RESEARCH ARTICLE

How hydroperiod and species richness affect the balanceof resource flows across aquatic-terrestrial habitats

Tiffany A. Schriever • M. W. Cadotte •

D. Dudley Williams

Received: 17 February 2013 / Accepted: 9 October 2013 / Published online: 18 October 2013

� Springer Basel 2013

Abstract Ecosystem functioning is influenced by the

flow of nutrients, detritus, and organisms. Variation in

these flows, like that found in temporary ecosystems,

affects temporal and spatial patterns of community diver-

sity and secondary production. We evaluated the influence

of hydroperiod and ecosystem size on the bi-directional

flow of subsidies from intermittent ponds and surrounding

forests by quantifying litter deposition and the abundance

and biomass of emerging insects and amphibians. In

addition, we assessed whether amphibian and insect

diversity influenced the magnitude of cross-habitat

resource flux. We found substantial spatial and temporal

variation in the magnitude, composition, and timing of

cross-habitat resource subsidies. Overall, deposition into

ponds far exceeded biomass exported via insect and

amphibian emergence. We found a negative association

between resource flux and the diversity of amphibians and

insects. Different species groups contributed to flux pat-

terns unequally, with insects having higher diversity but

lower flux compared to amphibians. Organismal flux varied

among ponds with amphibians having the highest flux in

the shortest hydroperiod pond and insect flux was highest

from an intermediate hydroperiod pond. This work reveals

how variation in pond size and permanence affects species

diversity and ecosystem flows. Species composition played

a major role in flux differences across ponds. Further, given

the general lack of research and conservation prioritization

of temporary ponds, uncovering how these ponds contrib-

ute to cross-habitat linkages is necessary to develop fully

integrated management strategies.

Keywords Amphibians �Biodiversity–ecosystem function � Biomass �Cross-habitat energy flow � Disturbance � Insects �Subsidy hypothesis

Introduction

Habitats are invariably open systems linked by the flow

of nutrients, detritus, and organisms, all of which have

the potential to influence population and community

dynamics, food webs, and diversity-stability relation-

ships in the recipient habitat (Polis et al. 1997; Nowlin

et al. 2008; McCoy et al. 2009). Spatial flows may also

have a large influence at the meta-ecosystem scale,

thereby strengthening connections between local eco-

systems (Loreau et al. 2003). Spatial subsidies are

resources that are spatially and temporally variable,

originating in one donor habitat and moving into another

(Polis et al. 1997). The flow of resources across habitat

boundaries is influenced by the distance separating

habitats, habitat size, perimeter-to-area ratio of the focal

habitat, adjacent habitat type, and the behavior of fluxes

Electronic supplementary material The online version of thisarticle (doi:10.1007/s00027-013-0320-9) contains supplementarymaterial, which is available to authorized users.

T. A. Schriever � M. W. Cadotte � D. D. Williams

Ecology and Evolutionary Biology, University of Toronto,

25 Willcocks Street, Room 3055, Toronto M5S 3B2, Canada

T. A. Schriever � M. W. Cadotte � D. D. Williams

Department of Biological Sciences, University of Toronto

Scarborough, 1265 Military Trail, Toronto, ON M1C1A4,

Canada

Present Address:

T. A. Schriever (&)

Department of Zoology, Cordley Hall 3029, Oregon State

University, Corvallis, OR 97331-2914, USA

e-mail: tiffany.schriever@gmail.com

Aquat Sci (2014) 76:131–143

DOI 10.1007/s00027-013-0320-9 Aquatic Sciences

123

that move between habitats (Cadenasso et al. 2004).

Spatial subsides can enter the food web at multiple tro-

phic levels (Polis et al. 1997) and vary in quantity and

quality (Massol et al. 2011).

Aquatic food webs are intimately linked with the sur-

rounding terrestrial landscape through a number of energy-

flow pathways. For example, aquatic systems contribute to

the terrestrial food web via aquatic insect emergences

(Benke 1993; Stagliano et al. 1998; Gratton et al. 2008) and

the reciprocal return of terrestrial material to aquatic food

webs through litterfall, such as organic matter deposition in

streams (Fisher and Likens 1973; Hutchens and Wallace

2002; Rubbo and Kiesecker 2004; Rubbo et al. 2006),

insect infall (periodical cicada deposition into streams and

ponds; (Pray et al. 2009), and nitrogen deposition (geese

deposit nitrogen-rich waste into wetlands gained from

foraging in agricultural land; (Post et al. 1998; Kitchell

et al. 1999). Prey subsidies, such as emerging freshwater

insects, have been shown to increase the abundance of

riparian arthropod assemblages (Hoekman et al. 2011;

Dreyer et al. 2012), or in a case where the aquatic insect

subsidy has been restricted, the abundance and biomass of

riparian lizards and arthropods was reduced (Sabo and

Power 2002). In addition, recent work has revealed that not

only the magnitude, but also the functional role of the

spatial prey subsidy is important in determining the impact

on recipient food webs (Leroux and Loreau 2008; Wesner

2010, 2012).

Woodland ponds have diverse communities of insects

and amphibians. Species in both groups undergo meta-

morphosis and leave the aquatic habitat and enter the

terrestrial realm, making them important sources of sub-

sidies between habitats (Schreiber and Rudolf 2008). These

ponds are often relatively small, have high perimeter-to-

area ratios, and are abundant in the landscape (Palik et al.

2001, 2006). In aquatic systems, environmental variation in

the form of duration (i.e., hydroperiod) influences species

richness, diversity, community composition, and survivor-

ship and has direct consequences for secondary production

(Welborn et al. 1996; Sabo and Post 2008). Freshwater

habitats along the hydroperiod gradient range from tem-

porary habitats with relatively few small-bodied predators,

transitioning to permanent fishless habitats with large

invertebrate predators, and finally habitats with fish (Wel-

born et al. 1996). It is unclear how environmental

variability interacts with species richness to influence

cross-habitat energy flow. A considerable accumulation of

published research has found that greater biodiversity

results in higher ecosystem functioning (i.e. productivity

and energy flow; Loreau et al. 2001). The question remains

of how resource subsidies and environmental variability

affects the biodiversity–ecosystem function relationship

(Romanuk et al. 2010).

There are few studies that have quantified the impor-

tance and reciprocal connection of temporary ponds to the

surrounding landscape (but see Palik et al. 2006; Palik and

Kastendick 2010). Given the extensive occurrence of

temporary ponds in temperate regions, ignoring their con-

nections to surrounding habitat hinders our ability to

adequately conserve and manage biota and the resulting

ecosystem services provided by these systems.

Here we present a case study that examined the envi-

ronmental (hydroperiod and ecosystem size) and

community (species richness, diversity) influences on the

magnitude of subsidy flows across habitat boundaries

(strength of habitat coupling). We conducted a natural

experiment using seasonally intermittent freshwater ponds

(classification follows Williams 2006) varying in hydro-

period and size. We quantified the aquatic export of insect

emergence and amphibian metamorphosis leaving ponds

and the reciprocal terrestrial input of organic matter

deposition of litterfall entering the ponds. Using this data,

we tested three hypotheses: First was that the predictable

shift in community composition across a hydroperiod

gradient will have a noticeable and positive effect on the

magnitude and/or composition of the cross-habitat resource

flows. Second, we hypothesized that pond hydroperiod and

size would influence the magnitude of the aquatic-terres-

trial linkage. In particular, we predict that longer

hydroperiods and larger ecosystem size will have higher

amphibian and insect emergence and biomass than shorter

hydroperiod ponds. Finally, we tested for a relationship

between diversity and productivity among ponds. Little is

known about how energy flow influences these biodiversity

effects. Since temporary woodland ponds are often small

and their food webs rely on detrital leaf litter, it is likely

that litter fall will scale with pond size, and we predict the

amount of litter deposition will be directly proportional to

the amount of animal export.

Methods

Study sites and physical habitat

We conducted this investigation at the Queen’s Biological

Station north of Kingston, Ontario, Canada (44.565977 N,

-76.324223 W). The area is predominantly mixed conif-

erous-deciduous forest and supports numerous types of

ponds. Ponds are filled by snowmelt in early spring and

lose water as summer progresses, resulting in some ponds

drying completely. We used four temporary ponds that

varied in hydroperiod and size (hereafter, Short, Interme-

diate 1, 2, 3; Table 1). Our experimental design took

advantage of a natural environmental gradient, capitalizing

on spatial and temporal variation; but this resulted in a

132 T. A. Schriever et al.

123

trade-off in replication. We were most interested in mea-

suring responses to natural environmental variation and

laying a foundation for an unknown system. We used

similar, non-replicated natural studies to check the gener-

ality of our results.

Hydroperiod was measured as the consecutive number

of days with water following ice off to complete drying in

2010. We measured pond area each month by mapping the

perimeter of each using a Trimble TSC1 GPS connected to

a ProXRS satellite and RTCM receiver. All mapping was

completed in one or two consecutive days within each

month. The area (m2) of each pond was then calculated in

ArcMap10 (Esri, Redlands, CA, USA) and averaged to

yield an average pond area for the entire study period.

Every month, we measured dissolved oxygen (mg/L),

water depth (m) and percent canopy cover measured at the

center deepest point in each pond. We used a hand-held

Hydrolab Quanta multiparameter probe (Hach Environ-

mental, Loveland, CO) to measure dissolved oxygen. We

took four canopy readings facing each cardinal direction,

counting the number of open squares on a spherical den-

sitometer (Forest Densiometers Bartlesville, OK, USA).

The four readings were averaged, multiplied by 4.17, and

the product subtracted from 100 % to obtain percent can-

opy density (California Department of Pesticide

Regulation, Environmental Monitoring Branch, SOP

number: FSOT.002.01). Data loggers placed at the center

of each pond *3 cm from pond bottom (StowAway Tid-

bit; Onset computer, Pocasset, MA, USA) recorded water

temperature every 5 h. These were averaged for a 24 h

period to acquire a daily temperature reading during the

period of 7 April to 14 October 2010 or until the pond

dried.

Field and laboratory methodology

To quantify emerging aquatic insects, we placed three

floating emergence traps (total trap area 0.63 m2) in each

pond. Traps were deployed in each pond from May to

August 2010 on the same day in each pond for *48 h

collection periods. We made four collections in May, three

collections each in June and July, and two in August for a

total of 2032.5 trap hours. Insects were aspirated from the

traps and placed into vials containing 80 % ethanol. All

aquatic insects were counted and identified to family.

Insects of non-aquatic origin (e.g. ants) were removed and

not used in further analyses. We stopped trapping for

emergent insects after the pond dried. Since ponds did not

dry on the same day, the number of trapping days per pond

varied. We measured the body length (mm) of a subsample

of random individuals in each family (refer to Online

Resource Table 1 for sample numbers) using a Nikon

SMZ1500 microscope connected to a computer using the

NIS Elements D3.1 software (Shinagaw-ku, Tokyo, Japan).

We used published length-mass conversions (Online

Resource Table 2) to estimate dry mass (DM) for each

individual and calculated biomass for each family by

summing biomass of all constituent individuals (Online

Resource Table 1). Studies have found preservation in

alcohol reduces invertebrate dry mass due to leaching

(Leuven et al. 1985; von Schiller and Solimini 2005;

Edwards et al. 2009), but this loss of dry mass was cor-

rected by using length-mass equations from non-preserved

specimens (von Schiller and Solimini 2005). Aquatic insect

emergence (Ei; g C m-2) was calculated for each pond

using dry mass estimates divided by trap area.

To capture metamorphosing amphibians, on June 4–6

we encircled each pond and sank pitfall traps (five gal

buckets) every 10 m flush with the fence and the ground

along the pond side of the fence (Gibbons 1974). Ponds

were encircled once breeding was over and before meta-

morphs emerged. We monitored tadpole development in

ponds on a bi-weekly schedule starting April 5, 2010 to

ensure we did not miss metamorph emergence. Each trap

contained a wet sponge to provide moisture to sustain

trapped amphibians. This method allowed us to capture

every amphibian moving out of the ponds, although tree-

frogs can climb over fences, and therefore their abundance

Table 1 Physical characteristics of the study ponds measured in 2010

Measurement Pond

Short Intermediate 1 Intermediate 2 Intermediate 3

Mean depth (m) 0.26 (0.20) 0.47 (0.33) 0.25 (0.16) 0.36 (0.17)

Mean water temperature (�C) 14.43 (5.23) 18.32 (4.86) 14.67 (4.29) 16.01 (4.14)

Mean dissolved oxygen (mg/L) 6.68 (1.06) 6.69 (2.15) 5.43 (1.34) 7.18 (1.40)

Mean canopy cover (%) 74.54 (5.57) 31.67 (11.59) 87.31 (12.72) 26.48 (8.93)

Mean pond area (m2) 182.68 (76.47) 987.48 (343.97) 249.13 (157.30) 188.52 (84.56)

Mean perimeter:area (m) 0.43 (0.13) 0.20 (0.03) 0.61 (0.34) 0.37 (0.11)

Hydroperiod (days) 78 145 168 168

Standard deviation is in parentheses

Spatial subsidies between pond and forest habitats 133

123

may have been underestimated. Pitfall traps were opened

and checked daily for all ponds during the periods of 8–10

and 20–29 June, 7–13 and 19–23 July, 9–12 and 16–18

August. Intermediate 3 pond traps were also open 28–29

August because it was the only pond remaining with water.

In total, traps were open for 99 collection days. We stopped

trapping for metamorphs once the pond completely dried

and we collected zero metamorphs for several trapping

days following drying. All amphibians collected were

counted, identified to species, weighed, and the snout-to-

vent length (SVL) measured with digital calipers. We

euthanized a small sample of individuals for biomass

estimation (sample sizes listed in Online Resource

Table 3). Euthanized individuals were dried at 80 �C for

5–7 days, weighed (DM), then ashed at 500 �C for 1–2 h

and reweighed to determine ash-free dry mass (AFDM).

We established species-specific length–DM and length-

AFDM relationships using linear regression to estimate dry

mass and AFDM from the SVL of all individuals collected

(Online Resource Table 3). The DM values were summed

to provide the total biomass of amphibians from each pond.

Amphibian emergence production (Ea; g C m-2) was cal-

culated for each pond using dry mass estimates divided by

pond area.

Litterfall is the main source of terrestrial organic matter

and is a main food resource for many invertebrates and

amphibians within these woodland ponds. Therefore, we

collected overhead inputs of coarse terrestrial particulate

organic carbon (i.e., litterfall) once a month from two litter

traps (0.5 m2 each) placed in each pond from August 29

through November 14, 2010. Trap locations were haphaz-

ardly selected within the wetted area of the ponds and

remained in the initial location throughout the 3 months of

sampling. Litterfall was removed from each trap and

placed in paper bags. In the laboratory, bags of collected

litter were emptied onto a white enamel tray, sorted into

deciduous leaf (mostly Quercus sp., Acer sp., Betula sp.),

coniferous needle (Pinus sp.), woody material (twigs,

branches, bark, cones, etc.), and miscellaneous (bud scales,

pollen, etc.) components, dried at 55 �C for at least 96 h,

and weighed (g DM). We quantified litter dry mass depo-

sition for each trap (Dareal; g m-2). Total litter deposition

(i.e., biomass) for each pond was the sum of litter dry mass

collected over the sampling period.

We used Gratton and Vander Zanden’s (2009) measure

of flux because it incorporates habitat geometry into the

calculation of cross-habitat resource flow. They define ‘‘the

flux (F) of emerging aquatic insects to land as the amount

of insect production annually leaving the body of water per

meter of aquatic shoreline (Fi; g C m-1 year-1)’’. We used

this metric of flux for amphibians (Fa) as well since they

are equivalent in terms of leaving the aquatic habitat. The

calculation was: F ¼ E � ðA=pÞ, where E is the emergence

in g C m-2 year-1of insects (Ei) or amphibians (Ea), A is

the pond area and p is pond perimeter. C represents grams

of dry mass carbon. The study ponds were not perfectly

circular, therefore a shoreline development factor ðDL ¼p=½ pAð Þ1=2�Þ corrected for irregularities in pond shape. We

used the equation F ¼ Eðr=2DLÞ, where r is pond radius.

Litter input per meter of shoreline (Fl), and flux to pond

was calculated as Dareal r=2DLð Þ. We also used a measure

of total flux or productivity to estimate the total aquatic to

terrestrial carbon flux (Etotal) and total terrestrial to aquatic

carbon flux (Dtotal) (Vander Zanden and Gratton 2011). We

calculated total emergence numbers and total production

for each site for insects (Eitotal) and amphibians (Eatotal)

separately because we were interested in the magnitude of

each subsidy type (Ei total ¼ Ei � A ; Ea total ¼ Ea � AAn

Etotal was calculated for each insect family and amphibian

species per pond as well as the Etotal for each pond. Lit-

terfall total was calculated in a similar fashion

(Dtotal = Dareal 9 A; g C y-1).

Statistical analyses

To test if the magnitude of resource flows changed over the

study period, we used general linear models. We calculated

Friedman rank sum test using friedman.test function in R, a

nonparametric version of one-way ANOVA with repeated

measures to evaluate differences in biomass export and

litterfall input among the ponds across repeated sampling

events. We did not use one-way ANOVA because the

assumption of equal variances of the differences between

repeat observations was not met and because of the small

sample size (four ponds). For these data, emergence and

litterfall are the dependent variables, the day of year

(DOY) the sample was collected was the repeated mea-

sures factor (‘‘block’’) and pond was the group factor. An

analysis of covariance (ANCOVA) tested for differences in

total flux between species. We calculated Bray–Curtis

distances on a matrix of insect family flux and amphibian

species flux values by pond. Dissimilarities were multiplied

by 100 and the result subtracted from 100 to get a measure

of percent similarity. We used simple linear correlations to

calculate Pearson product-moment correlation coefficients

in R using the cor and cor.test functions (R Development

Core Team 2009) to compare the relationship between

pond size and hydroperiod on the export fluxes of insects

(i.e., Eitotal, Fi) and amphibians (i.e., Eatotal, Fa), and the

input flux of litterfall (i.e., Dtotal, Fl). Ponds vary monthly

in size and shape, so we used mean pond area (m2) as

calculated from the available wetted months as the pond-

size variable.

We calculated species richness and Shannon diversity

(H0) for amphibians and insects (familial richness and

diversity) using the Biodiversity Calculator provided by J.

134 T. A. Schriever et al.

123

Danoff-Burg and C. Xu (http://www.columbia.edu/itc/cerc/

danoff-burg/Biodiversity%20Calculator.xls) using abun-

dance data (total numbers of each taxon collected per site).

We tested for differences between emergent amphibian and

insect communities in richness, diversity, and evenness

with t tests. We also looked at the relationship between

litterfall and pond canopy cover. Our rule of thumb for

interpreting strengths of correlations was very strong if r =

±0.70 or higher, strong if r = ±0.40–0.69, moderate if

r ± 0.30–0.39, weak if r = ±0.20–0.29, and there was no

relationship if r = ±0.01 to 0.19. Our results are reported

as r (df) = correlation coefficient, P = P value.

Results

Timing of resource flows and environmental influences

Different forms of resource flows occurred at distinctly

different times. Insects emerged prior to amphibian meta-

morphosis from the ponds (Fig. 1) and litterfall input

occurred after emergence (day 263, 20 September).

Emergent insect export began before and lasted almost as

long as the export from amphibians. Insect emergence

abundance peaked first in Short pond (14–16 May) fol-

lowed by Intermediate 1 (19–21 July) and then

Intermediate 2 and 3 ponds (16–18 August). Insect emer-

gence increased over the four-month sampling period

(R2 = 0.1, F48 = 5.34, P = 0.025; Fig. 1) with higher

emergences just prior to pond drying. Amphibians had two

major waves of emergence that were responsible for 43 %

of the total abundance collected (23–25 June and 8–10

July) which corresponded to the drawdown in one pond and

the increase in temperature in another (Online Resource

Fig. 1). The greatest number of amphibians (n = 154)

leaving the ponds at one time occurred on 24 June with

most individuals leaving from Short pond (121 wood frog

metamorphs, Lithobates sylvaticus, and 1 newt, Notoph-

thalmus viridescens) and the remainder from Intermediate

1 (33 wood frog metamorphs). The number of metamor-

phosing amphibians declined over the sampling period

(R2 = 0.09, F98 = 9.331, P = 0.003; Fig. 1).

Litter fall amounts were similar across sampling months

(mean g m-2 ± SD; 430.60 ± 147.68 in Aug.–Sep.,

413.93 ± 169.50 in Sept.–Oct., and 474.05 ± 166.54 in

Oct.–Nov.). However, the timing of maximum litter fall

drop varied among ponds.

Magnitude of fluxes

We collected 1,050 individuals representing a total insect

flux (Eitotal) of 3418.85 g C year-1 from aquatic to terrestrial

habitat. This constituted a flux of 498.24 g C m-1 year-1 or

the emergent insect flux per meter of pond margin. Inter-

mediate 1 accounted for 95 % of the total insect flux with

the highest abundance of emerging insects (791 individuals

m-2), emergence production (Ei = 3.32 g C m-2) and flux

(Fi = 474.07 g C m-1 year-1) (Table 2). High flux was

driven by the presence of odonate taxa. In terms of abun-

dance, Chironomidae midges made up 68 % of collected

individuals, but contributed only 2.4 % to insect flux, Fi

(6.7 % of Eitotal) of all insect families (Fig. 2a). Lestid

damselflies were important contributors to the emergent

insect flux (83 %; Eitotal = 2840.63 g C year-1) even

though they represented only 0.004 % of collected individ-

uals (Fig. 2b). Limnephilidae caddisflies were collected

from all ponds and, although not abundant (n = 26), they

contributed a flux of 25.27 g C m-1 year-1 (Fig. 2a) or 5 %

of Fi and Eitotal (172.16 g C year-1). The number of

emerging insects from each pond was heterogeneous across

sampling events (Friedman rank sum test: v2 (3,

N = 12) = 12.897, P = 0.005). We found that the simi-

larity of insect total flux was the lowest for Short and

Intermediate 1 ponds (0.5 %), whereas the highest similarity

was between Short and Intermediate 2 ponds (31 %). Dis-

similarity between Intermediate 1 and Short ponds resulted

from differences in the flux from Chironomidae, Lestidae,

and Limnephilidae which made up 95 % of total flux

0

50

100

150

200

250

300

Day of year

Em

erge

nce

(no.

/day

)

amphibianinsect

124145

157172

174176

178180

189191

193200

202221

223228

230

Fig. 1 Temporal variation in insect and amphibian emergence

throughout the study period. Insect collection was from day 124 to

230 (May 4th–August 18th), amphibian collection occurred between

day 171 and 241 (June 8th–August 29th). Each point represents the

mean number of captures and the SD across ponds per collection day.

Overall, amphibian peak emergence occurred between days 174 and

176 (23–25 June), while insect peak emergence occurred between

days 200 and 202 (19–21 July)

Spatial subsidies between pond and forest habitats 135

123

leaving Intermediate 1 pond and only 0.5 % from Short. It

also is important to note Intermediate 1 had the highest

insect familial richness (25) and Short the lowest (9).

The total amphibian flux and flux per meter of pond

margin was nearly 6 9 higher than that from insects

(Eatotal = 18,833.9 g C year-1; Fa = 2899.33 g C m-1

year-1, Fig. 2c) and the emergence 18 9 higher (mean

Ea = 18.03 g C m-2) than insect emergence (mean

Ei = 1.00 g C m-2). The maximum number of amphibians

emerging from the four ponds ranged from 123 individuals

at Intermediate 3 pond to 571 individuals from Short pond

between late June and late August 2010 (total number

collected, n = 1197). A Friedman test was conducted to

evaluate differences in the number of individuals emerging

among the ponds across sampling events. The test was

significant, indicating that the number of emerging

amphibians varied among ponds and time steps (v2 (3,

N = 28) = 10.1774, P = 0.02), and signifying emergence

is heterogenous across ponds and time. The pond with the

shortest hydroperiod (78 days) had the densest (3 individ-

uals/m2) and highest emergence (Ea = 41.9 g C m-2

year-1) of amphibians. Emerging wood frog metamorphs

(L. sylvaticus) represented 71 % of all individuals collected

and 63 % of the amphibian flux (Fa) across all ponds to the

terrestrial landscape (Fig. 2c). Intermediate 1 and 2 were

most similar (49 %) in terms of amphibian flux and that

Intermediate 3 and Short ponds were the least similar

(1 %). These differences are because Intermediate 3 was

dominated by P. crucifer, whereas Short pond had very

little biomass from this species. The majority of flux from

Short ponds originated from L. sylvaticus, which was not

present in Intermediate 3, indicating species composition

played a major role in flux differences across ponds. Eatotal

significantly differed among species (ANCOVA: df = 7,

F = 3.66, P = 0.01). A post hoc holm adjustment showed

pairs of L. sylvaticus–A. maculatum (P = 0.035), L. syl-

vaticus–Ambystoma sp. (P = 0.016) and N. viridescens–L.

sylvaticus (P = 0.016) were significantly different from

each other in their contribution to the total flux of energy

from aquatic to terrestrial habitats.

Overall, there was a carbon deposition into ponds that

exceeded that exported via insect and amphibian emer-

gence (Fig. 3). The average net energy gain [Dtotal-

(Eatotal ? Eitotal)] across ponds was 41,279.76 g C year-1

(range 21,019.83–88,312.94 g C year-1). This subsidy

showed spatial variation across ponds (32.5 % difference

in Dtotal from lowest input to highest input), but resource

subsidy input did not differ across sampling events

(Friedman rank sum test: v2 (3, N = 12 = 0.2, P = 0.98).

Deciduous leaves were the most important litter resource

for all ponds (Fig. 4). Coniferous needles were of sec-

ondary importance to one pond (Short). As assumed,

litterfall deposition into ponds was directly related to

canopy cover over the pond (GLM, poisson: P \ 0.0001,

df = 3, Dtotal = 9.795 ? 0.0159mean percent cover).

Diversity–ecosystem function response

We found that insects and amphibians had different

diversity–ecosystem function relationships. Total insect

flux (Eitotal) and flux (Fi) per meter of pond margin were

highest from the least diverse pond (Fig. 5). However, total

insect flux increased with increasing taxa richness (Fig. 5).

The amphibian flux (Fa) per meter of pond margin did not

show a consistent pattern with higher flux amounts from

the least and most diverse ponds (Fig. 5). Amphibian total

flux (Eatotal) also failed to produce a pattern with amphibian

richness, though ponds varied little in amphibian richness.

We also evaluated the flux by diversity relationships using

family level resolution for amphibians because that is the

level of resolution we have for insects, and found the

relationships were the same regardless of using amphibian

species richness or family richness.

Ponds with longer hydroperiod tended to have higher

amphibian diversity (r (2) = 0.56, P = 0.44) and evenness

(r (2) = 0.68, P = 0.32). Amphibian species richness was

not linearly related to hydroperiod (r (2) = -0.11,

P = 0.89), but rather peaked at intermediate hydroperiod.

Measures of amphibian diversity showed strong positive

relationships with pond area (richness: r (2) = 0.90,

Table 2 Summary of flux

values for insects and

amphibians

Emergent flux per m pond

margin (F) takes into

consideration the shoreline

developmental factor (DL), total

flux is the total aquatic to

terrestrial flux of carbon (Etotal)a Eshore in Vander Zanden and

Gratton (2011)b Eareal in Vander Zanden and

Gratton (2011)

Short Intermediate 1 Intermediate 2 Intermediate 3

Insects

Emergent flux (Fi, g C m-1 year-1)a 1.23 474.07 3.58 19.35

Emergence (Ei, g C m-2)b 0.04 3.32 0.12 0.53

Total flux (Eitotal, g C year-1) 7.4 3281.2 30.2 100.04

Amphibians

Emergent flux (Fa, g C m-1 year-1)a 1273.26 806.01 420.11 399.94

Emergence (Ea, g C m-2)b 41.91 6.17 13 11.06

Total flux (Eatotal, g C year-1) 7655.35 5569.05 3541.57 2067.93

136 T. A. Schriever et al.

123

P = 0.10; diversity: r (2) = 0.88, P = 0.12; evenness:

r (2) = 0.81, P = 0.20). Insect family richness showed a

strong positive correlation with hydroperiod (r (2) = 0.87,

P = 0.13) and pond area (r (2) = 0.56, P = 0.44). How-

ever, insect diversity (r (2) = -0.62, P = 0.38) and

evenness (r (2) = -0.73, P = 0.27) declined from small to

larger ponds. Insect evenness declined from shorter to

longer hydroperiod ponds (r (2) = -0.49, P = 0.51), but

diversity was only weakly related to hydroperiod

(r (2) = 0.27, P = 0.73).

Influence of pond area and hydroperiod

The subsidy flows of amphibians and insects differed in

their relationships to pond hydroperiod and area. Hydro-

period and the total amphibian flux to terrestrial habitat

were strongly negatively correlated (Eatotal: r = -0.92).

The longer the hydroperiod the lower the magnitude of

amphibian export (Fig. 6a). Pond area had very little

relation to the total amphibian flux to the terrestrial eco-

system (r = 0.21). Conversely, total insect flux (Eitotal)

showed a very strong positive relationship with pond area

(r = 0.99; Fig. 6b), however the positive pattern with hy-

droperiod was less clear (log Eitotal; r = 0.44). Ponds very

small and very large received similar amounts of litter

input. The intermediate sized pond had the highest P/A

ratio, and therefore received more litter biomass per unit

area (Online Resource Fig. 3).

Discussion

Aquatic food webs are not isolated from the surrounding

terrestrial landscape, as they are coupled through multiple

flow paths from multiple trophic levels. This is especially

true for temporary woodland ponds. We have shown that

Coe

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Lest

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Libe

llulid

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100

200

300

400

Odonata family

B C

A. l

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A .m

acul

atum

Am

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oma

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B. a

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ican

us

H. v

ersi

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r

L. s

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ticus

N. v

iride

scen

s

P. c

ruci

fer

0

200

400

600

800

1000

1200

Amphibian speciesInsect family

0

5

10

15

20

25

30

Flu

x (g

C m

−1

yr−

1 )

Bae

tidae

Bra

coni

dae

Cae

nida

eC

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opog

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aeC

haob

orid

aeC

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Cic

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Hyd

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arid

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idae

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anid

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idae

A

Fig. 2 Flux (mean and standard error) attributed by each a insect

family and b Odonata family and c amphibian species collected.

Odonata are separated from other insect families because of

drastically different values compared to other taxa. No variation

exhibited in flux by some families because individuals were collected

from1 pond only

Fig. 3 Comparison of total carbon flux from pond-to-land (Etotal)

(upper portion of figure) and land-to-pond (Dtotal) (lower portion of

figure) attributed from each subsidy type across study ponds. The

inset zooms in on the contribution from insects

Spatial subsidies between pond and forest habitats 137

123

amphibians and insects contribute a considerable amount

of biomass to the terrestrial ecosystem, though litterfall

input was substantially higher than reciprocal flows to

adjacent terrestrial habitat. But these flows also depended

on the physical and biological aspects of the ponds them-

selves. We predicted that longer hydroperiods and larger

ecosystem size would have higher amphibian and insect

emergence and biomass than smaller ponds with shorter

hydroperiod. We did not find this to be the case for

amphibians. Instead, the smallest pond with the shortest

hydroperiod boasted substantially more emergent amphib-

ians and higher biomass export than the largest and longer

hydroperiod ponds. The flux of amphibians was *30 %

lower from longer hydroperiod ponds (168 days) compared

to the short hydroperiod pond (78 days). The shortest hy-

droperiod pond had very low predator richness and

abundance (both insect and salamander) and low

amphibian species richness, thereby providing an optimal

habitat for growth and survival that could account for the

high biomass. Emergent insects followed our predicted

pattern for pond area; larger ponds contributed more indi-

viduals and higher flux to the terrestrial landscape.

However, an intermediate hydroperiod pond had much

higher biomass and more insects emerging than longer

hydroperiod ponds. The ratio of perimeter to area explained

the importance of passive allochthonous input of litterfall

into woodland ponds.

It is generally accepted that amphibians and insects

account for a large proportion of energy flow within

aquatic and terrestrial food webs (Burton and Likens 1975;

Pough 1980). However, the transfer of energy through

amphibians between aquatic and terrestrial habitats has

largely been ignored (Ballinger and Lake 2006) and was

only recently quantified (Gibbons et al. 2006; Regester

Fig. 4 Variation in the magnitude of litter components received in

each pond. Litter components are: deciduous leaves, miscellaneous

(bud scales, pollen, etc.), aquatic plants (Equisetaceae; horsetail),

coniferous needles, and woody material (twigs, branches, bark, cones,

etc.). Note y-axis is not on the same scale for each pond

138 T. A. Schriever et al.

123

et al. 2006). Organisms of aquatic origin represent a spatial

subsidy to the terrestrial habitat in three ways: (1) as a prey

source for terrestrial vertebrate and insect predators; (2) as

a nitrogen and phosphorus source via excretion for soil and

plant uptake; and (3) as a release of nutrients, through

death, for decomposers, primary producers, and scaveng-

ers. Thus, spatial subsidies provided by emerging insects

and amphibians could influence recipient food webs by

directly and indirectly affecting predators, decomposers,

herbivores, and plants.

Few studies have simultaneously quantified multiple

types of resource flow from a single habitat, and here we

provide a unique picture of the energy exchange between

the aquatic and terrestrial ecosystem by quantifying the

magnitude of resource flows from obligate habitat couplers

(amphibians and emergent aquatic insects) and a passive

allochthonous subsidy (litterfall) input across a gradient of

environmental variation. Regester et al. (2006) provided

one of the first estimates of energy flow for amphibians

using ambystomatid salamanders from woodland ponds to

the surrounding forest. From one temporary and two per-

manent natural ponds, they collected 662 emerging

juvenile salamanders contributing 210 g ash-free dry mass

to the adjacent forest ecosystem (no anurans were reported

in their study). In comparison, we collected almost twice

the number of amphibians (n = 1197) emerging from four

ponds, but this represented half as much export (68 g ash-

free dry mass). In our study, cross-habitat energy flow was

dominated by anurans (64.5 %), which generally weigh

less on a per individual basis and thus contribute less

biomass than salamanders. Other studies like that of Gib-

bons et al. (2006) found anurans made up *95 % of

exported amphibian biomass from one large open canopy

pond in South Carolina, USA. Differences in amphibian

community composition can have a profound effect on the

amount of cross-habitat resource flow and striking conse-

quences for, and influences on, the terrestrial ecosystem.

This is another aspect of cross-habitat linkages that is

understudied.

In addition to hydroperiod, canopy cover also strongly

influences the performance and composition of freshwater

communities (Schiesari 2006). In general, closed canopy

ponds have lower water temperature and dissolved oxygen

(DO) than open canopy ponds (Skelly et al. 2002).

Increased canopy cover has been shown to decrease insect

abundance (Palik et al. 2001) and amphibian abundance

(Halverson et al. 2003; Binckley and Resetarits 2007) and

performance (Schiesari 2006). The ponds in our study with

higher canopy cover (Short and Intermediate 2) had lower

average water temperatures, lower salamander abundance

and total flux, and lower insect flux than open canopy

ponds (Intermediate 1 and 3; Table 2). Anurans on the

other hand showed the opposite pattern. We collected over

2.5 times more anuran metamorphs and total flux (Eatotal) in

closed canopy ponds. Anuran abundance and flux was

dominated by wood frog metamorphs. Rubbo et al. (2008)

found increasing leaf litter inputs into mesocosms

increased developmental rate and survival of larval wood

0.0 0 .5 1.0 1 .5 2.0

2000

3000

4000

5000

6000

7000

8000

Amphibian diversity (H)

Eto

tal

0 2 4 6 8 10

Amphibian species richness

0.0 0 .5 1.0 1 .5 2.0

1

10

100

1000

5000

0 5 10 15 20 25 30

Insect diversity (H) Insect richness

log

(Eto

tal)

A B

C D

Fig. 5 The relationship between a amphibian diversity, b amphibian

species richness, c insect diversity, and d insect family richness with

the magnitude of cross habitat energy flow (Etotal; g C year-1)

Fig. 6 The influence of

a hydroperiod and b pond size

on the total flux (g C year-1) of

three types of spatial subsidies.

Insect flux values were log

transformed for better

visualization of relationship

Spatial subsidies between pond and forest habitats 139

123

frogs. Our observed abundance and flux of wood frogs

from natural ponds was highest in the pond with higher

canopy cover and litter input. The few studies that have

investigated salamander performance in response to can-

opy cover have shown mixed results. In a mesocosm study,

Earl et al. (2011) found ambystomatid salamanders had

greater biomass export from high shade ponds compared to

low shade pond treatments. However, other studies have

found reduced growth and developmental rates in newts

from closed canopy ponds (Van Buskirk 2009, 2011) which

could translate to lower biomass. Our results are in contrast

to Earl et al. (2011), but follow Van Buskirk (2009, 2011)

results in that salamander abundance and flux to terrestrial

habitat was highest from more open canopy ponds (Inter-

mediate 1 and 3). Canopy cover, litter input, and pond

water chemistry influenced amphibian and insect biomass

in different ways. These results highlight the need for more

studies examining environmental influences on cross-hab-

itat energy flow.

Magnitude, rate, and functional role of spatial subsidies

can differentially impact the recipient food web (Wesner

2010) and influence the variation and strength of trophic

cascades (Leroux and Loreau 2008; Wesner 2010). Some

insects (odonates) and amphibians act as predators in the

terrestrial food web and can have a high consumptive affect

on arthropods as they disperse from the natal pond (McCoy

et al. 2009). Wesner (2010) found the majority of aquatic

insects emerging from small streams were non-consumers

as adults, thus contributing as a major prey subsidy to the

terrestrial food web. Ponds have emerging insects and

amphibians. Salamanders play vital roles in forest-floor

detrital food webs and nutrient cycling (Davic and Welsh

2004 and citations within) while anurans consume a wide

variety of arthropods on the forest floor (e.g. wood frog)

and in tree canopies (e.g. grey treefrog). The most abun-

dant emerging insect in our study was chironomid midges,

which are non-consumers as adults and therefore act as a

prey subsidy to terrestrial food webs. However, emergent

aquatic insects made up only a small proportion (on aver-

age 39 % compared to 89 % for amphibians) of the total

pond to land flux. This illustrates the trophic structure of

the flux is dominated by consumer biomass. Intermittent

ponds, therefore, have a dual role as providers of prey

subsidies to terrestrial predators and a seasonal contributor

of intermediate predators to terrestrial food webs, making a

case that these aquatic habitats form multiple connections

within discrete food webs.

Studies on the flows of invertebrate subsidies to adjacent

terrestrial habitats have been strongly stream focused

(Jackson and Fisher 1986; Nakano and Murakami 2001;

Sabo and Power 2002; Kato et al. 2003; Baxter et al. 2005;

Ballinger and Lake 2006; Whiting et al. 2011) and some-

times limited to one or two insect families (Fisher and

Likens 1973; Alvarez and Pardo 2005; Runck 2007). There

are surprisingly few studies that have measured the entire

insect assemblage production from wetlands. We collected

individuals representing 36 insect families. Insect flux was

two times higher in our study ponds compared to that found

by Whiles and Goldowitz (2001) from open canopy wet-

lands with and without fish in Nebraska. Gratton and

Vander Zanden (2009) estimated insect emergence rates

across a range of ecosystem size and habitat types. Of the

three ponds used in Gratton and Vander Zanden (2009), the

largest pond (800 m2) had comparable emergence pro-

duction rates to a similar size pond (987.48 m2) in our

study (our empirical value 3.32 vs. theirs 3.29 g C m-2

year-1). Our estimates of emergence numbers and magni-

tude of flux are likely conservative given that insect

trapping started in May and ice off was in March, odonates

are not trapped with the same efficiency as other aquatic

insects, and treefrogs can climb drift fences avoiding

capture (Gibbons et al. 2006); therefore we are most likely

underestimating flux.

Terrestrial primary production in the form of leaf litter is

a significant resource in woodland ponds forming the base

of the aquatic food web (Rubbo et al. 2006; Batzer and

Palik 2007). Litterfall is a donor controlled seasonally

recurring resource subsidy to woodland ponds. We pre-

dicted the amount of litter deposition would be

proportional to the amount of animal export. However, the

highest input of litter was deposited in a pond with high

P:A and resulted in fairly low export via amphibian and

insect emergence. We measured almost 8.5 times more

litter input (Dtotal = 187,371.8 g C year-1) than the com-

bined total export flux of amphibian and insect resources

(Etotal = 22,252.75 g C year-1) from four ponds, resulting

in an overall net carbon input. These results are consistent

with Vander Zanden and Gratton’s (2011) modeled results

of lake-to-land flux (Dtotal) that exceed insect emergence

(Etotal) in small lakes (\2,000 ha). This net carbon input

may be common to most lakes and streams (Jackson and

Fisher 1986; Leroux and Loreau 2008).

A positive diversity and ecosystem function response is

typically found (Balvanera et al. 2006), especially for pri-

mary producers from terrestrial ecosystems (Schlapfer and

Schmid 1999; Bouchard et al. 2007), but more complex

responses are possible when studying multiple trophic

levels of consumer groups (Duffy 2002; Hooper et al.

2005). We found that communities with higher insect or

amphibian diversity result in lower flux to the terrestrial

ecosystem (Etotal and emergent flux; Fig. 5). We studied

four ponds, and therefore are limited in our ability to

generalize our results to pond ecosystems as a whole. That

said, we found an interesting pattern exhibited by two

different taxonomic groups, which indicates further

research on a multiple consumer groups and pond systems

140 T. A. Schriever et al.

123

is warranted. The amphibian export flux was dominated by

one species, wood frog, from one small, intermittent

woodland pond with low insect family and amphibian

species richness. This result is contrary to the sampling

effect model that states there is a higher chance of sam-

pling a productive species in a species rich assemblage

(Tilman et al. 1997; Aarssen 1997), but is consistent with

the dominance effect (Norberg 2004). It is important to

note that we looked across multiple trophic levels, whereby

most experiments look at single trophic levels. Wood frogs

had higher biomass than the cumulative biomass of a more

diverse amphibian community in an intermediate hydro-

period pond; perhaps because its development is best suited

to temporary ponds (Paton and Crouch 2002). We specu-

late that predation, presence of salamander larvae, and

insect predators, as well as competition from other larval

anurans decreased wood frog productivity in other ponds.

Mechanisms controlling community structure change along

the hydroperiod gradient (i.e., abiotic forces in short hy-

droperiod ponds versus strong biotic factors in longer

hydroperiod ponds) and thus can have strong and possibly

different, influences on the biodiversity–ecosystem func-

tion relationship. Further, perhaps the positive relationship

between biodiversity and ecosystem function may not be a

general pattern for aquatic vertebrates (Cardinale et al.

2006).

Conclusions

Intermittent ponds had a tight coupling with the terrestrial

landscape in two reciprocal ways: ponds are heterotrophic

depending on allochthonous carbon inputs to sustain the

aquatic food web and aquatic emergences of amphibians

and insects become part of the terrestrial food web as

consumers (e.g., amphibian juveniles) and as prey subsidy

(e.g. chironomids) to terrestrial organisms. Cross-habitat

resource flows were variable in time, space, and taxonomic

group. Environmental variation in hydroperiod had a strong

influence on amphibian flux to the adjacent terrestrial

habitat, but less influential to insect flux. Diversity was

higher in ponds with lower productivity, thus providing a

‘non-experimental’ view of the BEF relationship. The

magnitude of spatial subsidies in this system is a complex

product of interactions among pond hydroperiod, size,

perimeter-to-area relationships, and species diversity.

Therefore, landscapes comprised of ponds varying in hy-

droperiod are essential to conserving insect and amphibian

diversity and to maximize aquatic-terrestrial linkages.

Acknowledgments Construction of the drift fences and setting

pitfall buckets could not have been completed without the help of:

Katherine Bannar-Martin, Devin Bloom, Kristen Brochu, Kirsten

Comberford, Maria Modanu, Stephen Pynn, David Stitt, and Caroline

Tucker. We also thank Mark Conboy, Klara Jaspers-Fayer, Tristan

Willis and Monica Candelaria for helping with field work. The tre-

mendous dedication of Siao Ryan Yang and Ruby Sambi in

processing samples is greatly appreciated. We thank Karen Pope for

generously donating the emergence traps and Nathan Lovejoy for

microscope and laboratory use. This research was supported by a

Natural Sciences and Engineering Research Council of Canada Dis-

covery Grant awarded to DDW.

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