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Forest disturbance and human encroachment have the potential to influence intestinal parasite communities in animal hosts by modifying nutritional health, physiological stress, host densities, contact rates, and ranging patterns. Anthropogenic disturbances also have the ability to affect the ecological landscape of parasitic disease, potentially impacting the health of both wildlife and people. Our research investigated the association of forest disturbance and human encroachment on intestinal parasite communities in mantled howler monkeys, Alouatta palliata aequatorialis. We found that individual parasite species prevalence was associated with group size and forest disturbance. Proximity to people was not a direct factor influencing intestinal parasitism; rather, several human proximity indices were related to group size which was in turn related to overall species richness and the presence of specific parasite species. These results, coupled with previous findings, suggest that anthropogenic disturbances are likely influencing intestinal parasite communities. Though no one study has definitively explained all relationships between anthropogenic disturbances and intestinal parasitism, we propose that our models are appropriate for meta-analysis testing across other species and environments in other studies.
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Original Research Article
Folia Primatol 2017;88:307–322
DOI: 10.1159/000479687
Association of Anthropogenic Disturbances and
Intestinal Parasitism in Ecuadorian Mantled Howler
Monkeys, Alouatta palliata aequatorialis
WilliamD.Helenbrook a,b StephenV.Stehman b WilliamM.Shields b
ChristopherM.Whipps
b
a School for Field Studies, Urubamba , Peru;
b SUNY-ESF, State University of New York College
of Environmental Science and Forestry, Environmental and Forest Biology, Syracuse, NY , USA
Keywords
Alouatta palliata · Primates · Parasitism · Logging · Anthropogenic disturbance ·
Ecuador
Abstract
Forest disturbance and human encroachment have the potential to influence intes-
tinal parasite communities in animal hosts by modifying nutritional health, physiologi-
cal stress, host densities, contact rates, and ranging patterns. Anthropogenic distur-
bances also have the ability to affect the ecological landscape of parasitic disease,
potentially impacting the health of both wildlife and people. Our research investigated
the association of forest disturbance and human encroachment on intestinal parasite
communities in mantled howler monkeys, Alouatta palliata aequatorialis . We found that
individual parasite species prevalence was associated with group size and forest distur-
bance. Proximity to people was not a direct factor influencing intestinal parasitism; rath-
er, several human proximity indices were related to group size, which was in turn related
to overall species richness and the presence of specific parasite species. These results,
coupled with previous findings, suggest that anthropogenic disturbances are likely in-
fluencing intestinal parasite communities. Though no single study has definitively ex-
plained all relationships between anthropogenic disturbances and intestinal parasitism,
we propose that our models are appropriate for meta-analysis testing across other spe-
cies and environments. © 2017 S. Karger AG, Basel
Received: March 20, 2017
Accepted after revision: July 23, 2017
Published online: September 29, 2017
William D. Helenbrook
760 Parkside Trail NW
Marietta, GA 30064 (USA)
E-Mail wdhelenb @ syr.edu
© 2017 S. Karger AG, Basel
www.karger.com/fpr
E-Mail karger@karger.com
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308 Helenbrook/Stehman/Shields/Whipps
Introduction
Studies assessing parasitism in wild primate populations have emanated from an
interest in zoonotic pathogen transmission [Muriuki et al., 1998; Howells et al., 2011;
Ghai et al., 2014; Cibot et al., 2015], conservation of host species [Wallis and Lee, 1999;
Gillespie et al., 2005; Barelli et al., 2015], and ecosystem health [Marcogliese, 2005;
Kowalewski et al., 2011]. Modified landscapes due to anthropogenic disturbances have
led several researchers to test whether intestinal parasites of host populations vary in
response to habitat alteration [e.g., Gillespie et al., 2005; Trejo-Macías et al., 2007;
Kowalewski and Gillespie, 2009; Raharivololona and Ganz horn, 2009; Lane et al.,
2011; Junge et al., 2011; Jones et al., 2013]. However, despite various efforts to under-
stand the impact of forest degradation on parasite communities in primates and other
wildlife, there still remains the unanswered question of whether specific anthropo-
genic disturbances affect parasite communities and the degree to which they do so.
Both zoonotic transmission of pathogens and composition of parasite commu-
nities in mammals have been linked to human disturbances, as previously outlined
[Patz et al., 2000; Chapman et al., 2005; Puttker et al., 2008; Zommers et al., 2013;
Helenbrook, 2014; Parsons et al., 2015; reviewed by Han et al., 2016]. In particular,
parasite species richness and prevalence of several intestinal parasites have been re-
ported to be higher in selectively logged and forest fragments compared to primary
habitat [Gillespie et al., 2005; Salzer et al., 2007; Trejo-Macías et al., 2007; Schwitzer
et al., 2010]; however, there are several cases where no effect was found – particu-
larly as it relates to primates [Kowalewski and Gillespie, 2009; Young et al., 2013].
Results of a meta-analysis of howler monkey parasitism research revealed that human
proximity and habitat disturbance were not found to have an effect on parasite spe-
cies richness and the presence of most intestinal parasites [Kowalewski and Gillespie,
2009], though this finding may be due to these studies not assessing habitat distur-
bance quantitatively. Rather, forest types were characterized dichotomously (e.g.,
continuous vs. fragmented forest types).
Interactions between the environment, parasites, and hosts are dynamic and
complex. Two major anthropogenic factors likely driving intestinal parasitism are
forest fragmentation as a result of logging, and human encroachment. In the case of
forest fragmentation, there are several hypotheses. First, travel routes can be limited,
bringing individuals into greater contact with contaminated foliage or forcing them
onto the ground where they are more likely to come into contact with soil-transmit-
ted pathogens [Stoner, 1996; Gillespie et al., 2005; Trejo-Macías et al., 2007; Mbora
and McPeek, 2009; Pozo-Montuy et al., 2013]. Groups may also be forced into small-
er areas, thereby increasing density, contact rates and subsequent parasite exposure
[Stoner, 1996; Nunn et al., 2003; Gillespie et al., 2005; Vitazkova and Wade, 2007;
Wells et al., 2007; Mbora and McPeek, 2009; Arroyo-Rodriguez and Dias, 2010]. Edge
effects along logged forests can also inhibit or enhance parasitism through altered
environmental conditions such as temperature or moisture levels [Chapman et al.,
2006a], or increased physiological stress associated with degraded habitats could
make individuals more prone to infection [Martinez-Mota et al., 2007]. Fragmenta-
tion can also limit individual dispersal between groups or populations which could
conceivably force groups into “genetic islands” – limiting their ability to recruit new
members and ultimately leading to inbreeding depression [Estrada et al., 2002]. The
effect of inbreeding depression on wild primate parasite communities is unknown,
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though evidence from captive studies suggests that there is an association between
loss of genetic diversity and higher prevalence and abundance of certain parasites
[Charpentier et al., 2008; Oklander et al., 2010]. Forest fragmentation has also been
shown to affect food availability [Gillespie et al., 2005; Chapman et al., 2006b; Weyher
et al., 2006; Dunn et al., 2009]. As a result, host nutrition could be impacted by re-
duced habitat quality which increases the propensity for infection [Junge et al., 2011].
Likewise, lower density food resources are associated with larger home ranges, forc-
ing individuals to travel further, increasing their encounters with more parasites, and
in turn leading to higher parasite abundance and prevalence [Nunn et al., 2003; Gil-
lespie et al., 2005].
The second major factor driving intestinal parasitism in wildlife is human en-
croachment. As people and communities expand into largely untouched ecosystems,
chances for zoonotic transmission are likely to increase. For example, primates living
on the forest edge of fragmented habitats see increased contact with agricultural plots
and areas with domestic animals – both potential sources of intestinal parasites [Tre-
jo-Macías et al., 2007]. Wildlife may also be more likely to acquire parasites from lo-
cal human communities due to physical proximity [Graczyk et al., 2002; Goldberg et
al., 2007; Davies and Pedersen, 2008; Goldberg et al., 2008; Pedersen and Davies,
2010; Kowalewski et al., 2011]. For example, humans and nonhuman primates inter-
acting in the wild have been shown to share genetically similar Escherichia coli [Gold-
berg et al., 2007, 2008]. Colobus and guenons are also possible reservoirs for zoonot-
ic transmission of Giardia and Cryptosporidium species [Salzer et al., 2007].
There is evidence that howler monkeys ( Alouatta palliata ), a New World pri-
mate species, are negatively affected by anthropogenic disturbances, despite being
found in close proximity to people and able to withstand limited habitat degradation
[Martinez-Mota et al., 2007; Arroyo-Rodriguez and Dias, 2010]. Howler monkeys
provide an excellent example of a primate species that is routinely found near people,
providing an opportunity to test hypotheses related to zoonotic transmission. We
focused on 2 types of anthropogenic disturbances and their potential impact on par-
asite communities in mantled howler monkeys, including: (1) forest disturbance as
quantified from basal area (cross-sectional area of all trees 1.3 m from the ground
within a specified forested area) and percent of trees >40 cm diameter at breast height
(DBH; indicative of forest quality [Cottam and Curtis, 1956; Arroyo-Rodriguez et al.,
2007]), and (2) human encroachment as defined by proximity to the research station,
roads, agricultural plots, and local communities. Our aim was to help unravel the re-
lationship between identified intestinal parasites in a New World primate and indices
of anthropogenic disturbance. Understanding how anthropogenic disturbances in-
fluence intestinal parasite communities in primate populations is important for man-
agement and conservation planning purposes. Further, there is a clear benefit to the
people who live near these tropical forests if we can better understand how environ-
mental changes may impact parasitic disease dynamics and zoonotic transmission.
Methods
The Bilsa Biological Station (00°21’33” N, 79°42’02” W; 300–750 m; Fig.1 ) is located in
northwestern Ecuador, roughly 60 km from the Pacific Ocean. The reserve spans 3,300 ha and
is surrounded on 3 sides by the Mache Chindul Ecological Reserve which is adjacent to 2 local
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310 Helenbrook/Stehman/Shields/Whipps
communities. This field site is ideal for our study because varying levels of forest disturbance
are present, including 20% secondary forest, 2 nearby communities of less than 200 people in
total, and a single road that splits the reserve into 2 similarly sized areas [Ortega-Andrade et
al., 2010; Helenbrook, 2014]. The reserve is also devoid of recent hunting pressures, which
means that wildlife – particularly monkeys – are often found living in close proximity to
people.
To sample primate groups, teams of 2–5 people systematically traversed two 5-km transects
collecting 96 fresh fecal samples from 23 primate groups ranging in size from solitary to 10 indi-
viduals from June until August 2010. Five sampled groups had 2–3 individuals, 7 groups had 4–5
individuals, 6 groups had 6–7 individuals, and 1 group had 10 individuals. Four solitary indi-
viduals were found and were not included in our modeling analysis because a minimum of 2 in-
dividuals is needed to account for sampling effort, as described below. Primate groups were sys-
tematically sampled along each transect. We sampled the entire group, then moved to the next
group further along the transect. Towards evening, the location of the last sampled group would
be noted, and sampling would begin the following morning prior to howler movement, with the
next group encountered further along the transect from the previous group. It is possible that
groups were missed, but our main purpose was to avoid repeat sampling of either individuals or
groups [Helenbrook et al., 2015b]. Every effort was made to sample all individuals within a group,
though some individuals may have escaped detection or did not defecate during the collection
period.
The location of each monkey was recorded using a global positioning system. This informa-
tion, coupled with satellite imagery, was used to define 4 indicators of human encroachment
(proximity), distance from each monkey to the nearest road, distance to the nearest agricultural
field, distance to the Bilsa Biological Station, and distance to the nearest human settlement. A
forest disturbance estimate was calculated based on where each howler group was located. The
first individual howler sampled within a group was used as the central location of the forest plot
for that group. Two methods were used to assess forest disturbance: basal area was estimated us-
ing data from point-centered quarter methods in 10-m circular plots and percentage of trees
a
b
c
Fig. 1. Field research took place in northwestern Ecuador ( a ) at the Bilsa Ecological Station –
highlighted with diagonal lines ( b ). Two 5-km transects were used, running through both sec-
ondary and primary forest, along which mantled howler monkey groups were systematically
sampled (triangles). Groups with more than 1 individual are shown as solid triangles. There are
numerous homes, represented by circles, in 3 communities surrounding the reserve ( c ).
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greater than 40 cm DBH was calculated for each plot – both of these measures are inversely as-
sociated with forest disturbance [Cottam and Curtis, 1956; Arroyo-Rodriguez et al., 2007]. All
trees with a DBH ˰ 10 cm were measured. We were not able to set up a forest plot for 1 howler
group because of steep terrain so there were 22 forest plot locations in total.
Fecal samples were collected using disposable gloves and sterile tongue depressors to ma-
nipulate a portion of the sample into 50-mL tubes containing 10 mL RNAlater® (Qiagen Inc.,
Valencia, CA, USA), and some into zinc polyvinyl alcohol fixative [Helenbrook et al., 2015b].
These samples were then examined for intestinal parasites using trichrome stain on fecal smears,
centrifugal flotations, and sedimentations at the Fish and Wildlife Disease Laboratory at SUNY-
ESF, Syracuse, New York. Slides were scanned at 20× objective lens using a Nikon 80i compound
microscope with Nomarski and phase objectives. Images were captured at 40× objective lens with
a 3MP IDEA digital camera and analyzed with photomicrography software (Diagnostic Instru-
ments Inc. Spot RT Software 4.6, Sterling Heights, MI, USA). Samples preserved in RNAlater
®
were used for PCR-based detection of Blastocystis spp. because they are cryptic but common in-
testinal parasites found in primates [Stensvold et al., 2009]. DNA was extracted from approxi-
mately 200 mg of feces using the QIAamp DNA Stool Mini Kit (Qiagen, Hilden, Germany). Blas-
tocystis spp. individuals were confirmed using protocols and PCR primers BH1F/BHRDr, and
BLF/BLR and protocols described in Helenbrook et al. [2015a]. Results from fecal smears, flota-
tions, and sedimentation were combined to confirm presence or absence, and are subsequently
reported as a single value for all calculations. Thirteen intestinal parasite taxa were recovered as
previously described in Helenbrook et al. [2015b].
Statistical Analysis
Parasite species richness within each group of howler monkeys was adjusted for sampling
effort by controlling for group size using 3 nonparametric species richness estimators (Jack-
knife, ICE, and Chao 2 in EstimateS 9.1.0) [Chao, 1987; Colwell, 2013; Helenbrook et al.,
2015b]. To be included in this analysis, groups needed 2 or more individuals, resulting in a
sample size of n = 19 groups. All 3 richness estimates were examined in the group analyses be-
cause of their precision, minimally biased predictions, and relative performance compared to
other methods [Walter and Morand, 1998]. A significant positive relationship was previously
reported between group size and nonparametric species richness estimates [Helenbrook et al.,
2015b]. Additionally, the Spearman rank correlation was used to quantify the association be-
tween parasite species richness (at the group level) with each measurement of human proxim-
ity and forest structure. Multiple regression was then used to assess relative contributions of
the anthropogenic factors on group level parasite species richness. Multiple regression models
evaluated all explanatory variables regardless of statistical significance in single variable mod-
els. Analyses were also conducted focusing on the response variable presence/absence of indi-
vidual parasite species within each group. A Mann-Whitney U test was used to evaluate wheth-
er differences between the medians of sites with parasite species present and sites with the spe-
cies absent for each of the measures of human proximity and forest disturbance were significant.
The Spearman correlation was used to quantify the association between the proportion of mon-
keys in a group infected with a particular parasite with each of the individual human proxim-
ity and forest disturbance measurements.
Finally, structural equation modeling was used to test various relationships between factors
using linked regression equations [Mbora and McPeek, 2009; Grace et al., 2010]. Some advan-
tages of structural equation models (SEM) include the ability to use both continuous and binary
data, greater statistical power than conventional multiple regression analyses, and the use of la-
tent variables – multiple measurements for a single conceptual variable which allows for estimates
and removal of measurement error [Beran and Violato, 2010]. Two sets of SEMs were created
based on univariate results. One set of models was created to test the relationship between eco-
logical disturbances and parasite species richness in monkey groups, while the other was devel-
oped to test the relationship of these same ecological indices with the presence of specific parasite
species. Empirical data were tested against hypothesized path models to determine path coeffi-
cients and their standard errors using generalized least squares, a method particularly conducive
to smaller sample sizes [Beran and Violato, 2010]. A χ
2 statistic, Akaike information criterion
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312 Helenbrook/Stehman/Shields/Whipps
(AIC) and root mean square error of approximation (RMSEA) were used to assess model fit.
Lower AIC values are associated with best fit models, and RMSEA values <0.05 are generally con-
sidered to be indicative of good fit. AIC penalizes model complexity, weighting those most par-
simonious models the highest. Residual plots were examined to evaluate whether model assump-
tions were satisfied. All statistical analyses were done with STATISTICA 10 for Windows (Stat-
Soft Inc., Tulsa, OK, USA).
Ethical Note
All methods reported in this manuscript were noninvasive and adhered to guidelines set
forth by the Institutional Animal Care and Use Committee at SUNY-ESF in New York, and this
research adhered to Folia Primatologica principles for the ethical treatment of primates. Permis-
sion to import and transfer biological samples was approved by the Center for Disease Control
(Permit No. 2009-06-089), and research was approved in country according to guidelines and
permit No. 033-FAU-DPE-MA approved by the Director Provincial de Esmeraldas, Lic. Guill-
ermo Oleas Zabala from the Ministerio del Ambiente in Quito, Ecuador.
Results
Thirteen intestinal parasite genera were recovered across the 23 mantled howler
monkey groups sampled ( Table1 ). Nineteen groups had more than 1 individual and
were included in the subsequent analyses. Neither actual parasite species richness nor
any of the 3 species richness estimators were strongly correlated with human proxim-
ity or forest structure variables at the group level (Appendix 1) as the highest Spear-
man correlation was 0.42 between Jackknife species richness and percentage of trees
Table 1. Parasite prevalence among 96 individuals and 23 groups using presence/absence data
from pooled results of fecal smears, flotations, and sedimentations
Parasite species Percentage of
individuals positive
Percentage of
groups positive
Apicomplexa Cyclospora sp. 18 46
Isospora sp. 3 9
Other protozoa Balantidium sp. 9 26
Blastocystis spp. 6011001
Chilomastix sp. 4 17
Dientamoeba sp. 3 13
Entamoeba spp. 56 87
Iodamoeba sp. 5 22
Nematoda Enterobius sp. 3 9
Capillaria sp. 78 100
Strongyloides spp. 88 100
Trypanoxyuris sp. 12 13
Platyhelminthes Controrchis sp. 15 39
1 Blastocystis confirmation using PCR-based detection.
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>40 cm DBH. Species richness was not significantly associated with the suite of hu-
man proximity and forest disturbance variables when evaluated via a multiple regres-
sion model. All explanatory variables together yielded an R 2 of only 0.21, and the full
model with all explanatory variables was not statistically significant (i.e., p = 0.70 for
the test of the null hypothesis that all regression coefficients were simultaneously
equal to 0, F statistic with 6 and 14 degrees of freedom).
Presence of specific parasite species was associated with several of the human
proximity and forest disturbance measurements. Although none of the associa-
tions were statistically significant at α = 0.05, we examined effect sizes (i.e., mag-
nitude of differences in medians or size of Spearman correlation) to describe
trends observed from the data. The strongest associations observed were that mon-
key groups with individuals infected with Chilomastix sp. tended to be found clos-
er to human settlements than those that showed no infection (Appendix 2 and 3),
and a higher prevalence of Blastocystis spp. infections was found in groups nearest
to the research station (Appendix 3). This same trend was found in these 2 species
across all anthropogenic measurements – Chilomastix sp. and Blastocystis spp.
were much more common in disturbed areas. Conversely, groups with Controrchis
spp. were more likely to be found in forest that had lower levels of disturbance.
None of the Spearman correlations between proportion of infections within a
group and a human proximity or forest disturbance variable exceeded 0.40 in ab-
solute value (Appendix 3).
a
b
Fig. 2. Best-fit SEMs for overall species richness ( a ) and adjusted species richness ( b ) taking into
account group sample size. Similar to species richness models, Controrchis sp. infections were
dependent on a combination of group size, human proximity (indirectly), and forest disturbance
measurements (directly). Forest disturbance was not significantly associated with the presence of
individuals infected with Entamoeba spp.; rather, group size was a driving force which was influ-
enced by human proximity measurements.
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Structural Equation Models
Though there were no significant direct relationships between risk factors and
specific parasites or species richness, there were several SEMs that were created and
tested using univariate and multivariate results ( p < 0.10) from individual monkey
analyses. The first set of SEMs focused on predicting species richness ( Fig.2 ). Best-fit
SEMs for overall species richness were similar to individual parasite species models
in that anthropogenic disturbances were not directly related to the number of parasite
species found in a group. Rather, group size was an intermediate factor associated
with both measured parasite species richness ( Fig.2 a) and estimated parasite species
richness which takes into account sampling effort ( Fig.2 b). Likewise, those groups
with both lower basal area and lower percentage of trees >40 cm DBH tended to have
more parasite species. Forest structure and human proximity were not significantly
correlated with one another – though groups further from people tended to be found
in areas with slightly higher forest structure. A goodness of fit test for the fitted SEM
indicated that the data matched the model well. Further, model fit was deemed ac-
ceptable based on RMSEA and AIC values, while Schwarz and Brown measurements
suggest that observed parasite species richness was a better model fit than any of the
adjusted species richness estimates.
The second set of SEMs focused on predicting the presence of specific parasite
species ( Fig.3 , Entamoeba spp. and Controrchis sp.). Groups found with Entamoeba
spp. were significantly larger, and these larger groups were more likely to be found
closer to people than uninfected groups. Alternatively, the presence of Controrchis sp.
a
b
Fig. 3. Best-fit SEMs for a specific intestinal parasite, including the presence of Entamoeba spp.
( a ) and the presence of Controrchis sp. ( b ). In both models, group size plays a significant role in
predicting both measured species richness and estimated species richness which takes into ac-
count sampling effort. A χ
2 statistic, AIC and RMSEA were used to assess model fit. Lower AIC
values are associated with best-fit models, and RMSEA values <0.05 are generally considered to
be indicative of good fit. AIC penalizes model complexity, weighting those most parsimonious
models the highest. Bold arrows signify significant contributions to the model, though nonsig-
nificant (not bold) arrows are still instrumental to the overall model design.
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was dependent upon a combination of both human proximity and forest disturbance
variables. Groups infected with Controrchis sp. tended to be larger than uninfected
groups and tended to be found in areas with higher forest complexity (e.g., basal area
and percentage of trees >40 cm DBH). Every group was infected with Strongyloides
spp., Capillaria sp., and Blastocystis spp., making this approach uninformative for
predicting these parasite species.
D i s c u s s i o n
Anthropogenic disturbances such as increasing human encroachment of tropi-
cal forests, logging (and subsequent fragmentation, habitat degradation, and habitat
loss), ecotourism, livestock introduction, and agriculture are all likely drivers of
changing intestinal parasite communities and emerging infectious diseases. We
found that parasite species richness was not solely related to any one environmental
risk factor, but was rather dependent on the interplay of forest complexity (e.g., bas-
al area and percentage of trees >40 cm DBH), group size, and human proximity mea-
surements. Similarly, the presence of Controrchis sp. was dependent on these same
interactions, while Entamoeba spp. were more likely to be found in larger groups
which were closer to people.
Species Richness Models
No direct association was found between overall species richness and environ-
mental risk factors, nor was the presence of specific parasite species found to be di-
rectly dependent on any measured risk factors. Proximity to people was not a direct
factor influencing intestinal parasitism; rather, several human proximity indices were
related to group size which was in turn related to overall species richness and the pres-
ence of specific parasite species.
With a sample size of 19 groups, low statistical power is a possible contributing
factor to such associations not being detected as statistically significant. The SEM re-
sults support a causal chain in which we found that larger groups saw an increase in
parasite species richness, previously attributed to higher contact rates and subsequent
transmission rates of directly transmitted parasites [Freeland, 1976; Côté and Poulin,
1995; Arneberg et al., 1998; Nunn et al., 2003; Cross et al., 2009]. This trend is also
relevant for the vast majority of intestinal parasite species found in this study: groups
infected with 10 different parasite species were larger than those groups that were
uninfected, though this relationship was not statistically significant at the group level.
The other 3 parasite species in this study were found in all monkey groups and as a
result could not be compared.
Finding larger groups near human settlements is at first unexpected, because
these areas are often degraded. And as previously reported, secondary forests and
disturbed habitat are largely associated with reduced primate group size [Gonzalez-
Kirchner, 1998; McCann et al., 2003]. Primary forest normally harbors larger groups
because of increased carrying capacity associated with more available food types and
better quality food sources, whereas in disturbed forests primates are limited to a
lower quality and quantity of food [Clarke et al., 2002; Gillespie et al., 2005; Marshall
et al., 2005; Cristóbal-Azkarate and Arroyo-Rodriguez, 2007; Martinez-Mota et al.,
2007; Gillespie and Chapman, 2008]. However, in this case, primate groups living on
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316 Helenbrook/Stehman/Shields/Whipps
the edge of agricultural plots or near fruit-bearing tree plantations may actually be
larger due to higher-density food sources, which results in increased reproductive
rates and carrying capacity [Chapman et al., 1995].
The forests in our sampled area have had nearly 15 years to recover from previ-
ous anthropogenic disturbance such as logging. These secondary forests may actu-
ally provide an ideal heterogeneous habitat for howlers [Ortega-Andrade et al., 2010],
thus explaining why those groups near people were larger. The fact that degraded
habitats might not necessarily mean poor-quality food sources might explain why
changes in forest structure were not as strongly related to parasite species richness as
expected. Mantled howler monkeys are also well adapted to changing food sources as
a result of past deforestation [Peres, 1997; Pinto et al., 2003]. Perhaps more recent ef-
fects of logging would have resulted in a larger impact on ranging patterns, diet, and
stress, subsequently providing a more distinct contrast with undisturbed primary for-
est. Pinto et al. [2003] outline other possibilities why howler behavior might not
change with modified forest structure, including low logging intensity, untouched
forest adjacent to harvested areas, and extended time period between last harvest and
sampling.
Individual Parasite Species Models
Forest complexity was not significantly associated with overall species richness
at the group level in univariate analysis. However, when using structural equation
modeling, which controls for group membership at the individual level, forest struc-
ture measurements were found to be a contributing factor – along with group size – to
overall species richness. Specifically, groups with Cyclospora sp., Enterobius sp., Try-
panoxyuris sp., Dientamoeba sp., Chilomastix sp., Entamoeba spp., and Iodamoeba
sp. were all found in areas with lower basal area and a lower percentage of trees >40
cm DBH, though none were statistically significant. Additional testing is needed to
confirm this pattern; however, if several parasite species are more prevalent in de-
graded habitat, this pattern could be explained by associated changes in group size,
edge effects, or nutritional stress. The exceptions to this rule included Controrchis sp.,
Blastocystis spp., and Balantidium sp. which tended to be found in areas with higher
basal area and a higher percentage of trees >40 cm DBH (though again, no significant
statistical differences were found).
Two SEMs were developed for specific parasite species: Entamoeba spp. and
Controrchis sp. ( Fig. 3 ). The presence of Entamoeba spp. was dependent solely on
group size, which was in turn related to human encroachment indices. Previous
studies have suggested that Entamoeba spp. infections were more common in de-
graded habitat [Gillespie et al., 2005; Chapman et al., 2006a, b]. However, though this
was a trend in our study as well, it was not a statistically significant contributing fac-
tor in any of our models. Like many of the described species found in this study,
Entamoeba spp. are transmitted through the fecal-oral route. Thus, one would ex-
pect higher transmission rates of contagious parasite species in larger groups [Cóté
and Poulin, 1995; Patterson and Ruckstuhl, 2013]. The Controrchis sp. SEM was
similar to overall species richness design: group size was a driving force and forest
disturbance secondarily. Other studies have found a greater prevalence of Contror-
chis sp. in fragmented habitat. Cecropia trees, which harbor ants infected with Con-
trochis sp., are primarily found in disturbed habitat and are often eaten by howler
monkeys. The monkeys may be ingesting the ants and thereby becoming infected
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with the parasites [Kowalzik et al., 2010]. One explanation for this apparent contra-
diction is that Cecropia trees are still found in open gaps in primary forest [Sanford
et al., 1986], suggesting that gaps could create a higher prevalence of Controrchis sp.
in monkeys living in these areas. Further investigation of tree species utilized by
monkeys and testing of ants in areas of varying habitat disturbance could likely con-
firm or refute this hypothesis.
I m p l i c a t i o n s
Expanding human populations, particularly in tropical countries where many of
the world’s primate species live, will likely lead to greater interspecies interactions and
subsequent expansion of infectious diseases throughout both primate and human
populations as reported in previous studies [Wolfe et al., 2007; Vitazkova, 2009; Ped-
ersen and Davies, 2010]. Likewise, anthropogenic disturbances such as logging have
the potential to modify intestinal parasite communities in primate populations, sub-
sequently impacting host health. These results are likely to be applicable to other dis-
turbed systems, suggesting a potential threat to other wildlife populations.
Additional studies looking at more recent and more severe levels of anthropo-
genic disturbance would prove insightful, and increased sampling of individuals
across populations would expand the breadth of the inference achievable from the
sample data. We are cautious in suggesting a universal model applicable to all pri-
mates. Rather, we anticipate our models can be tested using empirical data from oth-
er primate studies or through meta-analysis.
A c k n o w l e d g m e n t s
We thank the codirectors at Bilsa Biological Station, Juliet Bermingham and Carlos Aulestia,
who were invaluable in conveying their extensive knowledge about the Bilsa rain forests. We
could not have completed the field research without assistance from Yelena Prusakova, Sara Glen,
and Torin Heavyside. Thanks go to James Gibbs for input on the field methodologies. Funding
was provided in part from the Program on Latin America and the Caribbean (PLACA) at Syra-
cuse University, the Leroy C. Stegeman Award from SUNY-ESF, and Sigma Xi (G2009150493).
Additional support was graciously provided by Linda and Edward Garwol, Dennis and Sandy
Suarez, Jennifer and Aaron Woloszyn, William and Sandy Helenbrook, and Jeff and Jennifer
Drozdowski.
Statement of Ethics
All methods reported in this manuscript were noninvasive and adhered to guidelines set
forth by the Institutional Animal Care and Use Committee at SUNY-ESF in New York, and were
approved in country according to guidelines and permission from the Ministerio del Ambiente
in Quito, Ecuador.
Disclosure Statement
None of the authors have any conflicts of interest to declare.
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318 Helenbrook/Stehman/Shields/Whipps
A p p e n d i x 1
Matrix of Spearman Correlations between Disturbance Variables, Measured Species
Richness Using Fecal Smear, Flotation and Sedimentation Techniques, and Three Species
Richness Estimators (Chao2, Incidence Coverage-Based Estimator, and Jackknife), Which
Take into Account Sampling Effort
Appendix 2
Summary Table of Anthropogenic Disturbances Associated with Gastrointestinal Parasites
Measured
species richness
Chao2 ICE Jack
Roads 0.08 –0.06 –0.04 0.06
People 0.03 –0.10 –0.18 –0.16
Agriculture 0.29ȗȗȗ 0.06 0.11 0.08
Bilsa –0.20 0.15 0.04 0.08
Basal area 0.04 0.20 0.16 0.23
Percent trees >40 cm DBH 0.00 0.35ȗ0.36ȗ0.42ȗ
ȗȗȗ p < 0.05, ȗ p < 0.20.
Roads, m People, m Agriculture,
m
Bilsa, m Basal area,
m2/ha
Trees >40
cm DBH, %
Cyclospora sp. 369, 681 1,253, 1,524 1,038, 805 1,003, 2,513ȗ 7.2, 9.8 0.1, 0.1
Isospora sp. 465, 569 1,015, 1,281 903, 885 837, 1,697 8.7, 8.1 0.1, 0.1
Balantidium sp. 681, 393 1,496, 1,281 755, 885 1,233, 1,697 11.1, 8.0 0.1, 0.1
Chilomastix sp. 269, 681 456, 1,506ȗȗ 423, 921 442, 1,860ȗ 7.2, 9.8 0.1, 0.1
Dientamoeba sp. 884, 332 1,597, 1,147 1,356, 859 988, 1,449 5.4, 10.4ȗ0.0, 0.1
Entamoeba spp. 453, 1,236 1,415, 1,075 910, 750 1,124, 3,103ȗ8.0, 9.8 0.1, 0.1
Iodamoeba sp. 884, 453 2,305, 1,147 859, 910 2,271, 1,124 9.7, 7.8 0.1, 0.1
Enterobius sp. 1,152, 565 1,657, 1,281 982, 885 1,394, 1,287 3.6, 9.8ȗ 0.1, 0.1
Trypanoxyuris sp. 691, 332 1,901, 1,147 910, 859 1,449, 1,124 11.1, 7.2 0.1, 0.1
Controrchis sp. 453, 677 2,305, 988ȗ 1,145, 733ȗ 2,271, 966 11.7, 7.8 0.12, 0.07ȗ
Values shown are medians for 2 groups defined by whether the howler monkey group is infected
(value on the left for each pair) or not infected with each parasite. p values are provided from Mann-Whit-
ney U tests testing the null hypothesis of no difference in medians. Strongyloides spp., Capillaria sp., and
Blastocystis spp. were found in every group. ȗȗp < 0.10, ȗp < 0.20.
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A p p e n d i x 3
Proportion of Individuals Infected within a Group versus Disturbance Variables
Roads People Agriculture Bilsa Basal area Trees
>40 cm
DBH
Cyclospora sp. –0.05 –0.13 0.20 –0.28 –0.28 –0.07
Isospora sp. –0.07 –0.14 –0.01 –0.14 0.01 0.01
Balantidium sp. –0.02 0.00 0.09 –0.09 0.18 0.00
Blastocystis spp. –0.12 –0.23 –0.06 –0.39ȗȗ –0.16 –0.03
Chilomastix sp. –0.19 –0.38ȗȗ –0.23 –0.35ȗ–0.07 –0.04
Dientamoeba sp. 0.23 0.09 0.21 –0.08 –0.35ȗ–0.19
Entamoeba spp. 0.20 –0.11 –0.15 –0.39 –0.31ȗ–0.30ȗ
Iodamoeba sp. 0.26 0.13 0.16 0.02 0.20 –0.02
Enterobius sp. 0.11 –0.03 0.02 –0.09 –0.33ȗ–0.23
Trypanoxyuris sp. 0.29ȗ0.16 0.06 0.01 0.16 –0.09
Controrchis sp. 0.01 0.27 0.26 0.06 0.17 0.26
ȗ ȗ p < 0.10, ȗ p < 0.20.
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... Primate health can also be negatively affected by changes in intestinal parasite communities attributed to these same anthropogenic disturbances (e.g., Gillespie et al. 2005;Junge et al. 2011). For example, degraded habitat can modify ranging patterns and changes in group size and contact rates in response to forest fragmentation can affect parasite communities (e.g., Helenbrook et al. 2017;Mbora and McPeek 2009;Valdespino et al. 2010). Other factors associated with changes in intestinal parasite prevalence or richness include changes in travel routes, edge effects, physiological stressors, inbreeding depression, and nutritional health . ...
... Changes in intestinal parasite communities could have far-reaching consequences for primates, affecting mortality and morbidity attributed to spontaneous abortions, birth defects, blood loss, tissue damage, increased energy expenditure, decreased nutrient absorption, and decreased ability in everyday tasks such as travel and feeding (Helenbrook et al. 2017). There is mounting evidence that habitat degradation and fragmentation are associated with nutritional deficits (Estrada et al. 2002), changes in population density Mbora and McPeek 2009), behavioral changes , changes in group dynamics (Helenbrook et al. 2017), and zoonotic transmission (Estrada 2006). ...
... Changes in intestinal parasite communities could have far-reaching consequences for primates, affecting mortality and morbidity attributed to spontaneous abortions, birth defects, blood loss, tissue damage, increased energy expenditure, decreased nutrient absorption, and decreased ability in everyday tasks such as travel and feeding (Helenbrook et al. 2017). There is mounting evidence that habitat degradation and fragmentation are associated with nutritional deficits (Estrada et al. 2002), changes in population density Mbora and McPeek 2009), behavioral changes , changes in group dynamics (Helenbrook et al. 2017), and zoonotic transmission (Estrada 2006). Intestinal parasitism can thus have a strong regulatory effect on wildlife populations, which makes understanding ecological patterns and forces driving infectious disease an important management tool (Tompkins and Begon 1999). ...
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Thesis
Understanding the relationship between anthropogenic disturbances and wildlife gastrointestinal parasite communities is important to both human health and conservation efforts. Forest logging and fragmentation, burgeoning human population growth, wildlife extraction, and expansion of livestock into formerly undisturbed landscapes can affect and compound the transmission of various pathogens between wildlife and people. This study therefore aims to further understand the relationship between two types of anthropogenic disturbance (forest degradation and human encroachment), and gastrointestinal parasite communities in both humans and mantled howler monkeys, Alouatta palliata aequatorialis by addressing the following: 1) chronicle primate parasitism, 2) investigate association of environmental degradation and parasitism, and 3) assess human attributes and actions associated with parasitism and potential transmission between human and howler monkey populations. Human and monkey endoparasite communities were characterized using morphological and genetic analyses, and people from surrounding communities were administered demographic surveys to evaluate risk factors associated with parasitism. Of 96 howler monkey fecal samples collected, 2 species of apicomplexan, 6 other protozoa, 4 nematodes, and 1 platyhelminth were detected. Four congeners were found in howlers and people: Entamoeba sp., Balantidium sp., Blastocystis sp., and Strongyloides spp. Several key parasites were non-randomly distributed throughout the sampled population. Proximity of agricultural plots and a local biological research station were both associated with the presence of Strongyloides spp. Individuals were more than four times likely to harbor Strongyloides spp. if they lived in areas considered disturbed forest. Individuals infected with Controrchis sp. were found further from human settlements than uninfected individuals and nearly ten times more likely to be found in primary forest. No evidence of shared Blastocystis subtypes were found between howlers and people, though Capillaria sequence types were similar, suggesting either zoonotic transmission or a common source. Several significant human factors were associated with parasite communities. The results from this study support the hypothesis that anthropogenic disturbances can place both primate populations and humans at risk of select gastrointestinal parasites. Aside from the various direct impacts of anthropogenic disturbances, additional focus should be placed on the indirect effects changing ecological systems have on parasite communities in threatened hosts.
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Emerging infectious diseases have become an important challenge for wildlife ecologists and managers. Management actions to control these diseases are usually directed at the parasite, the host population, or a key component of the environment, with the goal of reducing disease exposure and transmission. Control methods directed at the host population, however, remain limited in approach (e.g. vaccination, population reduction, test-and-remove) and scope, by financial, logistical, ethical and political constraints. Furthermore, these control methods have often been implemented without due consideration of how host ecology and behaviour may influence disease dynamics. This chapter highlights how host population structure and social organisation affect parasite transmission and prevalence. Traditionally, variation in disease prevalence among species, genders, and ages may have been explained by immunological differences in susceptibility. However, ecological and behavioural factors can also affect the rates and routes of parasite transmission and potential control options. Using this information, future control efforts may be improved by focusing on subsets of individuals, areas, environmental factors, or times of year that are most important in the propagation and persistence of a pathogen.