles, rinsed with 1 N HCL solution to remove any radioactive material adhering to the plants, sealed in scintillation vials, and frozen until analysis. Water samples were collected from each incubation depth at the start of the incubations for analysis of dissolved inorganic C (DIC) by infrared gas analysis following stripping with N. Alkalinity was determined by titration. Total surface irradiance (quanta) was logged over the 2-d period with a Licor LI1000 data logger and quantum cosine sensor. Incubations took place during near-cloudless conditions. On return to the laboratory, the materials were freeze-dried, homogenized with a mortar and pestle, and weighed into scintillation vials. Plant material was digested in vials with 0.2 mL of NCS (Amersham Biosciences, Piscataway, New Jersey) tissue solubilizer incubated in a water bath at 60⬚C for 4 h. Plastic strips were not pretreated. Ten mL of Hisafe scintillation cocktail mix (Wallac, Turku, Finland) were added to vials and radioactivity was assayed on a LKB1217 Rackbeta scintillation counter (Wallac). Counts of radiation were corrected for quenching and background, and blanks were prepared from plant materials incubated in dark bottles. C uptake was estimated as the proportion of radiolabelled C incorporated multiplied by the DIC content of the incubation bottle and normalized to plant freeze-dried mass (macro-

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phytes) or area (plastic strips). The sums of the two 3-h incubations (morning and afternoon) were converted to daily photosynthetic rates in each plant bed type by dividing by the proportion of total daily surface irradiance incident during the incubations, and to per unit area by multiplying by mass/unit area. In exotic plant beds, where photosynthetic rates were measured at 3 points through the canopy, daily photosynthetic rates for the whole plant canopy were estimated by integration of the 3 canopy zones, assuming a uniform distribution of plant biomass through the canopy. Phytoplankton and zooplankton Suspended particulate organic matter (POM) and zooplankton samples were collected at 6 stations associated with littoral sites that were ⬃200 m offshore from littoral transects. Phytoplankton biomass was sampled 4 times in 2001 at 5 offshore stations associated with littoralzone sites by filtering 1 L of near-surface water onto Whatman GF/F filters. Chl a extraction and quantification was as described for periphyton samples. POM samples for isotopic analysis were collected in the same manner as for biomass and filtered onto 25-mm precombusted Whatman GF/F filters. Zooplankton were collected in 10-m vertical hauls using an 80-␮m plankton net and stored at 4⬚C until transported to the lab for identification and counting at the species level. Representative individuals were retained for isotopic analysis. Benthic macroinvertebrate density Surveys of benthic macroinvertebrates were done at the 6 sites (3 native and 3 exotic) on the same dates as plant surveys. Three replicate Wisconsin grabs were taken from within each 16-m2 plot. The contents of grabs were stored separately in watertight plastic bags and transported to the laboratory for analyses. Benthic macroinvertebrates from each replicate grab sample were counted separately. Macrophytes from grabs were washed and removed; the remaining contents (invertebrates and detritus) were split in half using a barrel splitter. Half of the contents were sieved (250-␮m mesh) and preserved in 70% ethanol for taxonomic analyses, and the remainder was sorted live for stable isotope analyses or gut content analyses (see be-

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low). Counts of invertebrates were conducted at the lowest taxonomic level possible. In some cases, this level was species, but for some taxonomic groups identification was only to phylum (nematodes), class (oligochaetes and mites), or family (dipterans). Taxonomic samples were subsampled using a barrel splitter, and ⱖ25% of the entire sample was identified until ⱖ100 individuals of the most abundant taxa were counted. Invertebrates were identified according to keys by Winterbourn (1973), Winterbourn et al. (2000), and Thorp and Covich (1991) using dissecting microscopes. Dietary analyses of benthic macroinvertebrates Invertebrates sorted live from Wisconsin grabs were frozen in vials until analysis. The digestive tracts of the dominant invertebrate species, P. antipodarum (Gastropoda:Prosobranchia), Paroxyethira tillyardi (Trichoptera:Hydroptilidae), and Chironomus zealandicus (Diptera: Chironomidae), from each site were surgically removed under a dissecting microscope. Digestive tracts were placed on microscope slides and broken open using a scalpel and forceps. The contents were mounted in cornstarch solution with 4% formalin. Gut contents were counted under 200⫻ magnification and sorted into 6 categories: diatoms, cyanobacterial filaments, other filaments, animal, amorphous organics, and inorganic materials. The number of 100-␮m2 field areas (10 ⫻ 10-␮m grids) occupied by each item was used as the count unit to correct for variability in the sizes of different items. Randomly chosen fields were examined until ⱖ200 units were counted; once counting was initiated the entire field was counted. Fish abundance Abundances of common bullies were estimated from overnight sets of Gee-minnow trap lines in the 2 types of plant beds in December 2002. Trap lines that extended vertically through the plant canopy were used to correct for differences in canopy density between exotic and native plant beds and to prevent burying of traps in dense weeds. Three trap lines were set within the 16-m2 plot at each of the 6 sites. Each line consisted of 3 traps spaced ⬃50 cm apart running vertically through the plant canopy. The lines were suspended from a surface buoy

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and tightly anchored to the bottom, with traps at distances of 0, 50, and 100 cm from the lake bottom. Traps were set early in the evening (1900–2000 h) and retrieved the following morning (1000–1200 h). Bullies in each trap were counted and measured (fork length). Bully abundances were expressed as catch per unit effort (CPUE), calculated as the number of bullies (trap line)⫺1 h⫺1. Bullies were kept alive in buckets and brought back to the laboratory for feeding experiments (see below). Bullies for stable isotope analyses were trapped at the same times as macroinvertebrate surveys were conducted (July 2001) and were killed immediately and frozen until analysis. Stable isotopes Plant and macroinvertebrate specimens from replicate Wisconsin grabs were pooled from each site for the analyses. Macrophytes were washed, freeze-dried, and homogenized with a mortar and pestle before being weighed into individual tin capsules for isotopic analyses on a mass spectrometer. Because CaCO3 concentrations are low in New Zealand lakes, macrophytes did not require acid-rinsing to remove marl deposits prior to isotopic analysis. Invertebrates were left at 4⬚C for 24 h to allow gut contents to evacuate. Whole macroinvertebrates (typically ⱖ5 individuals, fewer for some rare taxa) were freeze-dried, homogenized, and lipid-extracted before being weighed into capsules for isotopic analysis. Invertebrates such as snails and caddisflies were removed from cases or shells prior to processing. Surface sediments beneath macrophytes were collected separately in 20-mL vials by divers, treated with 5% HCL to remove carbonate deposits (e.g., snail shells), freeze-dried, and processed as for macrophytes. Common bully muscle tissue was removed from the rear quarter of the fish above the anal fin, freeze-dried, and processed as for invertebrates. Lipids were extracted from all animal tissues on a Dionex ASE 200 Accelerated Solvent Extractor (Dionex Corporation, Sunnyvale, California) using 100% dichloromethane. Samples were extracted 2⫻ in cells heated to 70⬚C at 2000 psi with static hold periods of 5 min and a 100% flush volume, after which they were dried at 40⬚C for 12 h to evaporate residual solvent. Stable isotopes of C and N were analyzed by mass spectrometry on a Finnigan MAT Delta Plus

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Continuous Flow Mass Spectrometer (Thermo Finnigan, San Jose, California). Results are reported as the relative difference in ‰ between the sample and international standards of C and N (Pee Dee Belemnite for C and air for N) with an analytical precision of 0.1‰. Feeding rates of common bullies Feeding rates of common bullies on the snail P. antipodarum in differing plant types was tested experimentally in laboratory aquaria. Potamopyrgus antipodarum was observed in the guts of bullies during dissections for stable isotope analyses, so a clear feeding relationship was already established. Moreover, P. antipodarum was the most abundant and most robust invertebrate in all habitat types, making it a logical choice for this experiment. Six common bullies obtained from fish trapping in Lake Wanaka were placed in each aquarium (22 ⫻ 44 ⫻ 22 cm, W ⫻ L ⫻ H). Aquaria contained either I. alpinus, E. canadensis or L. major, and 3 replicate aquaria of each plant type were used (total of 9 aquaria). Plants were washed thoroughly to remove any invertebrates, and then anchored firmly to the substrata (pebbles) in the aquaria at uniform densities intended to mimic field densities for the particular plant species. The 2 hydrocharitacean species were shorter in aquaria (⬃20 cm) than in the field (mean ⫽ 88 cm) for logistic reasons. The sides of each aquarium were screened with black plastic to prevent fish from seeing into neighbouring aquaria, and the plant type in each aquarium was randomly determined. Lighting in the room was operated on a 12:12 h light:dark cycle. Each aquarium was stocked daily with 15 snails during the 5-d feeding trial. The snails were the only food supplied to the bullies. Snails were randomly scattered over the water surface at the time of their introduction. On the 6th day, the fish were removed from the aquaria, and the plants and substrata were checked carefully for the number of live snails remaining. Empty snail shells were presumed to indicate that the occupants had been consumed and their shells had passed through the bully guts intact. Feeding rates were expressed as the number of snails consumed (missing or empty shells) bully⫺1 d⫺1.

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Statistical analyses Mixed-model 2-way analysis of variance (ANOVA) was used to test for differences in several variables between plant types (native or exotic) and sites, with subreplicate measures of variables at a particular site nested as a random factor within plant type. Separate ANOVAs were run on macrophyte biomass, macrophyte surface area, epiphyton biomass, epiphyton productivity, macrophyte productivity, total invertebrate density, invertebrate species richness, invertebrate Shannon–Wiener diversity, common bully CPUE, and common bully fork length. A 2-way multivariate analysis of variance (MANOVA) was used to test for differences in log10(n ⫹ 1)-transformed densities of the major invertebrate taxonomic groups (Diptera, Trichoptera, Odonata, Mollusca, and others) between plant types and sites, with subreplicate site measures nested within vegetation type. Major taxonomic groupings were used because the model had insufficient degrees of freedom to analyze all taxa independently. Two-way MANOVA was used to test for differences in gut contents (diatoms, organics, inorganics, and animal) of invertebrate taxa between plant types and sites, with replicate site measures nested within plant type. Separate MANOVAs were run for each of the 3 taxa (C. zealandicus, P. tillyardi, and P. antipodarum), and all data were arcsine-transformed proportions of the total gut contents. One-way ANOVAs were used to test differences in predation rate, phytoplankton biomass, and sediment organic content between vegetation types. Relationships between invertebrate communities and plant productivity indicators were investigated using full stepwise multiple regression analysis. Invertebrate-community-dependent variables included densities of the major invertebrate taxonomic groups (as for MANOVA), total invertebrate density, species richness, and Shannon–Wiener diversity. Independent predictor variables included macrophyte biomass, macrophyte surface area, epiphyton biomass, epiphyton productivity, macrophyte productivity, and common bully CPUE. Autocorrelated variables (tolerance coefficient, p ⬍ 0.001) were removed before the regression analysis. All analyses were performed using SYSTAT statistical software (version 10, SPSS, Chicago, Illinois).

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TABLE 2. Macrophyte and littoral-zone variables in 3 native and 3 exotic plant beds in Lake Wanaka, New Zealand. – ⫽ species not present. Plant beds Native

Cover (%) Elodea canadensis Lagarosiphon major Isoetes alpinus Myriophyllum triphyllum Potamogeton cheesemanii Nitella pseudoflabellata Macrophyte canopy height (cm) Macrophyte biomass (g/m2) Littoral slope (0–6 m (drop/run) Sediment organics (% C)

Exotic

Beacon Point

Dublin Bay

Glendhu Bay

– – 95 3 2 – 32 69 0.29 1.1

– – 90 8 – 2 24 63 0.21 1.6

5 – 75 15 5 – 30 73 0.39 2.8

Results Plant biomass and productivity All native beds were dominated by I. alpinus (mean cover ⫽ 87%) with some intermixed M. triphyllum (mean cover ⫽ 7%) and P. cheesemanii (mean cover ⫽ 2%), and occasionally Nitella spp. and E. canadensis (Table 2). Exotic beds consisted exclusively of the 2 adventives, L. major and E. canadensis, but varied in their % composition (mean L. major cover ⫽ 60%, mean E. canadensis cover ⫽ 40%). On average, plants in exotic beds were ⬎3⫻ taller (88 cm vs 29 cm; difference not tested) and had ⬎3⫻ greater biomass (221.2 g/ m2 vs 68.3 g/m2 littoral zone, p ⫽ 0.005; Fig. 2A) than plants in native beds. Plant biomass was more variable in exotic than in native beds, but ANOVA for site nested within bed type showed no significant intersite differences in any vegetation variable. The surface area for epiphyton colonization was ⬎2⫻ greater in exotic than in native beds (6.8 m2/m2 vs 3.0 m2/m2 littoral zone, p ⫽ 0.017; Fig. 2B) and, therefore, the biomass of epiphyton was higher (77 mg chl a/m2 vs 16 mg chl a/m2 littoral area, p ⬍ 0.017; Fig. 2C) in exotic than in native beds. Epiphyton primary productivity was 41% higher (148 mg C d⫺1 m⫺2 vs 105 mg C d⫺1 m⫺2 littoral zone, p ⬍ 0.05; Fig. 2D) in exotic than in native beds, and macrophyte primary productivity was 10⫻ greater (2380 mg C d⫺1 m⫺2 vs 216 mg C d⫺1 m⫺2 littoral zone, p ⫽ 0.0001; Fig. 2E). Phytoplankton biomass did not

Paddock Bay

Parkins Bay

Glendhu Bluff

35 65 – – – – 57 111 0.42 2.0

70 30 – – – – 68 236 0.25 1.9

15 85 – – – – 140 537 0.30 5.5

differ between exotic and native beds (0.56 mg chl a/m3 vs 0.43 mg chl a/m3, p ⫽ 0.45; Fig. 2F). Organic content of sediments (% C) was 72% higher beneath exotic beds than native beds (3.1% vs 1.7%), but the difference was not significant because intersite variability was high (p ⫽ 0.33; Table 2). Invertebrate density and fish abundance Densities and taxonomic composition of benthic communities differed between exotic and native beds (MANOVA, p ⫽ 0.003; Table 3, Fig. 3). Variability in the densities and taxonomic composition of invertebrates collected from beds of the same plant type was low, and no significant effect of site nested within bed type was detected for any of the taxonomic groups (Table 3). Densities of chironomid taxa, mainly C. zealandicus, were markedly greater in exotic than in native beds (874/m2 vs 7/m2 littoral zone, p ⬍ 0.001; Fig. 3). Densities of Trichoptera, mainly Paroxyethira spp., were also significantly greater in exotic than native plant beds (603/m2 vs 112/m2 littoral zone, p ⫽ 0.003; Fig. 3), as were densities of Odonata (32/m2 vs 3/m2 littoral zone, p ⬍ 0.001; Fig. 3). The most abundant invertebrate taxon in both bed types was P. antipodarum, which made up ⬃30% of total density in both bed types. Density of P. antipodarum was almost 2⫻ greater in exotic than in native beds (908/m2 vs 512/m2 littoral zone; Fig. 3), but density of Physastra variabilis was greater in

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FIG. 2. Mean (⫹1 SE, n ⫽ 3) macrophyte productivity indicators in native and exotic beds in Lake Wanaka. A.—Macrophyte biomass. B.—Macrophyte surface area. C.—Epiphyton biomass. D.—Epiphyton productivity. E.—Macrophyte productivity. F.—Phytoplankton biomass. Bars sharing the same letter are not significantly different (p ⱖ 0.05). chl a ⫽ chlorophyll a.

native than in exotic beds, so Mollusca density did not differ between bed types (p ⫽ 0.45; Table 3). Densities of nematodes, oligochaetes, and mites (Others) were greater in native than in ex-

otic beds (p ⫽ 0.006; Table 3). Together, these other taxonomic groups made up ⬃50% of the total invertebrate density. Total invertebrate density was, on average,

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TABLE 3. Nested analysis of variance (ANOVA) comparing total invertebrate density, species richness (richness), Shannon–Wiener diversity (diversity) between native and exotic plant beds (plant) at 6 sites (site nested within plant bed type), and nested multivariate analysis of variance (MANOVA) comparing densities (no./m2 littoral area) of major invertebrate groups between native and exotic plant beds (plant) at 6 sites nested within plant bed type (Site[plant]). Mean square values (n ⫽ 3) apply to ANOVA results and the univariate test for each of the 5 main taxonomic groups. Hotteling’s H statistic applies to the MANOVA for all groups. All data were log10 (x ⫹ 1)-transformed with the exception of diversity indexes. Others ⫽ Oligochaeta, Nematoda, Hydracarina, and Copepoda.

Group ANOVA Total density

Richness

Diversity

MANOVA

Diptera

Trichoptera

Odonata

Mollusca

Others

Source

df

Mean square/ Hotteling’s H

Plant Site (plant) Error Plant Site (plant) Error Plant Site (plant) Error Plant Site (plant)

1 4 12 1 4 12 1 4 12 5 20

0.21 0.1 0.1 0.30 0.48 0.1 0.1 0.2 0.1 6.3 2.4

Plant Site (plant) Error Plant Site (plant) Error Plant Site (plant) Error Plant Site (plant) Error Plant Site (plant) Error

1 4 12 1 4 12 1 4 12 1 4 12 1 4 12

82% greater in exotic than native beds (3357/m2 vs 1839/m2 littoral zone, p ⫽ 0.009; Table 3, Fig. 4A). Taxonomic richness and Shannon–Wiener diversity were marginally higher in exotic than native beds, but neither difference was statistically significant (Table 3, Fig. 4B, C). Most of the taxa observed in native plant beds also were observed in areas invaded by exotic plants, and no taxa were specific to exotic beds. Foodweb dynamics and invertebrate diet Differences in C and N isotopic ratios in primary producers made it possible to trace foodweb pathways in both exotic and native beds.

18.4 0.1 0.8 1.9 0.1 0.1 7.2 0.3 0.3 0.1 0.1 0.1 7.5 1.5 0.7

F

p

9.6 0.5

0.009 0.750

3.7 5.8

0.080 0.008

0.6 2.9

0.463 0.065

1.1 0.8

0.003 0.730

24.1 0.2

⬍0.001 0.955

13.4 1.0

0.003 0.583

24.7 1.0

⬍0.001 0.441

0.6 0.6

0.449 0.680

11.4 2.4

0.006 0.112

In native beds, primary producers were best segregated by ␦13C ratios, and ␦15N ratios allowed less discrimination (Fig. 5A). In general, ␦13C ratios of vascular plants other than I. alpinus (⫺15.5‰) were higher than those of other primary producers. However, ␦13C ratios of I. alpinus (⫺21.3‰) were more similar to those of epiphyton (⫺21.3‰) than those of other vascular plants. ␦13C ratios of POM (⫺28.0‰) were the most depleted among the primary producers in both bed types. In general, ␦15N ratios of other vascular plants and POM were similar (2– 3‰), whereas ␦15N ratios of epiphyton (1.51‰) were lower than those of other vascular plants or POM.

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Mean (⫹1 SE, n ⫽ 3) density of zoobenthic invertebrate taxa in native and exotic beds in Lake

Trends in ␦13C ratios of primary producers in exotic beds were similar to those in native beds (Fig. 5B). ␦13C ratios of vascular plants (⫺14.2‰) were higher than those of other primary producers, and epiphyton ␦13C ratios (⫺22.3‰) were intermediate between those of vascular plants and POM (⫺28.3‰). Unlike in native beds, where I. alpinus differed in its ␦13C ratio from the rest of the vascular plants present, all macrophytes in exotic beds had similar ␦13C ratios. ␦15N ratios did not differ among vascular plants (1.2‰), epiphyton (2.0‰), and POM (1.8‰). ␦13C ratios of fine benthic organic matter (FBOM) collected from below macrophyte beds were approximately midway between those of the macrophytes and POM and close to those of I. alpinus and epiphyton (Fig. 5A, B). In native plant beds, ␦13C ratios of FBOM (⫺21.3‰) were nearly identical to those of I. alpinus, the dominant macrophyte species present. In exotic plant

beds, ␦13C ratios of FBOM (⫺22.5‰) were 8 to 9‰ lower than those of macrophytes. ␦15N ratios of FBOM in native beds (3.0‰) were higher than those in exotic beds (0.9‰). A trophic pathway linking POM to zooplankton, and zooplankton to juvenile bullies was evident in both bed types (Fig. 5A, B). Both C and N isotope ratios were enriched with each trophic link, though the degree of N enrichment varied. The dominant benthic invertebrate in both bed types, P. antipodarum, had similar isotopic ratios in native (␦13C ⫽ ⫺18.0‰, ␦15N ⫽ 2.3‰) and exotic (␦13C ⫽ ⫺19.0‰, ␦15N ⫽ 3.1‰) beds. Both ratios suggested a diet consisting mainly of epiphyton. Similarly, the snail P. variabilis and the caddisfly Paroxyethira spp. also had isotopic ratios indicative of an epiphyton-dominated diet in both bed types. The small amount of variability in the isotopic signatures of these 3 epiphyton grazers between exotic and native beds

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FIG. 4. Mean (⫹1 SE, n ⫽ 3) total invertebrate density (A), taxon richness (B), and Shannon–Wiener diversity (H’) (C) in native and exotic beds in Lake Wanaka. Bars sharing the same letter are not significantly different (p ⱖ 0.05).

may have been influenced by ingestion of basal macrophyte material in addition to epiphyton. The isotopic ratios of oligochaetes and nematodes collected from native beds and oligochaetes from exotic beds indicated diets consisting mainly of FBOM. Greater enrichment of ␦15N for C. zealandicus (␦15N ⫽ 2.7‰) over oligochaetes (␦15N ⫽ 1.8‰) in exotic beds indicated that C. zealandicus consumed a combination of FBOM and epiphyton. Isotopic ratios of Odonata, mainly Procordulia smithi and Xanthonemis zealandica, indicated that, in native beds, these taxa fed on a combination of oligochaetes, Paroxyethira, and nematodes, whereas in exotic beds, C. zealandicus probably was the dominant prey. The results of stomach content analyses of 3 of the dominant invertebrate taxa, C. zealandicus, Paroxyethira tillyardii, and P. antipodarum, were consistent with dietary patterns suggested by the isotope analyses. Bed type did not influence the gut contents of any of the taxa (MANOVA, p ⫽ 0.64; Fig. 6A, B). Chironomus zealandicus, found mostly in exotic beds, consumed amorphous organic material (74%), some diatoms (12%), and inorganic material (14%). Paroxyethira tillyardii consumed a greater proportion of diatoms (41% in native and 61% in exotic beds)

than C. zealandicus, and amorphous organic and inorganic materials and cyanobacterial filaments made up the remainder of their diets. Potamopyrgus antipodarum consumed mainly diatoms (51% in native and 37% in exotic beds) and amorphous organic material (43% in native and 59% in exotic beds), with small amounts of inorganic material and cyanobacterial filaments. ␦15N ratios of adult common bullies were enriched 2.5 to 3‰ above the most-enriched benthic invertebrate taxa (Odonata) (Fig. 5A, B). These ratios suggested that bully diets consisted of a mix of primary and secondary consumers. Low ␦13C ratios of adult common bullies (⬃⫺20‰ in both bed types) suggested that adult bullies may have had a trophic connection to the open-water production pathway. Juvenile bullies (⬍22 mm) had isotopic ratios that clearly were indicative of zooplanktivory. Benthic fish abundances and predation rates Abundances of common bullies were higher in exotic than in native beds (F ⫽ 38.1, p ⬍ 0.001; Fig. 7A). CPUE was ⬎3⫻ greater in exotic (3.0 fish trap⫺1 h⫺1) than in native beds (1.1 fish trap⫺1 h⫺1), and the fork length of bullies was slightly, but not significantly, larger in exotic

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FIG. 6. Mean (n ⫽ 3) % composition of gut contents of 3 common invertebrates collected from native (A) and exotic (B) beds in Lake Wanaka.

than in native beds (56 mm vs 52 mm, respectively; Fig. 7B). Most of the difference in CPUE between native and exotic beds was from the traps suspended above the lake bottom. In exotic weed beds, ⬃33% of the total CPUE (⬃1.0 fish trap⫺1 h⫺1) were in traps 50 to 100 cm above the bottom, whereas in native beds, bullies were caught almost exclusively in traps on the bottom of the lake. Feeding rates of bullies on P. antipodarum were significantly lower in aquaria with exotic (E.

canadensis, L. major) than native (I. alpinus) plants (F ⫽ 17.2, p ⫽ 0.003; Fig. 8A). Feeding rates were 38% lower in L. major treatments (stem density ⫽ 416/m2) and 26% lower in E. canadensis treatments (stem density ⫽ 306/m2) than in native I. alpinus treatments (stem density ⫽ 161/m2). The relationship between predation rate and plant stem density was strongly negative (r2 ⫽ 0.87, p ⬍ 0.001; Fig. 8B), whereas the relationship between predation rate and plant biomass was weaker (r2 ⫽ 0.46, p ⫽ 0.043).

← FIG. 5. Dual plots of mean (⫾1 SE, n ⫽ 3) ␦13C and ␦15N ratios in biota collected from native (A) and exotic (B) beds in Lake Wanaka. Filled symbols indicate consumers and open symbols indicate primary producers, fine benthic organic matter (FBOM), and suspended particulate organic matter (POM). Ratios for Isoetes alpinus were calculated separately from ratios of other vascular plants.

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FIG. 7. Mean (⫹1 SE, n ⫽ 3) catch per unit effort (CPUE using Gee-minnow traps) for common bullies in native and exotic beds at varying depths through the plant canopy (measured vertically from the lake bottom) (A), and fork length of bullies trapped in the 2 bed types (B). Bars sharing the same letter are not significantly different (p ⱖ 0.05).

Discussion Invasive macrophyte effects on littoral-zone zoobenthos Major changes in productivity, composition, and foodweb dynamics of benthic invertebrate and fish communities occurred in a New Zealand high-country lake following invasions by hydrocharitacean adventives. Exotic beds consisting of hydrocharitaceans were, on average, 3⫻ greater in biomass and canopy height than native plant beds consisting mainly of I. alpinus. Exotic beds had 5⫻ greater standing stocks of epiphyton, 2⫻ greater benthic invertebrate density, and 3⫻ greater CPUE of common bullies than native beds. Macrophyte invasions elsewhere have similarly affected biomass, canopy architecture, and productivity of macrophyte beds, as well as the zoobenthic communities that inhabit them (Keast 1984, Johnson et al. 2000, Cheruvelil et al. 2001, 2002). Two recent studies reported declines in invertebrate densities and reduced fish foraging capacity with increasing dominance by Eurasian milfoil (Myriophyllum spicatum) (Cheruvelil et al. 2002, Valley and Bremigan 2002). Our study documented

significant differences between native and exotic beds in the composition and density of zoobenthic communities and in energy flow through the food web. The productivity of exotic beds was significantly greater than that of the lower-growing, native beds (2375 mg C d⫺1 m⫺2 vs 216 mg C d⫺1 m⫺2 littoral zone, respectively). Epiphyton productivity was 40% greater in exotic than in native beds (primarily because of increased surface area for attachment), but much of the increased primary productivity in exotic beds was from the macrophytes themselves. The difference observed between native and exotic plant productivity probably was slightly overestimated because I. alpinus was used to calculate photosynthetic rates in native beds, and this species has notably lower photosynthetic rates than cooccurring milfoils and pondweeds (Madsen et al. 2002). However, I. alpinus typically made up ⬎90% of total plant biomass at native sites, so this underestimate was probably small. The standing stocks of hydrocharitaceans observed in Lake Wanaka (111–537 g/m2) were similar to those reported for other oligotrophic lakes in New Zealand (e.g., Lake Taupo: 294–424 g/m2,

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FIG. 8. Mean (⫹1 SE) predation rates of common bullies on Potamopyrgus antipodarum in laboratory trials in aquaria containing different plant species (Isoetes alpinus, Elodea canadensis, or Lagarosiphon major) (A), and the relationship between predation rate and plant stem density in these trials (B). In panel A, bars sharing the same letter are not significantly different (p ⱖ 0.05).

Lake Waikeremoana: 63–537 g/m2; HowardWilliams and Davies 1988). The standing stocks were similar to those documented for invasive species such as Eurasian watermilfoil in North American lakes (Howard-Williams and Davies 1988, Johnson et al. 2000, Lillie 2000). However, the biomass of L. major was much lower than that reported for eutrophic New Zealand lakes such as Rotoma (⬎3 kg/m2) and Rotoiti (1–1.5 kg/m2) (Clayton 1982). Thus, the level of productivity from these adventives reported in our

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study probably is at the lower end of the scale for New Zealand lakes (Clayton 1982, Schwarz and Howard-Williams 1993). Densities of benthic invertebrates, especially chironomids (C. zealandicus), purse-case caddisflies (Paroxyethira spp.), and snails (P. antipodarum), were higher in the highly productive exotic beds than in native beds. However, taxonomic richness and diversity were similar for both bed types because P. antipodarum was the dominant invertebrate taxon in both habitats (⬃30% of total density), and no taxon was specific to either exotic or native beds. Our findings were based on a single survey of invertebrates collected in midwinter. Previous work in highcountry lakes in New Zealand has demonstrated that seasonal variability in both plant and invertebrate populations is not high compared with that of Northern Hemisphere temperate systems, where climate variability is much greater (Talbot and Ward 1987, Schwarz and Howard-Williams 1993). Moreover, winter generally is recognized as a good season for sampling in New Zealand lakes because most zoobenthic taxa, even those that can be affected by periodic spring and summer insect emergences, are present (Talbot and Ward 1987). The strong predictive relationships between plant DM/SA and densities of almost all major taxa, taxon richness, and diversity suggest that higher invertebrate densities in hydrocharitacean beds than native beds is the result of greater habitat availability in exotic than in native beds (Table 4). Epiphyton biomass was not a strong predictor of the density of any invertebrate group, but this variable frequently was removed from the model because it was highly correlated with plant surface area. However, the importance of epiphyton as a food source for 2 of the most abundant benthic invertebrate taxa, P. antipodarum and Paroxyethira spp., and lower standing stocks of epiphyton per unit plant area in native vs exotic plant beds suggest that resource limitation may be driving differences in grazer densities between bed types. Resource limitation also may explain the greater density of chironomid taxa, particularly C. zealandicus, below exotic beds than below native beds because this species fed primarily on organic material and diatoms. Our study supports the findings of others that demonstrate the importance of macro-

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TABLE 4. Stepwise multiple regression analysis for densities (no./m2 littoral zone) of zoobenthos in Lake Wanaka (n ⫽ 18) showing environmental variables that best predicted densities of invertebrate groups, total density, taxon richness, and Shannon–Wiener diversity in native and exotic beds. Others ⫽ Oligochaeta, Nematoda, Hydracarina, and Copepoda. CPUE ⫽ catch per unit effort. Group Diptera Trichoptera Odonata Mollusca Others Total density Taxon richness Shannon–Wiener diversity

Source

Coefficient

t

p

Model F

Model r2

biomass surface area biomass CPUE CPUE biomass surface area biomass surface area biomass

32.0 ⫺1016 1.3 ⫺13.0 ⫺163.9 ⫺20.5 689.6 28.5 ⫺832.2 0.01

5.24 ⫺4.95 3.92 ⫺3.32 ⫺3.43 ⫺2.84 2.85 3.06 ⫺2.66 4.35

⬍0.001 ⬍0.001 0.001 0.004 0.003 0.012 0.012 0.008 0.018 ⬍0.001

27.4 24.5 15.4 11.0 11.8 8.09 8.11 9.36 7.10 18.9

0.73

Plant surface area

0.03

3.27

0.005

10.7

0.41

Plant Plant Plant Bully Bully Plant Plant Plant Plant Plant

phyte biomass to the abundances of littoral zoobenthos (Brown et al. 1988, Cyr and Downing 1988, Lalonde and Downing 1992, Weatherhead and James 2001). Other New Zealand studies have reported differences in the composition of invertebrate communities inhabiting exotic and native macrophyte communities (including hydrocharitacean weeds), but no significant positive correlations between macrophyte biomass and invertebrate abundances have been observed (Biggs and Malthus 1982, Talbot and Ward 1987). However, these studies included charophyte communities, and charophytes have low plant biomass in relation to surface area and canopy height and tend to have high invertebrate abundances (Talbot and Ward 1987). Northern Hemisphere studies have, in some cases, reported significant declines in both abundance and diversity of macroinvertebrates with increasing dominance by Eurasian milfoil (Cattaneo et al. 1998, Cheruvelil et al. 2001). However, these studies reported densities per unit of macrophyte biomass rather than per unit area of the littoral zone. In our study, when densities were expressed per unit of macrophyte biomass, densities of several invertebrate taxa, including the dominant grazer P. antipodarum, were lower in exotic than in native beds, even though densities per unit area of the littoral zone were higher in exotic than native beds for most taxa. CPUE of common bullies was greater in exotic than in native beds. In exotic beds, bullies

0.49 0.41 0.42 0.43 0.65 0.54

were often caught in traps that extended 50– 100 cm above the sediments into the plant canopy. Conversely, in native plant beds, almost no bullies ventured into traps off the lake bottom. The height and complexity of the plant canopy in exotic beds appeared to provide more habitat for zoobenthic prey, more resting area for benthic fish such as bullies, and greater refuge from top predators than in native beds (Gilinsky 1984, Keast 1984, Gotceitas 1990, Schriver et al. 1995, Valley and Bremigan 2002). However, field experiments on fish habitat selection in estuarine ecosystems suggest that midlevel predators (such as juvenile fish and bullies) tend to select habitats based mainly on food availability (Connolly 1994). Our observations suggest that exotic beds offer more prey per unit area of the littoral zone but little, if any, more per unit volume, and that predation rates of common bullies on P. antipodarum actually decrease with increasing plant stem density. These observations suggest that increased food availability is unlikely to be the dominant factor leading to high fish abundance in exotic beds. In freshwater and marine systems, plant density and fish predation rates are inversely related (Stoner 1980, Gotceitas 1990, Schriver at al. 1995, Valley and Bremigan 2002). Thus, overall primary and secondary productivity was higher in exotic than in native beds, but common bullies probably were responding to the greater habitat availability and refuge from predation in exotic beds.

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Effects of invasive macrophytes on littoral foodweb dynamics We originally hypothesized that differences in the composition and physical structure of macrophyte beds would result in differences in resources available to zoobenthic consumers, thereby affecting pathways of C flow through the littoral food web. Our results provide only limited support for this hypothesis. Intersite variability in community composition and pathways of C flow were not high enough to mask relationships between plant resources and zoobenthic consumers, as has been found in some littoral foodweb studies (Boon and Bunn 1994). Moreover, trends in isotopic ratios seemed consistent with the results of direct dietary analyses on some of the dominant zoobenthic taxa. The densities of epiphytic grazer taxa such as C. zealandicus, Paroxyethira spp., and P. antipodarum were enhanced in exotic beds, whereas detritivorous oligochaete and nematode taxa were favored in native beds. Isotopic ratios and dietary analysis of C. zealandicus, the most abundant sediment-dwelling taxon beneath exotic beds, suggested that epiphyton made up a significant proportion of its diet. Thus, it appears that epiphyton was the dominant pathway for flow of C to consumers in exotic beds, whereas FBOM was more important in native beds where oligochaetes and nematodes made up a larger proportion of the community. However, the most abundant taxon in both bed types was P. antipodarum, indicating that, as in other studies, C flow was primarily from epiphyton to consumers (Hecky and Hesslein 1995, James et al. 2000b). Slight enrichment (2–3‰) of ␦13C of some grazers (particularly P. antipodarum) relative to epiphyton resources in both vegetation types indicated that a small portion of their energetic requirements must have come from macrophyte sources of C. The degree of fractionation of ␦13C and ␦15N probably varied among primary consumer taxa (Vander Zanden and Rasmussen 2001), making it difficult to precisely ascertain the importance of the macrophyte contribution. The contribution of macrophytes to C flow generally appeared low. These findings are consistent with other studies of littoral foodweb dynamics, indicating that live macrophytes contribute little C to higher trophic levels (Engle 1988, James et al. 2000b). Despite a massive in-

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crease in macrophyte C pools available to consumers in exotic beds, only a very small proportion of this C makes its way directly into the food chain. Macrophyte C could enter the food web via the FBOM pools (Newman 1991, Kornijo´w et al. 1995), but the low ␦13C of FBOM (⫺21 to ⫺23‰) suggests that C contributions by macropyhytes must have been considerably diluted by other sources such as POM, epiphyton, or epipelon with lower enrichments. This dilution was not evident in native beds where the relatively low enrichment of I. alpinus (possibly reflecting its ability to use CO2 from sediments) meant that no clear distinction could be made between macrophyte-derived and other sources of C. It was difficult to determine from stable isotope analyses whether bed type affected the diet of common bullies. Adult and juvenile bullies had almost identical isotopic signatures in both bed types, despite large differences in the density of prey items available between bed types and in prey isotopic ratios. However, isotopic ratios of bullies from both habitats consistently indicated an ontogenic switch in their foraging habits from zooplanktivory as juveniles (⬍22 mm fork length) to benthivory as adults (⬎30 mm). This shift also has also been observed in dietary studies (Rowe and Chisnall 1996). The slightly depleted ␦13C (⬃1‰) of adult bullies relative to benthic prey probably indicated a trophic connection to the open-water production pathway, through zooplankton or larval bullies. Pelagic foodweb contributions to adult bully diets also were indicated by similar ␦15N signatures for adult bullies in both bed types, even though ␦15N ratios of benthic prey items were lower in exotic than in native beds. Moreover, the isotopic signatures of fishes reflect their diets integrated over several months (Vander Zanden and Vadeboncoeur 2002). If fish moved among habitats and consumed prey from more than one bed, detection of an effect of bed type could be difficult. Little is known about longterm (monthly or yearly) movements of bullies in New Zealand lakes, but they are known to move into near-shore marginal habitats at night to feed in some systems (Rowe 1999). Thus, the spatial scale of our study may have limited our ability to detect differences in bully feeding patterns between bed types.

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Effects of macrophyte invasions on whole-lake productivity and management implications Invasive plants affect a wide range of ecosystem attributes. Their impact on native plant diversity and their ability to degrade recreational and aesthetic values when they reach nuisance proportions are well known (Howard-Williams et al. 1987, Howard-Williams and Davies 1988). However, our study indicates that invasion of oliogotrophic lakes by hydrocharitaceans can cause significant increases in littoral-zone primary and secondary productivity with no adverse effect on invertebrate diversity. In the case of invasion by L. major and E. canadensis, this effect is confined to a band of the littoral zone between 2 and 7 m depth, where sediment and exposure conditions permit colonization. In Lake Wanaka, charophytes typically extend to ⬃17 m (Hawes et al. 2003a), and characeans also support a high invertebrate biomass (Weatherhead and James 2001), so the enhanced productivity associated with exotic macrophytes probably is diluted on a lake-wide basis. It is uncertain from our study whether zoobenthic productivity would be enhanced or diminished in the denser canopies that form in more eutrophic lakes. However, our study clearly demonstrates the importance of macrophytes to lake productivity through their influence on habitat complexity, epiphytic productivity, and area available for colonization by zoobenthic and fish communities. Acknowledgements Stable isotope analyses were conducted by Sarah Bury and Jill Parkyn of the National Institute of Water and Atmospheric Research (Wellington). Neil Blair and Greg Kelly assisted with field surveys; Isla Croft and Donna Sutherland assisted with invertebrate taxonomy. We thank Clive Howard-Williams and Barry Biggs for comments on versions of the manuscript. This study was funded by the New Zealand Foundation for Research, Science and Technology NSOF award contract number C01X0221. Literature Cited BIGGS, B. J. F., AND T. J. MALTHUS. 1982. Macroinvertebrates associated with various aquatic macrophytes in the backwaters and lakes of the upper

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Clutha Valley, New Zealand. New Zealand Journal of Marine and Freshwater Research 16:81–88. BROWN, C. L., T. P. POE, J. R. P. FRENCH, AND D. W. SCHLOESSER. 1988. Relationships of phytomacrofauna to surface area in naturally occurring macrophyte stands. Journal of the North American Benthological Society 7:129–139. BOON, P. I., AND S. E. BUNN. 1994. Variations in the stable isotope composition of aquatic plants and their implications for food web analysis. Aquatic Botany 48:99–108. CARTER, V., N. B. RYBICKI, AND R. HAMMERSCHLAG. 1991. Effects of macrophytes on dissolved oxygen, pH and temperature under different conditions of wind, tide, and bed structure. Journal of Freshwater Ecology 6:121–133. CATTANEO, A., G. GALANTI, S. GENTINETA, AND S. ROMO. 1998. Epiphytic algae and macroinvertebrates on submerged and floating leaved macrophytes in an Italian lake. Freshwater Biology 39: 725–740. CHERUVELIL, K. S., P. A. SORANNO, AND J. D. MADSEN. 2001. Epiphytic macroinvertebrates along a gradient of Eurasian watermilfoil cover. Journal of Aquatic Plant Management 39:67–72. CHERUVELIL, K. S., P. A. SORANNO, J. D. MADSEN, AND M. J. ROBERSON. 2002. Plant architecture and epiphytic macroinvertebrate communities: the role of an exotic dissected macrophyte. Journal of the North American Benthological Society 21:261– 277. CLAYTON, J. S. 1982. Effects of fluctuations in water level and growth of Lagarosiphon major on the aquatic vascular plants in Lake Rotoma, 1973–80. New Zealand Journal of Marine and Freshwater Research 16:89–94. CONNOLLY, R. M. 1994. The role of seagrass as preferred habitat for juvenile Sillaginodes punctata (Cuv. & Val.) (Sillaginidae, Pisces): habitat selection or feeding? Journal of Experimental Marine Biology and Ecology 180:39–47. CYR, H., AND J. A. DOWNING. 1988. Empirical relationships of phytomacrofaunal abundance to plant biomass and macrophyte bed characteristics. Canadian Journal of Fisheries and Aquatic Sciences 45:976–984. ENGLE, S. 1988. The role and interactions of submersed macrophytes in a shallow Wisconsin lake. Journal of Freshwater Ecology 4:329–341. GILINSKY, E. 1984. The role of fish predation and spatial heterogeneity in determining benthic community structure. Ecology 65:455–468. GOTCEITAS, V. 1990. Variation in plant stem density and its effects on foraging success of juvenile bluegill sunfish. Environmental Biology of Fishes 27:63–70. HANSON, J. M., AND W. C. LEGGETT. 1982. Empirical prediction of fish biomass and yield. Canadian

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WEATHERHEAD, M. A., AND M. R. JAMES. 2001. Distribution of macroinvertebrates in relation to physical and biological variables in the littoral zone of nine New Zealand lakes. Hydrobiologia 462:115– 129. WELLS, R. D. S., J. S. CLAYTON, AND M. D. DE WINTON. 1998. Submerged vegetation of Lakes Te Anau, Manapouri, Monowai, Hauroko, and Poteriteri, Fiordland, New Zealand. New Zealand Journal of Marine and Freshwater Research 32:621–638. WETZEL, R. G., AND G. E. LIKENS. 1991. Limnological analyses. 2nd edition. Springer–Verlag New York. WINTERBOURN, M. J. 1973. A guide to the freshwater mollusca of New Zealand. Tuatara 20:141–159. WINTERBOURN, M. J., K. L. D. GREGSON, AND C. H. DOLPHIN. 2000. Guide to the aquatic insects of New Zealand. 3rd edition. Bulletin of the Entomological Society of New Zealand 13. Entomological Society of New Zealand, Auckland, New Zealand. Received: 10 October 2003 Accepted: 17 November 2004