Ecology of the invasive New Zealand mud snail, Potamopyrgus antipodarum (Hydrobiidae), in a mediterranean-climate stream system

Abstract

The New Zealand mud snail, Potamopyrgus antipodarum, is a widely distributed non-native species of management concern on four continents. In a southern California stream, P. antipodarum abundance, which ranged from ca. <10 to nearly 150,000 snails m⁻², was related to discharge and temperature patterns. Laboratory experiments indicated that P. antipodarum (1) survivorship decreased from 13 to 27°C, but its growth rate was higher at 13 and 20°C than 27°C; (2) grazing rates were similar to those of native algivores in short-term trials; (3) grazing impact was greater than that of a native hydrobiid snail in longer-term trials; (4) ingested different diatom sizes than some other grazers; (5) reduced the abundances of medium-sized and large diatoms, and several filamentous cyanobacteria and chlorophytes, while increasing the relative abundances of tough filamentous chlorophytes (e.g., Cladophora); (6) impact on other grazing invertebrates was species specific, ranging from competition to facilitation; (7) reduced the survivorship of Anaxyrus boreas tadpoles; and (8) was consumed by non-native Procambarus clarkii and naiads of Aeshna and Argia. Ecological effects of introduced P.antipodarum are subtle, occurring primarily at transitory high densities, but flow regulation may enhance their effects by eliminating high flows that reduce their population sizes.

Introduction

Like running water systems in many parts of the world, streams in Mediterranean climates are often invaded by non-native plant and animal taxa (Dudley & Collins, 1995; Vila-Gispert et al., 2005; Ribeiro et al., 2008). Also, as in other systems, the impacts of non-native species on native communities will depend on the characteristics of the invader and the invaded community, and on environmental conditions (Poff, 1996). Because climatic, hydrological, thermal, and other environmental conditions will vary across regions, the success and impacts of potentially invasive species also are likely to vary across regions (Dukes & Mooney, 1999), as mediated through the effects of environmental conditions on disturbance regimes and the life histories, and birth, death, and growth rates, of invasive and native species. Knowledge of the success and impacts of invasive species across different regions and ecosystem types can be useful in developing a more nuanced understanding of invasive species effects on ecosystems, guiding management activities (Meekins & McCarthy, 2001; Vilà et al., 2011). For example, a hydrological regime with periodic flooding can prevent extirpation of a native topminnow (Poeciliopsis occidentalis) that is displaced by introduced mosquitofish (Gambusia affinis) in stable streams that no longer flood (Meffe, 1984).

The New Zealand mud snail, Potamopyrgus antipodarum (Gray, 1843), has been introduced to stream systems throughout the world, including in Europe, Australia, North America, and eastern Asia (Ponder, 1988; Bowler, 1991; Richards et al., 2001; Alonso & Castro-Diez, 2008; Davidson et al., 2011). In the continental U.S., this non-native mollusk is established in a wide range of habitats, encompassing arid desert to high-elevation temperate streams, coastal waterways, and the Great Lakes (ANS, 2007; Levri et al., 2008). As a consequence, P. antipodarum’s broad distribution provides an opportunity to compare its ecological relationships and impacts across regions to better understand and predict where and when its impacts rise to the level of management concern (Murria et al., 2008). This diminutive snail (ca. 3–6 mm from shell apex to base in adults) is an opportunistic grazer-detritivore in its native and adventive range (Schreiber et al., 1998; James et al., 2000; Broekhuizen et al., 2001; Hall et al., 2003) and is found in a variety of habitat types, including stream riffles and pools, lentic habitats, drainage ditches, and even high estuarine zones (Schreiber et al., 2003; Brown et al., 2008; Brenneis et al., 2011; Benson, 2011). Because P. antipodarum can reach densities of over 300,000 individuals per m² (Hall et al., 2003), it has a high potential for affecting benthic food resources and co-occurring organisms (Kerans et al., 2005; Evans, 2012; Alonso & Castro-Diez, 2012). Because of their similarity, invasive P. antipodarum might be expected to affect closely related native hydrobiid snails, including threatened species (Richards et al., 2001). Although P. antipodarum is now widely distributed in California, few ecological studies have examined its dynamics, success, environmental constraints, and impacts on native communities (Twardochleb et al., 2012; Moore et al., 2012; Krist & Charles, 2012; Kolosovich et al., 2012), although such studies might inform management activities, not only in California but also in other regions with Mediterranean climates that P. antipodarum has invaded (e.g., southern Europe, Alonso & Castro-Diez, 2008; Murria et al., 2008).

The objectives of the field and laboratory studies described here were to examine the life history and population dynamics of P. antipodarum in a southern California stream, to evaluate its responses to environmental conditions, particularly temperature and flow, and to examine its interactions with native algal communities, native grazer species, and native and non-native predators. This comprehensive battery of studies allowed comparisons to P. antipodarum’s dynamics, success, and impacts in other regions, as well as provided scientific evidence to determine if efforts to mitigate P. antipodarum’s impacts through traditional or biological control practices were justifiable.

We addressed the following questions:

• 1. What are the characteristics of P. antipodarum’s life cycle, and reproductive and population dynamics, in a southern California stream, and how are its population dynamics related to seasonal changes in abiotic factors? Aquatic invertebrate life histories and population dynamics can be highly variable, depending on climatic and hydrological conditions (Li et al., 2011). Because mediterranean systems are characterized by winter rains, summer droughts, flashy hydrographs, and benign temperatures, the life histories and dynamics of organisms in mediterranean streams are often different from those in other biomes (Gasith & Resh, 1999; Moffitt & James, 2012; Bonada & Resh, 2013).

• 2. How are the growth and survivorship of P. antipodarum affected by temperature? Consistently warmer temperatures in mediterranean compared to other temperate streams can accelerate the growth, development, and reproduction of P. antipodarum, potentially leading to more continuous life cycles (Nebeker, 1971). In some cases, P. antipodarum may grow faster in shaded habitats with lower temperatures, independent of food quality (Liess & Lange, 2011).

• 3. How do the grazing rates and diet composition of P. antipodarum compare to those of native grazers and how does P. antipodarum alter the physiognomy and species composition of algal assemblages? Owing to its small size, relatively sedentary feeding behavior, and sometimes very high densities, P. antipodarum can have substantial impacts on algal biomass and composition, as well as nutrient cycling (McCormick & Stevenson, 1989; Riley et al., 2008; Liess & Kahlert, 2009).

• 4. Does P. antipodarum affect the birth, death, and population growth rates of native grazers and vice versa? By altering algal assemblages, reducing algal biomass, and affecting the behavior of competitors, P. antipodarum can affect native grazers, including insect and amphibian larvae, and particularly ecologically similar snails (Cross & Benke, 2002; Riley et al., 2008; Kerans et al., 2010).

• 5. Can different native and non-native invertebrate predators or omnivores feed on P. antipodarum and, if so, do predators prefer particular P. antipodarum sizes? Predation has often inhibited the establishment and population growth of non-native organisms, so pre-existing predators, including non-native species, may reduce P. antipodarum population size and ecosystem impacts (Baltz & Moyle, 1993; Brenneis et al., 2011; Twardochleb et al., 2012).

Methods

Field studies

Study site

We monitored the densities and size structure of Potomopyrgus antipodarum in Piru Creek, a 3rd order tributary of the Santa Clara River in Los Angeles County, California, from 2009 to 2011 (densities: February 2009 to February 2011; size structure: March 2009 to January 2010). Our primary study site on Piru Creek was Frenchman’s Flat (118° 44′ 51″W, 34° 37′ 50″ N; elevation ca. 680 m) located approximately 4 kms downstream from Pyramid Reservoir in an area where P. antipodarum has been established since at least 2006 (Fig. 1).

Fig. 1

Map of study reach of Piru Creek, CA, USA, below Pyramid Reservoir and above Piru Reservoir, and location within California (inset). The dot represents the location of the New Zealand mud snail (Potamopyrgus antipodarum) monitoring site at Frenchman’s Flat and the square represents the location of the Blue Point collection site, where mud snails and tadpoles were collected for experiments. Map data ©2014, Google maps, terra matrics

This site has a mediterranean climate with a mean annual rainfall of about 40 cm yr⁻¹, which mostly falls between November and April, with hot, dry summers indicative of its inland location (ca. 75 km from the Pacific Ocean). Middle Piru Creek is surrounded by Angeles and Los Padres National Forests and is used primarily for recreation. Pyramid Dam controls water releases into the study section, although some effort is made to simulate natural discharge patterns, with high winter and low summer flows. The study reach flows through a narrow rocky canyon, but has an open canopy with its sandstone banks lined with Fremont cottonwood (Populus fremontii), willows (Salix spp.), poison oak (Toxicodendron diversilobum), and giant reed (Arundo donax). The creek between Pyramid and downstream Piru Dams contains arroyo and western toads (Anaxyrus californicus and A. boreas), California and Pacific tree frogs (Pseudacris cadaverina and P. regilla), bullfrogs (Rana catesbeiana), the pond turtle (Actinemys marmorata), and a variety of fish species, including Santa Ana suckers (Catostomus santaanae), sticklebacks (Gasterosteus aculeatus), arroyo chubs (Gila orcuttii), sculpins (Cottus asper), largemouth bass (Micropterus salmoides), and introduced and wild rainbow trout (Oncorhynchus mykiss), among others (Bell, 1978; Swift et al., 1993). The section of Piru Creek extending from Pyramid Dam downstream to near the Frenchman’s Flat site is a California-designated Wild Trout area and part of the Heritage and Wild Trout Program.

Field methods

The surface densities of P. antipodarum (number of snails m⁻²) were estimated monthly from February 2009 to February 2011, on soft substrata (mud, sand, and gravel) and hard substrata (rock, cobble, boulder) in riffles and pools, separately, by recording the number of snails observed under a ‘viewing bucket’ (plastic box with a clear plexiglass bottom gridded into 15 squares (5 × 5 cm each, 375 cm² per quadrat)). Five hard and five soft substrata quadrat counts were taken across two riffles and two pools on each sampling date. In addition, we took 5-cm-deep core samples (core surface area = 19.6 cm²) monthly from five soft and five macroalgae (primarily Cladophora, but with occasional vascular plants) substrata from riffles and pools, separately, from March 2009 to January 2010. On each sampling date, we combined the five cores from each substrate type (macroalgae vs. soft substrata) and habitat (pools vs. riffles) in one container and preserved each composite sample in 75% ethanol.

The size structure of P. antipodarum populations was determined from core samples in March 2009 from only riffle macroalgae habitats, in April 2009 from riffle macroalgae and soft sediment habitats, and from June 2009 to January 2010 from macroalgae and soft sediment habitats in both riffles and pools. On each date and for each habitat (soft and macroalgae substrate types in riffles vs. pools), we measured the lengths of 0–150 snails (depending on the numbers in samples), from the shell’s apex to the anterior tip of the aperture, under a microscope at ×10 magnification or by analyzing photographs using “Image J” software. In this paper, we primarily distinguish adult snails (length ≥ 2 mm) from juvenile snails (length < 2 mm), because P. antipodarum reaches reproductive maturity at a length of 2.5–3 mm (Zaranko et al., 1997).

Water temperature, conductivity, pH, and total dissolved solids (TDS) were measured in a pool and a riffle on each date using an Extech digital handheld meter (ExStik MULTI Meter - PH100). Monthly average discharge data were obtained for station PYRAMID (PYM) on Piru Creek below Pyramid Lake from the California Department of Water Resources (California Data Exchange Center: http://cdec.water.ca.gov/selectQuery.html).

Experimental studies

Laboratory experiments were performed to examine the effects of temperature on P. antipodarum survivorship and growth and to quantify interactions between P. antipodarum and their algal resources, potential competitors, and potential predators. Experiments were conducted in glass or plastic containers filled with stream water collected either from Piru Creek, Ventura County, or from Mission Creek, Santa Barbara County, then filtered through a 50-µm sieve. Experimental units were exposed to fluorescent light on a 12-h light/12-h dark regime. With the exception of the experiment where we manipulated temperatures, most temperatures remained between 17 and 20°C. All containers were aerated using air pumps and 2 mm tubing. Potamopyrgus and Anaxyrus tadpoles used in these experiments were collected from Piru Creek, whereas other stream insects and snails were collected from coastal streams in Santa Barbara County (Mission, Arroyo Hondo, Carneros, Romero Creeks). With the possible exception of native Pyrgulopsis, all taxa co-occur with Potamopyrgus in Piru Creek or other southern California streams, and Pyrgulopsis was used because it is similar to Potamopyrgus and occurs in streams vulnerable to Potamopyrgus invasion.

Temperature experiment

In this experiment, 24 glass experimental units (container volume = 473 cm³, bottom surface area = 0.018 m²) contained filtered stream water at temperatures of 13, 20, and 27°C, effected using aquarium water heaters, with eight units assigned to each temperature level. Thirty P. antipodarum juveniles (average length = 1.3 mm) were placed into each container at the beginning of the experiment. Because snails only grew to a length of 2.1 mm by the end of this 70-day experiment, no reproduction was observed. Snails were fed Spirulina throughout this experiment (1.2 teaspoons of powder week⁻¹). Dead snails were removed and replaced during the first 48 h of the experiment, but dead snails were removed but not replaced thereafter. The number of live and dead snails, and average snail size, in each container were determined weekly.

Grazing experiments

Grazing experiments were conducted in clear plastic containers (container volume = 131 cm³, bottom surface area = 0.0025 m², water depth = 4 cm) with tiled bottoms covered by diatoms and unicellular green algae. In these experiments, we measured the response of algal biomass (as chlorophyll a concentration) to the manipulated densities of Potamopyrgus antipodarum (mean length = 3.6 mm) and the snail Physella gyrina (Physidae; mean length = 4.8 mm), the snail Pyrgulopsis cf. stearsiana (a native snail from the same Hydrobiidae family as P. antipodarum, mean length = 2.0 mm), late instar nymphs of the mayfly Centroptilum sp. (Family Baetidae), and mid- to late-instar larvae of the caddisfly Gumaga nigricula (Family Sericostomatidae). In the first 2-day experiment, densities of Potamopyrgus, Physella, Pyrgulopsis, and Centroptilum were 5, 10, 15, and 20 per container, respectively, when each species was alone, and were 10 (5 Potamopyrgus + 5 of each of the other species), 15 (10 or 5 Potamopyrgus + 5 or 10 of each of the other species) and 20 (10 Potamopyrgus + 10 of each of the other species) per container when Potamopyrgus was combined with each of the other species. In the second grazing experiment, we measured the effects of four densities (2, 4, 6, 8 individuals per container) of Potamopyrgus and Gumaga, when each was alone, as well as four density combinations of these two species (2 Potamopyrgus + 2 Gumaga, 2 Potamopyrgus + 4 Gumaga, 4 Potamopyrgus + 2 Gumaga, and 4 Potamopyrgus + 4 Gumaga), on algal biomass over 2 days. Finally, in a third experiment, we measured the effects of four monospecific densities of Potamopyrgus and Pyrgulopsis (5, 10, 15, 20 per container, same sizes as in the 1st experiment) on algal biomass over 17 days. In all grazing experiments, control containers contained no grazers and three replicate experimental units were assigned to each treatment level (i.e., controls and each monospecific density and each density combination of two species).

At the end of these trials, we sampled algae in each grazing chamber by scraping algae from each tile, stirring the suspension, and then removing 10 ml of algal slurry with a syringe. Syringe samples were filtered through glass fiber filters (Whatman GFF), and chlorophyll on each filter was extracted for 24 h with 10 ml of 90% acetone in a dark freezer. Chlorophyll solutions were centrifuged then read on a Turner 10-AU Digital Fluorometer, both with and without acidification. After the first set of grazing trials, grazers were preserved in 70% ethanol, then guts of individual grazers, except for Gumaga, were removed and gut contents were extruded and examined at ×400 under a microscope to enumerate the numbers of recognizable algal species in each gut.

Additionally, we determined the effects of P. antipodarum (mean size = 3.6 mm) and Western toad tadpoles (Anaxyrus boreas, initial mean body length = 9.8 mm) on algal composition, algal density (number of cells cm⁻²), and filamentous algal cover over 27 days by manipulating both initial algal community composition (three types: diatoms, Cladophora, Zygnema) and the presence or absence of tadpoles (0 or 5 T) and mud snail density (0, 100, 500 S) with the following treatments: 0T/0S, 5T/0S, 5T/100S, 0T/500S, 5T/500S. This experiment was conducted in plastic boxes (container volume = 6,000 cm³, bottom surface area = 0.0175 m², water depth = 12 cm; three replicates per treatment) whose bottoms were covered with cleaned beach sand (3 cm deep). Three initial algal community types were used in this experiment: (1) algal communities dominated by single-celled algae, mostly diatoms (Bacillariophyceae; called the diatom culture), (2) algal communities dominated by filamentous Cladophora sp. (Chlorophyceae, called the Cladophora culture), and (3) algal communities dominated by other filamentous algae, including Zygnema (Chlorophyceae) and Melosira varians (Bacillariophyceae) with some Mougeotia and Spirogyra (Chlorophyceae, hereafter called the Zygnema culture). Algal assemblages were sampled by determining percent cover of filamentous algae using a gridded quadrat, then removing overlying filamentous algae and taking three cores (each 2.5 cm diameter, 1 cm deep) from the sand covering each experimental container’s bottom at the beginning and end of the experiment. Core samples from each container were agitated thoroughly, the algal suspension was mixed, and then a syringe was used to collect 10 ml of suspension, which was preserved in 10% formalin. Subsamples of preserved samples were identified and enumerated at ×400 under a microscope. Initial characteristics of each algal culture community can be found in Supplementary Materials Table 1.

Competition experiments

In the first competition experiment, we examined the effects of different densities of Potamopyrgus, Pyrgulopsis, Centroptilum, Physella, and Gumaga, both alone and together, on the birth, death, and population growth rates of each potential competitor species. We used the same sizes of grazers as used in the grazing trials. We set up three monospecific densities of Potamopyrgus (25, 50, 75 individuals/container), Pyrgulopsis (15, 30, 45/container), Centroptilum (6, 12, 18/container), Physella (10, 20, 30/container), and Gumaga (9, 18, 27/container), as well as specific combinations of Potamopyrgus (Po) (25 and 50 Po/container) with each of the other species (25 and 50 Po with 6 and 25 Po with 12 individuals of Centroptilum; 25 and 50 Po with 10 and 25 Po with 20 individuals of Physella; 25 and 50 Po with 9 and 50 Po with 18 individuals of Gumaga; and 25 and 50 Po with 15 and 25 Po with 30 individuals of Pyrgulopsis). Each experimental container had a volume of 1,770 cm³ and a ceramic tile bottom surface area of 0.023 m², three replicate units were assigned to each experimental treatment, and the experiment lasted 55 days. Bottom tiles were colonized primarily by diatoms collected from Piru and Mission Creeks and introduced to containers for 2 weeks before the experiment started. To compensate for mayfly nymph deaths due to handling and losses to adult emergence, we replaced emerged or dead mayflies daily over the first 7 days to maintain preset densities. Snails and Gumaga showed no initial mortality and Gumaga remained in the larval stage throughout the experiment. We counted the number of dead and live individuals in each container during the experiment, as well as the number of emerged Centroptilum in the Centroptilum treatments.

In the second experiment, we examined the effects of Potamopyrgus and Western toad tadpoles on each other’s growth and mortality across three different algal community types (diatoms, Cladophora, Zygnema, described in grazing impact section, above). In this experiment, the following five treatments were set up in each algal culture type: no tadpoles or snails, 5 tadpoles and no Potamopyrgus, 5 tadpoles and 100 Potamopyrgus, no tadpoles and 500 Potamopyrgus, and 5 tadpoles and 500 Potamopyrgus. Each experimental container had a volume of 6,000 cm³ and a bottom surface area (covered by 3 cm of cleaned sand) of 0.0175 m², Three replicate containers were assigned to each treatment, and the experiment ran for 27 days. The number of live and dead tadpoles in treatments with tadpoles was counted weekly, whereas the numbers of dead and live Potamopyrgus were counted at the end of the experiment, and the lengths of tadpoles and Potamopyrgus were determined at the beginning and end of the experiment using digital images and “Image J” software.

Predation experiments

We conducted predation trials to determine if native and non-native predators commonly found in southern California streams would consume Potamopyrgus of different sizes. We examined possible predation by the ephemerillid mayfly Drunella sp. (n = 4 replicate trials), the perlodid stonefly Isoperla sp. (n = 1), the dragonflies Aeshna walkeri (n = 6), Cordulegaster dorsalis (n = 1), Octogomphus specularis (n = 7), and Paltothemis lineatipes (n = 3), the damselfly Argia vivida (n = 19), the hemipteran back swimmer Notonecta hoffmanni (n = 3), the hellgrammite Protochauliodes (n = 2), the caddisflies Rhyacophila sp. (n = 1), Lepidostoma sp. (n = 6), and Gumaga nigricula (n = 3), and three size classes (<5 cm, n = 2; 5–10 cm, n = 2, and >10 cm, n = 3) of the non-native red swamp crayfish (Procambarus clarkii) on two (crayfish trials, <2 mm, >2 mm) or three (insect trials, <2 mm, 2–4 mm, and >4 mm) P. antipodarum size classes. A single, starved (72 h) predator was used in each trial, and initial densities of P. antipodarum were 6 per container (2 of each of 3 size classes) in insect trials and 200 per container (100 of each of 2 size classes) in crayfish trials. Because P. antipodarum mortality was negligible in the competition and grazing experiments, we did not include control units with no predators present. Each predation trial container had a volume of 1,770 cm³ and a tile bottom surface area of 0.023 m², and no food was provided to snails during these trials. After 2 days, the number of live snails remaining in each container was counted. Fecal analysis was performed at 10X under a microscope to determine if snails passed viably through predator guts.

Statistical analysis

For the field sampling, we determined if there were consistently different densities of P. antipodarum over sampling times between different habitat types (riffle vs. pool, hard vs. soft substrata for quadrat samples, macroalgae vs. soft substrata for core samples) using paired t tests. To examine relationships between P. antipodarum densities in different habitats and abiotic factors, we performed linear and quadratic (only for water temperature) regressions of P. antipodarum densities versus conductivity, water temperature, and discharge, independently, as well as multiple regressions using snail densities as dependent variables and the abiotic factors listed above, including a quadratic term for temperature, as independent variables with the backwards stepwise elimination of non-significant independent variables. Intercorrelations among all abiotic factors were also examined (Pearson’s r). In all cases, snail densities were log₁₀(x + 1)-transformed and discharge data were log₁₀-transformed before analyses.

For laboratory experiments, we used one-way ANOVAs to examine treatment effects (when designs were asymmetrical) and two-way ANOVAs to examine the effects of manipulated variables (in cross-classified designs), followed by either t tests (two categories) or Tukey’s HSD tests (> two categories) to determine which contrasts or treatments were significantly different. Data for all proportionate response variables (proportion surviving, proportionate increases in size, i.e., final length—initial length/initial length, proportion of population in small size classes, proportion filamentous algae cover) were logit-transformed before analyses, with adjustments for extreme 0 and 1 values (Warton & Hui, 2011).

Response variables for the competition, grazing, and predation experiments were derived from an exponential population growth and loss model (similar to Riley et al., 2008), i.e., final biomass or density = initial biomass or density X e^rt with r = b−d, which well fit our time series data; r, b, and d represent the per capita population growth, birth, and death rates of each population, respectively, and t represents experiment duration in days, with all vital rates having units of d⁻¹. We used data on total numbers of live and dead individuals of each taxon at the beginning and end of the first competition experiment to calculate r, b, and d rates for each competitor in each experimental container. Because there was no Centroptilum and Gumaga recruitment in treatments with these species, this collapsed to a calculation of d rates for Gumaga; however, because we also tracked emergence numbers for Centroptilum, we could partition population losses into per capita losses owing to emergence versus death for Centroptilum. Similarly, grazing impact or grazing rate indices in the short-term (2 d) grazing experiments were calculated as [ln (Chl a in control containers without grazers/Chl a in containers with grazers)]/[(time in days) × (number of grazers)], with final dimensions per grazer per day, and predation impact indices were calculated as [(ln (beginning P. antipodarum number/final P. antipodarum number)]/(time in days), with final dimensions per predator per day because a single predator was used in each trial. In the long-term grazing experiment with P. antipodarum and Pyrgulopsis in monospecific culture, grazer impact was calculated as [ln (chl a without grazers/chl a with grazers]/(grazer number) to obtain a per grazer impact index for the duration of the experiment. Contrary to typical calculations, we reversed the numerator and denominator of grazer or predator vs. control trial values in grazing and predation impact calculations, so that they would scale positively with increasing grazer or predator impact or feeding rates. Because the short-term grazing experiments used substitutive designs, we compared observed grazing impacts when Potamopyrgus was present with another competitor species to predicted grazing impacts derived from grazing rates when each species was alone, at the same total grazer density, weighted by the proportion of total grazer numbers comprised of each species. The means and indices of dispersion for predicted grazing rates were calculated using bootstrapping techniques, where random subsamples with replacement were taken from the population of all combinations of the weighted averages of the grazing rates of each species when alone, then compared to observed grazing rates in mixed culture using t tests. In the competition experiments, we also calculated competition coefficients as growth (r), birth (b), or death (d) rates when a competitor was alone—r, b, or d rates when competitors were added, including conspecific or heterospecific competitors, divided by the added number of competitors. Because the analysis of competition coefficients produced similar conclusions as the analysis of competitor treatment effects on r, b, and d rates, they are only briefly mentioned in the Results.

In examining the effects of P. antipodarum and Anaxyrus tadpoles, alone and together, on algal communities, we classified individual algal taxa by life form (unicells, colonies, filaments), taxonomy (diatoms, cyanobacteria, chlorophytes), and size (small = less than 10 µm, medium = 10–100 µm, large = > 100 µm) following Dussart (1965), Patrick & Reimer (1966), Hillebrand et al. (1999), Krammer & Lange-Bertalot (1986, 1988, 1991a, b), Kociolek & Stoermer (1993), and Potapova & Hamilton (2007). We then performed two-way ANOVAs using the population growth rates (r = ln (algal density at the end of the 27-day experiment/algal density at the beginning of the experiment)) for total algae and each algal group as dependent variables and algal culture type and tadpole-snail treatment (0T/0S, 5T/0S, 5T/100S, 0T/500S, 5T/500S), and their interaction, as independent variables. In all cases, the algal culture type X tadpole-snail treatment interaction effect was significant, so we then conducted analyses to determine the effects of competitor treatment on algal community structure for each algal culture type, independently. To examine the effects of competitor treatment on multivariate community structure for each algal culture type, we first performed non-metric multidimensional scaling (NMS) on the matrix of experimental containers by the relative abundances of algal taxa. We then examined the effects of competitor treatment on NMS axes scores, as well as on the population growth rates of total algae and each algal group, using one-way ANOVAs followed by Tukey’s HSD tests. To interpret each NMS axis, we determined the percentage of multivariate community variation accounted for by each NMS axis and significant correlations between NMS axes and common (occurring in ca. 50% or more of samples) algal taxa. We also performed multiple response permutation procedures (MRPP) to examine the effects of competitor treatment on multivariate algal community distances between experimental containers.

Results

Population dynamics, life history in Piru Creek, and experimental effects of temperature

Potamopyrgus antipodarum populations reached their highest densities during the summer dry season and occurred at low densities during the winter wet season in Piru Creek (Figs. 2 and 3). Although juvenile (<2 mm) mud snails were collected throughout the year, they reached highest densities in late spring and early summer (Fig. 3). For both quadrat and core samples, there were no consistent differences in Potamopyrgus densities or size structure between riffles and pools, or between different substrata types, throughout the year, with the exception that Potamopyrgus densities in core samples, primarily the densities of large (>2 mm) Potamopyrgus, were higher in macroalgae than soft sediment substrata in riffles (paired t test, t _{1, 10} = 2.3, P < 0.05).

Fig. 2

Discharge (dotted line) and Potamopyrgus densities in riffles (dashed line) and pools (solid line), based on quadrat sampling, in Piru Creek from February 2009 to February 2011

Fig. 3

Potamopyrgus densities (number m⁻², histograms), based on core sampling, and water temperature (line) from March 2009 through January 2010. Histograms are shaded to show densities of juvenile (<2 mm, black) and adult (>2 mm, white) Potamopyrgus on each date. For each date, the histogram on the left represents Potamopyrgus density for macroalgal substrata and the histogram on the right is Potamopyrgus density for soft sediment substrata. No data on snail densities were collected from soft sediment substrata in March 2009

Because environmental variables were intercorrelated, it was difficult to disentangle the effects of different environmental variables on Potamopyrgus densities or per capita population growth rates. Piru Creek discharge was negatively related to conductivity (r = −0.63, P < 0.05), water temperature (r = −0.76, P < 0.0001), and total dissolved solids (r = −0.70, P < 0.05) and was positively related to current velocity in both riffles and pools (rs = + 0.87 and +0.72, both ps < 0.001), but was unrelated to pH (n = 10–22). Nevertheless, multiple backwards stepwise regression analysis consistently produced models containing only a negative relationship between log Potamopyrgus densities and log stream discharge for both quadrat and core samples from both riffles and pools (R ²s = 0.30–0.55, ps < 0.05 to < 0.001).

Although Potamopyrgus densities often had positive linear or quadratic relationships with water temperature, these relationships became non-significant when discharge was also entered as an independent environmental variable. For example, the relationship between log Potamopyrgus density and water temperature was best fit by a quadratic model, with snail densities peaking at around 17°C (Fig. 4). Only one (marginally) significant relationship was observed between Potamopyrgus population growth rates and water temperature, for riffle quadrat samples, and this relationship was also best described by a quadratic model (Potamopyrgus growth rate = 0.08 + 0.008* temperature−0.0015 temperature^2x; * = P < 0.05, ^x = P < 0.10, peak at ca. 20°C). In the laboratory temperature experiment, Potamopyrgus individual growth rates were higher at 13 and 20°C than 27°C, whereas Potamopyrgus survival decreased as temperature increased from 13 to 27°C (Fig. 5).

Fig. 4

The relationship between log₁₀ Potamopyrgus density (from quadrat sampling) on hard, riffle substrata and water temperature. This relationship was best fit with a quadratic equation, which is shown along with the associated coefficient of determination and probability level. Significant equation terms are indicated as ** = P < 0.01 and * = P < 0.05

Fig. 5

Responses of Potamopyrgus survivorship (% over 70 days) and individual growth (as % length increment over initial length over 70 days) at three experimental temperatures (13, 20, 27°C). Histograms overlain by the same letter in each graph are not significantly different (P > 0.05, Tukey’s HSD test)

Grazing trials and grazer effects on algal communities

In the first set of grazing trials, different grazer species (Centroptilum (Ce), Pyrgulopsis (Py), Potamopyrgus (Po), and Physella (Ph)) showed similar per capita grazing rates per day (no significant grazer species effect; means in monospecific culture ±1 SE: Ce = 0.014 ± 0.005, Py = 0.014 ± 0.004, Po = 0.019 ± 0.004, Ph = 0.008 ± 0.006; n = 12 per species). Grazing rates declined with increasing total grazer density, being higher at the lowest grazer density (5 per container) than at higher grazer densities (10, 15, 20 per container) (grazer density effect: F _1,82 = 17.6, P < 0.0001; 5 > 10, 15, 20 grazers per container, Tukey’s HSD test, P < 0.05; Fig. 6a–c). In the second grazing trial, grazing rates of both Gumaga and Potamopyrgus were low and variable, and there were no significant effects of grazer species or total grazer density on grazing rates (Fig. 6d).

Fig. 6

Grazing impact ((ln (Chl a no grazers/Chl a with grazers))/grazers per container/time in days) of five grazer species (Potamopyrgus (Po), Pyrgulopsis (Py), Physella (Phys), Centroptilum (Cen), Gumaga (Gum)), both alone and with Potamopyrgus, across four total grazer densities (5, 10, 15, 20 per container for Py, Phys, and Cen, alone and with Po, and for Po alone; 2, 4, 6, 8 per container for Gum and Po alone and together). Values are means ± 1 SE. Grazer impacts for Po (black circle) and other grazers (white circle) alone are compared to values for species when together (hatched square for equal densities of Po and other grazer and for 2/3 Po and 1/3 other grazers at the third highest grazer density; gray square for 1/3 Po and 2/3 other grazers at third highest density). Asterisks denote significant differences between observed grazer impacts when different grazer species were together versus additive expectations based on grazer impacts when each species was alone (two-tailed t _1,4, > 3.0, * = P < 0.05)

At specific total densities, additive expectations based on grazing rates of each grazer species when alone differed from observed grazing rates when both species were together; however, the direction of these differences in expected and observed grazing rates varied with total density and grazer species combination. Observed grazing rates were lower than additive expectations at a total density of 10 grazers per container for Potamopyrgus and Pyrgulopsis when together and for Potamopyrgus and Centroptilum when together (Figs. 6a,b) and at a total density of 8 per container for Gumaga and Potamopyrgus (Fig. 6d) when together, but the reverse was observed at a total density of 10 per container when Potamopyrgus and Physella were together (Fig. 6c). In the Physella–Potamopyrgus trials, differences between observed and expected grazing rates were associated with large declines in Physella grazing rates from 5 to 10 Physella per container (see Fig. 6c).

In the longer-term experiment examining the impacts of Pyrgulopsis and Potamopyrgus in monospecific culture on algal biomass (as chlorophyll a), there were significant effects of species (F _1,20 = 3.5, P < 0.01) and grazer density (F _1,20 = −4.2, P < 0.001) on per capita grazer impacts, with a marginally significant species X density interaction effect (F _1,20 = −1.8, P < 0.10). The interaction effect was owing to a significant difference between Pyrgulopsis and Potamopyrgus grazing impacts at a density of 10 snails per container, but not at other grazer densities (Fig. 7).

Fig. 7

Grazer impact (means ± 1 SE) per grazer for two snail species (Potamopyrgus = black dots, Pyrgulopsis = white dots) in monospecific culture across four snail densities (5, 10, 15, 20 per container) over a 17-day experiment. Asterisk denotes a significant difference between species at a snail density of 10 per container (ANOVA, F _1,4 = 10.4, P < 0.05)

Gut content analysis of grazers from the first set of grazing trials revealed differences in the diets of different grazer species (Fig. 8). Although all grazers consumed diatoms (97% of cells ingested), they differed in the sizes of diatoms eaten. Centroptilum consumed a greater proportion of large diatoms and a smaller proportion of small diatoms than Pyrgulopsis, Potamopyrgus, and Physella (P’s < 0.05, Tukey’s HSD test), whereas Potamopyrgus consumed a greater proportion of medium-sized diatoms than Pyrgulopsis (P < 0.05, Tukey’s HSD), with Physella and Centroptilum consuming intermediate proportions of medium diatoms.

Fig. 8

Percentage contribution of different algal sizes (small, medium, large; see text) to the gut contents of four different grazer species (Centroptilum, Pyrgulopsis, Potamopyrgus, Physella) at the end of the first grazing experiment

We also examined the long-term (27 days) impacts of western toad larval stages (tadpoles) and Potamopyrgus antipodarum, both alone and together, on algal community structure across three initial algal community types (dominated by diatoms, Cladophora or Zygnema). In general, the presence of tadpoles had little effect on algal community structure, but there were large changes in algal community structure with increasing Potamopyrgus densities. Further, because initial ANOVAs showed significant interaction effects of grazer treatment and algal community culture type on the population growth rates of different algal groups, we examined the effects of grazer treatment on algal growth rates for each algal culture type independently. Total algal population growth rates declining with increasing snail density in the Zygnema culture and were lowest in the 5 tadpole and 100 Potamopyrgus treatment in the diatom culture, but were unaffected by treatment in the Cladophora culture (Supplementary Materials Table 2). Percent cover by filamentous algae was not significantly different among grazer treatments in the Cladophora and Zygnema cultures (mean = 87 and 37%, respectively), but was affected by treatment in diatom culture, being significantly higher where Potamopyrgus was present (17%) than where it was absent (0%). Among algal categories, the population growth rates (r) of (1) diatoms and other unicellular algae declined with increasing Potamopyrgus density across all algal culture types, (2) colonial algae and cyanobacteria were lowest in the 5 tadpole/100 Potamopyrgus treatment in diatom culture but not affected by treatment in the other algal cultures, and (3) filamentous algae were highest in the 5 tadpole/500 Potamopyrgus treatment in Cladophora culture and lowest in the same treatment in Zygnema culture, and tended to decline with snail density in diatom culture. Regarding algal size categories, population growth rates of large algae tended to increase with increasing snail density in Cladophora culture, were lowest in the 5 tadpole/500 Potamopyrgus treatment in Zygnema culture, and were unaffected by treatment in diatom culture. Medium-sized algal rs declined with snail density in the diatom and Zygnema cultures, and small algal rs increased with snail density in the diatom culture and were higher in the 0 tadpole/500 snail treatment than 5 tadpole/500 snail treatment in the Zygnema culture (Supplementary Materials Table 2, all P’s < 0.05, Tukey’s HSD test). Growth rates of medium and small algae were unaffected by grazer treatment in the Cladophora culture.

Potamopyrgus treatments had significant effects on algal community structure in all algal culture types (MRPP test, As = 0.09–0.54, ts = −3.0 to −4.4, P's < 0.01 to < 0.001). Non-metric multidimensional scaling (NMS) on the matrix of the relative abundances of algal species across experimental containers in the diatom culture produced a two-dimensional solution with a stress of 0.005, with scores for both NMS axes being significantly affected by grazer treatment (Figs. 9a, b). The scores for NMS axis 1, which accounted for 76% of the variation in the multivariate dataset, were much greater at high snail densities than where snails were absent (Fig. 9a). Correlations between common algal species and NMS axis 1 scores indicated that high snail densities were associated with high relative abundances of the small, colonial, planktonic green alga Ankistrodesmus falcatus and low relative abundances of the medium-sized, filamentous cyanobacterium Cylindrospermum spp. and several medium to large benthic diatom species, with the reverse observed in containers where snails were absent (Fig. 9a). Algal taxa relationships with, and grazer treatment effects on, NMS axis 2 scores indicated that Cladophora and a number of benthic diatom species showed greatest relative abundances in the 5 tadpole/100 snail than other treatments, and that the relative abundances of these algal taxa also were higher at the highest snail densities than where snails were absent (Fig. 9b). For algal data from the Cladophora culture, the two-dimensional NMS solution had a stress of 8.7, with scores for only NMS axis 1, which accounted for 82% of the variation in the multivariate dataset, being significantly affected by grazer treatment (Fig. 9c). Correlations between algal species and NMS axis 1 scores indicated that containers containing snails had higher relative abundances of Cladophora and lower relative abundances of the filamentous diatom Melosira varians and a number of unicellular, medium-sized benthic diatoms than containers lacking Potamopyrgus (Fig. 9c). Finally, for the algal data from the Zygnema culture, the two-dimensional NMS solution had a stress of 12.1, with significant grazer treatment effects on only NMS axis 1, which accounted for 47% of the variation in the multivariate dataset (Fig. 9d). Associations between the relative abundances of algal species and NMS axis 1 scores indicated that the filamentous green algae Spirogyra and Mougeotia and the medium-sized benthic diatom Nitzschia linearis/gracilis had higher relative abundances, while Ankistrodesmus, the filamentous green alga Zygnema, and the small benthic diatom Amphora pediculus had lower relative abundances, where snails were absent than where they occurred at high densities (particularly in the 0 tadpole/500 snail treatment, Fig. 9d). In general, these collective results suggest that Potamopyrgus reduces filamentous cyanobacteria, a variety of medium and large benthic diatoms, and the green filamentous algae Spirogyra and Mougeotia, depending on initial algal conditions, while increasing the relative contributions of small colonial planktonic (Ancistrodesmus) and large filamentous benthic green algae (Cladophora, Zygnema).

Fig. 9

Mean NMS axis scores (± 1 SE) significantly affected by grazer treatment in the tadpole and Potamopyrgus experiment. For each algal culture type (diatoms, Cladophora, Zygnema) a NMS was conducted on the matrix of experimental containers by the relative abundances of algal species. Final stresses for two-dimensional NMS solutions were: diatom culture = 0.005, Cladophora culture = 8.7, Zygnema culture = 12.1. Grazer treatments were 0 tadpoles/0 snails (0T/0S), 5 tadpoles/0 snails (5T/0S), 5 tadpoles and 100 snails (5T/100S), 0 tadpoles and 500 snails (0T/500S), and 5 tadpoles and 500 snails (5T/500S). The percentage of the overall variation in the multivariate dataset accounted for by each NMS axis and the algal species significantly (P < 0.05, Pearson’s r) correlated with each axis are shown. In each graph, bars overlain by the same letter are not significantly different (P > 0.05, Tukey’s HSD test)

Competition experiments

The competition experiments revealed large differences in the birth and death rates of different competitor species, with Physella having much higher vital rates (b = 0.045 and d = 0.011) than Pyrgulopsis (0.002, 0.006) and Potamopyrgus (0.0006, 0.0012). There was no recruitment to Gumaga and Centroptilum populations and no Gumaga emergence during the experiment. Gumaga showed little mortality and remained at constant densities throughout the experiment, whereas Centroptilum densities had declined to 0 in most containers by the end of the experiment, owing to emergence and death.

There were no effects of competitor treatments on the population growth (r) and death (d) rates of Gumaga and Potamopyrgus when these two species were manipulated alone and together. Death rates of Centroptilum and Potamopyrgus also were unaffected by competitor treatment, but Centroptilum per capita emergence rates were higher at the highest monospecific Centroptilum density than in the 6 Centroptilum/25 Potamopyrgus treatment (Fig. 10a). Orthogonal contrasts showed that Centroptilum emergence rates were higher in containers where Potamopyrgus was absent than where it was present (F _1,16 = 8.1, P < 0.02). On the other hand, Potamopyrgus birth rates (b) were higher in the 25 Potamopyrgus/12 Centroptilum treatment than in other treatments (Fig. 10b).

Fig. 10

Effects of competitor treatments on the emergence rate of Centroptilum (Ce) (a) in the Centroptilum–Potamopyrgus experiment, on the birth rates of Potamopyrgus (Po) in the Centroptilum–Potamopyrgus experiment (b) and of Physella (Ph) (e) and Potamopyrgus (f) in the Physella–Potamopyrgus experiment, and on the death rates of Pyrgulopsis (Py) (c) and Potamopyrgus (d) in the Pyrgulopsis–Potamopyrgus experiment and of Physella (g) and Potamopyrgus (h) in the Physella–Potamopyrgus experiment. All vital rates have units of d⁻¹. Treatment code letters represent grazer species, as outlined in the first sentence of this legend, and grazer numbers (number per container). Trt denotes the effects of competitor treatment on competitor vital rates (ANOVA: * = < 0.05, ** = < 0.01, *** = < 0.001). Bars overlain with the same letter in each graph are not significantly different (P > 0.05, Tukey’s HSD test)

The most intense competitive and facilitative interactions, however, were observed among the snail species. Orthogonal comparisons showed that Pyrgulopsis r and b rates (F _1,16 = 11.4 and 10.6, P < 0.005 in both cases) were higher and that d rates were marginally lower (F _1,16 = 3.4, P < 0.10) where they were alone than with Potamopyrgus, whereas Potamopyrgus death rates were higher and growth rates lower, where Pyrgulopsis was present versus absent (F _1,16 = 11.6 and 7.2, P < 0.005 and <0.02). Further, in the Pyrgulopsis–Potamopyrgus experiment, significant treatment effects were found on Pyrgulopsis r and d rates and on Potamopyrgus d rates, with d rates tending to be highest at the highest density of heterospecific competitor (Figs. 10c,d). Significant competitor treatment effects also were found in the Physella–Potamopyrgus experiment with orthogonal comparisons showing significant negative effects of Potamopyrgus on Physella death rates (facilitation) and significant positive effects of Physella on Potamopyrgus birth and death rates (F _1,16 = 8.8–15.6, P < 0.01 to < 0.005; Figs. 10f–h). Physella b rates were strongly, negatively affected by its own densities (Fig. 10e). In examining per capita competition coefficients, Physella had the largest competitive effect on its own growth and birth rates and Physella had positive per capita effects on both the b (facilitative) and d (competitive) rates of Potamopyrgus, translating into little net effect on Potamopyrgus population growth rates.

In the tadpole–Potamopyrgus competition experiment, there were no effects of Potamopyrgus density on tadpole individual growth rates across algal culture types, but there was a significant algal type X snail density interaction effect on tadpole survivorship (Fig. 11b). Tadpole survivorship declined with increasing Potamopyrgus density and was lower in Cladophora than other algal cultures at Potamopyrgus densities of 0 and 500 per container. There were no effects of algal culture type on snail population and individual growth rates, and the proportions of snail populations that were dead or composed of small individuals (Fig. 12). On the other hand, the individual growth rates and proportions of populations composed of small individuals for Potamopyrgus were higher in the 5 tadpole/100 snail treatment than in the 0 tadpole/500 snail treatment (Figs. 12b,c). The proportions of snail populations that were composed of small individuals and snail population growth rates were positively correlated (r = + 0.54, P < 0.005, n = 27), indicating that higher growth rates translated into higher reproductive rates. The proportions of snail populations that were dead were significantly higher in the 0 tadpole/500 snail and 5 tadpole/500 snail treatments than in the 5 tadpole/100 snail treatment, indicating that Potamopyrgus mortality rates increased with increased Potamopyrgus density (Fig. 12d).

Fig. 11

Effects of algal culture type (dominated by diatoms, Cladophora, Zygnema) and Potamopyrgus density (number per container) on the percentage individual growth (length increment divided by initial length × 100) and percentage survivorship (over 27 days) of Anaxyrus boreas tadpoles. The note in the bottom graph indicates significant effects of snail density and algal culture type (**** = P < 0.001 and ** = P < 0.01), and a marginally significant effect of their interaction (* = P < 0.10), on tadpole survivorship (two-way ANOVA)

Fig. 12

Effects of competitor treatment (0 tadpoles/500 Potamopyrgus (0T/500S), 5 tadpoles/100 Potamopyrgus (5T/100Po), 5 tadpoles/500 Potamopyrgus (5T/500S)) and algal culture type (diatoms, Cladophora, Zygnema dominated) on the population and individual growth rates, percentages of populations composed of individuals <2 mm, and percentages of populations that were dead for Potamopyrgus. Heights of bars are means ± 1 SE. Algal type had no effect on Potamopyrgus responses, but competitor treatment did, as indicated by the letters next to the competitor treatment code. For each response variable, competitor treatments with the same following letter were not significantly different (P > 0.05, Tukey’s HSD tests)

Predation trials

Large crayfish (Procambarus clarkii >10 cm) had the highest predation rates on New Zealand mud snails followed by medium Procambarus (5–10 cm), the dragonfly Aeshna walkeri, and small Procambarus (<5 cm) (Fig. 13). The damselfly Argia vivida also had predator impact indices significantly different from zero, but other taxa only consumed mud snails sporadically (Protochauliodes, Cordulegaster dorsalis, Octogomphus specularis, Notonecta hoffmanni, Drunella sp., Paltothemis lineatipes) or not at all (Rhyacophila sp., Isoperla sp., Lepidostoma sp. Gumaga nigricula). Some of the predators that consumed substantial numbers of mud snails were size-selective, with Argia naiads, for example, largely consuming small (<2 mm) and medium-sized (2–4 mm) snails, but not large snails (>4 mm) (paired t tests on proportions consumed, small ~ medium > large, two-tailed t ₁,₁₉ = −4.1 and −4.4, P’s < 0.001). Predator impact indices for small Procambarus were significantly higher for small (<2 mm) than large (>2 mm) snails, but medium and large Procambarus preyed preferentially on large snails (paired t test for medium and large crayfish combined, large snails > small snails, two-tailed t _1,5 = −9.7, P < 0.001) (Fig. 14). On the other hand, Aeshna naiads preyed on all snail size classes at similar rates.

Fig. 13

Predator impact indices (means ± 1 SE, per predator per day) for different invertebrate taxa and, in the case of Procambarus clarkii, different predator size classes. Invertebrate taxa with the same following alphabetical letters were not significantly different (P > 0.05, Tukey’s HSD test)

Fig. 14

Size-selective predation on Potamopyrgus by different size classes of Procambarus clarkii. Predator impact indices (means ± 1 SE) are shown for each predator size class feeding on each of two Potamopyrgus size classes (small = < 2 mm, black bar; large = > 2 mm, white bar). Different letters next to bars for each predator size class indicate significant differences in predator impact on different snail size classes

Discussion

The success and impacts of an invasive species in a new environment depend on its capacity to tolerate novel ecosystem conditions and then attain abundances that alter trophic relationships and/or disturbance regimes (Vitousek et al., 1997). Potamopyrgus antipodarum has become established in a wide variety of freshwater and estuarine habitats on four continents, including aquatic systems in mediterranean climates (Ponder, 1988; Alonso & Castro-Diez, 2008, 2012). Although P. antipodarum is present in some waterways in California, its impacts are largely unknown and management prescriptions are uncertain (ANS, 2007; Davis and Moeltner, 2010; Brenneis et al., 2011; Moore et al., 2012). To address this knowledge gap, we used a combination of field monitoring and laboratory experiments to evaluate P. antipodarum dynamics in a southern California stream and its potential impacts on the native stream biota.

Population dynamics and abiotic regulation of life cycles

Invasive P. antipodarum can achieve extremely high densities during summer months in some temperate streams, which has been attributed to the high primary production, stable discharge, and benign summer temperatures found in invaded habitats (Richards et al., 2004; Hall et al., 2006; Kolosovich et al., 2012). In most temperate streams, life cycles are marked by high summer reproductive and growth rates, but little or no production during the winter, presumably owing to low temperatures (Moffitt & James, 2012). In our southern California study stream, P. antipodarum populations reached densities as high as 150,000 individuals/m² in late spring and summer, but growth and reproduction declined in the winter. In contrast to studies in other temperate regions, some reproduction was observed in winter, presumably because water temperatures remained above 9°C. Population dynamics reflected stream discharge patterns with declines in abundance being associated with winter floods. Although P. antipodarum thrives in thermally stable streams with moderate flow variability (Cox & Rutherford, 2000), flow rather than thermal regimes largely drove P. antipodarum’s dynamics in Piru Creek, consistent with previous results (Holomuzki & Biggs, 1999; Holomuzki & Biggs, 2006; Sepulveda & Marczak, 2011). Most Californian streams known to support high abundances of P. antipodarum are at least partially regulated with relatively stable discharge patterns, so we postulate that flashy, unregulated streams in mediterranean climates may be at lower risk for P. antipodarum invasion than regulated streams.

In addition, P. antipodarum population densities declined during low flow periods in August, when temperatures exceeded 20°C. Population densities tended to peak between 15° and 20°C, congruent with our laboratory results showing that growth rates and survivorship declined as temperatures approached 27°C. Studies in other regions reported P. antipodarum at temperatures ranging from 0 to 34°C, with population growth rates falling below ca. 12°C and above 25°C (Cox & Rutherford, 2000; Dybdahl & Kane, 2005; Cejka et al., 2008). Because temperatures are benign during southern California’s winters, P. antipodarum shows lower seasonality in growth and reproduction, near-continuous reproduction with indistinct cohorts, and greater annual production than in other temperate regions (Evans, 2012).

In Piru Creek, P. antipodarum densities were similar in pools and riffles, and on hard versus soft substrata. On the other hand, P. antipodarum was most abundant in filamentous algae, primarily Cladophora glomerata, and juveniles were found during peak reproduction periods primarily on macroalgae. Suren (2005) similarly found that P. antipodarum showed a preference for filamentous green algae. Cladophora provides increased surface area, reduced flow conditions, and microhabitat and epiphytic resources for many stream invertebrates (Dudley et al., 1986a, b; Dudley, 1992), perhaps accounting for the increased abundance of P. antipodarum in macroalgal habitats.

Beyond these relationships with macroalgae, stream discharge and, perhaps, temperature, no other environmental variables were associated with P. antipodarum densities. The conductivities recorded in Piru Creek (>447 µS/cm) were far higher than those proposed to limit P. antipodarum‘s distribution and abundance (Herbst et al., 2008) and similar to the conductivities recorded in other streams with high P. antipodarum densities (Hall et al., 2003, 2006; Kerans et al., 2005).

Grazing effects

Potamopyrgus antipodarum is an opportunistic feeder on fine detritus and microalgae (James et al., 2000; Broekhuizen et al., 2001; Suren, 2005). Given its potential to achieve very high densities, we anticipated that snails would alter algal composition, biomass, and production, with indirect impacts on native grazer species. Prior results have been mixed, however, with some studies showing substantial reductions in primary producers ( Hall et al., 2003; Cada, 2004) and changes in algal assemblage composition (Arango et al., 2009), and others showing little impact on algal resources (Welch et al., 2000; Murria et al., 2008; Riley et al., 2008). In our controlled studies, P. antipodarum significantly reduced algal biomass at per capita rates comparable to those of native insect and snail grazers. Grazing rates were highest at low grazer densities and declined as grazer density increased. The net impacts of P. antipodarum on algal levels were generally low, perhaps owing to grazer excretion of ammonium which would stimulate algal growth (Dorgelo & Leonards, 2001; Riley, 2002). One study also suggested that P. antipodarum impacts on algae were modest because of its low energetic requirements and morphological specialization for low profile, adnate algae (Broekhuizen et al., 2002). Conversely, others have reported that high densities of P. antipodarum consumed 75% of gross primary production and significantly altered nitrogen dynamics and food web structure (Hall et al., 2003, 2006). In our studies, we found little evidence that P. antipodarum has higher grazing rates or can more efficiently convert food to biomass than do native grazers (but see Aberle et al., 2005).

Dietary analysis showed some differences in the algae ingested by different grazers. Nymphs of the mayfly Centroptilum consumed a greater proportion of medium-sized and large diatoms than did snails. Although the snail species had similar diets, P. antipodarum ingested a greater proportion of medium-sized diatoms than Pyrgulopsis, which favored small diatoms. Despite some evidence for slightly different food habits, grazing rates of different grazer species when together were sometimes lower than those predicted from grazing rates when each species was alone. These results for Potamopyrgus with Pyrgulopsis, Centroptilum, or Gumaga could be attributed to interference between species, a contention supported by Kerans et al.’s (2010) observations of mutual interference between P. antipodarum and baetid mayflies. They also found that Brachycentrus, a filter-feeding caddisfly larva, reduced the use of tile surfaces by P. antipodarum, illustrating that interference interactions can occur between species with different feeding modes or food habits (Dudley et al., 1990), similar to what we observed between Gumaga, a consumer of macroalgae and detritus (Feminella et al., 1989), and P. antipodarum at high densities.

In our studies, high densities of P. antipodarum altered the composition and physiognomy of algal assemblages consistently reducing the abundances of larger unicellular diatoms and several soft-tissue filamentous taxa including cyanobacteria (e.g., Cylindrospermum), green algae (e.g. Spirogyra, Mougeotia,), and the colonial diatom, Melosira varians (Holomuzki & Biggs, 2006; Arango et al., 2009). On the other hand, the relative and absolute abundances of tough macroalgae, such as Cladophora and Zygnema, increased with increasing P. antipodarum density, presumably by snails removing epiphytic diatoms which can compete with host algae for common resources (Dudley, 1992). Explanations for increases in planktonic Ankistrodesmus with increasing P. antipodarum densities are more problematic, but also could involve nutrient regeneration by the snails. We conclude that high densities of P. antipodarum have substantial effects on algal assemblages, but that those effects will depend on the structure of the primary producer community and the relative importance of snail effects mediated through direct feeding versus nutrient regeneration.

Grazer–Grazer interactions

Literature results indicate that the effects of P. antipodarum on native grazers can range from reductions (Richards, 2004; Kerans et al., 2005, 2010; Murria et al., 2008; Riley et al., 2008; Schmidlin et al., 2012) to little or no impact (Cada, 2004; Brenneis et al., 2010) to increases in native grazer densities (Schreiber et al., 2002; Cope & Winterbourn, 2004). In our study, P. antipodarum competitive impacts depended on the identity and density of competitors. P. antipodarum and Gumaga had no effects on each other’s vital rates, presumably because they had different diets. Centroptilum mayflies have greater dietary overlap with P. antipodarum and, as expected, showed reduced per capita emergence rates when P. antipodarum was present, but this could be attributed to both interference and exploitative competition (Kerans et al., 2010). Conversely, P. antipodarum birth rates were greatest at high mayfly densities, presumably because dead mayflies augmented P. antipodarum food supplies either directly or via nutrient regeneration leading to enhanced algal growth.

Competitive interactions between P. antipodarum and native snails, particularly with the hydrobiid Pyrgulopsis, were expected to be intense owing to similarity in their diets. The death rates of both P. antipodarum and Pyrgulopsis were higher and birth rates of Pyrgulopsis were lower, at the highest densities of the heterospecific competitor, indicating competition between these species. Riley et al. (2008) reported that P. antipodarum and native Pyrgulopsis robusta reduced periphyton to the same degree, but that the introduced snail depressed the population growth of the native species, consistent with our results, and that the native species promoted the growth of Potamopyrgus, contrary to our results. Riley et al. (2008) suggested that these asymmetrical interactions accounted for the dominance of the exotic species where it was present and the restriction of native Pyrgulopsis to reaches without Potamopyrgus.

We also observed complex interactions between P. antipodarum and Physella not readily explained by overlap in their food habits or microhabitat use. Physella increased both the birth and death rates of P. antipodarum at intermediate densities, but Physella mortality declined in the presence of P. antipodarum. Furthermore, Physella had the greatest negative impacts on its own feeding, birth, and population growth rates, and greatest positive effect, when alone, on its own death rates. Previous studies indicated that P. antipodarum had weak competitive interactions with other native gastropods in New Zealand, but that it enhanced the growth and fecundity of the introduced Physella acuta (Cope & Winterbourn, 2004).

Interactions between P. antipodarum and Anaxyrus boreas larvae (western toad tadpoles) were also complex. Over two months, tadpoles had little effect on algal community structure, but high Potamopyrgus densities reduced the abundances of medium to large diatoms and soft filamentous taxa while promoting the growth of filamentous Cladophora. Tadpole development tended to be slowed by high P. antipodarum densities, particularly when fed Cladophora and diatom food sources, although this trend was not statistically significant. On the other hand, tadpole survivorship decreased with increasing P. antipodarum density and was lower with Cladophora than other food types. In general, then, P. antipodarum appeared to negatively affect tadpole survivorship by facilitating inedible Cladophora growth, which also inhibited tadpole movement, and reducing the growth rates of edible algae. Conversely, P. antipodarum individual growth, reproductive, and death rates were related more strongly to their own densities than to the presence of tadpoles. Competitive interactions between snails and tadpoles have been documented in other systems (e.g., Holomuzki & Hemphill, 1996; Smith et al., 2012). These results carry management implications because Piru Creek provides critical habitat for the endangered arroyo toad, (A. californicus); a congener of the western toad also present in this system. Because high P. antipodarum densities can reduce edible algae while enhancing Cladophora cover in the sand-silt feeding habitats of arroyo toad tadpoles (Sweet & Sullivan, 2005), they may affect arroyo toad populations, posing problems for the management of this species.

Interactions with predators

The introduction of a new species into an ecosystem can provide a new food resource for secondary consumers (predators) and alter predator–prey relationships, and native or non-native predators may inhibit the establishment and expansion of introduced species, sometimes below thresholds of management concern (ANS, 2007; Dzialowski, 2007; Ricciardi & MacIsaac, 2011; Montserrat et al., 2012; Twardochleb et al., 2012). Because of the possible importance of predators in regulating the population growth of invasive species, we examined the potential for non-native crayfish (Procambarus clarkii) and a variety of native insects to consume P. antipodarum. Crayfish were the most effective predators on P. antipodarum, with small individuals (<5 cm) consuming primarily small snails (<2 mm) and larger (>5 cm) crayfish consuming large snails (>2 mm). Other studies have shown that crayfish can consume P. antipodarum at high rates, with the signal crayfish, Pacifastacus leniusculus, consuming over 900 snails in 12 h (Twardochleb et al., 2012; also see Brenneis et al., 2011).

Predatory insects, namely naiads of the dragonfly Aeshna walkeri and the damselfly Argia vivida, consumed snails at significant rates, in the case of Aeshna at rates comparable to those for small crayfish. Aeshna is a large-bodied, active predator found in stream pools, whereas Argia is found primarily in slow-flowing sections of small and medium-sized streams, a microhabitat favored by the snail. Argia also can be found in macroalgal mats and prefers small snails, so may pose a predatory bottleneck for developing P. antipodarum. Although not addressed in this study, fish impacts on P. antipodarum populations in southern California are likely to be minimal, either because fish are uncommon in many P. antipodarum habitats or because fish rarely ingest and often cannot digest P. antipodarum, obtaining little nutritional benefit from this invasive species (Haynes et al., 1985; Cada, 2004; Bersine et al., 2008; Vinson & Baker, 2008; Brenneis et al., 2011). Generalist predators are often poor regulatory agents for invasive species, so the control of P. antipodarum populations may require introduction of specialized parasites, such as the trematode Microphallus, which infects P. antipodarum in its native habitat (Winterbourn, 1970; Jokela & Lively, 1995; Dybdahl et al., 2005; Hechinger, 2012).

The New Zealand mud snail as an invasive species

P. antipodarum has successfully invaded aquatic ecosystems in many regions of the world, engendering concern from conservationists, water managers and wildlife protection agencies (Ponder, 1988; ANS, 2007; Davis & Moeltner, 2010). Of particular concern is that this invasive species is capable of achieving very high densities (Hall et al., 2006), with possible impacts to fisheries via reductions in the quantity and quality of fish food resources (Vinson & Baker, 2008). As a consequence, considerable effort has been devoted to preventing its establishment and spread, although success may be low because this species is already widespread and easily dispersed (Richards et al., 2004; California Department of Fish and Game, 2005).

Potamopyrgus antipodarum possesses a number of traits (Kolar & Lodge, 2001) that make it an effective invader, including early reproductive maturity, high dispersal and population growth rates, phenotypic plasticity and tolerance of a broad range of environmental conditions. It also uses a variety of food types and benefits from humans as dispersal vectors. Although fecundity for this ovoviviparous species is low (ca. several dozen embryos per female (Richards, 2002)), P. antipodarum’s ability to reproduce parthenogenetically and produce multiple generations in a season enables high population growth under suitable conditions. Furthermore, by attenuating peak flows in streams and rivers, humans create conditions conducive to the establishment of non-native species not adapted to flashy hydrographs (Baltz & Moyle, 1993).

The reasons for the invasion success of P. antipodarum have been reviewed by Alonso & Diez-Castro (2008), but we lack guidelines for determining when its impacts reach levels necessitating control efforts (National New Zealand Mudsnail Conference, 2011). Our experimental results indicate that P. antipodarum and native grazing invertebrates affect primary producers and consumers in similar ways under laboratory conditions. Because these experiments often dealt with two trophic levels or pairs of species, it is not clear how P. antipodarum will affect stream communities with multiple trophic levels and many species (Murria et al., 2008).

The impacts of P. antipodarum invasion on stream communities in this region may be short in duration, because P. antipodarum populations reached high levels in Piru Creek for fairly brief periods owing to the effects of periodic high flow events. Accumulating evidence (Hall et al., 2006, Moore et al., 2012) also indicates that P. antipodarum may exhibit the ‘boom-bust’ population dynamics observed in numerous invasive species populations, with initially high densities during early establishment followed by a decline to more moderate, sustained levels (Simberloff & Gibbons, 2004). Recent surveys of Piru Creek seem to corroborate this trend (Bennett & Wood, unpublished data), although it is difficult to differentiate long-term declines from the effects of the current drought in the western U.S. If P. antipodarum declines are sustained, their impacts may be transitory. This population expansion–contraction pattern can be driven by several mechanisms, including depletion of food resources, alterations in environmental conditions, and increased predation or parasite pressure (Sakai et al., 2001; Simberloff & Gibbons, 2004; Strayer & Malcolm, 2006). Given the breadth of P. antipodarum’s food habits and environmental tolerances (temperature, salinity, water quality), it is unlikely that snail populations would be reduced in the long term by these factors. Invertebrate predators, particularly crayfish, consumed large numbers of P. antipodarum (Brenneis et al., 2011; Twardochleb et al., 2012; this study), but could not account for long-term snail declines because they were absent from some invaded systems (Moore et al., 2012) and, as generalist predators, were unlikely to provide sustained population regulation. It is possible that naturally occurring parasites or pathogens are infecting P. antipodarum, as observed in Europe (Gerard et al., 2003), thereby causing snail population declines which may preclude the need for biocontrol development.

Although our investigations indicate significant impacts of P. antipodarum on native algae and invertebrates, the likely vulnerability of P. antipodarum to high flows, drought and high temperatures suggest that it might not have large, long-term impacts on natural stream ecosystems in southern California. In streams where moderate flows are maintained by dams or other structures, however, P. antipodarum’s populations may thrive and pose risks to co-occurring native species. The need for management actions may be mitigated by the moderate effects of P. antipodarum on native species, which are similar to the effects of other native species, and their long-term population declines; however, their likely impacts on endangered species, such as the arroyo toad, suggest management action may be needed in special cases.