El Colegio De La Frontera Sur

El Colegio de la Frontera Sur

Deforestación y fragmentación de selvas en el sur de la Península de Yucatán, México (1990 –2006)

TESIS presentada como requisito parcial para optar al grado de Maestría en Ciencias en Recursos Naturales y Desarrollo Rural

por

Juan Pablo Ramírez Delgado

2012

Dedicatoria y agradecimientos

Agradezco al Consejo Nacional de Ciencia y Tecnología, Conacyt, por la beca otorgada para la realización de los estudios de maestría. A la Dra. Birgit Schmoock, por su amistad, consejos, ideas, sugerencias y revisiones durante todas las etapas de este trabajo de investigación. Al Dr. Zachary John Christman, tanto por todo su apoyo como por las repetidas revisiones al documento. Al Dr. Héctor A. Hernández Arana y al Dr. Gerald A.

Islebe, por sus revisiones y sugerencias al documento final. Y a mis padres y hermanos,

María Teresa, Carlos Arturo, Natalia y Carlos Andrés, por la constante motivación y consejos durante mi estancia en México.

Dedico este trabajo en memoria de mi padre quien siempre me mantuvo motivado y me brindo todo su apoyo hasta el último momento compartido.

Índice

Introducción ...... 1 Deforestation and fragmentation of seasonal tropical forests in the southern Yucatán, Mexico (1990 –2006) ...... 7 Keywords ...... 7 3. Methods ...... 11 3.1. Study area ...... 11 3.2. Preparation of the spatial data ...... 13 3.3. Deforestation and fragmentation analysis ...... 13 4. Results ...... 14 4.1. Deforestation rates ...... 14 4.2. Changes in forest cover: 1990 –2000 ...... 15 4.3. Changes in forest cover: 2000 –2006 ...... 16 4.4. Trends in forest fragmentation: 1990 –2000 ...... 16 4.5. Trends in forest fragmentation: 2000 –2006 ...... 17 5.1. Changes in forest cover ...... 18 5.2. Trends in forest fragmentation ...... 20 6. Conclusions and management implications ...... 22 Acknowledgements ...... 23 Highlights ...... 33 Artwork ...... 35 Tables ...... 40 Conclusiones ...... 42 Referencias ...... 43

Introducción

La deforestación y la fragmentación de bosques han sido reconocidas como las principales amenazas para los ecosistemas de todo el mundo (Iida y Nakashizuka 1995,

Dale y Pearson 1997, Noss 2001, Armenteras et al . 2003). La deforestación ocurre cuando la cantidad total del hábitat original se reduce, mientras que la fragmentación sucede cuando el hábitat remanente es dividido en fragmentos de diversos tamaños y grados de aislamiento (McGarigal y Cushman 2002; Fahrig 2003). Ambos procesos pueden tener efectos negativos sobre la biodiversidad, ya que aumentan el aislamiento entre hábitats

(Debinski y Holt 2000), modifican la dinámica de las poblaciones e incrementan el número de especies tanto de animales como de plantas en riesgo (Watson et al . 2004). La fragmentación, la cual reduce la probabilidad de establecimiento y dispersión de las especies que ocurren en hábitats naturales (Gigord et al . 1999), así como la capacidad de los parches de hábitat para sostener una población residente (Iida y Nakashizuka 1995), ha afectado la riqueza de las aves frugívoras en algunas selvas de la Península de Yucatán

(Weterings et al . 2008), ha favorecido la expansión del helecho Pteridium aquilinum (L.)

Kuhn 1 y ha proliferado la vegetación intolerante a la sombra a lo largo de los bordes de los fragmentos de dichas selvas (Hernández-Stefanoni y Dupuy 2008, Schneider 2008,

Weterings et al . 2008).

Entre los temas ambientales más importantes que enfrentan los países tropicales en vías de desarrollo están la pérdida y fragmentación de sus selvas (Laurance 1999). La preocupación relacionada con la destrucción de éstas ha venido ampliándose a nivel mundial (Lugo y Brown 1992, Myers et al . 2000, Tucker y Townshend 2000), debido principalmente a que son el hogar de muchos pueblos indígenas (Alcorn 1993), proveen

1 El helecho Pteridium aquilinum (L.) Kuhn es una especie invasora que se establece en áreas que fueron anteriormente dominadas por fuego, deforestación y actividades agrícolas (ver Schneider 2008). 1

servicios vitales de los ecosistemas tales como el mantenimiento de agua, la reducción en la magnitud y frecuencia de las inundaciones (Eisenbies et al . 2007) y la conservación de suelos (Costanza et al . 1997), suministran recursos naturales maderables y no maderables

(Arnold y Ruiz Pérez 2001), son proveedoras de medicinas naturales (Balick y Mendelsohn

1992) y tienen una gran influencia en el almacenamiento de carbono y del clima a escala global y regional (Malhi y Phillips 2004, Bonan 2008).

A nivel del paisaje, el cual se define como un área espacialmente heterogénea (Turner

2005), las consecuencias de la fragmentación por efecto de la deforestación incluyen la pérdida de hábitats para algunas especies de animales y plantas (e.g. Santos et al . 2002,

Honnay et al . 2005, Chushman 2006), la generación de hábitats para otras (e.g. Didham et al . 1998, Sekercioglu et al . 2002, Feeley y Terborgh 2006), la disminución de la conectividad entre vegetación remanente, el decremento en el tamaño de parches de selva, el aumento de la distancia entre parches de selva y el crecimiento de bordes a expensas del hábitat original (Skole y Tucker 1993, Mace et al . 1998).

Algunos estudios han demostrado que la configuración espacial del paisaje y la estructura de una comunidad pueden afectar significativamente la riqueza de especies

(Steiner y Köhler 2003). Otros enfatizan la necesidad de incorporar la configuración espacial y los atributos de conectividad a un nivel de paisaje para así proteger la integridad ecológica de ensambles de especies (Herrmann et al . 2005, Piessens et al . 2005). Esto demuestra que los efectos ecológicos de la fragmentación pueden ser diferentes en función de la configuración o de los patrones espaciales impuestos sobre un paisaje y en cómo este varia temporal y espacialmente (Ite y Adams 1998, Armenteras et al . 2003, Cayuela et al .

2006). Por lo tanto, es necesario entender la relación entre los patrones del paisaje y los procesos ecológicos que influyen en la distribución de especies para mejorar los programas

de manejo de recursos naturales y la toma de decisiones sobre el uso del suelo (Ranta et al . 1998, Turner et al . 2001a).

La aplicación de sistemas de información geográfica (SIG) y el análisis de imágenes de satélite se han convertido en un valioso conjunto de técnicas para estimar la pérdida y la fragmentación de los bosques (Luque 2000, Franklin 2001, Imbernon y Branthomme 2001,

Armenteras et al . 2003, 2006, Echeverria et al . 2006, Teixido 2010). Con respecto a esto, un gran número de estudios se han llevado a cabo en selvas húmedas (e.g. Skole y Tucker

1993, Turner y Corlett 1996, Imbernon y Branthomme 2001, Sader et al . 2001, Steininger et al . 2001, Armenteras et al . 2006), principalmente en las que se encuentran en el Amazonas, región considerada como una de las más estudiadas en el mundo con relación a la deforestación y la fragmentación de sus selvas (Jorge y Garcia 1997, Pedlowski et al . 1997,

Ranta et al . 1998, Laurance 1999, Laurance et al . 2000, Sierra 2000). Sin embargo, aún no han sido reportados análisis temporales con SIG e imágenes de satélite sobre los patrones espaciales de deforestación y fragmentación de los bosques de México, con la notable excepción de Cayuela y colaboradores (2006), quienes reportan tasas de pérdida y patrones de fragmentación del bosque mesófilo de montaña en los Altos de Chiapas,

México, durante un periodo de 25 años.

Una de las selvas que han presentado altas tasas de deforestación son las que se encuentran en el sur de la Península de Yucatán (Turner et al . 2001b, 2004), al sur occidente de Quintana Roo y sur oriente de Campeche, en México. Esta región cuenta con una temperatura media anual de 25 oC (Whigham et al . 1990), un rango de precipitación total anual que va de los 900 mm en el noroccidente a los 1400 en el suroriente (Lawrence y Foster 2002, Martínez y Galindo-Leal 2002, Pérez-Salicrup 2004), una estación seca que se extiende de diciembre a mayo (Lawrence y Foster, 2002), una topografía kárstica que se caracteriza por tener tierras altas ondulantes con elevaciones que van desde los 100 a los 3

360 m.s.n.m (metros sobre el nivel del mar) (Lawrence et al . 2004, Vester et al . 2007), un clima tropical húmedo y seco (Aw en la clasificación climática de Köppen) (Peel et al . 2007), y suelos de tipo rendzina en las tierras de alta elevación y de tipo vertisol en las de baja elevación (Turner et al . 2001b, 2004).

Estas condiciones dan soporte a un número de tipos de selva que se diferencian por la cantidad de sus plantas caducifolias, su estatura y su abundancia relativa de especies: selva baja inundable, selva baja, selva mediana subperennifolia, selva alta perennifolia, y selva baja y mediana subcaducifolia (después de Miranda y Hernández 1963, Pérez-

Salicrup 2004, Rzedowski 2006, Schmook et al . 2011). 2

El sur de la Península de Yucatán se destaca por ser la extensión más grande y continua de selvas que están desapareciendo más rápidamente en México (Pérez-Salicrup

2004, Vester et al . 2007), así como también por formar parte de la mayor área continua de bosques que queda en Mesoamérica (Carr 1999). Esta región comprende un importante gradiente geográfico de vegetación que es relativamente heterogéneo pero ambientalmente estable (van der Maarel 1990), el cual va de sur a norte y conecta el extremo norte del

Petén guatemalteco, donde se encuentra un alto nivel de endemismos dentro de la selva húmeda allí presente (Espadas-Manríquez et al . 2003), con las selvas secas en el norte

(Vester et al . 2007).

En los últimos 3000 años, el sur de la Península de Yucatán ha pasado por diferentes etapas de cambio en el uso del suelo. Estas incluyen una fuerte deforestación durante el periodo Clásico de la cultura Maya (ca. 800 –1000 a.C.) (Klepeis y Turner 2001, Turner et al .

2001b, Turner et al . 2003, Klepeis 2004) y un posterior periodo de recuperación de selvas

2 Para más detalles sobre la vegetación presente en el sur de la Península de Yucatán, consultar a Pérez-Salicrup y/o Schmook y colaboradores 2011. 4

(80 –260 años) que originó cambios significativos en la abundancia de las especies de esta región (Lambert y Arnason 1981, Gómez-Pompa 1987, Mueller et al. 2010).

Desde finales de 1800’s, México ha llevado a cabo diferentes políticas de uso y desarrollo relacionadas con las tierras forestales que se encuentran en el sur de la

Península de Yucatán. Estas se han centrado en la producción y comercialización de chicle

(resina de goma de mascar) (Snook 1998), la extracción de madera, el desarrollo agrícola, la conservación de selvas y los programas arqueo eco turísticos (Klepeis y Turner 2001,

Turner et al . 2001b, Turner et al . 2003, Foster y Turner 2004, Klepeis 2004).

Un intenso periodo de asentamientos agrícolas patrocinados por el Estado Mexicano durante lo s 1970’s y 1980’s aumentó rápidamente la población de pequeños agriculto res a lo largo del sur de la Península de Yucatán, llevándola a obtener altas tasas de deforestación (Klepeis y Turner 2001, Turner et al . 2001b, Turner et al . 2003, Klepeis 2004,

Vester et al . 2007). Estas hicieron que dicha región fuera designada como hotspot de deforestación tropical durante 1990 y 1997 (Achard et al . 1998, 2002).

Después de varias décadas de uso agropecuario dentro del sur de la Península de

Yucatán, y gracias al continuo aumento de la preocupación internacional y regional en relación a la magnitud de la deforestación de las selvas de esta región (Achard et al . 1998,

2002, Myers et al . 2000, Folan y García 2001), el Estado Mexicano designó, en 1989,

7225.15 km 2 del sur de esta península, cerca de un tercio de esta región, como Reserva de la Biosfera Calakmul, para que actuara como parte fundamental del Corredor Biológico

Mesoamericano y contribuyera con la preservación de la biodiversidad y el almacenamiento de carbono en México (Boege 1995, Primack et al . 1998, Vester et al . 2007). A pesar de los esfuerzos de preservación y conservación, el sur de la Península de Yucatán continúa enfrentando amenazas debidas a la conversión de selvas a tierras de cultivo y de pastizal,

lo que ha disminuido la superficie de sus selvas y aumentado la fragmentación de las mismas (Lawrence et al . 2004).

El objetivo de este trabajo de investigación es aportar información para el entendimiento de los patrones espaciales de deforestación y fragmentación de las selvas presentes en el sur de la Península de Yucatán, México, durante el periodo que va de 1990 a 2006. En particular, se pretende estimar las tasas de deforestación durante dicho periodo a través del uso de imágenes de satélite clasificadas a diferentes intervalos de tiempo, así como también evaluar los cambios temporales en la configuración espacial de estas selvas mediante el uso seleccionado de métricas de fragmentación a un nivel de paisaje.

Debido a que una gran cantidad de comunidades han ido deforestando sus tierras para usos agrícolas en los limites orientales y occidentales en el sur de la Península de Yucatán, y a lo largo del límite sur oriental de la Reserva de la Biosfera Calakmul, se plantea como hipótesis una deforestación focalizada en estas partes de la región, así como también que dicha deforestación esté relacionada con la fragmentación, debido a los cambios en la configuración espacial en términos de tamaño, forma y grado de aislamiento de los parches de selva remanente en esta región.

J.P. Ramírez-Delgado, et al. / submitted article in Forest Ecology and Management (2012)

1 Deforestation and fragmentation of seasonal tropical forests in the

2 southern Yucatán, Mexico (1990 –2006)

3 Juan Pablo Ramírez-Delgado a, Birgit Schmook a,*, Zachary John Christman b

4 aEl Colegio de la Frontera Sur-Chetumal, Carretera Chetumal-Bacalar, Km 2, AP 424, CP 77000, Chetumal, Quintana Roo, Mexico 5 bDepartment of Geography and Environment, Rowan University, Glassboro, New Jersey, USA

6 Abstract 7 The southern Yucatán (SY), with the largest and most rapidly disappearing contiguous tract of 8 seasonal tropical forest in Mexico, has been recognized as a hotspot of biodiversity at great risk 9 of deforestation, with high rates of forest loss between 1990 and 1997. However, the rate of 10 deforestation has recently been slowing across the region. Three land cover maps, derived from 11 satellite imagery acquired over a 16 year period (1990, 2000, and 2006), were used to assess both 12 annual deforestation rates and spatial patterns of forest fragmentation around and within 13 Mexico’s largest tropical forest reserve, the Calakmul Biosphere Reserve (CBR), which covers 14 about one third of the SY. Results indicate a slowing in annual deforestation rates in the SY, 15 including inside and outside the CBR, but with significant sub-regional variations in the quantity 16 and rate of forest loss. Deforestation was concentrated mostly in those land management units 17 located in the western and eastern sides of the region, and along the southeastern border of the 18 CBR. Although deforestation declined during the two study periods, it caused considerable 19 fragmentation both inside and outside the reserve. Given the necessary conditions of sustainable 20 forest management and that most timber harvesting has been carried out without formal 21 management plans, permanent forest areas, over which a community management logging plan 22 would operate, could enable forest conservation in this region. 23 Keywords: Forest fragmentation; Deforestation; Seasonal tropical forests; Mexico; Southern 24 Yucatán 25 26 1. Introduction 27 Deforestation has been recognized as the most serious threat to tropical biodiversity 28 worldwide ( Klepeis and Turner, 2001 ; Wright, 2005 ; Brook et al., 2008 ; Laurance et al., 2011 ).

* Corresponding author. Tel.: +52 9838350440x4407 E-mail address: [email protected] (B. Schmook).

J.P. Ramírez-Delgado, et al. / submitted article in Forest Ecology and Management (2012)

29 One of the most alarming aspects of forest loss is the fragmentation of continuous forest habitats 30 (Saunders et al., 1991 ), contributing to species extinction and habitat destruction, with impacts on 31 the carbon budget in an increasingly variable climate ( Sala et al., 2000 ; Houghton, 2003 ; Malhi 32 and Phillips, 2004 ; Foley et al., 2005 ; Bonan; 2008). Deforestation, which often causes forest 33 fragmentation ( Saunders et al., 1991 ; Brook et al., 2008 ), may have negative effects on 34 biodiversity by decreasing the size of habitat patches ( Brook et al., 2008 ), increasing isolation of 35 habitat patches ( Debinski and Holt, 2000 ), modifying species dynamics ( Watson et al., 2004 ) and 36 trophic interactions ( Sala et al. 2000 ), endangering species ( Sala et al., 2000 ), reducing the 37 probability of successful dispersal and establishment of individuals ( Gigord et al., 1999 ), 38 reducing the capacity of habitat patches to sustain a resident population ( Iida and Nakashizuka, 39 1995), and increasing susceptibility to logging, fires and invasive species ( Sala et al., 2000 ). 40 The ecological consequences of fragmentation may differ depending on the patterns or spatial 41 configuration inflicted on a landscape, and how it varies both spatially and temporally (Ite and 42 Adams, 1998 ; Armenteras et al., 2003 ; Arce-Nazario, 2007 ; Fearnside, 2008 ). Therefore, and in 43 response to growing concerns over biodiversity loss caused by deforestation and fragmentation 44 globally ( Achard et al., 1998 , 2002; Myers et al., 2000 ), the analysis of the relationship between 45 landscape patterns and ecological processes should be a priority in forest management programs 46 (Ranta et al., 1998 ; Turner et al., 2001 ). 47 The application of geographic information systems (GIS) and the analysis of satellite imagery 48 have become a valuable set of techniques for estimating the quantity of forest loss and 49 fragmentation (e.g. Sader et al., 2001 ; Armenteras et al., 2003 ; Viña et al., 2004 ; Echeverria et al. 50 2006; Etter et al., 2006 ). Few studies, however, have identified the spatial pattern or the 51 arrangement, position, or orientation of deforestation and fragmentation, and analyzed the 52 temporal dynamics of the landscape in Mexico’s tropical regions (for an exception see Cayuela et 53 al., 2006 ), where forests are disappearing at a rapid pace with variations in annual deforestation 54 figures ranging from 0.4 to 1.4% across the country ( Trejo and Dirzo, 2000; Velázquez et al., 55 2002; FAO, 2006 ). 56 The southern Yucatán (SY), a study region defined by several research projects, surrounding 57 the Calakmul Biosphere Reserve (CBR) and covering much of the southwestern Quintana Roo 58 and southeastern Campeche ( Turner et al., 2004 ), has contributed significantly to overall 59 deforestation in Mexico ( Achard et al., 1998, 2002). This region is notable and merits special

J.P. Ramírez-Delgado, et al. / submitted article in Forest Ecology and Management (2012)

60 attention since it is a development frontier in which Mexico placed its largest protected reserve of 61 semi-tropical forests, the CBR, which covers about one third of the SY ( Primack et al., 1998 ; 62 Vester et al., 2007 ). Acting as a pivotal part of the Mesoamerican Biological Corridor (MBC) 63 (Vester et al., 2007 ), the CBR aims to protect and facilitate the movement of biota throughout 64 Central America and preserve Mexican carbon stocks ( Vester et al., 2007 ; Turner et al, 2004 ). 65 Despite conservation and preservation efforts, the SY faces continuing threats due to ongoing 66 forest conversion for cultivation and cattle ranching ( Lawrence et al., 2004 ; Turner et al., 2004 ; 67 Vester et al., 2007 ; Busch and Geoghegan, 2010 ), which have diminished its extent and increased 68 forest fragmentation ( Lawrence et al., 2004). 69 In addition to human impacts on forests, the SY is a region with a long history of natural 70 disturbance ( Boose et al., 2003 ). The most recent category 5 hurricanes to strike the region were 71 Hurricane Dean in 2007, Hurricane Gilbert in 1988, and Hurricane Janet in 1955 ( Boose et al., 72 2003; Vandecar et al., 2011 ). A long legacy of hurricane and human disturbance in the SY has 73 created a mosaic of landscape of secondary and mature forest fragments of varying sizes and 74 ages, interspersed with pasture and agricultural plots ( Vandecar et al., 2011 ). 75 Various definitions and methods of assessing deforestation across the SY have generated 76 different estimates of forest losses (e.g. Bray et al., 2004 ; Turner et al., 2004 ; Rueda et al., 2010 ). 77 Bray and colleagues (2004) indicate an annual deforestation rate of 0.1% for the forests of the 78 Mayan zone, in the center of the state of Quintana Roo, over the period 1984 –2000, subsuming 79 late successional growth in the forest category. The Southern Yucatán Peninsula Region (SYPR) 80 project ( Turner et al., 2004 ), which has examined the interactions and consequences of tropical 81 deforestation and agricultural expansion around and within the CBR and MBC in southeastern 82 Quintana Roo and southwestern Campeche in Mexico since 1997, reports an annual rate of 83 deforestation of 0.29% for the years 1984 –1993, decreasing to 0.21% for 1987/1988 –2000 84 (Turner et al., 2004; Vester et al., 2007 ) and considering successional growth of 25 years and 85 older as forest. Another study on upland deforestation in the SY reports an annual deforestation 86 rate of 0.5% over the period 1984 –1993, localized on the communal land management units in 87 the western and eastern sides of the SY, and on those located in the southwestern border of the 88 CBR ( Rueda, 2010 ). This study also noted an upland forest recovery during the 1990s on the 89 communal land management units that had experienced large deforestation rates in the previous 90 decade.

J.P. Ramírez-Delgado, et al. / submitted article in Forest Ecology and Management (2012)

91 Despite different estimates of forest losses in the SY, this region was designated by a panel of 92 experts as a global hotspot of tropical deforestation between 1990 and 1997 (see Achard et al., 93 1998, 2002). While the deforestation of upland forest began to decrease over time, the 94 intensification of uses on extant opened land began to increase, largely by reducing the period of 95 fallow and therefore the age of secondary vegetation to be cut ( Schmook, 2010 ; Turner, 2010 ). 1 96 Forest fragmentation has been rising in the SY due to ongoing deforestation. Lawrence and 97 colleagues (2004) reported that forest fragmentation in this region rose by 107% from 1987 to 98 2000, defining forests as vegetation of 50 years or more, indicating that fragmentation due to 99 forest conversion into agricultural lands and pasture declined between 1995 and 2000, while 100 increasing in secondary or successional forest. The same study suggests that forest edge density 101 (the proportion of all pixels in the study area located at the forest edge) increased from 1.15% in 102 1987 to 3.14% in 2000, and mean forest compactness (square root of ratio of edge perimeter to 103 perimeter of circle encompassing the same area of forest) decreased from 0.0326 to 0.0151 for 104 1987 to 2000 respectively. 105 Although some ecological consequences of forest fragmentation have been studied in the SY 106 (e.g. Evelyn, 1996 ; Lawrence et al., 2004; Rivera and Calmé, 2006 ), spatial patterns of 107 deforestation and fragmentation at the landscape level have not been reported yet. 108 The main objective of this study is to analyze the spatial patterns of deforestation and 109 fragmentation at the landscape level of seasonal tropical forests in the SY, México, over the 110 1990–2006 period. In particular, we seek to: (1) provide estimates of deforestation rates over the 111 period from 1990 to 2006 within and around the CBR; and (2) assess changes in the spatial 112 configuration of seasonal tropical forests within and around the CBR through time, by using 113 selected fragmentation metrics at the landscape level. This study was conducted in collaboration 114 with the Environmental Disturbance Greater Yucatán (EDGY) project 115 (http://landchange.rutgers.edu/ ), which examines the role of hurricanes, a persistent disturbance 116 on the forests of the SY, on ecosystems and people, in hopes of generating information that may 117 be directly applied to management strategies of tropical seasonal forests across the MBC in 118 Mexico.

1 Crop fallow cycles vary, but commonly involve tree years of agricultural cultivation followed by nine or twelve years of forest fallow, after which the plot is cut and burned for subsequent reuse in agriculture ( Turner et al., 2004 ). 10

J.P. Ramírez-Delgado, et al. / submitted article in Forest Ecology and Management (2012)

119 3. Methods

120 3.1. Study area 121 The SY study area occupies approximately 18,247 km2 of southwestern Quintana Roo and 122 southeastern Campeche, north of the Mexican Guatemala border ( Fig. 1 ). It is dominated by a 123 tropical wet and dry climate (Aw in the Köppen climate classification) (Peel et al., 2007 ), with a 124 seasonal dry period extending from December to May (Lawrence and Foster, 2002 ) and a mean 125 annual temperature of 25 oC ( Whigham et al., 1990 ). A significant difference in total annual 126 precipitation is observed, ranging from 900 mm in the northwest to 1400 mm in the southeast 127 (Lawrence and Foster, 2002 ; Martínez and Galindo-Leal, 2002 ; Pérez-Salicrup, 2004 ). Its 128 topography is characterized by a karstic terrain composed of rolling limestone hills and ridges 129 that range from about 100 to 360 m amsl (above mean sea level), interspersed with large solution 130 sinks or bajos that retain water during the wet season ( Vester et al., 2007 ). 131 132 Fig. 1 133 134 These conditions support a number of forest types in the SY that are differentiated by stature, 135 deciduousness, and the relative abundance of species (wetland upland transition forest, seasonally 136 inundated wetland forest, medium stature semi evergreen upland forest, tall stature evergreen 137 upland forest, and low and medium stature semi deciduous upland forest) (Schmook et al., 138 2011),2 capturing an important ecocline between the xeric forest of the northern peninsula and the 139 humid forests of El Petén, Guatemala ( Vester et al., 2007 ). The different forest types, 140 characterized mainly by the abundance of different species, apparently in response to variations 141 in soil and moisture and past human disturbances ( Ibarra-Manríquez et al., 2002 ; Martínez and 142 Galindo-Leal, 2002 ; Pérez-Salicrup, 2004 ), play important roles in maintaining the biodiversity 143 along the ecocline. 144 After substantial deforestation during the apex of the Classic Maya civilization (ca. AD 800– 145 1000) (Klepeis and Turner, 2001 ; Turner et al., 2004 ), followed by an extended period (80 –260 146 years) of forest recovery (Mueller et al., 2010 ) and subsequent development and use ( Klepeis and 147 Turner, 2001 ; Turner et al., 2004 ), much of the SY was opened to agricultural occupation in the

2 These forest types are described in Schmook and colleagues (2011) and represent a synthesis of several regional classification schemas ( Miranda and Hernández, 1963 ; Pérez-Salicrup, 2004 ; Rzedowski, 2006 ). 11

J.P. Ramírez-Delgado, et al. / submitted article in Forest Ecology and Management (2012)

148 1960s (Klepeis and Turner, 2001 ; Turner et al., 2004 ), raising regional and international concerns 149 about threats to carbon stocks, biota, and forest habitats ( Achard et al., 1998 , 2002; Myers et al., 150 2000; Folan and García, 2001 ). This led to the establishment of the CBR in 1989 (Boege, 1995 ; 151 Primack et al., 1998 ; Folan and García, 2001 ; Vester et al., 2007 ), which is made up of a core and 152 buffer zone ( Roy Chowhury and Turner, 2006 ; Vester et al., 2007 ). While cultivation and 153 activities focused on sustainable forest extraction are permitted in the buffer zone, timber cutting 154 and most wildlife extraction are prohibited in the core zone ( Roy Chowhury and Turner, 2006 ). 155 Four types of land management units exist in the SY: forest ejidos , agricultural ejidos , private 156 ranches, and state controlled lands. These management units and the CBR are the focus of our 157 study (Fig. 1 ). Ejidos are the dominant land tenure units in the region, and access is usually 158 through usufruct rights given to member households ( Vance et al., 2004 ). Ranches and 159 agricultural ejidos dominate the eastern side of the SY and the major north south and east west 160 roads ( Turner et al., 2010 ). Most private ranches are in the southern side of the SY. State 161 controlled lands 3 are restricted to the central north side and central south side of the CBR, which 162 has large portion of its periphery composed of forest extraction and agricultural ejidos (Turner et 163 al., 2010 ). 164 The temporal variation in deforestation of the SY appears to be linked to a combination of 165 slash and burn agriculture (corn, bean, and squash crops) ( Vester et al., 2007 ; Schmook, 2010 ), 166 commercial chili production ( Keys, 2004 ), and pasture livestock activities ( Schmook and Vance, 167 2009; Busch and Geoghegan, 2010; Busch and Vance, 2011 ), challenging the conservation 168 efforts for maintenance of its mature forest ( Turner et al., 2004 ). 169 In the SY, the amount of land devoted to crop is intensifying and the amount of land devoted 170 to pasture is expanding. This dominant trend has given a “hollow frontier ” economic status to the 171 SY (see Busch and Geoghegan, 2010 ), in which increasingly intensive land use activities are 172 conducted by relatively fewer people, leading to an expansion of needs but lower population. 173 Together, reserve agents, local communities, government agencies and nongovernment 174 organizations, are working together to make land management decisions, obtaining mixed 175 success between the management units of the region ( Klepeis and Vance, 2003 ; Abizaid and 176 Coomes, 2004 ; Vester et al., 2007 ).

3 There are three stated controlled land units in the SY, one in the north and the other two in the south. All of them are inside the CBR but do not occupy the entire CBR area (see Fig. 1 ). 12

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177 3.2. Preparation of the spatial data 178 For this study we used a land cover classification from Vaca and colleagues (in press/2012) to 179 analyze forest loss and fragmentation in the SY. The data were derived from Landsat-5 Thematic 180 Mapper (TM) imagery from 1987 –1996 for 1990s, Landsat-7 Enhanced Thematic Mapper-Plus 181 (ETM+) imagery from 1999 –2001 for 2000, and ETM+ imagery from 2005 –2007 for 2006, at a 182 28.5 m pixel –1 spatial resolution. Images were acquired during the dry season, from December to 183 May. The data were classified together in a single multi-date image to produce a direct estimate 184 of change, which reduces false change errors caused by differences between images dates in 185 vegetation phenology, illumination conditions and atmospheric interference, rather than 186 classifying single-date images individually and then combining them to derive changes estimates 187 (Harper et al., 2007 ). A supervised classification method was used to classify each two-date 188 image pair. The classification consists of two classes: forest (areas of vegetation dominated by 189 tree cover at least five meters in height, with at least 10% of the canopy closed, including old 190 growth forest, secondary and degraded forest, as well as forest plantations) and nonforest 191 (pastures, crops, and infrastructure, water bodies, and cloud/shade regions of no data). Images 192 were classified using the Sequential Maximum A-Posteriori (SMAP) algorithm in the GRASS 193 GIS framework ( GRASS Development Team, 2005 ). For the classification, the methodology 194 described by Leimgruber and colleagues (2005), Christie and colleagues (2007) , and Harper and 195 colleagues (2007) was used, which is based on evaluating the quality of the map using field data 196 and visually checking the map with aerial photographs and high resolution images.

197 3.3. Deforestation and fragmentation analysis 198 Land cover maps were analyzed to quantify land cover change and configure them for the 199 application of fragmentation metrics by using the IDRISI Taiga GIS platform (Eastman, 2009) . 200 The forest category was isolated to perform deforestation and fragmentation analysis of the 201 forests in SY. Annual deforestation rates were calculated for each of the land management unit 202 using the relationship of Puyravaud (2003) : 203 ଵ଴଴ ஺మ 204 ܲwhere ൌ ௧మ ି௧A1భ and݈݊ ஺ భA2 are the forest cover at time t1 and t2 respectively, and P is the percentage of 205 forest loss per year.

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206 We decided to estimate deforestation rates among the land management units around and 207 within the CBR since the decisions on land use and land cover changes in the SY depend on the 208 intrinsic characteristics of these land management units as much as external pressures, which 209 affects the CBR conservation planning. 210 Quantification and temporal comparison of the spatial configuration of forest fragments were 211 conducted using selected fragmentation metrics at the landscape level to represent different 212 characteristics of forest configuration (after Armenteras et al., 2003 ; Fitzsimmons, 2003 ; 213 Millington et al., 2003 ; Echeverria et al., 2006 ; Teixido et al., 2010 ): (a) patch density (number of 214 patches per 100 ha), as a measure of forest loss and division; (b) total edge length (km), as a 215 measure of patch shape and the forest to nonforest interface; (c) total core area (total patch size in 216 km 2) for a distance to edge of 100 m (minimum distance deﬁned according to previous studies 217 that deal with edge effect) ( Laurance et al., 2002 ; Millington et al., 2003, Porej et al., 2004), as a 218 measure of high quality forest interior habitat; and (d) mean proximity index (ratio between the 219 size and proximity of all patches whose edges are within a 1 km search radius of the focal patch), 220 as a measure of isolation. All metrics were computed using Fragstats version 3.3 ( McGarigal et 221 al., 2002 ). 222 We used an analytical kernel of 7.6 km pixel –1 to enable the identification and generalization 223 of forest loss and recovery spots in the land management units of the SY during the 1990 –2000 224 and 2000–2006 periods, and to identify statistically significant differences at the 95% confidence 225 level across the entire study area and inside and outside the CBR. A systematic sampling design 226 and a Kruskal Wallis test were used for these purposes ( Dytham, 2003 ). Statistical analyses were 227 carried out using the SPSS software package version 15.0.

228 4. Results

229 4.1. Deforestation rates 230 From 1990 to 2006, the annual rate of deforestation was 17.65 km 2 year –1 (0.12% year –1) for 231 the whole study area, 1.90 km2 year –1 (0.03% year –1) for inside the CBR, and 15.75 km 2 year –1 232 (0.18% year –1) for outside the CBR. The changes in annual rate of deforestation between 1990 233 and 2006 were especially high in three locations: in the ejidos of the far west, in the ejidos of the 234 east, and among the ejidos along the southeastern border of the CBR ( Fig. 2 ). A total of 127 land 235 management units experienced a reduction in their forests, probably due to conversion of forests

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236 into lands for cultivation and pasture ( Lawrence et al., 2004 ; Turner et al., 2004 ; Vester et al., 237 2007; Busch and Geoghegan, 2010), while the remaining 14 exhibited no significant 238 deforestation. Most of the forest loss was concentrated in the first 10 years of the study period, at 239 a deforestation rate of 23.30 km 2 year –1 (0.15% year –1), while it decreased to 8.23 km 2 year –1 240 (0.06% year –1) over the 2000 –2006 period. 241 242 Fig. 2

243 4.2. Changes in forest cover: 1990 –2000 244 Between 1990 and 2006, forest cover decreased from 15,181 km 2 (83.2% of the study area) to 245 14,947 km2 (81.92%) in the entire study area, from 6465 km 2 (92.1% of the CBR) to 6437 km2 246 (91.71%) within the CBR, and from 8715 km 2 (77.63% of outside the CBR) to 8510 km2 247 (75.80%) around the CBR ( Table 1 ). Although overall deforestation was not so high between 248 1990 and 2000, there were, however, considerable forests losses in some specific sites across the 249 SY. We identified six deforestation hotspots during this time period: in land management units 250 located in the far west, far northeast, central northeast, far east, central east, and southeast ( Fig. 251 3a ). Forty six land management units were responsible for almost 90% of all the deforestation 252 during 1990 –2000 (Fig. 4a ), coinciding with those land management units located in the west, 253 east, central south and central north sides of the SY, and with the ejidos of the zona chilera 4, 254 along the southeastern border of the CBR ( Fig. 4b ). 255 256 Table 1 257 258 Fig. 3 259 260 Fig. 4 261 262 Although there was no significant difference in forests area between 1990 and 2000 across the 263 study area, neither inside not outside the CBR ( Table 2 ), the absolute changes in it indicated that 264 forests in the SY became increasingly reduced during this time period ( Table 1 ).

4 The zona chilera is an area spanning the southeastern border of the CBR where commercial cultivation for chili is enabled ( Keys, 2004 ). 15

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265 Table 2

266 4.3. Changes in forest cover: 2000 –2006 267 Over the period 2000 to 2006, there was a decrease in deforestation rates with respect to the 268 previous decade. Forest cover decreased slightly, from 14,947 km 2 (81.92% of the study area) to 269 14,898 km2 (81.65%) in the entire study area, from 6437 km 2 (91.71% of the CBR) to 6435 km2 270 (91.67%) within the CBR, and from 8510 km 2 (75.8% of outside the CBR) to 8463 km2 (75.38%) 271 around the CBR ( Table 1 ). Deforestation was, however, heterogeneous across the region ( Fig. 272 3b ). Only one deforestation hotspot on the far western side of the SY could be detected during 273 2000–2006 (Fig. 3b ). Twenty ejidos were responsible for almost 90% of all deforestation from 274 2000 to 2006 (Fig. 5a ), coinciding with the ejidos of the western, northwestern and eastern 275 reaches of the SY ( Fig. 5b ). 276 277 Fig. 5 278 279 Although there was no significant difference in forests area between 2000 and 2006 across the 280 study area, neither inside not outside the CBR ( Table 2 ), the absolute changes in it indicated that 281 forests in the SY became increasingly reduced during this time period ( Table 1 ).

282 4.4. Trends in forest fragmentation: 1990 –2000 283 As noted by Lawrence and colleagues (2004) , the decade of 1990s experienced an increase in 284 forest fragmentation throughout the SY. However, mature forest fragmentation caused by forest 285 conversion into agricultural lands and pasture declined since the mid 1990s, and instead increased 286 for secondary forests ( Lawrence et al., 2004 ). Fragmentation, as a result of deforestation 287 (Saunders et al., 1991 ; Brook et al., 2008 ), had significant sub-regional variations over the 1990 – 288 2000 period (see Fig. 4a ). During this decade, patch density presented a decrease of 5.77% in the 289 entire study area and of 6.84% outside the CBR, but a slight increase inside the CBR (0.28%) 290 (Fig. 6a ); total edge length presented a decrease of 2.52% in the entire study area and of 4.18% 291 outside the CBR, but an increase of 1.85% inside the CBR ( Fig. 6b ); total core area presented a 292 decrease of 1.55% in the entire study area, of 2.21% outside the CBR, and a slight fall inside the 293 CBR (0.76%) ( Fig. 6c ); and mean proximity index, or measure of isolation, presented a decrease

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294 of 4.97% in the entire study area, of 6.82% outside the CBR, and of 1.48% inside the CBR ( Fig. 295 6d ). 296 297 Fig. 6 298 299 Although there were no significant differences in fragmentation metrics between 1990 and 300 2000 across the study area, neither inside not outside the CBR ( Table 2 ), the absolute changes in 301 each one of them indicated that forests in the SY became increasingly fragmented during this 302 time period (Fig. 6 ).

303 4.5. Trends in forest fragmentation: 2000 –2006 304 Overall, the period between 2000 and 2006 demonstrated ongoing forest fragmentation 305 throughout the SY, but intensity was less compared to previous years. It appears that this 306 fragmentation was due to conversion of forests into cultivated and pasture lands ( Lawrence et al., 307 2004; Turner et al., 2004 ; Vester et al., 2007 ; Busch and Geoghegan, 2010 ), with sub-regional 308 variations in amount and rates of forest loss ( Fig. 5a ). From 2000 to 2006, patch density 309 presented an increase of 6.07% in the study area, of 7.07% outside the CBR, but a slight rise 310 inside the CBR (0.85%) ( Fig. 6a ); total edge length presented a slight increase in the study area 311 (0.7%), the same as around (0.86%) and within the CBR (0.31%) ( Fig. 6b ); total core area 312 presented a slight fall in the study area (0.48%), as well as around (0.81%) and within the reserve 313 (0.09%) ( Fig. 6c ); and mean proximity index presented a decline of 1.48% in the study area, an 314 increase of 1.45% outside the CBR, and remained practically unchanged within the CBR (0.01%) 315 (Fig. 6d ). 316 Although there were no significant differences in fragmentation metrics between 2000 and 317 2006 across the study area, neither inside not outside the CBR (Table 2 ), the absolute changes in 318 each one of them indicated that forests in the SY became increasingly fragmented during this 319 time period ( Fig. 6 ). 320 321 5. Discussion 322 Our results show a strong spatial and temporal variation in forest losses across the SY, 323 producing an increasingly fragmented landscape given changes in the spatial configuration 324 because of size, shape and degree of isolation of forest patches.

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325 5.1. Changes in forest cover 326 Population pressures and the accompanied demand of lands for cultivation and pasture, largely 327 in response to government programs, have been the main causes of deforestation in the SY 328 (Klepeis and Turner, 2001 ; Lawrence et al., 2004 ; Turner et al., 2004 ; Vester et al., 2007 ; Busch 329 and Geoghegan, 2010 ). A nearly 15-fold increase in population since the 1960s has caused an 330 increase in agricultural lands, thereby reducing forests areas ( Lawrence et al., 2004 ). Despite the 331 rapid pace of deforestation between 1984 and 1993, with variations in annual figures ranging 332 from 0.2% to 0.5% (see Turner et al., 2004 ; Vester et al., 2007 ; Rueda, 2010 ), forest loss began to 333 decrease from 1990 to 2006 (Table 1 ) but with significant sub-regional variations in amount and 334 rates of deforestation ( Fig. 4 and 5). 335 Although annual deforestation rates were low during the years from 1990 to 2000 (23.30 km2 336 year –1 or 0.15% year –1) and from 2000 to 2006 (8.23 km 2 year –1 or 0.06% year –1), these forest 337 losses were concentrated mostly in the western and eastern flanks of the SY, and along the 338 southeastern border of the CBR ( Fig. 2 ), sections of the SY that are located outside the 339 northwestern xeric zone, where annual rainfall increased on average, especially in the more 340 humid regions of the east ( Rueda, 2010 ). These three locations with higher deforestation rates, 341 ranging from 0.4 to 1.8% annually (Fig. 2 ), partially overlap with those identified by Rueda 342 (2010) during the 1984 –1993 period, suggesting a temporal continuity of deforestation. However, 343 that study also detected forest recovery scenarios during the 1993 –2000 period in ejidos that had 344 experienced high deforestation from 1984 to 1993, indicating a deceptive comparison in 345 deforestation due to the use of different criteria for vegetation classification (see Foody and Hill, 346 1996) and methods to assess deforestation (see Puyravaud, 2003 ). 347 High forest losses in the western and eastern edges of the SY might have emerged since the 348 very implementation of several large scale, government sponsored rice project in the late 1970s 349 and early 1980s ( Lawrence et al., 2004 ; Turner et al., 2004 ). Although these projects failed, they 350 left a large growing population and land clearing equipment ( Klepeis and Turner, 2001 ; Turner et 351 al., 2004 ). Most households in the ejidos located here had large amounts of land, up to 80 ha per 352 household (Rueda, 2010 ). For that, they have been able to preserve a significant portion of their 353 forest holdings. However, they usually abandon already cultivated lands due to bracken fern 354 invasion but instead cut down forest lands to grow their crops ( Geoghegan et al., 2004 ; Schneider 355 and Geoghegan, 2006 ; Rueda, 2010 ), diminishing more and more their forest lands over time.

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356 Large forest ejidos such as Tres Garantías and Zoh Laguna (eastern side of the region), that 357 had a high annual deforestation rate during the 1990s (see Fig. 4 ), were established on land 358 previously granted to timber concessions ( Klepeis, 2000 ), allowing smallholders from these 359 ejidos to take advantage of the acquired machinery and roads left by the timber companies 360 (Galetti, 1999 ). Most occupants in these two ejidos had come to work in timber extraction and 361 continue to do so under the new ejidal regime (see Bray et al., 2005 ). However, logging of older 362 growth forests in these two large ejidos continued until mid 1990s and with less intensity than in 363 the previous years ( Flachsenberg and Galleti, 1999 ). 364 High forest losses along the southeastern border of the CBR during the 1990s might have been 365 caused by the emergence of market jalapeño chili pepper production, which inspired the small 366 ejidos from this location to invest in jalapeño cultivation, and led smallholders from these ejidos 367 to change forest lands into chili plantations ( Keys 2004 ; Keys and Roy Chowdhury 2006 ). Given 368 the small size of the ejidos located here, forest losses in this side of the region were very 369 impactful in relative terms, especially between 1990 and 2000 ( Fig. 4 ). 370 Deforestation continued between 2000 and 2006 across the SY, but less than in the previous 371 time period ( Table 1 ). In spite of that, the deforestation was concentrated mostly on ejidos in the 372 west and east ( Fig. 5 ), suggesting a temporal continuity of high deforestation in these two 373 locations. Forest loss might have occurred mainly for cattle pasture establishment as this activity 374 has been spreading across the region since at least 1997 ( Busch and Geoghegan, 2010 ; Busch and 375 Vance, 2011 ). According to a land cover classification of the SY in 2000, pasture zones were 376 concentrated mostly on those land management units located in the western and eastern flanks of 377 the region (see Schmook et al., 2011 ), coinciding with those locations where deforestation was 378 high during the 2000 –2006 period ( Fig. 5 ). Smallholders apparently began to focus its cultivation 379 practices on successional forest growth throughout the region, with minimal effects on older 380 growth forest ( Lawrence et al., 2004; Schmook, 2010 ). This probably occurred around 1995 381 (Lawrence et al., 2004 ), suggesting that the regulations of the CBR against harvesting old growth 382 forests might have played some role (see Primack et al., 1998 ; Roy Chowdhury and Turner, 383 2006). However, upland forests continued to be cleared for cultivation and pasture in the 384 northwestern and eastern reaches of the SY until the late 1990s as suggested by Rueda (2010) . 385 A growing number of households, especially in the zona chilera , have started to use their 386 profits from chili cultivation to support international labor migration ( Radel and Schmook, 2008 ).

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387 They also appear to be investing in off farm activities with the aim of getting a bigger and more 388 secure income ( Radel et al., 2010 ). These may be the main factors that contributed to a slowing in 389 deforestation rates during the 1990 –2000 and 2000–2006 periods (see Fig. 3a, b ). Other possible 390 factors include the agricultural intensification of land use ( Busch and Geoghegan, 2010 ; Radel et 391 al., 2010 ; Busch and Vance, 2011 ) and the CBR rules about cutting forest lands (Primack et al., 392 1998; Roy Chowdhury and Turner, 2006 ).

393 5.2. Trends in forest fragmentation 394 Although deforestation rates decreased in the SY during the 1990 –2000 and 2000 –2006 395 periods, this phenomenon inevitably caused forest fragmentation across the region (Fig 3). No 396 significant changes were detected in the spatial patterns of the forest from 1990 to 2006 in the 397 study area, including inside and outside the CBR ( Table 2 ). However, the absolute changes in 398 such patterns indicate an increase in forest fragmentation ( Fig. 6 ). Forest fragmentation due to the 399 conversion of forests into lands for cultivation and pasture apparently declined since the mid 400 1990s for mature forests but increased for secondary forests, as suggested by Lawrence and 401 colleagues (2004) . The general trend was towards a decrease in the connectivity of forest 402 fragments as a result of increasing forest fragmentation through time, which dissected the forest 403 into smaller and more isolated fragments. Similar fragmentation trend has been reported on other 404 tropical forest studies (e.g. Imbernon and Branthomme, 2001 ; Millington et al., 2003 ; Cayuela et 405 al., 2006 ). 406 Ongoing deforestation commonly leads to a decline in patch density, as also observed in other 407 forest fragmentation studies (e.g. Zipperer et al., 1990 , in the central New York, USA; Trani and 408 Giles, 1999 , in George Washington and Jefferson National Forests, Virginia, USA; Echeverria et 409 al., 2006 , in Rio Maule Cobquecura, Chile). In the SY, this process was observed only in the 410 earliest stage of forest loss and fragmentation, with an increase in the numbers of fragments 411 during the later stages, suggesting a subdivision of forest patches that remained after the first 412 study period (between 1990 and 2000) ( Fig. 6a ). The changes in patch density were associated 413 with the total edge length values ( Fig. 6b ), meaning that the absolute decrease of patch edge 414 length during the first study period was due to the elimination of forest patches, rather than the 415 reduction in size and geometric modification of forest fragments. Echeverria and colleagues 416 (2006) also observed this trend in Rio Maule Cobquecura, Chile, together with a reduction in 417 patch density between 1990 and 2000. The absolute increase in the total edge length between

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418 2000 and 2006 involved a rise in the number of forest fragments, including the reduction in size 419 and the modification of their geometric shape, a key component of ecosystem fragmentation (see 420 Armenteras et al., 2003 ). 421 The steady decrease in the interior forest area as a result of forest loss and fragmentation 422 indicated that forest patches became smaller and more elongated trough time ( Fig. 6c ). This 423 tendency was also observed by Cayuela and colleagues (2006) in the Highlands of Chiapas, 424 Mexico, between 1975 and 2000, in which interior areas were also defined by a distance of 100 m 425 from the edge. The mean proximity index, or measure of isolation, experienced a steady decrease 426 due to a division of forest patches over time, meaning that forest patches became spatially 427 separated and less contiguous in distribution through time (Fig. 6d ). Such an increase in the patch 428 isolation may suggest a loss of forest connectivity (see Ochoa-Gaona et al., 2004 ), an effect of 429 deforestation that was also observed by Imbernon and Branthomme (2001) and Cayuela and 430 colleagues (2006) in Brazilian and Mexican landscapes, respectively. 431 Similar temporal variation was observed outside the CBR for patch density, total edge length, 432 and total core area ( Fig. 6a, b, c ). The mean proximity index ( Fig. 6d ) decreased in the earliest 433 stage of forest loss and fragmentation, meaning that forest patches became spatially separated and 434 less contiguous in distribution between 1990 and 2000. The subsequent increase of this metric 435 outside the CBR suggests a subdivision of forest patches that remained after the first study 436 period, leaving closer but smaller forest parches. 437 Within the CBR, an increase in patch density and in the total edge length was observed during 438 the 1990 –2000 and 2000–2006 periods (Fig. 6a, b ). These changes were by the reduction in size 439 and the modification of the geometrically complex shaped fragments rather than the elimination 440 of forest patches as in all study area and outside the CBR between 1990 and 2000. The total core 441 area and the mean proximity index constantly decreased inside the CBR as in all study area (Fig. 442 6c, d ). This indicates that forest patches became smaller and more elongated, and the 443 surroundings were areas of different land cover types. The forest thereby became spatially 444 separated and less contiguous through time in the CBR. 445 Reductions in patch size and core area, and an increase in edge length and spatial isolation of 446 patches may have negative ecological impacts on species occurring on forest habitats over time. 447 Edge effects, for instance, alter many aspects of the structure, microclimate, dynamics, and 448 species composition of fragmented ecosystems ( Laurance et al., 2002 ; Lehtinen et al., 2003 ; Ries

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449 et al., 2004 ; Wirth et al., 2007 ; Brook et al., 2008 ; Laurance et al., 2011 ). Edge effects impact 450 ecological process such as plant regeneration, seed dispersion or plant growth and survival 451 (Bruna, 1999 ; Echeverria et al., 2006 ; Hernández-Stefanoni, 2005 ; Hernández-Stefanoni and 452 Dupuy, 2008 ). Another consequence of forest fragmentation is the isolation of forest fragments, 453 which may have negative impacts if the population of species occurring in forest habitats have no 454 capacity to survive in isolated fragments or to move through the surrounding modified vegetation 455 matrix ( Reed and Levine, 2005 ; Echeverria et al., 2007 ). 5 456 As mentioned before, the SY have been subject to frequent hurricane disturbances 457 contributing to forest loss and fragmentation. Over the 16 year study period, at least five major 458 hurricanes made landfall in the region: Roxanne (1995), Mitch (1998), Keith (2000), Isidoro 459 (2002), and Emily (2005) ( Boose et al., 2003 ; Rogan et al., 2011 ). These hurricanes may blow 460 down significant areas of forest, especially in the eastern part of the region ( Boose et al., 2003 ). 461 Hurricane Dean, a category 5 hurricane, made landfall on the SY on 21 August 2007, altered 462 forest structure and composition in the region, favoring certain wind resistant species ( Vandecar 463 et al., 2011 ). Because our data do not allow us to investigate the impacts of this category 5 464 hurricane on forest loss and fragmentation in the region, we recommend that future research 465 addresses this issue. 466 Although an investigation on the effect of forest fragmentation on biodiversity across the SY 467 was outside the scope of this study, the description of spatial landscape patterns provide a basis 468 for future research to analyze the effects of forest fragmentation on biodiversity.

469 6. Conclusions and management implications 470 Our results indicate slowing deforestation in the SY, inside and outside the CBR, but with 471 significant sub-regional variations in the amount and rates of forest loss. Land management units 472 in the SY, especially those located in the western and eastern sides of the region and along the 473 southeastern border of the CBR, were transforming forests into agricultural lands (crops and 474 pasture), particularly during 1990 –2000. The main factors that contributed to a slowing of 475 deforestation in the SY during the 1990 –2000 and 2000 –2006 periods that we could identify were 476 the agricultural intensification of land use (Busch and Geoghegan, 2010 ; Radel et al., 2010 ;

5 In the Amazon, forest fragments surrounded by cattle pastures have suffered considerably greater species losses than those surrounded by secondary forest ( Stouffer and Bierregaard, 1995 ; Gilgert and Setz, 2001 ), which might be happening in the SY landscape. 22

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477 Busch and Vance, 2011 ), the increasing diversification of household income generating activities 478 (Radel and Schmook, 2008 ; Radel et al., 2010 ), and the CBR rules against cutting old growth 479 forests (Primack et al., 1998 ; Roy Chowdhury and Turner, 2006 ). Notwithstanding that the entire 480 SY experienced low rates of deforestation during 1990 –2000 and 2000–2006, even these low 481 deforestation rates caused forest fragmentation. The general fragmentation trend in the SY, 482 including the reserve, led to decreasing connectivity of forest fragments areas resulting from 483 increasing forest fragmentation through time, dissecting the forest into smaller and more isolated 484 fragments. 485 Given the necessary conditions of sustainable forest management in the SY, that timber 486 extraction represents an important economic activity in this region ( Primack et al., 1998 ; Roy 487 Chowdhury and Turner, 2006 ), and that a lot of timber harvesting has been carried out without 488 formal management plans ( Porter Bolland et al., 2006 ), permanent forest areas, over which a 489 sustainable community management logging plan operates, would favor forest conservation in 490 the SY, especially on those land management units that do not have formal constraints on forest 491 land uses, which has occurred in some other similar areas near the region (see Bray et al., 2004 ; 492 Dúran-Medina et al., 2007 ; Bray et al., 2008 ).

493 Acknowledgements 494 This research was supported by the Gordon and Betty Moore Foundation under Grant 1697. 495 We sincerely thank to Tijl Essens, who provided insights on the statistical methods, to Raúl Vaca, 496 who provided satellite imagery, and to CONACYT, for the fellowship awarded to Juan Pablo 497 Ramírez-Delgado. 498 499 References 500 Abizaid, C., Coomes, O.T., 2004. Land use and forest fallowing dynamics in seasonally dry 501 tropical forest of the southern Yucatan peninsula. Land Use Policy 21 (1): 71 –84. 502 Achard, F., Eva, H., Glinni, A., Mayaux, P., Richards, T., Stibig, H.J., 1998. Identification of 503 deforestation hot spot areas in the humid tropics. European Commission, Luxemburg. 504 Achard, F., Eva, H.D., Stibig, H.-J., Mayaux, P., Gallego, J., Richards, T., Malingreau, J.P., 505 20 02. Determination of deforestation rates of the world’s humid tropical forests. Science 506 297, 999–1002.

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Highlights · We have assessed annual deforestation rates and spatial patterns of forest fragmentation. · Study focuses on the area within and surrounding Mexico’s largest tropical reserve . · Results indicate a slowing in deforestation with significant sub-regional variations. · Despite decreases in deforestation, the landscape became increasingly fragmented. · Results affirm the need for careful management of forested resources.

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Figure captions Fig. 1. Map of study area showing the Calakmul Biosphere Reserve, principal roads and land management units included in the study. Fig. 2. Annual deforestation rate by land management unit in the southern Yucatán between 1990 and 2006. Fig. 3. Forest loss and recovery spots in the land management units of the southern Yucatán between (a) 1990 –2000, and (b) 2000 –2006. Fig. 4. Annual deforestation rates (a) and locations (b) of forty six land management units responsible for 90% of forest loss in the southern Yucatán between 1990 and 2000. P: annual deforestation rate (% year –1), R: annual change (km 2 year –1). Fig.5. Annual deforestation rates (a) and locations (b) of twenty ejidos responsible for 90% of forest loss in the southern Yucatán between 2000 and 2006. P: annual deforestation rate (% year – 1), R: annual change (km 2 year –1). Fig. 6. Changes in fragmentation patterns applied to forest cover in the southern Yucatán and within and around the Calakmul Biosphere Reserve for the periods 1990 –2000 and 2000 –2006. Values of fragmentation metrics were derived from maps of 1990, 2000, and 2006. Fragmentation metrics: (a) patch density (number of patches per 100 ha), (b) total edge length (km), (c) total core area (for a distance to edge of 100 m), and (d) mean proximity index (for a distance of 1 km).

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Tables % 0.04 0.04 0.55 0.33 2006 2006

2 2.47 2.47 2000- km 46.89 46.89 49.37 Net change Net change % 0.43 0.43 2.35 1.53 2000 2000 the Calakmul the Calakmul

1990- km 27.98 27.98 205.04 205.04 233.02

% 8.33 8.33 24.62 24.62 18.35

2 Nonforest Nonforest km 584.66 584.66 2763.81 2763.81 3348.47 2006 2006 % 91.67 91.67 75.38 81.65

Forest Forest 2

km 6434.74 6434.74 8463.39 14898.12 14898.12

% 8.29 8.29 24.20 24.20 18.08

2 Nonforest Nonforest km 582.19 582.19 2716.92 2716.92 3299.11 2006 – 2000 2000 % 91.71 91.71 75.80 81.92

Forest Forest 2

km 6437.21 6437.21 8510.28 14947.49 14947.49 2000 and 2000 –

% 7.90 7.90 22.37 22.37 16.80

2 Nonforest Nonforest km 554.21 554.21 2511.87 2511.87 3066.08 1990 1990

% 92.10 92.10 77.63 83.20 Forest Forest 2 2

km 6465.19 6465.19 8715.32 15180.51 15180.51

Table 1 southern in the in forest around change within cover of and and and net and nonforest Yucatán forest Area the 1990 for periods Reserve Biosphere CBR Inside CBR Outside area All study

J.P. Ramírez-Delgado, et al. / submitted article in Forest Ecology and Management (2012)

Table 2 Effects of time on the variables used to characterize forest loss and fragmentation patterns in the southern Yucatán and within and around the Calakmul Biosphere Reserve. P-value < 0.05 means significant effect of time. N = 65 for outside the CBR in 1990 and 2000, N = 61 for outside the CBR in 2006, N = 31 for inside the CBR in 1990, N = 30 for inside the CBR in 2000, N = 32 for inside the CBR for 2006, N = 96 for all study area in 1990, N = 95 for all study area in 2000, and

N = 93 for all study area in 2006

1990 –2000 2000 –2006 Inside CBR Outside CBR All study area Inside CBR Outside CBR All study area Variable X2 P-value X2 P-value X2 P-value X2 P-value X2 P-value X2 P-value Forest area 0.19 0.66 0.48 0.49 0.72 0.40 0.03 0.87 0.10 0.75 0.02 0.97 Patch density 0.04 0.84 0.85 0.36 0.69 0.41 0.82 0.36 0.03 0.86 0.17 0.68 Total edge length 0.02 0.90 2.79 0.10 1.87 0.17 0.33 0.56 0.28 0.60 0.01 0.91 Total core area 0.02 0.88 0.80 0.37 0.88 0.35 0.17 0.68 0.08 0.78 0.01 0.93 Mean proximity index 2 x 10 -4 0.99 0.87 0.35 0.39 0.53 0.31 0.58 0.02 0.89 0.42 0.52

Conclusiones

Los resultados obtenidos muestran una reducción de la deforestación en el sur de la

Península de Yucatán con variaciones sub-regionales en cantidad y pérdidas de selvas. Las unidades de manejo localizadas en el occidente y oriente de la región, y aquellas ubicadas a lo largo del borde sur oriental de la Reserva de la Biosfera Calakmul, fueron las que presentaron mayores pérdidas de áreas forestales, entre 0.4 y 1.8 % al año, especialmente durante el periodo 1990 –2000.

De acuerdo a estudios previos, los principales factores que se pudieron identificar, y los cuales contribuyeron a la reducción de la deforestación en el sur de la Península de

Yucatán durante los periodos 1990 –2000 y 2000 –2006, fueron el aumento en el rendimiento de los cultivos por unidad de área (Busch y Geoghegan 2010, Radel et al .

2010, Busch y Vance 2011), la creciente diversificación de actividades generadoras de ingresos fuera del campo (ver Radel y Schmook 2008; Radel et al . 2010) y las normas de manejo de tierras forestales impuestas por la Reserva de la Biosfera Calakmul (ver Primack et al . 1998, Roy Chowdhury y Turner 2006, Rueda 2007).

Aunque las tasas de deforestación en el sur de la Península de Yucatán fueron bajas durante estos dos periodos, estas causaron inevitablemente la fragmentación de sus selvas. La tendencia general de la fragmentación en esta región, incluyendo la reserva, fue hacia una disminución en la conectividad de los fragmentos de selva como resultado de la división y disminución en el tamaño de áreas forestales a través del tiempo.

Teniendo en cuenta la necesidad de un manejo sustentable de las selvas en el sur de la

Península de Yucatán, que la extracción de madera representa una actividad económicamente importante en esta región (Primack et al . 1998, Roy Chowdhury y Turner

2006) y que la mayor parte de la extracción de madera ha sido llevada a cabo sin planes de

manejo formales dentro de la misma (Porter Bolland et al . 2006), las áreas permanentes de selvas, sobre las cuales opere un plan sustentable de manejo comunitario para la extracción de madera, podrían favorecer la conservación de las selvas de esta región, especialmente en aquellas unidades de manejo que no tienen restricciones formales sobre el uso de tierras forestales, tal y como ha ocurrido en otros lugares cerca del sur de la

Península de Yucatán (ver Bray et al . 2004, Durán-Medina et al . 2007, Bray et al . 2008).

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