Aquatic Sciences

, Volume 74, Issue 3, pp 397–404 | Cite as

Species loss in the brown world: are heterotrophic systems inherently stable?

  • Michael J. Rubbo
  • Lisa K. Belden
  • Sara I. Storrs-Mendez
  • Jonathan J. Cole
  • Joseph M. Kiesecker
Research Article


Determining the effects of species loss on ecosystems has received considerable attention given the current threats many ecosystems are facing. A significant body of research has yielded many insights to this question, but this work has been limited by a focus on ecosystems where primary production plays a significant role in energy transfer. As many ecosystems rely on energy sources that are not derived from in situ production, there is a need to better understand how species loss will affect ecosystems of varying trophic states. To examine the effects of species loss on an ecosystem that is not reliant on in situ primary production, we manipulated the larval amphibian community of temporary forest ponds. These ponds are heterotrophic systems that rely on allochthonous inputs of detritus as a basal energy source. The larvae of two amphibian species that are prone to local extinction, wood frogs (Lithobates sylvatica) and spotted salamanders (Ambystoma maculatum), were removed from ponds and net ecosystem production was monitored. We found no effects of the removal of these top consumers on ecosystem functioning or on lower trophic groups (i.e., zooplankton, algae, bacteria). While amphibians can influence food web dynamics in other systems, their influence on system processes in temporary forest ponds appears to be limited. We hypothesize that the lack of any effects is due to the microbial degradation of detritus “swamping” the system, providing more than enough energy to maintain the food web, and/or due to omnivory dampening species interactions. These data indicate that the functioning of heterotrophic systems may be inherently stable due to internal dynamics that minimize interaction strengths among trophic groups.


Heterotrophic Biodiversity Ecosystem function Detritus Amphibian 


As factors such as urbanization and climate change continue to threaten biodiversity, there is a pressing need to better understand how the loss of species, specifically those sensitive to alterations in the environment, will affect the services and functions ecosystems provide (Thebault and Loreau 2003; Cardinale et al. 2006). Typically, research investigating the relationship between community composition and ecosystem functioning operates by removing species and then monitoring ecosystem responses such as primary productivity (Duffy 2003; Balvanera et al. 2006). Primary production has been a focal ecosystem response because the fixation of CO2 via photosynthesis is generally considered to be indicative of the total amount of carbon available to higher trophic levels. This work has led to the development of a rich field of literature; however, as many ecosystems are not solely reliant on primary production (Hairston and Hairston 1993; Polis and Strong 1996), a more holistic examination of the responses of ecosystems to the loss of species is necessary (Moore et al. 2004; Reiss et al. 2009). While primary production has historically been viewed as the dominant pathway of energy flow in many ecosystems (Hairston et al. 1960), the importance of detritus in structuring food webs has also been recognized for quite some time (Lindeman 1942; Hairston and Hairston 1993; Wetzel 1995; Polis and Strong 1996). Food webs that rely on detritus are quite common and this energy source is a significant metabolic pathway in many ecosystems such as headwater streams (Wallace et al. 1997), temporary forest ponds (Rubbo et al. 2006a), shallow lakes (Jansson et al. 2007), and terrestrial soil communities (de Ruiter et al. 1995). Detritus can play a key role in ecosystems by stabilizing food webs, altering habitat complexity, and influencing the flow and transfer of energy and nutrients (Wetzel 1995; Moore et al. 2004; Shurin et al. 2006).

In many ecosystems, detritus is commonly subsidized by allochthonous inputs of dissolved or particulate C which allows secondary production to exceed levels that would otherwise be limited by primary production (Polis et al. 1997; Pace et al. 2004). In these systems, the microbial catabolism of detritus drives secondary production and is a major pathway of energy flow (Wetzel 1995; Pace et al. 2004). Microbes (bacteria and fungi) either respire the C provided in detritus or transfer it to higher trophic levels (Tranvik 1992; Hall and Meyer 1998); however, either pathway will result in enhanced levels of community respiration.

The trophic state of respiration exceeding primary production is referred to as heterotrophy (Dodds and Cole 2007) and is quite common among aquatic communities (Cole et al. 2000; Duarte and Prairie 2005). Despite its ubiquity, little is known of the factors that control heterotrophic systems (Cole et al. 2000; Dodds and Cole 2007). These systems may be resistant to the ecosystem impacts of species loss due to compensatory mechanisms such as the high turnover rates of microbes (Pace et al. 1999), but we have yet to develop a comprehensive understanding of the manners in which systems based on secondary production will respond to alterations in their food web structure.

To explore the influence of species loss on the functioning of a heterotrophic system, we manipulated the food web of temporary forest ponds. These ponds are heterotrophic systems that rely on allochthonous inputs of leaf-litter as a primary energy source (Rubbo et al. 2006a). Temporary ponds are an ideal system in which to study the effects of species loss on ecosystem functioning as the top consumers in these ponds in the northeastern US, larval wood frogs (Lithobates sylvatica) and larval spotted salamanders (Ambystoma maculatum), are prone to local extinction (Rubbo and Kiesecker 2005; Harper et al. 2008). Therefore, the simulated loss of these species from the food web represents a deterministic pattern of species loss that occurs under natural conditions (sensu, Duffy 2003).

We removed either wood frogs or spotted salamanders or both of these species from natural ponds and monitored the ecosystem response of net ecosystem production (NEP). As integrating multiple pathways of energy flow is a necessary approach to determining the influence of species loss on ecosystems (Reiss et al. 2009), NEP is an ideal measure because it incorporates both gross primary production (GPP) and respiration (R). NEP describes the metabolic state of the entire ecosystem and is defined as the difference between GPP and the sum of autotrophic (RA) and heterotrophic respiration (RH). Positive values of NEP indicate that the system is net autotrophic while negative values indicate that the system is net heterotrophic.


Study sites

We used 12 temporary woodland ponds in the State Game Lands 176, Centre County, Pennsylvania, USA in this experiment. The maximum distance between ponds was 6 km and the minimum 0.05 km. Ponds were chosen based on the overall similarity in both abiotic characteristics and amphibian breeding effort. The ponds used in this study all lacked vegetation in their basins, had similar hydroperiods and canopy coverage (~80%), and ranged in size from 25 to 412 m2 and in maximum depth from 0.2 to 1.1 m. Spotted salamanders and wood frogs breed in all of the ponds used in this experiment. These species have complex life cycles, whereby adult and juvenile stages are terrestrial and egg and larval stages are aquatic. Jefferson’s salamanders (A. jeffersonianum) also breed in these ponds, but as their breeding effort is rather heterogeneous, we excluded this species from this study. These ponds typically fill in late winter/early spring (i.e., February–March) and dry by late summer (i.e., June–July).

Food web structure

The food web of temporary forest ponds is quite complex containing a diverse community of organisms. Higher consumers in these ponds can include invertebrates such as Aeshnidae and Libelluidae (Odonata), Notonecta (Heteroptera) and Dytiscidae (Coleoptera: Bonner et al. 1997; Batzer et al. 2004) and vertebrates such as larval salamanders and anurans (Wilbur 1997). While a number of amphibian species may breed in these ponds, spotted salamanders and wood frogs are the most abundant amphibians locally; thus, we focused on these species. The role of macroinvertebrates as consumers in these ponds is not clear. Prior data in this system indicate that macroinvertebrates are more abundant in open-canopy ponds than in the forested ponds (Rubbo et al. 2006b), suggesting that their role in the food web of closed-canopy may be less significant than it is in more open sites. Data on macro-invertebrate biomass in a similar system (Leeper and Taylor 1998) combined with our data on amphibian biomass (Table 1) suggest that larval salamanders and tadpoles are the top consumers in this system. These patterns are consistent with our observations in local ponds as well as with prior studies on this food web (Bonner et al. 1997; Wilbur 1997; Petranka 1998). Larval spotted salamanders feed upon a variety of invertebrates including zooplankton, chironomids, odonates, trichopterans, and dipterans (Petranka 1998). Lower consumers in these ponds consist of invertebrates such as Chironomidae and Culicidae (Diptera) and Limnephilidae (Trichoptera). Zooplankton (Cladocera, Copepoda, Ostracoda) make up a large proportion of the primary consumer pool and are a major food resource for larval salamanders (Bonner et al. 1997). Wood frog tadpoles are also a significant lower consumer in these ponds, grazing upon algae and detritus (Skelly and Golon 2003). Basal energy sources are derived from primary producers (algae) and microbes (bacteria and fungi).
Table 1

Biomass and generalized trophic status of selected organisms from temporary ponds


Biomass (mg/m2)




Primary consumer (detritivore)



Primary consumer (omnivore)



Higher consumer (predator)



Higher consumer (predator)



Primary consumer (omnivore)



Higher consumer (predator)



Secondary consumer (predator)



Primary consumer (omnivore)



Primary consumer (omnivore)



Primary consumer (omnivore)

Spotted salamander


Higher consumer (predator)

Wood frog


Primary consumer (omnivore)

aLeeper and Taylor (1998)

bRubbo, M.J., unpublished data

As larval spotted salamanders and wood frogs are species susceptible to local extinction and as they constitute large components, in terms of both numbers and biomass, of the consumer pool (Table 1), we chose to manipulate these species to determine if the loss of consumers had any effects on ecosystem functioning in this system.

Species manipulation

On April 13, 2003 we removed all amphibian egg masses from the 12 ponds. We chose this date because no new breeding activity (presence of adults or new egg masses) was observed in the ponds for 7 days prior. We pooled all of the egg masses that we removed from the 12 ponds and haphazardly selected the appropriate number of egg masses to add back to each pond based on our four treatments: no amphibians (n = 3); wood frogs only (n = 3); spotted salamanders only (n = 3); and wood frogs and spotted salamanders (n = 3). Treatments were randomly assigned to ponds. Based on prior data from ponds at this location (Rubbo et al. 2006b), we used densities of 0.11 egg masses/m2 for spotted salamanders and densities of 0.03 egg masses/m2 for wood frogs. Egg masses were placed in the ponds at depths similar to those from which they were collected. To ensure that our treatment designations accurately reflected the larval amphibian community of these ponds, we sampled the ponds on June 19, 2003 for larval amphibians using dip nets. The number of dip net sweeps in each pond was based on pond size and ranged from 10 to 30, 1-m sweeps. No wood frogs tadpoles or spotted salamander larvae were found in the amphibian removal ponds, while both species were found in the control ponds. We only found wood frogs in the salamander removal treatment and spotted salamander larvae in the wood frog removal treatment. Therefore, we are confident that our manipulations were successful. As one of the amphibian removal ponds (i.e., no wood frogs or spotted salamanders) dried before any embryos hatched, we removed this pond form the analysis leaving two replicates for that treatment.

Net ecosystem production

We used the continuous diel oxygen method (Cole et al. 2000; Rubbo et al. 2006a) to quantify NEP in these ponds. In brief, we deployed YSI sondes (model 600 XLM) in the upper mixed layer of these ponds. The sondes were deployed in the ponds for 24 h. We began placing sondes in the ponds on a weekly basis beginning April 16, 2003 and ending June 27, 2003. Each day, two sondes would be placed into two ponds and the next day the sondes would be moved to two other ponds. The order in which sondes were placed into the ponds was held constant each week. The sondes recorded dissolved oxygen (DO) every 15 min over the entire 24 h period. We then used the equations given by Cole et al. (2000) to determine NEP for each 15 min sampling interval which was then averaged for the entire 24 h period to determine the daily value for NEP. GPP and community respiration were then derived from the NEP values, oxygen saturation levels, temperature, and estimates of gas diffusion rates (Cole et al. 2000).

Community composition

In addition to ecosystem functioning data, we also collected some food web data to help elucidate potential mechanisms for any responses of ecosystem functioning to the manipulation of the top consumers. However, as these data were meant only to be supportive of the larger ecosystem data set, their interpretation should be taken with caution. We sampled the ponds for chlorophyll a (Chl a, an estimate of phytoplankton biomass) and zooplankton on April 28, May 13, and May 28, 2003. Chl a was estimated by filtering surface water through GF/F filters. The filters were then covered with basic methanol for 24 h and we determined the Chl a content of the sample, after correction for phaeopigments, using a fluorometer. We sampled the macro-zooplankton of the ponds by collecting integrated water column samples using a tube (4.5 cm diameter PVC) sampler fitted with a check valve (Rubbo and Kiesecker 2004). We collected 15 L of water and filtered it through 75 μm Nytex mesh. Zooplankton were then preserved with 1% Lugol’s solution. We identified zooplankton to the major taxonomic groups of cladocera and cyclopoid, calanoid, and harpacticoid copepods.

To obtain an estimate of bacterial production, we used the micro-centrifugation method (Smith and Azam 1992). This method measures the uptake of 3H-leucine by bacteria in a given time period (i.e., 45 min). Uptake experiments determined that 25 nmol/L maximized uptake rates in this system (Rubbo, unpublished data). We analyzed three live replicates per sample and one killed control. We collected samples for bacterial production on April 28, May 13, and May 28, 2003.

Statistical analyses

To determine if the manipulation of the consumers in these ponds affected ecosystem functioning, we used repeated measures analysis of variance (rANOVA) to determine if treatment influenced NEP, R, and GPP. Community attributes were given by Chl a, zooplankton, and bacterial production and were also analyzed using rANOVA. Zooplankton, both total and individual groups, and Chl a data were natural log transformed to meet the assumptions of parametric statistics.


Ecosystem functioning

Net ecosystem production was not affected by the loss of top consumers from this ecosystem (F(3,2) = 0.80, P = 0.60). Sampling date had no effect on NEP (F(10,20) = 1.36, P = 0.27), and there was no interaction between NEP and sampling date (F(30,20) = 0.72, P = 0.79). All ponds were consistently net heterotrophic and displayed no patterns in response to the manipulation of top consumers (Fig. 1). The manipulation of the amphibian community also did not affect R (Fig. 1: F(3,2) = 1.69, P = 0.39), which was not affected by sampling date (F(10,20) = 1.07, P = 0.43), and there was no interaction between R and sampling date (F(30,20) = 0.64, P = 0.87). GPP did not differ between treatments (Fig. 1: F(3,2) = 0.96, P = 0.55). Sampling date had no effect on GPP (F(10,20) = 0.73, P = 0.69) and there was no interaction between GPP and sampling date (F(30,20) = 0.59, P = 0.90).
Fig. 1

Summary of mean (±SE) values of a net ecosystem production; b gross primary production, and c respiration per experimental treatment

Community attributes

The loss of amphibians from these ponds also did not affect the lower trophic groups (Table 2). Chl a levels did not differ between treatments (F(3,6) = 0.45, P = 0.73) or sampling date (F(2,12) = 2.68, P = 0.11), and there was no interaction between treatment and date (F(6,12) = 0.49, P = 0.80). Bacterial production also did not differ among treatments (F(3,6) = 0.30, P = 0.82). Production by bacteria did vary among sampling dates (F(2,12) = 4.53, P = 0.03) with the lowest values observed on May 14. There was no interaction between sampling date and treatment (F(6,12) = 0.76, P = 0.61).
Table 2

Summary of mean (±SE) values of total zooplankton, chlorophyll a (Chl a), and bacterial production (BP)



April 28

May 13

May 28

Chl a (μg L−1)

No amphibians

1.49 (0.31)

0.41 (0.02)

1.15 (0.09)

Wood frogs

2.02 (0.99)

3.85 (3.46)

4.44 (3.30)

Spotted salamanders

2.08 (0.79)

1.49 (0.49)

2.85 (0.21)

Frogs and salamanders

6.52 (5.66)

2.52 (2.19)

6.39 (1.56)

Zooplankton (L−1)

No amphibians

12.95 (0.95)

39.33 (4.53)

17.73 (9.47)

Wood frogs

19.40 (17.20)

43.70 (27.10)

5.20 (2.77)

Spotted salamanders

20.00 (12.50)

20.32 (6.57)

26.40 (21.10)

Frogs and salamanders

24.80 (9.26)

19.18 (9.59)

20.78 (6.41)

BP (μgC L−1 day−1)

No amphibians

137.40 (26.00)

104.60 (51.90)

134.90 (42.40)

Wood frogs

88.70 (14.40)

82.90 (43.20)

119.00 (17.30)

Spotted salamanders

132.60 (35.70)

81.70 (28.70)

171.80 (37.90)

Frogs and salamanders

122.80 (38.40)

95.70 (26.90)

158.20 (42.20)

The total number of zooplankton in the ponds also did not differ among treatments (F(3,7) = 0.26, P = 0.85). The individual groups of cladocera (F(3,7) = 0.24, P = 0.87), cyclopoid copepods (F(3,7) = 0.86, P = 0.51), calanoid copepods (F(3,7) = 0.88, P = 0.50), and harpacticoid copepods (F(3,7) = 0.07, P = 0.98) were not affected by the experimental treatments. None of these groups exhibited a response to sampling date or showed an interaction between date and treatment (P > 0.05).


The removal of top consumers from this ecosystem did not appear to have any effects on the functioning of this system or on lower trophic levels. NEP, GPP, and R all were unaffected by the removal of spotted salamander larvae and wood frog tadpoles (Fig. 1). All ponds were net heterotrophic, consistent with prior observations in this system (Rubbo et al. 2006a), but removing species did not alter the manner in which this system processed energy. In addition to the lack of an ecosystem response, we observed no responses by other trophic groups to the manipulation of the larval amphibian community. Zooplankton, Chl a, and bacterial production were all unaffected by the removal of these consumers (Table 2). While there may have been compositional shifts in these groups or compensatory dynamics among other members of the food web (e.g., macro-invertebrates) that we did not detect due to our sampling protocol, these factors did not affect overall ecosystem functioning. These data suggest that larval amphibians in this ecosystem do not exert any top-down control and that ecosystem functioning is controlled from bottom-up pathways.

While larval amphibians have been shown to affect community dynamics in a variety of systems (Wilbur 1997; Chase 2003), less is known of how these organisms influence ecosystem processes. A thorough series of studies on the loss of larval anurans from detritus-based neotropical streams, found that the loss of these organisms resulted in compositional shifts in primary producers, potentially affecting primary production (Whiles et al. 2006; Colon-Gaud et al. 2009). However, it is not clear if the loss of tadpoles affected system respiration or microbial production. Other studies have also found that larval amphibians can influence primary producers (Seale 1980; Harris 1995; Kupferberg 1997; Downing 2005), but their influence on secondary production was not quantified. Moreover, as the food resources exploited by tadpoles can vary based on the species in question or the physical characteristics of a specific pond (Whiles et al. 2010), more work is necessary to develop a better understanding of the species-specific traits and system properties that dictate when the loss of amphibians will affect ecosystem functioning. However, it is important to note that despite the lack of an affect of these amphibians on pond functioning, they do play important roles in energy flux from ponds to the terrestrial environment (Regester et al. 2006).

The lack of an ecosystem response to the loss of larval wood frogs and spotted salamanders may be due to the manner in which energy flows through this system. As respiration exceeds primary production in these ponds, energy must be supplied to support the food web. This subsidy is provided by allochthonous inputs of detritus (Rubbo et al. 2006a) which can stabilize the food web by providing a consistent source of energy to higher trophic levels (Wetzel 1995; Moore et al. 2004). Microbes (bacteria and fungi) serve as the initial consumers of detritus and transfer this energy to higher trophic level through a variety of pathways (Jansson et al. 2007). These microbes have high turnover rates which may buffer them from any top-down effects as they can compensate for grazing by increasing production (Pace and Cole 1994; Mikola and Setala 1998). Therefore, the loss of species at the top of the food web may have less of an impact on food webs that rely on the microbially-mediated conversion of C than on food webs that rely on pathways based on primary production.

In addition to the potential dampening of any effects by microbes, omnivory may further stabilize this food web. Obtaining energy from a variety of trophic levels buffers ecosystems from species losses by decreasing interaction strengths among species. (Moore et al. 2004; Petchey et al. 2004). Larval wood frogs (Skelly and Golon 2003) and zooplankton (Jansson et al. 2007) are both capable of using a variety of food resources; thus, they do not always obtain their resources from the trophic level directly below. As omnivory in this ecosystem appears common throughout the food web, this suggests that interactions among trophic groups may be weak.

An alternative explanation for the lack of any observed responses may be because of a lack of statistical power in the experiment. To evaluate this possibility we conducted a power test. As we expected significant variation in the data due to the system-level scale of the experiment we utilized an alpha-level of 0.1 for this test. We found that our design achieved a power value of 0.5. While this value is below the generally accepted value of 0.8 for power tests, we are confident that if any biologically-meaningful trends were present, they would have been detected. Moreover, we have used this identical experimental design for other ecosystem experiments (Rubbo et al. 2006a) and have detected strong experiment effects. However, additional work is warranted to further evaluate the lack of an ecosystem response to the loss of amphibians and to elucidate the roles that these organisms play in the flow of energy through this system.

The scale at which this study took place also likely played a role in determining the outcome of this work. The majority of experimental research studying the effects of species loss from ecosystems has taken place in highly controlled situations such as in the laboratory, microcosms, or mesocosms. These environments tend to over-simplify natural systems and underestimate the complex emergent effects of species on ecosystems. The venue of experimentation is critically important to this type of research as it can influence interaction strengths among species (Skelly 2002). For example, Holomuzki et al. (1994) found that while larval salamanders exerted strong top-down control in enclosures these effects were not apparent in surveyed ponds. The disappearance of species effects in accordance with the increased realism of experimental venue is frequently due to stabilizing mechanisms inherent in more complex systems (Romanuk et al. 2009). This highlights the need for more experimental work at the scale of whole-systems. While ecosystem experiments are logistically difficult and can suffer from high variation and low replication, they are necessary to evaluate ecological phenomenon under the most natural conditions.

This study suggests that the heterotrophic food web of temporary forest ponds may be inherently stable due to donor-controlled dynamics at the base of the food web. Bottom-up effects have been observed previously in this system and are mediated by the amount of allochthonous C available (Rubbo et al. 2006a). It appears that microbial production “swamps” the system providing more than enough energy to support the food web. Understanding the manners in which the loss of species will affect entire ecosystems is of crucial importance given the threats that many ecosystems are currently facing. Recent research has begun to disentangle the effects of species on ecosystem functioning in detritus-based food webs (reviewed in Lecerf and Richardson 2009; Dudgeon 2010; Gessner et al. 2010), but the majority of this work has taken place in stream ecosystems. Moreover, research in field settings is needed to accurately depict the effects of species loss on ecosystem functions (Dudgeon 2010; Gessner et al. 2010). Studies that incorporate greater realism, species vulnerable to extinction, as well as those that quantify a variety of ecosystem functions are necessary to further develop this rich field. We feel that this study is an initial step in this direction as it represents a realistic pattern of species loss in an ecosystem-level manipulation. Therefore, these data have value in that they can add to our understanding of how whole systems respond to species loss under natural conditions.



We thank J. Chase for reviewing an earlier draft of this manuscript and J. Falkenbach and L. Grove for assistance with field work. Financial support was provided by the NIH/NSF Ecology of Infectious Disease Program (1R01ES11067-01 to JMK) and NSF (IBN) Grant #0131229 to JMK, and the Department of Biology, Pennsylvania State University.


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Copyright information

© Springer Basel AG 2011

Authors and Affiliations

  • Michael J. Rubbo
    • 1
  • Lisa K. Belden
    • 2
  • Sara I. Storrs-Mendez
    • 3
  • Jonathan J. Cole
    • 4
  • Joseph M. Kiesecker
    • 5
  1. 1.Teatown Lake ReservationOssiningUSA
  2. 2.Department of Biological SciencesVirginia Polytechnic Institute and State UniversityBlacksburgUSA
  3. 3.Division of Biological SciencesUniversity of MissouriColumbiaUSA
  4. 4.Cary Institute of Ecosystem StudiesMillbrookUSA
  5. 5.The Nature ConservancyFort CollinsUSA

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