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Direct and indirect trophic effects of predator depletion on basal trophic levels

Honors Thesis

Steve Hagerty

Sc.B. Candidate

Environmental Science

Brown University

April 17, 2015

1

TABLE OF CONTENTS

Abstract 3

Introduction

Human impacts on coastal 4

Predator depletion in New England salt marshes 5

Consequences on meiofauna 6

Question and Hypothesis 7

Materials and methods

Study sites 8

Correlative sampling 8

Consumer removal 9

Sediment type experiment 9

Results 10

Discussion

Predator depletion impacts on lower trophic levels 14

Direct vs. indirect trophic effects 15

Indirect modification effects 15

Conclusion 15

Acknowledgements 17

Literature Cited 17

Appendix 22

2

INTRODUCTION

Human impacts on coastal ecosystems

Human impacts on ecosystems and the services they provide is one of the most pressing problems for conservation and (Vitousek et al. 1997a, Karieva et al. 2007). Understanding these impacts sufficiently to predict and/or mitigate them requires elucidating their mechanistic nature, interactions and cumulative effects.

All ecosystems have been impacted by human activity, whether on local scales by human development and point source (Vitousek et al. 1997a, Goudie 2013) or on global scales by and shifts in the global nitrogen supply (Hulme at al. 1999, Vitousek et al. 1997b,

Galloway et al. 2008). Accelerated by unchecked human population (Luck 2007, Goudie 2013) and economic growth (He et al. 2014), these impacts have led to the unprecedented collapses or major degradation of critical marine ecosystems, including coral reefs (Mora et al. 2008, Hughes et al.

2015), seagrasses (Lotze et al. 2006), mangroves (Kathiresan and Bingham 2001, Alongi 2002), benthic continental shelves (Jackson et al. 2001, Hall 2002), pelagic communities (Game et al. 2009) and salt marshes (Gedan et al. 2009, Bertness et al. 2002). A recent compilation of available data predicts that human impacts on marine ecosystems has globally pushed most marine ecosystems to tipping points of ecological collapse (McCauley et al. 2015). This makes elucidating the causes and effects of human impacts on ecosystems an ecological problem of large concern (Barbier et al. 2008,

Barbier 2012).

Some critical questions remain regarding human influence, such as whether impacts are additive, synergistic, or counteractive (Vitousek et al. 1997a, Crain et al. 2008), whether effects are largely direct or indirect (Wootton 1994, Werner and McPeek 1994, Schmitz and Suttle 2001, Vetter et al.

4 2005) and the cumulative impact of direct and indirect effects (Vitousek et al. 1997, Engelhardt and

Ritchie 2001, Vetter et al. 2005). Here we examine the combination of effects of human-driven

predator depletion in New England salt marshes. We test whether predator depletion is strengthening

lower control on meiofauna, and/or leading to salt marsh die-off conditions that promote meiofauna , by increasing carbon substrate supply through engineering.

Predator depletion and New England salt marshes

Salt marsh die-off was first reported in New England in 2004 (Smith

2005), but was later shown by historical reconstruction to have

started in the early 1980s (Coverdale et al. 2013a). Comparative and

experimental studies have shown that predator depletion leads to

increased densities of the herbivorous, burrowing purple marsh crab,

Sesarma reticulatum. At high densities, Sesarma denude the low

marsh of Spartina alterniflora (hereafter Spartina or cordgrass),

both above and belowground (Holdredge et al. 2010, Altieri et al.

2012). Along steep creek banks, barren, heavily burrowed peat calves off and gradually erodes away.

On less steep marsh banks, heavily burrowed peat erodes in situ, expanding abandoned burrows and

releasing dead debris and sediment no longer bound together by live plant roots. This release of

carbon-based peat sediment initiates a microbial and meiofauna production links (Lopez

and Levinton 1987). On Cape Cod (Altieri et al. 2012), Narragansett Bay (Bertness et al. 2014) and

Long Island Sound (Coverdale et al. 2013b), this has led to the conversion of hundreds of acres of

marsh habitat, which represents the legacy of over 200 years of peat accretion (Coverdale et al.

2014), to carbon rich peat sediment.

New England salt marsh die-off is not impacting all salt marshes. While still spreading (Bertness et

5 al. 2014), die-off almost exclusively occurs on marshes near or affected by heavy recreational fishing pressure (Altieri et al. 2012). Even on Cape Cod where die-off has been ongoing for 35 years, marshes protected from recreational fishing (e.g. in reserves, embedded within communities or inaccessible by boat or car) are largely free from die-off (Altieri et al. 2012, Bertness et al. 2014).

Recreational fisheries landings exceed 10 million tons globally (Cooke and Cowx 2004), and can rival or exceed commercial fishing in coastal (Beal et al. 1998, Schroeder and Love 2002).

Consequences on meiofauna

To date, mechanisms of the salt marsh recovery have been explored (Altieri et al. 2013, Bertness and

Coverdale 2014), but the consequences of salt marsh die-off on other trophic levels have yet to be examined. Lower or basal trophic groups, like meiofauna, have received little attention. Meiofauna are small benthic invertebrates in both marine and freshwater habitats that are operationally defined as that pass through a 1000 µm sieve, but are retained by a 63µm sieve. They are important in benthic habitats due to their high abundance, diversity, and essential role in benthic food webs, particularly in processing via the microbial (Coull, 1999). A limited number of studies have examined the response of benthic meiofauna to direct deposit feeder addition or removal (Hoffman et al. 1984, Schratzberger and Warwick 1999, Fleeger et al. 2008), most providing evidence for the meiofauna regulatory capacity of deposit feeders.

However, the underlying mechanisms of higher trophic level effects on meiofauna are not fully understood, especially for the indirect pathways such as altering habitat structure or nutrient cycling through ecological engineering effects. In fact, a recent research from freshwater realm (Majdi et al.

2013) has posited that the meiofauna colonization was indirectly improved by predator-flatworm through enhancing the deposition and retention of fine sediments. In New England salt marshes,

Sesarma and Uca crabs engineer their environment in numerous ways (Holdredge et al. 2010,

Coverdale et al. 2012), most notably through burrowing activity which increase rates of biological

6 and chemical exchange, and transport of detritus, which enhances microhabitat heterogeneity and

rates. At die-off sites, with more consumer activity from fewer top predators, the effects of engineering should be more pronounced. The direct and indirect impacts of this

anthropogenic collapse on meiofauna are unknown.

QUESTION AND HYPOTHESES

In this paper we examine the consequences of predator depletion in New England salt marshes on

meiofauna. What are the mechanisms and direction of impact of predator depletion on meiofauna?

Specifically, we experimentally test competing hypotheses (Fig. 1):

1) predator depletion directly decreases meiofaunal abundance and control by increasing deposit

feeder densities and consumer control of meiofauna

and/or

2) predator depletion indirectly increases

meiofaunal abundance by driving salt marsh die-off,

leading to more favorable conditions for meiofauna

- with fewer predators, Sesarma burrow and

engineer their environment more, releasing

accumulated peat and sequestered carbon, more

favorably supporting microbial and detritus food

sources for meiofauna.

Since the direct trophic consequences of human-

generated trophic cascades are relatively predictable

and well-studied, our work is focused on the less

studied indirect consequences of trophic cascades. Fig 1

7 This work can help assess the relative importance of these indirect effects in trophic interactions, as

these effects are often characterized as system specific and difficult to predict or generalize (Menge

1995).

MATERIALS AND METHODS

Study Sites

Our research was carried out on Cape Cod, MA and Narragansett Bay, RI at eleven previously

studied sites (Fig. 2; Coverdale et al. 2013c, Bertness et al. 2014). Five of these sites had never

experienced catastrophic creek bank die-off (> 50% of marsh edge entirely vegetated) and were

operationally designated as “healthy” marshes (Common Fence Point, Herring River, Red River,

Sippiwissett, and Waquoit Bay), three were operationally defined as “recovery” sites (Parker River,

Smith Cove, and Wing’s Neck) that had experienced severe die-off, but more than 50% of the marsh

creek bank edges had recovered, and three were operationally defined as “die-off” marshes (Colt

State Park, Hundred Acre Cove and Saquatucket Harbor) that had active die-off with >50% of their

creek banks currently affected. Detailed information and the location of these sites can be found

elsewhere (Coverdale et al. 2013c, Bertness et al. 2014).

Correlative sampling

At each of these sites we quantified pressure on fiddler crabs (Uca pugnax, hereafter Uca),

Uca burrow densities, meiofaunal densities, and substrate characteristics (% nitrogen and carbon).

Predation pressure on Uca was quantified in early August 2014 at each site by tethering 45 adult male crabs, 15 in open controls, 15 protected from predation in hardware cloth cages, and 15 exposed to predation in cage control treatments. Crabs were tied and glued to 15cm tethers (see Altieri et al.

2012 for methods) and pinned to the low marsh substrate at low tide with access to a shallow burrow

8 and scored for survivorship after two tidal cycles. Predation was easily assessed since crab predation

events left broken carapaces attached to the tethers or no remnants.

Uca burrow density counts were made in the low marsh at each site in late August 2014, by

randomly throwing a 1 × 1m quadrat >3m apart along creek banks and counting the Uca burrows (N

= 10/site). Uca burrows are easily distinguished from the burrows of the herbivorous crab Sesarma reticulatum, since Sesarma burrows are larger, have multiple surface openings and never associated with deposit feeding pellets. Uca burrows have a single surface opening, are typically in softer

substrate and are usually associated with feeding pellets.

Nitrogen and carbon content by weight was determined by taking 8 independent replicate samples

from each site. Each sample was obtained with a 2.5 cm diameter corer to a depth of 3cm from

sampling plots (0.25 × 0.25 m) randomly established at each site 50-100cm below the cordgrass

border. After drying to constant weight at 60 oC, sediment samples were ground and analyzed for

percent carbon and nitrogen with an elemental autoanalyzer (Model NC2100; ThermoQueat CE

Instruments, San Jose, CA, USA). Sediment carbon content was used as a proxy for organic content

or the contribution of eroded peat to the substrate (Coverdale et al. 2014).

Meiofaunal densities were evaluated by taking sediment from sampling plots (0.25 × 0.25 m) at three

healthy (Common Fence Point, Herring River and Waquoit Bay), recovery (Wing’s Neck, Smith

Cove and Parker River) and die-off (Saquatucket Harbor, Hundred Acre Cove and Colt State Park)

sites in August 2014. Six sampling plots (0.25 × 0.25 m) were randomly established at each site 50-

100cm below the cordgrass border. Three sediment cores were taken with a 2.5 cm diameter corer to

a depth of 3cm from each plot and combined into a single composite sample to reduce spatial

variance. These samples were immediately fixed with 10 % formalin for > 7 days, after which time

meiofauna were extracted by flotation with Ludox TM and counted under a dissecting .

9 Uca and meiofauna density were analyzed with nested ANOVA (sample sites nested within site types) to determine differences among habitat types (die-off, recovery and healthy). Tukey’s HSD test was used for post-hoc analysis. Since homogeneity of variances could not be achieved for predation pressure and substrate characteristics, non-parametric Kruskal-Wallis tests followed by a

Dunn’s post-hoc test were performed with site type as the main factor. We performed linear regression to examine the relationship between Uca density, substrate carbon content and predation rate.

Consumer removal experiment

To test the hypothesis that predator depletion at salt marsh die-off sites leads to increased Uca densities and concomitant reduction in meiofauna densities, we ran a total consumer removal caging experiment at two die-off (Colt State Park and Saquatucket Harbor) and two healthy sites (Common

Fence Point and Herring River). We predicted that if direct trophic interactions dictated meiofaunal densities, experimentally excluding consumers would have a greater effect at die-off than control sites. At each site we randomly marked 18 plots (25 × 25 cm) in sediment 50-100cm below the cordgrass border as: control plots (N = 6/site), consumer removal cages (20 × 20 × 5cm (L × W × H)

4 cm mesh stainless steel hardware cloth; N = 6/site), and cage controls (identical to consumer removal cages but with their sides removed to allow access to crawling benthic organisms like Uca;

N = 6/site).

These treatment replicates and ambient sediment samples were taken adjacent to each replicate the first week of June 2014. Plots were maintained weekly for 8 weeks after which time plot sediment was sampled for meiofaunal analysis with a 2.5 cm diameter corer to a depth of 3cm. Three cores from each plot were sampled and bulked into a composite sample. Total meiofauna density was analyzed with a split-plot ANOVA, with habitat (die-off and healthy) as the between-plot factor, site nested within habitat types, and caging treatment (cage, cage control, control) as a with-plot factor.

10 Data were log-transformed to meet the ANOVA assumptions. Tukey’s HSD test was used for post- hoc analysis.

Sediment-type experiment

To examine the hypothesis that carbon rich sediment at die-off sites leads to increased meiofaunal densities in comparison with healthy sites that have not been enriched with eroding carbon rich peat, we ran a sediment type manipulation experiment between a die-off (Saquatuket Harbor: % carbon =

18.6 ± 0.3; % nitrogen = 1.1 ± 0.02; C to N ratio = 16.8 ± 0.07) and a healthy (Herring River: % carbon = 0.2 ± 0.01; % nitrogen = 0.02 ± 0.001; C to N ratio = 8.3 ± 0.2) site. At both sites, we collected ~ 2L of surface (top 5cm) sediment and returned it to the laboratory. Sediments from each site were filled into 18 containers each (PVC tubes, 50mm in length and 2.6 mm in diameter) and retained with fishing net end caps (1 mm mesh nylon). The meiofauna in 12 containers of each sediment type were eliminated by freezing at -20 °C for 48hr, then incubating at 30 °C for 72hr, and freezing for another 48hr (Ronn et al. 1988; Blouin et al. 2005). The remaining 6 tubes of each type were not frozen and returned to the original site as a procedural control to test the tube effect on meiofauna communities. Sediment type was labeled with color-coded cable ties and fishing line.

Sediment tubes were deployed at each site (6 die-off meiofauna-eliminated sediment, 6 healthy meiofauna-eliminated, and 6 procedure control) 50cm below the intertidal cordgrass border at a depth of 3cm Tubes were buried horizontally and separated by > 25cm. Plots were checked and maintained weekly. After 8 weeks, sediment tubes were collected, and 6 control sediment samples of the same volume were taken near the tubes. Meiofauna in each replicate was washed and quantified.

Meiofauna densities were transformed as necessary and analyzed with a 2-way ANOVA, with treatment (die-off sediment, healthy sediment, procedural control and control) and site (Herring

River, Sacquatucket Harbor) as factors.

11 RESULTS

Correlative sampling

Predation intensity varied across habitat types (H2, 11 = 6.69, p = 0.035; Fig. 3A). The predation rate on tethered Uca at healthy sites (80 ±11%) was four times higher than at die-off sites (20 ± 7%).

Predation at recovery sites (58± 5%) was also greater than at die-off sites, but the difference was not significant. Uca abundance varied among marsh types in reflecting differences in predator control

(F2, 99 = 128.6, p < 0.001; Fig 2B). Uca density at die-off sites was twice that of healthy sites, and

intermediate at recovered marshes. Substrate carbon content (Fig. 3C) and nitrogen content (Fig. 3D)

also varied among marsh types with a similar pattern (H2, 88 = 66.5, p < 0.001; H2, 88 = 64.4, p <

0.001, respectively). Healthy sites had the lowest carbon and nitrogen content, recovery sites intermediate, and die-off sites the highest. Regressions between Uca densities, sediment carbon content and predation rate revealed that inter-site variation in predation depletion explained 82% of the inter-site variation in Uca density (p < 0.001; Fig. 4A) and 67% of the inter-site variation in sediment carbon content (p = 0.002; Fig. 4B). Recovery sites had higher meiofaunal density than healthy and die-off sites (F2, 45 = 10.49, p < 0.001; Fig. 3E), but no difference in meiofaunal density was detected between healthy and die-off sites (Tukey’s HSD post hoc, p > 0.05).

Consumer removal

After eight-weeks of consumer removal, meiofaunal density in cages was significantly higher than in

controls and cage controls at both die-off and healthy sites, suggesting that the consumer control by

Uca on meiofauna is significant (F2, 64 = 42.0, p < 0.001; Fig. 5). Excluding Uca led to a 75%

increase in meiofaunal density at die-off marshes, but only a 52% increase in healthy marshes,

reflecting differences in Uca densities between marsh types.

Sediment-type experiment

The sediment type manipulation experiment revealed differences in meiofaunal density across

12 treatments (F3, 40 = 9.73, p < 0.001; Fig. 6). Meiofaunal density increased by 164% more in die-off

sediment than in healthy sediment. Open control meiofaunal density was similar to die-off sediment

but was higher than healthy sediment. There was no significant difference between procedural

controls and other treatments. Site effect on meiofauna density in this experiment was not significant

(F1, 40 = 2.71, p = 0.11).

DISCUSSION

Predator depletion on New England salt marshes has led to increased Uca densities (Fig 3B, 4A) and

deposit feeding, directly decreasing meiofaunal abundance (Fig. 5) through a trophic cascade.

Released from predator control, however, burrowing and herbivory by Sesarma has led to salt marsh

die-off and the release of organic carbon substrate (Fig. 4B) that supports microbial and meiofauna

populations (Fig. 6). This process indirectly increases meiofaunal abundance by increasing the

and availability of sequestered carbon from centuries of accreted peat (Coverdale et al.

2014). Thus, both direct and indirect trophic effects may play important regulatory functions on

marsh basal trophic levels through counteractive mechanisms.

Predator depletion impacts on lower trophic levels

Our previous work in this system revealed that fishing pressure near recreational fishing access areas

leads to increased abundance of the nocturnal burrowing herbivorous crab, Sesarma reticulatum

(Holdredge et al. 2008, Altieri et al. 2010). This in turn leads to consumer-driven salt marsh die-off on a regional scale and the loss of hundreds of acres and over 200 years of accreted salt marsh peat on Cape

Cod (Coverdale et al. 2013b).

In this paper, we examined how predator depletion influences lower trophic levels directly and

indirectly. Inter-site surveys and Uca tethering predation pressure assays reveal that elevated Uca densities do occur at predator depleted die-off sites (Fig. 3A) decreasing meiofaunal abundance

13 through direct trophic interactions (Fig. 5). Our data also show that die-off site sediment is carbon

rich (Fig. 3C) due to peat erosion and that carbon rich sediment supports higher meiofauna densities

as an indirect effect of predator depletion (Fig. 6). Thus, our data show that the direct and indirect

trophic effects of predator depletion on meiofauna interact antagonistically, canceling out the

magnitude of each effect, and making them cryptic without explicit experimentation to separate and

isolate them.

Direct vs. indirect trophic effects

Our results suggest that human impacts may be difficult to generalize among communities because of

the idiosyncratic nature of indirect trophic effects. In communities where human impacts, like

predator depletion, result in direct trophic effects, human impacts may be predictable and

deterministic. Specifically, if affects an autogenic engineer () that

provides refuge for consumers, foundation species effects can impact consumers and direct effects

can predictably extend to lower trophic levels. If, on the other hand, disturbance indirectly affects

allogenic ecosystem engineers that convert a from one state to another, like Sesarma converting peat to carbon rich sediment, food web structure and trophic impacts will be indirect, unpredictable and lead to ecosystem-specific responses. Trophic cascades have been identified in a many ecosystems (Pace et al. 1999). Significant work elucidated the direct causal links of these cascades - for instance, sea otter populations control herbivorous sea urchins through consumptive effects. Sea otter removal leads to and loss of communities (Estes and

Palmisano 1974). However, less experimental and conceptual work has emphasized the indirect effects in these cascades or systems (Estes et al. 2004).

Indirect Habitat Modification Effects

How common are human impacts that lead to trophic structure shifts mediated by indirect trophic

14 effects? Indirect effects are ubiquitous in natural ecosystems (Menge 1995), suggesting that most human impacts have both significant direct and indirect consequences.

More recently, indirect trophic effects of predator depletion in marine systems have been studied

(Duggins et al 1989, Reisewitz 2002, Vicknair 1996, Irons et al 1986, Dayton 1975, Dulvy et al.

2004), showing large impact across functional groups and changing structure (Estes et al

2004). Other examples of coupled direct and indirect trophic effects have been observed in of Caribbean coral reefs, which led to coral loss and algal dominated food webs (Hughes

1994) and the extermination of wolves from North America which led to overgrazing, the loss of foundation species and the emergence of novel food webs (Ripple and Beschta 2004). Like our results, these are all examples of the direct and indirect effects of predator depletion impacting community organization and require experimental studies to understand.

Indirect trophic mechanisms are also ubiquitous structuring forces by themselves, in the absence of direct trophic mechanisms. , like the European herbivorous snail Littorina littorea clearing North American Atlantic estuarine shorelines of sediment and algal cover (Bertness 1984) and invasive zebra mussels filtering North American lakes of planktonic and shifting whole lake trophic structure (Strayer 2009), are examples of allogenic ecosystem engineers indirectly changing community composition and structure. These indirect effects also require experimental studies to quantify.

Nutrient pollution is an additional human impact that can shift estuarine pelagic and benthic food webs from seaweed, seagrass and coral-based ecosystems to rapidly growing phytoplankton and ephemeral algal-based ecosystems that increase sedimentation effects via indirect effects (e.g.

Schramm and Nienhuis 1996, Burkholder et al. 2007, McCook 1999). Nutrient loading in New

England salt marshes facilitates the invasion of non-native to invasive Phragmites genotypes and the

15 displacement of the native nitrogen fixing Phragmites (Holdredge et al. 2010b) and the entire native

vascular plant assemblage (Bertness et al. 2002). These are all examples of human impacts on

autogenic ecosystem engineers that led to shifts in trophic structure through ecosystem-specific

indirect effects, and these mechanisms are only shown through experimentation.

CONCLUSION

Human-driven predator depletion has significant cascading effects, through direct consumptive interactions and indirect habitat engineering effects. Here, we examined the effects of predator depletion on New England salt marshes on an ecologically important basal trophic group, meiofauna.

We tested the mechanisms of this impact - direct deposit feeder increase, or indirect sediment

modification - to predict the direction of community change on meiofauna.

Overall, we found correlative evidence that meiofauna densities did not vary significantly between

marsh types. Through experimentation, we found that both deposit feeder densities and sediment type

are important in regulating meiofauna density, and work in differing directions. Thus, the balance

between direct and indirect effects of human impacts among community types may rarely be

generalizable. Scientists will need to utilize experimentation to predict the mechanistic basis and

cumulative outcome of human impacts in communities.

ACKNOWLEDGEMENTS

Thank you to Steve Smith and Caitlin Chaffee for access to field sites, Elena Suglia for field support,

and Huili Chen, Mark Bertness, Sinead Crotty for their comments on the thesis, and Tyler Coverdale and Q He for comments on the manuscript for a paper draft that came out of this thesis project. Our research was supported by a NSFC fund (41201054) and Excellent Youth Scholars Training Project of

Hanghzou Normal University Fellowship to Huili Chen, Brown University Research and Teaching

Fellowships to me and Sinead Crotty and the Robert P. Brown Endowment to Mark Bertness.

16

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FIGURE LEGENDS

FIG. 1. Conceptual model

FIG. 2. Creekbank characteristics and site designation. Site was defined according to the dominant marsh bank characteristic (> 50%) of each marsh.

FIG. 3. Results of surveys that link predation depletion with meiofaunal density: (A) predation intensity, (B) Uca density, (C) carbon content, (D) nitrogen content, and (D) meiofaunal density.

Data are mean ± 1SE.

FIG. 4. Relationship between Uca density, carbon content and predation pressure. As predation pressure decreases, Uca density (A) and carbon content (B) increase. Data are mean ± 1SE.

FIG. 5. Consumer removal experiment. Data are mean ± 1SE for 6 replicates per treatment. Different letters denote significant differenced at P < 0.05 (Tukey HSD post hoc test).

FIG. 6. Sediment transplant experiment. Data are mean ± 1SE for 6 replicates per treatment.

Different letters denote significant difference at P < 0.05 (Tukey HSD post hoc test).

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FIG. 1.

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FIG. 2

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FIG. 3.

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FIG. 4.

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FIG. 5.

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FIG. 6.

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Thank you to the Bertness lab!

(From left): Matt Bevil, Elena Suglia, Sinead Crotty, Steven

Hagerty, Huili Chen. Photo taken by Mark Bertness.

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