Snake parasitism in an urban old-growth forest
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Urban-associated changes can have immediate or long-term consequences on animal populations. Such changes may be assessed through parasite prevalence and abundance in wildlife hosts, as urbanization can influence parasitism and disease transmission in wildlife. Snakes are widespread and diverse vertebrates that often persist in urban environments; however, parasitism of snakes in urban environments has yet to be studied, leaving the roles of snakes in parasite transmission uncharacterized. Field ecology, microscopy, molecular techniques, and geographic information science (GIS) were integrated to characterize parasitism of snakes in an urban old-growth forest park. The species, sex, mass, length, location, and prevalence of ecto-, hemo-, and fecal parasites were determined for 34 snakes of 6 species. Ectoparasites (mite), hemoparasites (Hepatozoon spp.), and fecal parasites (Entamoeba spp., Trichomonas spp., Strongloides spp., and an unidentified helminth) were detected in snakes and 64.7 % of snakes were infected by at least one of these parasites. Parasite infections were generally not related to the sex, age, or body condition of snakes. The locations of infected snakes were used to produce risk maps indicating where parasite prevalence is predicted to be greatest. The analysis of these maps indicated that snakes with fecal parasites were closer than non-infected snakes to the edge of the forest. This study confirms that snakes may be important parasite hosts or reservoirs in parasite transmission pathways in urban environments and it provides an integrative multidisciplinary approach that may be used to monitor parasitism dynamics in other urban wildlife areas.
KeywordsAgkistrodon Ectoparasite Fecal Forest fragment Hemoparasite Hepatozoon
Currently more than 50 % of the global human population lives in urban areas and urbanization is expected to increase as the global population climbs towards 9 billion by 2050 (United Nations Population Division 2010). As urbanization intensifies, habitat modifications and fragmentation will continue and as a result, local floral and faunal communities may experience decreased species richness (McKinney 2008; Walker et al. 2009), shifts in trophic dynamics due to reduced top-down controls and increased productivity (Faeth et al. 2005), or behavioral modifications (Riley et al. 2003; Ditchkoff et al. 2006), among other things. Mechanisms underlying such responses to urbanization have recently been examined (Shochat et al. 2006), yet effects of urbanization on many animal species remain poorly understood because the majority of published studies have examined conspicuous models, including birds (Chace and Walsh 2006; Slabbekoorn and Ripmeester 2008), relatively large mammals (LoGiudice et al. 2003; Riley et al. 2003; Randa and Yunger 2006; Page et al. 2008), and amphibians (Riley et al. 2005; Rubbo and Kiesecker 2005; Parris 2006). These studies have revealed tremendous interspecific variation in responses to urbanization. For example, urbanization may influence vocalizations in some birds (Leonard and Horn 2008; Slabbekoorn and Ripmeester 2008; Mendes et al. 2011), the timing of nesting in others (Schoech and Bowman 2001), and cause local extinction in others (Blair 2001; Marzluff 2001; McKinney 2008). Thus, comparative studies on a wider variety of species are needed.
Parasitism is a successful and well-known life strategy with numerous ecological and physiological impacts (Marcogliese 2005; Preston and Johnson 2010). Urbanization likely has ecologically relevant effects on host-parasite interactions because it alters the composition and distribution of suitable wildlife habitat, which can constrain species’ distributions and alter transmission rates of wildlife parasites (Bradley and Altizer 2007). Moreover, urbanization results in loss of biodiversity (McKinney 2008; Walker et al. 2009) and decreased biodiversity frequently increases transmission of animal and plant infectious diseases (Keesing et al. 2010). Finally, efforts to control arthropod vectors (e.g. mosquitos) in urban areas may further disrupt transmission of parasites among wildlife hosts. Parasites can be used as bio-indicators of environmental impact, including urbanization (Vidal-Martínez et al. 2010), yet a notable gap exists in our understanding of ecological and evolutionary significance of host-parasite interactions for most species (Deviche et al. 2001; Vitt and Caldwell 2009). Thus, research is needed to elucidate these relationships, particularly in urban environments.
Snakes are a diverse group of ectothermic predatory vertebrates that possess several characteristics which may make them important in urban parasite transmission pathways. Notably, snakes are known to host a variety of parasites (Brown et al. 2006; Telford 2009) and often persist and even thrive in urban environments, including urban parks (Shine and Koenig 2001; Burger et al. 2004; Pattishall and Cundall 2009; Smith et al. 2009; Vignoli et al. 2009). Moreover, snakes vary tremendously in size, fill numerous ecological niches, occupy a variety of trophic levels in ecosystems, and tend to live longer than most endothermic hosts (Greene 1997). Therefore snakes provide an excellent opportunity to study parasitism in an urban environment.
A multidisciplinary approach that combined field ecology, microbiology, molecular biology, and geographic information science (GIS) was used to provide an integrative assessment of snake parasitism in an old-growth urban forest in Memphis, Tennessee. The specific objectives of this study were to: 1) identify the community compositions of snakes and the ecto-, hemo-, and fecal parasites they host, 2) assess relationships among parasite prevalence and individual snake characteristics (i.e., sex, age, body condition), and 3) determine the spatial distribution of snakes and factors that may influence parasite prevalence (i.e., distance to the forest edge, the nearest neighboring snake, or nearest neighboring parasitized-snake). The results of this study provide a novel assessment of the parasite community in urban snakes, which may contribute to future assessments of the roles snakes play in parasite transmission in an urban environment.
Sample sizes, species characteristics, and parasitism prevalence of six snake species studied in an urban park
Parasite prevalence: individuals infected/total individuals (%)
bSex ratio M:F
bAge ratio A:J
149.7 ± 15.1
562.5 ± 25.2
6.16 ± 2.4
219.3 ± 30.3
16.5 ± 1.5
290.0 ± 10.0
1.6 ± 0.05
210.5 ± 74.5
Parasite collection and analyses
When each snake was processed, a handheld 10× magnifying lens was used to visually inspect the skin, particularly skin-folds around the eyes, mouth, and vent, for the presence of ectoparasites. Presence or absence of ectoparasites was determined for each individual and the percentage of infected individuals sampled, or parasite prevalence (Bush et al. 1997), was determined. When present, live ectoparasites were collected for subsequent species identification using published keys and then stored for possible future molecular analyses to determine invertebrate vectors involved in snake parasitism.
Hemoparasite prevalence in snakes was assessed using both microscopy and molecular techniques. A 1-mL heparinized syringe with a 25-gauge needle was used to collect <0.1 mL of blood from the caudal vein of each snake and a small drop of blood was immediately used to create a blood smear. Blood smears were allowed to air dry at ambient temperature in the field and then transported to the laboratory where slides were fixed in absolute methanol for 10 min and then stained using a modified Giemsa method (Bennett 1970). Stained smears were cleared with xylene, allowed to air dry, and then cover-slipped and sealed using Cytoseal 60 (VWR, San Francisco, CA). To confirm the presence or absence of hemoparasites, each blood smear was analyzed under 1000× magnification using a light microscope (Model BX40, Olympus, Center Valley, PA) attached to a digital camera (Infinity-1, Lumenera, Ottawa, ON, Canada). Hemoparasites were identified to genus using a published key (Telford 2009) and then hemoparasite prevalence was determined for each snake by surveying blood smears for 5 min under 600× magnification.
The remainder of each blood sample was placed onto a DNA collection card (model FP705, Fitzco, Inc., Spring Park, Minnesota) for stable storage until molecular analyses were completed. To confirm the genus and to identify the species of hemoparasites, genomic DNA was extracted from blood samples on DNA cards using DNeasy Blood and Tissue kit (Qiagen, Valencia, CA), following the manufacturer’s instructions. To detect the presence of hemoparasites of the Hepatozoon spp., a Polymerase Chain Reaction (PCR) was used to amplify the 18S RNA gene of Hepatozoon spp. as described previously (Harris et al. 2011). Briefly, a nested PCR reaction with primers HEMO1 and HEMO2 (Perkins and Keller 2001), and then primers HEPF300 and HEPR900 (Ujvari et al. 2004), were used for Hepatozoon species confirmation and identification. PCR products were isolated and purified using Ultra Clean GelSpin DNA purification kit (MO BIO, Carlsbad, CA) and PCR product sequencing was performed at the University of Tennessee Health Science Center Molecular Resource Center. A BLAST search was performed to match the product sequence to that of a known hemoparasite species.
In 2010 analyses of fecal samples were undertaken to identify fecal parasite genera and prevalence in snakes at Overton Park. Samples were collected from 14 snakes by gently palpating the intestines toward the vent until feces were produced and then placed in a plastic bag until fecal floats were produced. In the laboratory, a 0.15-mg sample of each fresh fecal sample was placed in a 2-mL centrifuge vial, 0.2 mL of a 1:10 ratio solution of Lugol’s iodine to deionized water was added (to improve visualization of parasites and their ova), and a sodium nitrate solution (Fecasol solution, Vetoquinol USA, Inc., Buena, New Jersey) was added to fill the vial. A cover-slip was placed on top of the vial in contact with the solution and after a 20-min incubation period, the cover slip was removed and placed on a glass microscope slide. Fecal floats were analyzed under 1000× magnification using light microscopy to determine the presence and absence of fecal parasites or their ova, which were identified using a published key (Klingenberg 1993). Fecal parasite prevalence for all genera detected in fecal samples was then determined for each snake by surveying fecal floats for 5 min under 200× magnification. Total fecal parasite prevalence reported herein reflects the sum prevalence of the types of fecal parasites; thus, presence of any fecal parasite is reported as positive for fecal parasites.
A Chi-Square test was used to test whether ectoparasite and hemoparasite infection prevalences were related in the snakes. Fisher’s exact tests were used to determine ectoparasite or Hepatozoon spp. infection differed between sexes, age classes (adult or juvenile), or BCI (positive or negative). Fecal parasitism was not evaluated statistically due to low sample size. Statistical significance was confirmed at P < 0.05 for all analyses.
Locations of snakes were plotted (Fig. 1c) using ArcGIS 9.3 (ESRI 2008). Satellite imagery from ArcGIS Online (www.arcgis.com) was used to delineate the forest boundaries of Overton Park and to measure distance from each snake to the forest edge, to the closest neighboring snake, and to the closest neighboring snake infected with ectoparasites, Hepatozoon spp., and fecal parasites, or a combination of the three. Mann-Whitney U-tests were used to test whether infected and non-infected snakes differed in their distance to the forest edge.
Risk maps (Ostfeld et al. 2005) were created by interpolating the presence of Hepatozoon spp. and ectoparasites across the entire area of the forest fragment, based on the presence of these parasites in the snakes, using the geo-statistical method of kriging with the Spatial Analyst extension for ArcGIS 9.3. Kriging is commonly used to map and predict the spread of disease (Moore and Carpenter 1999; Kleinschmidt et al. 2000; Booth and Dunne 2004). The data for fecal parasites were not interpolated due to low sample size.
We studied 34 snakes of six species including copperhead (A. contortrix), eastern worm (Carphophis amoenus), eastern garter (Thamnophis sirtalis), brown (Storeria dekayi), southern ringneck (Diadophis punctatus), and rat (Pantherophis spiloides) snakes (Table 1). The capture rate was 0.03 snakes per sampling hour, and only one individual copperhead was recaptured during the study period. Snakes in this urban forest spanned a five-fold range in SVL and a 318-fold difference in body mass from the smallest juvenile brown snake to the largest adult ratsnake. Overall, the snake community was male-biased with 1.8 males for every female and similarly, adult snakes were 1.8 times more common than juveniles in the community (Table 1). Copperhead snakes were the dominant species in the urban forest representing 73.5 % of the total, whereas no other species was detected more than three times (Table 1).
In the urban forest, 64.7 % of the snakes that were sampled had one or more type of parasites (Fig. 1c; Table 1). 76.0 % of copperheads were infected by at least one parasite type, but of the other five species, only 55.6 % of individuals were infected (Table 1). Of the 34 snakes, 12 (35.3 %) had Hepatozoon spp. and 9 (26.5 %) had ectoparasites. Of the 14 snakes sampled for fecal parasites, 10 (71.4 %) were positive. Nine snakes were infected with hemoparasites only, four with fecal parasites only, and two with ectoparasites only (Fig. 1c). There were seven instances of co-infections and of these, four had fecal and ectoparasites, one had hemo- and ectoparasites, and two had all parasites present (Fig. 1c). Hepatozoon spp. infection in a snake was not related to whether or not a snake had ectoparasites (χ2 = 0.021; P > 0.1), nor was Hepatozoon spp. infection different between sexes (P = 0.27), age classes (P = 0.14), or BCI (P = 0.14). Similarly, ectoparasite infection did not differ between sexes (P = 0.68) or age classes (P = 0.44), however, ectoparasites were more common in snakes with positive body condition (P = 0.019).
Temporal distribution of snake and parasite samples
Parasite prevalence: individuals infected/total individuals
Distance metrics for 34 snakes, indicating the snakes’ proximity to the forest edge and proximity to other snakes
Distance to forest edge (m)
Distance to nearest snake neighbor (m)
aDistance to nearest snake with Hepatozoon spp. (m)
aDistance to nearest snake with ectoparasites (m)
aDistance to nearest snake with fecal parasites (m)
The six species encountered during this study represented 46.2 % of the 13 species predicted to occur in the area (Scott and Redmond 2008). Snakes in this study were of normal body mass and SVL for their species (Gibbons and Dorcas 2005), suggesting that inhabiting the urban forest has no significant consequences on body size of these species. The sex and age demographics of the urban forest snake community were driven primarily by copperhead snakes, which represented 73.5 % of snakes encountered during this study. Copperheads were most likely encountered far more than other species because copperheads are relatively large sit-and-wait predators that spend more time on the substrate than other species (Gibbons and Dorcas 2005). For example, ratsnakes are also large but are arboreal, making them less conspicuous during walking surveys. Moreover, brown, eastern worm, and southern ringneck snakes are fossorial and relatively small, thus making them more difficult to detect (Gibbons and Dorcas 2005). Future studies will include additional surveys to acquire more even sampling amongst species, which will enable intraspecific comparisons of parasitism prevalence.
A type of snake mite was the only ectoparasite detected and it was found almost exclusively on copperhead snakes, the only pit-viper in the urban forest (Table 1). Surprisingly, no ticks were found on snakes in this study despite an abundance of these ectoparasites in the urban forest and the potential for ticks to parasitize snakes (Aubret et al. 2005; Pandit et al. 2011). Nearly 50 % of copperheads were infected with Hepatozoon spp. (Table 1), yet no other snake species studied was infected. However, Hepatozoon spp. infections have been documented in non-viperid snakes including water pythons (Liasis fuscus) (Ujvari et al. 2004) and Australian keelback snakes (Tropidonophis mairii) (Shilton et al. 2006) at much greater prevalence than reported here (90 to 100 %). Fecal parasites were detected in 3 of 5 species examined. Of the three parasite types measured (fecal, hemo- and ectoparasites), fecal prevalence was the greatest (71.4 %), and it was similar to fecal parasite prevalence (54 %) in the striped swamp snake (Regina alleni) (Franz 1977).
The factors that influence parasitism in snakes vary dramatically. For example, aside from ectoparasites and body condition, parasite prevalence was not related to sex, age, or body condition in this study, which corresponds to earlier studies that failed to demonstrate differences in parasitism due to age (Sperry et al. 2009), sex (Shine et al. 1998; Madsen et al. 2005; Sperry et al. 2009), or body condition (Sperry et al. 2009). In contrast, other studies have demonstrated greater hemoparasite prevalence in juvenile water pythons and those with below-average BCI (Madsen et al. 2005) and greater ectoparasite prevalence in adult male ball pythons (Python regius) (Aubret et al. 2005). The effects of snake parasites on their hosts are equally variable. Hepatazoon spp. infection decreases BCI of adult water pythons as well as growth rate and survival of juvenile water pythons (Madsen et al. 2005) and can impair immune function in the same species (Ujvari and Madsen 2005). In other snakes, parasitism has apparent severe effects on neither body condition (Shilton et al. 2006; Sperry et al. 2009) nor other measures of host fitness including growth rate, feeding rate, locomotor performance, and reproductive output (Shilton et al., 2006). Long-term population- or community-level studies may elucidate whether parasitism, or a combination of parasitism and the possible secondary stress of urbanization, affect host fitness in snakes inhabiting the urban forest.
Spatial analysis indicated that snakes infected with fecal parasites were closer to the forest edge than snakes that were not infected. Although mean distance to the forest edge for snakes infected with Hepatozoon spp. was 76.9 m and mean distance for those not infected was 104.6 m, the means were not statistically different; however, snakes with Hepatozoon spp. were at greater concentration in the northwestern section of the park. This area of the park is close to Rainbow Lake, a human-made, 1-ha lake that is directly below the Memphis Zoo. In order to determine whether these two variables contribute to the prevalence of Hepatozoon spp. in this area of the forest, we would need to collect additional snakes from the perimeter of the park, closer to Rainbow Lake, and on the eastern border of the Memphis Zoo. The risk maps for Hepatozoon spp. and ectoparsites (Fig. 4a and b) were not similar in interpolating the prevalence of these types of parasites in the urban forest. The lack of similarity suggests that these parasites are driven by different vectors (and potentially different environmental conditions), and that snakes infected by Hepatozoon spp. are not more likely to have ectoparasites.
Prior to this study, little was known about snake parasitism in an urban old-growth forest. The abundance and diversity of snakes in the forest, coupled with high overall parasite prevalence (64.7 % hosted at least one parasite), confirm that snakes can be important hosts, or possibly reservoirs, facilitating parasite transmission in urban environments. Parasites respond to environmental changes and, with the appropriate experimental design, can be used as bio-indicators of ecosystem health (Vidal-Martínez et al. 2010). This study provides 1) a baseline of understanding of the snake parasite community in an urban forest, and 2) an integrated multi-disciplinary approach that may be used for monitoring the health of the ecosystem in this and other urban forest parks.
Following this preliminary survey, further analysis is required to fully characterize the parasite community. For example, although important for the detection and control of helminth infections, fecal sampling seldom provides a definitive identification of the worm. Therefore, snake necropsies will be required for reliable helminth identification. Moreover, the molecular analysis performed on the hemoparasite could not determine the Hepatozoon species, suggesting that the copperheads in Overton Park may be infected with a yet to be described Hepatozoon spp. To confirm the species, greater coverage of the 18S rRNA gene is required, as well as amplification of other genomic regions such as the mitochondrial cytochrome b gene and the internal transcribed spacer (ITS) region or both. In addition, snake necropsies would be crucial to determine developmental stages and the identification of the vector required to fully characterize the Hepatozoon species. The ectoparasite remains to be identified. For this, better sample preservation will be required at the moment of capture.
It is important to integrate landscape ecology with epidemiology in order to better understand the interactions among vectors, hosts, potential reservoirs, and transmission (Ostfeld et al. 2005). A greater sample size, distributed across the entire park, would help us determine to what extent the different types of parasites are found in the park. Furthermore, a greater sample size would allow us to investigate the extent to which the urban landscape (e.g., forest disturbance, water bodies, vector prevalence) influences parasite prevalence. Future data will be used to: 1) test the reliability of our current risk maps, 2) determine the minimum sample size needed in order to accurately predict parasite prevalence, and 3) further investigate the landscape variables that influence parasite prevalence and transmission.
We thank the Rhodes Fellowship for funding this project, Kimber Jones, Matt McCravy, and Alex Yu for assistance in the lab, and one anonymous reviewer for feedback on an earlier version of this manuscript. Special thanks to the Tennessee Natural Resources Agency and the City of Memphis for site access. This research was conducted in accordance with Rhodes IACUC # 099 and Tennessee Wildlife Resources Agency Permit # 3506.
Conflict of interest
The authors declare that they have no conflict of interest.
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